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WATER ENCYCLOPEDIA

DOMESTIC, MUNICIPAL, AND INDUSTRIAL WATER SUPPLY AND WASTE DISPOSAL Jay Lehr, Ph.D. Editor-in-Chief Jack Keeley Senior Editor Janet Lehr Associate Editor Thomas B. Kingery III Information Technology Director

The Water Encyclopedia is available online at http://www.mrw.interscience.wiley.com/eow/

A John Wiley & Sons, Inc., Publication

Copyright  2005 by John Wiley & Sons, Inc. All rights reserved. Published by John Wiley & Sons, Inc., Hoboken, New Jersey. Published simultaneously in Canada. No part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, scanning, or otherwise, except as permitted under Section 107 or 108 of the 1976 United States Copyright Act, without either the prior written permission of the Publisher, or authorization through payment of the appropriate per-copy fee to the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400, fax 978-646-8600, or on the web at www.copyright.com. Requests to the Publisher for permission should be addressed to the Permissions Department, John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, (201) 748-6011, fax (201) 748-6008. Limit of Liability/Disclaimer of Warranty: While the publisher and author have used their best efforts in preparing this book, they make no representations or warranties with respect to the accuracy or completeness of the contents of this book and specifically disclaim any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives or written sales materials. The advice and strategies contained herein may not be suitable for your situation. You should consult with a professional where appropriate. Neither the publisher nor author shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. For general information on our other products and services please contact our Customer Care Department within the U.S. at 877-762-2974, outside the U.S. at 317-572-3993 or fax 317-572-4002. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print, however, may not be available in electronic format. Library of Congress Cataloging-in-Publication Data is available. Lehr, Jay Water Encyclopedia: Domestic, Municipal, and Industrial Water Supply and Waste Disposal ISBN 0-471-73687-2 ISBN 0-471-44164-3 (Set) Printed in the United States of America 10 9 8 7 6 5 4 3 2 1

WATER ENCYCLOPEDIA

WATER QUALITY AND RESOURCE DEVELOPMENT Jay Lehr, Ph.D. Editor-in-Chief Jack Keeley Senior Editor Janet Lehr Associate Editor Thomas B. Kingery III Information Technology Director

The Water Encyclopedia is available online at http://www.mrw.interscience.wiley.com/eow/

A John Wiley & Sons, Inc., Publication

Copyright  2005 by John Wiley & Sons, Inc. All rights reserved. Published by John Wiley & Sons, Inc., Hoboken, New Jersey. Published simultaneously in Canada. No part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, scanning, or otherwise, except as permitted under Section 107 or 108 of the 1976 United States Copyright Act, without either the prior written permission of the Publisher, or authorization through payment of the appropriate per-copy fee to the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400, fax 978-646-8600, or on the web at www.copyright.com. Requests to the Publisher for permission should be addressed to the Permissions Department, John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, (201) 748-6011, fax (201) 748-6008. Limit of Liability/Disclaimer of Warranty: While the publisher and author have used their best efforts in preparing this book, they make no representations or warranties with respect to the accuracy or completeness of the contents of this book and specifically disclaim any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives or written sales materials. The advice and strategies contained herein may not be suitable for your situation. You should consult with a professional where appropriate. Neither the publisher nor author shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. For general information on our other products and services please contact our Customer Care Department within the U.S. at 877-762-2974, outside the U.S. at 317-572-3993 or fax 317-572-4002. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print, however, may not be available in electronic format. Library of Congress Cataloging-in-Publication Data is available. Lehr, Jay Water Encyclopedia: Water Quality and Resource Development ISBN 0-471-73686-4 ISBN 0-471-44164-3 (Set) Printed in the United States of America 10 9 8 7 6 5 4 3 2 1

WATER ENCYCLOPEDIA

SURFACE AND AGRICULTURAL WATER Jay Lehr, Ph.D. Editor-in-Chief Jack Keeley Senior Editor Janet Lehr Associate Editor Thomas B. Kingery III Information Technology Director

The Water Encyclopedia is available online at http://www.mrw.interscience.wiley.com/eow/

A John Wiley & Sons, Inc., Publication

Copyright  2005 by John Wiley & Sons, Inc. All rights reserved. Published by John Wiley & Sons, Inc., Hoboken, New Jersey. Published simultaneously in Canada. No part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, scanning, or otherwise, except as permitted under Section 107 or 108 of the 1976 United States Copyright Act, without either the prior written permission of the Publisher, or authorization through payment of the appropriate per-copy fee to the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400, fax 978-646-8600, or on the web at www.copyright.com. Requests to the Publisher for permission should be addressed to the Permissions Department, John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, (201) 748-6011, fax (201) 748-6008. Limit of Liability/Disclaimer of Warranty: While the publisher and author have used their best efforts in preparing this book, they make no representations or warranties with respect to the accuracy or completeness of the contents of this book and specifically disclaim any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives or written sales materials. The advice and strategies contained herein may not be suitable for your situation. You should consult with a professional where appropriate. Neither the publisher nor author shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. For general information on our other products and services please contact our Customer Care Department within the U.S. at 877-762-2974, outside the U.S. at 317-572-3993 or fax 317-572-4002. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print, however, may not be available in electronic format. Library of Congress Cataloging-in-Publication Data is available. Lehr, Jay Water Encyclopedia: Surface and Agricultural Water ISBN 0-471-73685-6 ISBN 0-471-44164-3 (Set) Printed in the United States of America 10 9 8 7 6 5 4 3 2 1

WATER ENCYCLOPEDIA

OCEANOGRAPHY; METEOROLOGY; PHYSICS AND CHEMISTRY; WATER LAW; AND WATER HISTORY, ART, AND CULTURE Jay Lehr, Ph.D. Editor-in-Chief Jack Keeley Senior Editor Janet Lehr Associate Editor Thomas B. Kingery III Information Technology Director

The Water Encyclopedia is available online at http://www.mrw.interscience.wiley.com/eow/

A John Wiley & Sons, Inc., Publication

Copyright  2005 by John Wiley & Sons, Inc. All rights reserved. Published by John Wiley & Sons, Inc., Hoboken, New Jersey. Published simultaneously in Canada. No part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, scanning, or otherwise, except as permitted under Section 107 or 108 of the 1976 United States Copyright Act, without either the prior written permission of the Publisher, or authorization through payment of the appropriate per-copy fee to the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400, fax 978-646-8600, or on the web at www.copyright.com. Requests to the Publisher for permission should be addressed to the Permissions Department, John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, (201) 748-6011, fax (201) 748-6008. Limit of Liability/Disclaimer of Warranty: While the publisher and author have used their best efforts in preparing this book, they make no representations or warranties with respect to the accuracy or completeness of the contents of this book and specifically disclaim any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives or written sales materials. The advice and strategies contained herein may not be suitable for your situation. You should consult with a professional where appropriate. Neither the publisher nor author shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. For general information on our other products and services please contact our Customer Care Department within the U.S. at 877-762-2974, outside the U.S. at 317-572-3993 or fax 317-572-4002. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print, however, may not be available in electronic format. Library of Congress Cataloging-in-Publication Data is available. Lehr, Jay Water Encyclopedia: Oceanography; Meteorology; Physics and Chemistry; Water Law; and Water History, Art, and Culture ISBN 0-471-73684-8 ISBN 0-471-44164-3 (Set) Printed in the United States of America 10 9 8 7 6 5 4 3 2 1

WATER ENCYCLOPEDIA

GROUND WATER Jay Lehr, Ph.D. Editor-in-Chief Jack Keeley Senior Editor Janet Lehr Associate Editor Thomas B. Kingery III Information Technology Director

The Water Encyclopedia is available online at http://www.mrw.interscience.wiley.com/eow/

A John Wiley & Sons, Inc., Publication

Copyright  2005 by John Wiley & Sons, Inc. All rights reserved. Published by John Wiley & Sons, Inc., Hoboken, New Jersey. Published simultaneously in Canada. No part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, scanning, or otherwise, except as permitted under Section 107 or 108 of the 1976 United States Copyright Act, without either the prior written permission of the Publisher, or authorization through payment of the appropriate per-copy fee to the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400, fax 978-646-8600, or on the web at www.copyright.com. Requests to the Publisher for permission should be addressed to the Permissions Department, John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, (201) 748-6011, fax (201) 748-6008. Limit of Liability/Disclaimer of Warranty: While the publisher and author have used their best efforts in preparing this book, they make no representations or warranties with respect to the accuracy or completeness of the contents of this book and specifically disclaim any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives or written sales materials. The advice and strategies contained herein may not be suitable for your situation. You should consult with a professional where appropriate. Neither the publisher nor author shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. For general information on our other products and services please contact our Customer Care Department within the U.S. at 877-762-2974, outside the U.S. at 317-572-3993 or fax 317-572-4002. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print, however, may not be available in electronic format. Library of Congress Cataloging-in-Publication Data is available. Lehr, Jay Water Encyclopedia: Ground Water ISBN 0-471-73683-X ISBN 0-471-44164-3 (Set) Printed in the United States of America 10 9 8 7 6 5 4 3 2 1

PREFACE No one really questions that water is the life blood of humankind. We all remain amazed that the existence of water separates our planet from every other we have thus far viewed in our universe. We can arguably do without every naturally occurring molecule on the earth except water. Life was clearly formed within water and exists in one way or another on water. Few people in the developed world give this simple fact of life much thought. We have an abundance of water for most of our needs, although some agricultural areas, a few municipalities, and some rural families, at times are strapped for the full amount of water they desire. In the developing world, however, in some locations, the collection and distribution of water is a critical part of every day life, with many women devoting the major portion of their day to the provision of water for their family. In other villages, the construction and protection of a single well can be the primary focus of community needs. In general, the actual delivery of water for a myriad of uses followed by its disposal is taken for granted by all but the individuals charged with carrying out these often amazing tasks. In this volume of the Water Encyclopedia, a collaborative effort of hundreds of people from dozens of countries, we have tried to cover every conceivable topic of interest to people in every walk of life, be they students, researchers, professionals, or just plain folks with an intellectual curiosity about our elixir of life. We are concerned in this volume with the actual delivery of water to the home by the home owner (subjects include disinfection, corrosion control, nitrates, gray water, septic tanks, and windmills), from the municipal supplier (and their challenges, including distribution, filtration, zebra mussels, reverse osmosis,

cryptosporidium, arsenic, and public confidence), to industry and its special needs (such as microfiltration, effluent discharge, reuse, energy, nuclear reactor coolants, and even golf course irrigation), and of course the disposal of our used water in a safe and efficient manner (subjects such as air stripping, bioassays, flotation, sludge, bioavailability, and wetlands). We hope that no reader can stump the experts, which means that we have covered every area of interest. However, we know that this goal is not currently possible, but in coming years and in coming editions on paper and on the World Wide Web, we will more closely approach it. Let us know on our website where our information may be incomplete, and we will be sure to follow-up and fill in the gap in the future. The contributors to this volume have freely offered their expertise to this project. Some have focused their information on those in need of complete and often complex detail of their subject matter. Others have followed a middle road for a wider audience, and still others believed that a very simple approach to conveying information on their subject was best. The reader may find all approaches on the same subject matter because the editors frequently sought overlapping information presented from different points of view. We are confident that most people will find their needs met. Through this encyclopedia, which is the most comprehensive effort ever undertaken on behalf of this most important subject, we hope that we will collectively make a contribution that will enhance the distribution and use of our water supplies in ever safer and more efficient ways. Jay Lehr Jack Keeley

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PREFACE Cities, towns, states, and nations must manage their water resources wisely from both a quality and a quantitative perspective. If we do otherwise and manage them with a narrow perspective, the public’s needs will not be adequately met. In this volume of the Water Encyclopedia, authors from around the world have described a myriad of problems relating to individual water bodies as well as to geographic water resources and their management dilemmas. Humans and other living creatures contribute to our water quality problems. Neither can be fully controlled. Even the nature of contaminant sources and programs for their elimination can be difficult to design. This volume contains the best and brightest ideas and case studies relating to the areas of water quality and resource management problems. Quality problems deal with a diverse suite of subjects ranging widely from acid mine drainage to biosorption, colloids, eutrophication, protozoa, and recalcitrant compounds. Resource management features drought studies, flood control, river basin management, perennial overdraft, water banking, and a host of other subjects. The perspective of scientists from nearly every continent of the world offers a truly catholic view of

attitudes and biases harbored in different regions and how they affect scientific and regulatory outcomes. The editors cannot imagine what has been left out, but we know of course that readers will at times come up short of finding an exact match to a problem they face. We hope they will contact us at our website and allow us the opportunity of adding additional subjects to our encyclopedia. At the same time, the reader will understand that many subjects in the area of water quality may have been addressed in our Surface Water category. It was often difficult to determine where an investigator would be more likely to look for a piece of information. (The complete index of all five volumes appears in the Ground Water volume as well as on our website.) We trust all users of this encyclopedia will find it detailed, informative, and interesting. Not only are a wide range of subjects treated, but authors choose varying approaches to presenting their data to readers who may be professionals, students, researchers, as well as individuals simply satisfying their intellectual curiosity. We hope we are successfully serving all of these populations in some useful way. Jay Lehr Jack Keeley

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PREFACE Surface water and agricultural water are uniquely associated as they provide many of our basic needs, including food and fiber, power, transportation, and recreation. Like other volumes in the Water Encyclopedia, we have selected articles as varied in content as they are in technical sophistication. To this end, the reader will also recognize that single topics are occasionally duplicated at varying levels of scientific acumen. Articles are also provided that demonstrate that surface and agricultural water are associated in yet another way: They must be used efficiently and protected to assure their productiveness far into the future. For example, agricultural water use efficiency is discussed from several viewpoints with respect to irrigation technology. River basin planning is approached in diverse ways, including stream classifications, watershed hydrology, modeling, erosion control, and water conservation. We have necessarily included articles addressing issues of quality with respect to both surface and agricultural water. In addition to an assessment of pollution outflow from agricultural areas, the quality of reclaimed irrigation is addressed from both chemical and microbial standpoints. Watershed areas are examined according to their contribution and vulnerability to contamination, flooding, sediment transport, and trace elements.

Discourses on surface water would not be complete without articles related to fish. Accordingly, we have included articles on fish growth, fisheries, fishponds, and the use of fish scales in toxicological studies as examples. Another vital area of study in this volume is perhaps best described as the practical side. These areas are of less esoteric origins, including salt tolerance of plants, irrigation wells, weed control, tile drains, and moisture content in to agriculture. Similar topics in surface water include riparian systems, reservoir design, wetlands, lakes, levees, and the unit hydrograph. Finally, and appropriately, this volume of the Water Encyclopedia contains articles on specific water bodies and the consequences of their being. Included are the Aral Sea, the Ganga River of India, the Great Lakes, and the Yellow River in China, only to name a few. Here, too, the association of surface water and agricultural water are reinforced. This volume presents an important segment of the topic of water. We believe that the reader’s educational pursuits will be well met by its contents. Jay Lehr Jack Keeley

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PREFACE No natural molecule on the planet is more fascinating than water. It has unique properties ranging from the unusual angles formed between its two hydrogen ions and its single oxygen molecule, to the fact that unlike most substances, it expands when it freezes rather than shrinks and reaches a maximum density as a liquid 4◦ F above its freezing point. These and many other aspects of the special physics and chemistry are described in this volume, including the impact of a wide variety of chemicals occurring in water, osmosis, diffusion, hydration, isotope exchange, along with the fun physics of the mariotte bottle. Equally fascinating are the many unusual physical and chemical encounters in both the ocean and the atmosphere. Although oceanography and meteorology are frequently considered separate sciences from hydrology, their limited inclusion in the Water Encyclopedia was deemed necessary to tell the complete story. Tidal changes, benthic nutrients, the sea floor, el nino, sea level, and ocean/climate relationships make up but a few of the oceans fascinating stories, whereas water spouts, hurricanes, monsoons, droughts, sublimation, and barometric efficiency just touch the tip of what this volume has in store in the area of meteorology.

But water has a wonderful human face as well. We have reached around the world to describe the history of water and its role in the development of civilizations and the many beliefs held about it. As society developed, the distribution of water needed supervision, which lead to a wide variety of water laws we have attempted to categorize and describe in an interesting and meaningful way. We are equally proud of our open-minded effort to describe the role that water has played in art and culture. We have attempted not be judgmental, with stories of water forms and water intelligence along with some medical theories and, of course, the wonderful descriptions of early water clocks. This volume is a true intellectual cornucopia of water in the life of humankind on a personal level. We are confident that in the coming years and editions of the Water Encyclopedia, this volume will expand with more participation from individuals working in unusual fields relating to water. Jay Lehr Jack Keeley

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PREFACE from many American scholars, equally superb writings from such diverse countries as England, Nigeria, India, Iran, Thailand, and Greece are provided. As the origins of the selected articles are diverse, so are the subjects of discussion. Along with straightforward descriptions of basic groundwater concepts (drawdown around pumping wells, hydraulic head, field capacity, and flow), the reader is introduced to more complex subjects (isotope technologies, aquifer tests, in situ remediation, tritium dating, modeling, and geophysical properties). There are also articles for more practical applications (well maintenance, subsurface drainage, nitrate contamination, tracer tests, well yields, and drilling technologies). Of course, for the more fanciful reader, we have selected articles that remind us of the way windmills sounded at night, the ancient use of qanats in Persia to provide sustainable groundwater resources, and the development of Darcy’s Law. In the end, we feel that the information provided will afford an educational home for readers approaching the Water Encyclopedia from a variety of needs as well as different levels of scientific acumen. We are also confident that many readers will simply be expanding their knowledge base by these sets of enjoyable reading.

Throughout history, groundwater has played a major role in providing the resource needs of the world. It accounts for 97% of the world’s freshwater and serves as the base flow for all streams, springs, and rivers. In the United States, one half of the population relies on groundwater for its drinking water and is the sole source of supply for 20 of the 100 largest cities. Well over 90% of rural America is totally dependent on groundwater. An inventory of the total groundwater resources in the United States can be visualized as being equal to the flow of the Mississippi River at Vicksburg for a period of 250 years. One of the first groundwater scientists was a French engineer who was in charge of public drinking water in Dijon. In 1856, Henri Darcy conducted experiments and published mathematical expressions describing the flow of water through sand filters. His work remains one of the cornerstones of today’s groundwater hydrologists. At about the same time, a Connecticut court ruled that the influences of groundwater movement are so secret, changeable, and uncontrollable that they could not be subject to regulations of law, nor to a system of rules, as had been done with surface streams. In this volume of the Water Encyclopedia, we have attempted to erase the ignorance that existed in the early years of groundwater science by presenting the most current knowledge on the subject as provided by authors from around the globe. In addition to excellent articles

Jay Lehr Jack Keeley

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ACKNOWLEGMENTS The editors of this encyclopedia wish to acknowledge two special people without whose assistance this enormous undertaking could not have been initiated, administered, or completed. Bob Esposito, Executive Editor, placed the project before us and convinced us it would be as exciting and rewarding as time has proved. Jonathan T. Rose, Editorial Program Coordinator, was the true backbone of our team. His warm, accommodating, and skillful management helped us overcome each and

every problem. His interaction with contributors and his knowledge of the publishing process offered us the security and comfort required to persevere through the four years required to complete this work. We will forever be in his debt. Jay Lehr Jack Keeley Janet Lehr Thomas B. Kingery III

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CONTRIBUTORS Robert Bruce, National Center for Environmental Assessment, Cincinnati, Ohio, Health Effects of Microbial Contaminants and Biotoxins in Drinking Water, Health Effects of Commonly Occurring Disinfection Byproducts in Municipal Water Supplies Charlie Bryce, Napier University, Edinburgh, Scotland, United Kingdom, The Arsenic Drinking Water Crisis in Bangladesh Mario O. Buenfil-Rodriguez, National University of Mexico & Mexican Institute of Water Technology, Cuernavaca, Morelos, Mexico, Public Water Supply World, Water Distribution System Operation, Water Meter Zia Bukhari, American Water, Quality Control and Research Laboratory, Belleville, Illinois, Potential Risks of Waterborne Transmission of Escherichia coli O157:H7, Measuring Cryptosporidium Parvum Oocyst Inactivation Following Disinfection With Ultraviolet Light Michael A. Butkus, United States Military Academy, West Point, New York, Biochemical Oxygen Demand D. Butler, Imperial College, London, United Kingdom, Gray Water Reuse in Households Rebecca L. Calderon, Ph.D., National Health and Environmental Effects Laboratory Research Triangle Park, North Carolina, Improving Waterborne Disease Surveillance Christine M. Carey, University of Guelph, Guelph, Ontario, Canada, Molecular-Based Detection of Cryptosporidium Parvum in Water Peter S. Cartwright, P.E., Minneapolis, Minnesota, Water Reuse Frank J. Castaldi, Brown and Caldwell, Austin, Texas, Degradation of Chloro-Organics and Hydrocarbons, Aqueous Behavior of Elements in a Flue Gas Desulfurization Sludge Disposal Site ´ eric ´ Fred J.F. Chagnon, Massachusetts Institute of Technology, Cambridge, Massachusetts, Chemically Enhanced Primary Treatment of Wastewater Chavalit Chaliraktrakul, Thammasat University, Pathumthani, Thailand, A Real-Time Hydrological Information System for Cities Lena Ciric, Centre for Ecology and Hydrology, Oxford, United Kingdom, Molecular Biology Tools for Monitoring Biodiversity in Wastewater Treatment Plants Gunther F. Craun, P.E., M.P.H., D.E.E., Gunther F. Craun and Associates, Staunton, Virginia, Improving Waterborne Disease Surveillance James Crocker, Richland, Washington, Radioactive Waste Sukumar Devotta, Environmental Genomics Unit, National Environmental Engineering, Research Institute, CSIR, Nagpur, India, Water Security: An Emerging Issue John E. Dodes, Forest Hills, New York, Fluoridation L. Donald Duke, University of South Florida, Tampa, Florida, Effluent Limitations and the NPDES Permit Timothy J. Downs, Clark University, Worcester, Massachusetts, Municipal Watersheds, Integrated Capacity Building Needs for Water Supply and Wastewater Sanitation Jorg E. Drewes, Colorado School of Mines, Golden, Colorado, Wastewater Reclamation and Reuse Research James B. Duncan, Kennewick, Washington, Aqueous Reactions of Specific Organic Compounds with Ozone ´ ´ ´ Maria-del-Carmen Duran-de-Baz ua, UNAM, National Autonomous University of Mexico, M´exico D.F., M´exico, Sugarcane Industry Wastewaters Treatment, Use of Anaerobic-Aerobic Treatment Systems for Maize Processing Plants C. Erickson, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies T. Erwe, Saarland University, Saarbrucken, ¨ Germany, Application of Microfiltration to Industrial Wastewaters, Bonding of Toxic Metal Ions Susan L. Franklin, Tetra Tech MPS, Ann Arbor, Michigan, Consumer Confidence Reports From Drinking Water Regulation and Health, Wiley 2003, 1962 U.S. Public Health Service Standards Floyd J. Frost, Ph.D., The Lovelace Institutes, Albuquerque, New Mexico, Improving Waterborne Disease Surveillance Michael H. Gerardi, Linden, Pennsylvania, Microbial Foaming in the Activated Sludge Process, Biological Phosphorus Removal in the Activated Sludge Process, Nitrification in the Activated Sludge Process, Denitrification in The Activated Sludge Process

Segun Michael Ade Adelana, University of Ilorin, Ilorin, Kwara State, Nigeria, Water and Human Health, Nitrate Health Effects Franklin Agardy, Forensic Management Association, San Mateo, California, Chemical Drinking Water Standards, Past, Present, and Future Rasheed Ahmad, Khafra Engineering Consultants, Atlanta, Georgia, Filtration With Granular Media, Hydraulic Design of Water Distribution Storage Tanks, Ozone With Activated Carbon for Drinking Water Treatment, Particulate Removal, Filtration Water Treatment, Particulate Matter Removal by Filtration and Sedimentation, Synthetic and Natural Organic Removal by Biological Filtration, Water Filtration, Flocculation, Gravity Separation/Sedimentation, Particulate Matter Removal by Coagulation, Dechlorination, Wastewater Reclamation and Reuse, Wastewater Reclamation and Reuse Treatment Technology ´ de Moreno, Universidad Nacional de Mar del Plata, Julia E. Aizpun Mar del Plata, Argentina, Macrophytes as Biomonitors of Polychlorinated Biphenyls Imram Ali, National Institute of Hydrology, Roorkee, India, Wastewater Treatment and Recycling Technologies George R. Alther, Biomin, Inc., Ferndale, Michigan, The Role of Organoclay in Water Cleanup Pietro Argurio, Universita della Calabria, Rende, Italy, Ultrafiltration—Complexation in Wastewater Treatment Scott Arthur, Heriot-Watt University, Edinburgh, Scotland, United Kingdom, Roof Drainage Hydraulics Samuel C. Ashworth, Idaho National Engineering and Environmental Laboratory, Idaho Falls, Idaho, Polycyclic Aromatic Hydrocarbons, Water Treatment in Spent Nuclear Fuel Storage, Hydrocarbon Treatment Techniques, Metal Speciation and Mobility as Influenced by Landfill Disposal Practices Kwok-Keung Au, Greeley and Hansen, Chicago, Illinois, Removal of Pathogenic Bacteria, Viruses, and Protozoa, Granular Bed and Precoat Filtration Ann Azadpour-Keeley, United States Environmental Protection Agency, Ada, Oklahoma, Virus Transport in the Subsurface Christine L. Bean, University of New Hampshire, Durham, New Hampshire, Giardiasis Asbjorn Bergheim, RF-Rogaland Research, Stavanger, Norway, Waste Treatment in Fish Farms Vipin Bhardwaj, NDWC Engineering Scientist, Diatomaceous Earth Filtration for Drinking Water, Reservoirs, Towers, and Tanks Drinking Water Storage Facilities, Pumps, Water Meters, Preventing Well Contamination, Cross Connection and Backflow Prevention, Repairing Distribution Line Breaks William J. Blanford, Louisiana State University, Baton Rouge, Louisiana, Review of Parasite Fate and Transport in Karstic Aquifers ¨ C. Blocher, Saarland University, Saarbrucken, ¨ Germany, Application of Microfiltration to Industrial Wastewaters Patrick Bond, Kensington, South Africa, The Economics of Water Resources Allocation T.R. Bott, University of Birmingham, Birmingham, United Kingdom, Industrial Cooling Water—Scale Formation, Industrial Cooling Water—Corrosion, Industrial Cooling Water—Biofouling, Energy Dissipation Jeanine L. Boulter-Bitzer, University of Guelph, Guelph, Ontario, Canada, Molecular-Based Detection of Cryptosporidium Parvum in Water Brenda Boutin, National Center for Environmental Assessment, Cincinnati, Ohio, Health Effects of Microbial Contaminants and Biotoxins in Drinking Water, Health Effects of Commonly Occurring Disinfection Byproducts in Municipal Water Supplies Jacqueline Brabants, University of New Hampshire, Durham, New Hampshire, Cryptosporidium Alexander Brinker, Fischereiforschungsstelle des Landes BadenWurttemberg, ¨ Langenargen, Germany, Waste Treatment in Fish Farms B.M. Brouckaert, University of Natal, Durban, South Africa, Key Causes of Drinking Water Quality Failure in a Rural Small Water Supply of South Africa

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CONTRIBUTORS

Sanjiv Gokhale, Vanderbilt University, Nashville, Tennessee, Basics of Underground Water and Sewer Pipeline Assessment, Repair, and Rehabilitation, Trenchless Repair and Rehabilitation Techniques, Problems Encountered During Pipe Repair and Renewal Janusz Guziur, University of Warmia and Mazury in Olsztyn, Olsztyn, Poland, Role of Small Water Reservoirs in Environment Donald R.F. Harleman, Massachusetts Institute of Technology, Cambridge, Massachusetts, Chemically Enhanced Primary Treatment of Wastewater Arnim R.H. Hertle, GHD Pty Ltd, Wembley, Washington, Australia, Fine Bubble Diffused Air Aeration Systems, Aeration James Higgins, USDA-ARS, Beltsville, Maryland, Threat Agents and Water Biosecurity Arthur M. Holst, Philadelphia Water Department, Philadelphia, Pennsylvania, Windmills Robert M. Hordon, Rutgers University, Piscataway, New Jersey, Bottled Water Marsha A. Hosner, DHI, Newtown, Pennsylvania, Approaches for Securing a Water Distribution System Look Hulshoff Pol, Lettinga Associates Foundation, Wageningen, The Netherlands, Anaerobic Wastewater Treatment M. Eng Nguyen Quang Hung, Asian Institute of Technology, Pathumthani, Thailand, A Real-Time Hydrological Information System for Cities Herbert Inhaber, Risk Concepts, Las Vegas, Nevada, Water Use in Energy Production Enos C. Inniss, University of Texas, San Antonio, Texas, Use of Redox Potentials in Wastewater Treatment Th. A. Ioannidis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Solidification/Stabilization of Hazardous Solid Wastes Irena B. Ivshina, Russian Academy of Sciences, Perm, Russia, Microbial Foaming and Bulking in Activated Sludge Plants James A. Jacobs, Environmental Bio-Systems, Inc., Mill Valley, California, Water Impacts from Construction Sites Kauser Jahan, Rowan University, Glassboro, New Jersey, Detergents C.K. Jain, National Institute of Hydrology, Roorkee, India, Wastewater Treatment and Recycling Technologies H.A. Jenner, KEMA Power Generation and Sustainables, Arnhem, The Netherlands, Chlorine and Chlorine Residuals Dick de Jong, IRC International Water and Sanitation Centre, Delft, The Netherlands, Multistage Filtration: An Innovative Water Treatment Technology Mohammad R. Karim, American Water, Quality Control and Research Laboratory, Belleville, Illinois, Microbiological Concerns of Drinking Water Distribution Systems A. Katsoyiannis, Aristotle University of Thessaloniki, Thessaloniki, Greece, The Fate of Persistent Organic Pollutants (POPs) in The Wastewater Treatment Process I. Katsoyiannis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Technologies for Arsenic Removal from Contaminated Water Sources Absar A. Kazmi, Department of Civil Engineering, Roorkee, Uttaranchal, India, Water and Wastewater Properties and Characteristics, Reclaimed Water, Sewage, Wastewater Characterization Jack W. Keeley, Environmental Engineer, Ada, Oklahoma, Virus Transport in the Subsurface Jaehong Kim, Georgia Institute of Technology, Atlanta, Georgia, Municipal Water Supply: Ozonation K. Thomas Klasson, Oak Ridge National Laboratory, Oak Ridge, Tennessee, Mercury Removal From Complex Waste Waters Jeff Kuo, Cerritos, California, Air Stripping Maria S. Kuyukina, Russian Academy of Sciences, Perm, Russia, Microbial Foaming and Bulking in Activated Sludge Plants Zacharia Michael Lahlou, Technical Assistance Consultant, Valves, Water Hammer, Point-of-Use/Point-of-Entry Systems (POU/POE), Leak Detection and Water Loss Control, System Control and Data Acquisition (SCADA) Guenter Langergraber, BOKU—University of Natural Resources and Applied Life Sciences, Vienna, Austria, Ecological Wastewater Management, Wastewater Treatment—Small Scale, Constructed Wetlands Lars Chr Larsen, DHI Water and Environment, Hørsholm, Denmark, A Real-Time Hydrological Information System for Cities

Mark W. LeChevallier, American Water, Voorhees, New Jersey, Microbiological Concerns of Drinking Water Distribution Systems, Potential Risks of Waterborne Transmission of Escherichia coli O157:H7 Hung Lee, University of Guelph, Guelph, Ontario, Canada, MolecularBased Detection of Cryptosporidium Parvum in Water Gatze Lettinga, Wageningen University and Research Center, Wageningen, The Netherlands, Anaerobic Sewage Treatment Srinivasa Lingireddy, University of Kentucky, Lexington, Kentucky, Design of Water Distribution Systems Gerasimos Lyberatos, University of Patras, Patras, Greece, Nitrification of Potable Water Using Trickling Filters Bruce A. Macler, Toxicologist, U.S. Environmental Protection Agency, San Francisco, California, Application of Risk Assessments in Crafting Drinking Water Regulations N. Makala, University of Fort Hare, Alice, South Africa, Assessing the Bactericidal Efficiency of Polydex for the Disinfection of Drinking Water in Rural Areas of South Africa, Key Causes of Drinking Water Quality Failure in a Rural Small Water Supply of South Africa Babalola Makinde-Odusola, Public Utilities, Riverside, California, Well Head Protection, Drinking Water Quality Standards (DWQS)-United States Joe D. Manous, Jr., United States Military Academy, West Point, New York, Biochemical Oxygen Demand Ole Mark, DHI Water and Environment, Hørsholm, Denmark, A Real-Time Hydrological Information System for Cities Kostas A. Matis, Aristotle University, Thessaloniki, Greece, Bonding of Toxic Metal Ions, Application of Microfiltration to Industrial Wastewaters, Flotation as A Separation Process Paul Mavros, Aristotle University, Thessaloniki, Greece, Mixing and Agitation in Water Treatment Systems V. Mavrov, Saarland University, Saarbrucken, ¨ Germany, Application of Microfiltration to Industrial Wastewaters, Bonding of Toxic Metal Ions Steve Maxwell, TechKNOWLEDGEy Strategic Group,Boulder, Colorado, Ten Key Trends That Will Shape the Future of the World Water Industry, The State of the Water Industry—2004 Kevin S. McLeary, Pennsylvania Dept. of Environmental Protection, Harrisburg, Pennsylvania, Wastewater Treatment Processes and Water Reuse, Domestic Sewage Sue McLeod, William Forrest and Sons, Omoa Works, Newarthill, Motherwell, United Kingdom, Odor Abatement in Wastewater Treatment Plants Fayyaz A. Memon, Imperial College, London, United Kingdom, Gray Water Reuse in Households Mirta L. Menone, Universidad Nacional de Mar del Plata, Mar del Plata, ´ Argentina and Consejo Nacional de Investigaciones Cientificas y T´ecnicas (CONICET), Buenos Aires, Argentina, Macrophytes as Biomonitors of Polychlorinated Biphenyls Chris Metzgar, Graphic Designer, Reservoirs, Towers, and Tanks Drinking Water Storage Facilities J.G. Mexal, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies C. Mfenyana, University of Fort Hare, Alice, South Africa, Inadequate Treatment of Wastewater: A Source of Coliform Bacteria in Receiving Surface Water Bodies in Developing Countries—Case Study: Eastern Cape Province of South Africa Z. Michael Lahlou, Technical Assistance Consultant, Water Quality in Distribution Systems Karina S.B. Miglioranza, Universidad Nacional de Mar del Plata, Mar del Plata, Argentina and Consejo Nacional de Investigaciones ´ Cientificas y T´ecnicas (CONICET), Buenos Aires, Argentina, Macrophytes as Biomonitors of Polychlorinated Biphenyls Mel Mirliss J., International Diatomite Producers Association, Diatomaceous Earth Filtration for Drinking Water M. S. Mohan Kumar, Indian Institute of Science, Bangalore, India, Modeling Chlorine Residuals in Urban Water Distribution Systems Dinesh Mohan, Gomti Nagar, Lucknow, Uttar Pradesh India, Granular Activated Carbon, Competitive Adsorption of Several Organics and Heavy Metals on Activated Carbon in Water T. C. Molden, Magnatech Corporation, Fort Wayne, Indiana, Zebra Mussel Control Without Chemicals Raffaele Molinari, Universita` della Calabria, Rende, Italy, Photocatalytic Membrane Reactors in Water Purification, Ultrafiltration—Complexation in Wastewater Treatment

CONTRIBUTORS M.N.B. Momba, University of Fort Hare, Alice, South Africa, Key Causes of Drinking Water Quality Failure in a Rural Small Water Supply of South Africa, Assessing the Bactericidal Efficiency of Polydex for the Disinfection of Drinking Water in Rural Areas of South Africa, Inadequate Treatment of Wastewater: A Source of Coliform Bacteria in Receiving Surface Water Bodies in Developing Countries—Case Study: Eastern Cape Province of South Africa John E. Moore, Hydrologic Consultant, Denver, Colorado, Septic Tank Systems ´ Victor J. Moreno, Universidad Nacional de Mar del Plata, Mar del Plata, Argentina, Macrophytes as Biomonitors of Polychlorinated Biphenyls Chandrika Moudgal, National Center for Environmental Assessment, Cincinnati, Ohio, Health Effects of Microbial Contaminants and Biotoxins in Drinking Water, Health Effects of Commonly Occurring Disinfection Byproducts in Municipal Water Supplies G. R. Munavalli, Walchand College of Engineering, Sangli, India, Modeling Chlorine Residuals in Urban Water Distribution Systems Michael Muntisov, GHD Pty Ltd., Melbourne, Victoria, Australia, Guide to Selection of Water Treatment Processes Susan Murcott, Massachusetts Institute of Technology, Cambridge, Massachusetts, Household Drinking Water Treatment and Safe Storage National Drinking Water Clearinghouse, Treatment for Technologies for Small Drinking Water Systems, Disinfection, Filtration, Corrosion Control, Ion Exchange and Demineralization, Organic Removal, Package Plants, Water Treatment Plant Residuals Management, Lime Softening, Iron and Manganese Removal, Water Conservation Measures, Membrane Filtration, Ozone, Radionuclides, Slow Sand Filtration, Ultraviolet Disinfection Abid M. Nasser, Water Quality Research Laboratory, Ministry of Health, Tel-Aviv, Israel, Persistence of Pathogens in Water Louis H. Nel, University of Pretoria, Pretoria, Gauteng, South Africa, Emerging Waterborne Infectious Diseases Robert Y. Ning, King Lee Technologies, San Diego, California, Reverse Osmosis, Membrane Foulants, Arsenic in Natural Waters, Reverse Osmosis, Process Chemistry, Reverse Osmosis, Membrane Cleaning Office of Water—United States Environmental Protection Agency, EPA’s National Pretreatment Program, 1973–2003: Thirty Years of Protecting The Environment Oladele Ogunseitan, University of California, Irvine, California, Pharmaceuticals in Water Systems A. Okeyo, Programme Unit of Biochemistry and Microbiology, University of Fort Hare, Alice, South Africa, Assessing the Bactericidal Efficiency of Polydex for the Disinfection of Drinking Water in Rural Areas of South Africa Daniel A. Okun, (from Drinking Water Regulation and Health, Wiley 2003) University of North Carolina, Chapel Hill, North Carolina, Drinking Water and Public Health Protection Lindell E. Ormsbee, University of Kentucky, Lexington, Kentucky, Design of Water Distribution Systems Aisling D. O’Sullivan, University College Dublin, Belfield, Ireland, Using Ecosystem Processes in a Constructed Wetland to Treat Mine Wastewater in Ireland Marinus L. Otte, University College Dublin, Belfield, Ireland, Using Ecosystem Processes in a Constructed Wetland to Treat Mine Wastewater in Ireland David Lloyd Owen, Envinsager, Llangoedmor, Ceredigion, United Kingdom, Private Sector Participation, Marketing and Corporate Strategies in Municipal Water Supply and Sewerage L. Palmisano, Universita` di Palermo, Palermo, Italy, Photocatalytic Membrane Reactors in Water Purification E.N. Peleka, Aristotle University, Thessaloniki, Hellas, Bonding of Toxic Metal Ions Jim Philp, Napier University, Edinburgh, Scotland, United Kingdom, The Arsenic Drinking Water Crisis in Bangladesh, Odor Abatement in Wastewater Treatment Plants, Molecular Biology Tools for Monitoring Biodiversity in Wastewater Treatment Plants, Landfill Laurel Phoenix, Green Bay, Wisconsin, Extraterritorial Land Use Control to Protect Water Supplies G. Picchioni, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies

xv

Nicholas J. Pokorny, University of Guelph, Guelph, Ontario, Canada, Molecular-Based Detection of Cryptosporidium Parvum in Water Kelly Pollack, University of California, Irvine, California, Pharmaceuticals in Water Systems Christopher Polley, William Forrest and Sons, Omoa Works, Newarthill, Motherwell, United Kingdom, Odor Abatement in Wastewater Treatment Plants Rathnavel Ponnuswami, CARE2, Redwood City, California, Water Hammer: Quantitative Causes and Effects Prakhar Prakash, Pennsylvania State University, University Park, Pennsylvania, Selective Coagulant Recovery from Water Treatment Plant Residuals Using the Domain Membrane Process C.A. Prochaska, Aristotle University of Thessaloniki, Thessaloniki, Greece, Municipal Storm Water Management, Combined Sewer Overflow Treatment Hemant J. Purohit, Environmental Genomics Unit, National Environmental Engineering, Research Institute, CSIR, Nagpur, India, Water Security: An Emerging Issue S. Rajagopal, Radboud University Nijmegen, Nijmegen, The Netherlands, Chlorine and Chlorine Residuals D. Ramalingam, University of Kentucky, Lexington, Kentucky, Design of Water Distribution Systems Niranjanie Ratnayake, University of Moratuwa, Moratuwa, Sri Lanka, Water Disinfection Using UV Radiation—A Sri Lankan Experience Eugene R. Reahl, Ionics, Inc., Watertown, Massachusetts, Answering the Challenge Robin J. Reash, American Electric Power, Water & Ecological Resource Services, Columbus, Ohio, Electric Generating Plants—Effects of Contaminants Bethany Reed, NESC Graphic Designer, Pumps, Cross Connection and Backflow Prevention Steven J. Renzetti, Brock University, St. Catharines, Ontario, Canada, Economics of Residential Water Demands, Economics of Industrial Water Demands Susan Richardson, U.S. Environmental Protection Agency,, What is in Our Drinking Water? Ingrid Ritchie, Indiana University-Purdue University, Indianapolis, Indiana, Magnetic Water Conditioning Paul D. Robillard, World Water Watch, Cambridge, Massachusetts, Methods of Reducing Radon in Drinking Water D.S. Rodriguez, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies Stephen J. Rooklidge, Aurora, Oregon, Multistage Drinking Water Filtration, Slow Sand Filtration and the Impact of Schmutzdecke David L. Russell, Global Environmental Operations, Inc., Lilburn, Georgia, Introduction to Wastewater Modeling and Treatment Plant Design, Practical Applications of Wastewater Modeling and Treatment Plant Design Z. Samani, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies Petros Samaras, Chemical Process Engineering Research Institute, Thermi-Thessaloniki, Greece, Landfill Leachates: Part 2: Treatment, Landfill Leachates, Part I: Origin and Characterization Technological Educational Institute of West Macedonia, Kozani, Greece Evaluation of Toxic Properties of Industrial Effluents by on-Line Respirometry C. Samara, Aristotle University of Thessaloniki, Thessaloniki, Greece, The Fate of Persistent Organic Pollutants (POPs) in The Wastewater Treatment Process T. Sammis, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies Charles H. Sanderson, Magnatech Corporation, Fort Wayne, Indiana, Zebra Mussel Control Without Chemicals, Physical Water Conditioning Zane Satterfield, NDWC Engineering Scientists, Water Meters Lucas Seghezzo, Wageningen University and Research Center, Wageningen, The Netherlands, Anaerobic Sewage Treatment Arup K. SenGupta, Lehigh University, Bethlehem, Pennsylvania, Selective Coagulant Recovery from Water Treatment Plant Residuals Using the Domain Membrane Process William E. Sharpe, Pennsylvania State University, University Park, Pennsylvania, Methods of Reducing Radon in Drinking Water

xvi

CONTRIBUTORS

M. Siddiqui, University of Utah, Salt Lake City, Utah, Ultraviolet Irradiation, Ozone–Bromide Interactions Kunwar P. Singh, Dinesh Mohan and Kunwar P. Singh, Gomti Nagar, Lucknow, Uttar Pradesh IndiaGranular Activated Carbon, Competitive Adsorption of Several Organics and Heavy Metals on Activated Carbon in Water Marin Slunjski, Orica Watercare, Regency Park, SA, Australia, Ion Exchange—Use of Magnetic Ion Exchange Resin For DOC Removal Jo Smet, IRC International Water and Sanitation Centre, Delft, The Netherlands, Multistage Filtration: An Innovative Water Treatment Technology Stuart A. Smith, Smith-Comeskey Ground Water Science LLC, Upper Sandusky, Ohio, Evaluation of Microbial Components of Biofouling Mervyn Smyth, Centre for Sustainable Technologies, Newtownabbey, United Kingdom, Domestic Solar Water Heaters Muhammad Sohail, Loughborough University, Leicestershire, United Kingdom, Domestic Water Supply—Public–Private Partnership Ludovico Spinosa, National Research Council, BARI, Italy, Sludge Treatment and Disposal Fiona M. Stainsby James C. Philp Sandra Dunbar, Napier University, Edinburgh, Scotland, United Kingdom, Microbial Foaming and Bulking in Activated Sludge Plants Bradley A. Striebig, Gonzaga University, Spokane, Washington, Sewerage Odors—How to Control Patrick Sullivan, Forensic Management Association, San Mateo, California, Chemical Drinking Water Standards, Past, Present, and Future Bryan R. Swistock, Pennsylvania State University, University Park, Pennsylvania, Methods of Reducing Radon in Drinking Water David J. Tonjes, Cashin Associates, PC, Hauppauge, New York, New York City Harbor Survey Jack T. Trevors, University of Guelph, Guelph, Ontario, Canada, Molecular-Based Detection of Cryptosporidium Parvum in Water Konstantinos P. Tsagarakis, University of Crete, Rethymno, Greece, Wastewater Management for Developing Countries Izrail S. Turoviskiy, Izrail S. Turoviskiy, Jacksonville, FloridaProcessing of Sludge, Biosolids, Wastewater Sludge U.S. Environmental Protection Agency—Office of Wastewater Management, What Wastewater Utilities Can Do Now to Guard Against Terrorist and Security Threats, U.S. Geological Survey, Estimated Use of Water in The United States in 1990 Industrial Water Use, Source Water Assessment, Miguel A. Valenzuela, Instituto Politecnico Nacional—ESIQIE. MEXICO,, Wastewater Treatment Techniques—Advanced G. van der Velde, Radboud University Nijmegen, Nijmegen, The Netherlands, Chlorine and Chlorine Residuals D.V. Vayenas, University of Ioannina, Agrinio, Greece, Nitrification of Potable Water Using Trickling Filters

Raghuraman Venkatapathy, Oak Ridge Institute for Science and Education, Cincinnati, Ohio, Health Effects of Microbial Contaminants and Biotoxins in Drinking Water, Health Effects of Commonly Occurring Disinfection Byproducts in Municipal Water Supplies, Disinfectants V.P. Venugopalan, BARC Facilities, Kalpakkam, India, Chlorine and Chlorine Residuals Roger C. Viadero Jr., West Virginia University, Morgantown, West Virginia, Lime–Soda Ash Processes Christian J. Volk, Indiana-American Water, Richmond, Indiana, Corrosion Control in Drinking Water Systems Constantin Von Der Heyden, School of Geography and the Environment, Oxford, United Kingdom, Industrial Mine Use: Mine Waste Nikolay Voutchkov, Poseidon Resources Corporation, Stamford, Connecticut, Desalination, Settling Tanks Sutat Weesakul, Asian Institute of Technology, Pathumthani, Thailand, A Real-Time Hydrological Information System for Cities Uruya Weesakul, Thammasat University, Pathumthani, Thailand, A Real-Time Hydrological Information System for Cities Janice Weihe, American Water, Quality Control and Research Laboratory, Belleville, Illinois, Potential Risks of Waterborne Transmission of Escherichia coli O157:H7 June Weintraub, City and County of San Francisco, Department of Public Health, San Francisco, California, Disinfectants Andy Whiteley, Centre for Ecology and Hydrology, Oxford, United Kingdom, Molecular Biology Tools for Monitoring Biodiversity in Wastewater Treatment Plants Dan Wolz, City of Wyoming, Michigan, Wyoming, Michigan, Getting Our Clean Water Act Together Don J. Wood, University of Kentucky, Lexington, Kentucky, Design of Water Distribution Systems Grant Wright, Heriot-Watt University, Edinburgh, Scotland, United Kingdom, Roof Drainage Hydraulics J. Michael Wright, Harvard School of Public Health, Boston, Massachusetts, Disinfectants W. Zachritz II, New Mexico State University, Las Cruces, New Mexico, Land Applications of Wastewater in Arid Lands: Theory and Case Studies B. Zani, University of Fort Hare, Alice, South Africa, Key Causes of Drinking Water Quality Failure in a Rural Small Water Supply of South Africa Grietje Zeeman, Wageningen University and Research Center, Wageningen, The Netherlands, Anaerobic Sewage Treatment Anastasios Zouboulis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Landfill Leachates: Part 2: Treatment, Landfill Leachates, Part I: Origin and Characterization, Solidification/Stabilization of Hazardous Solid Wastes, Municipal Storm Water Management, Combined Sewer Overflow Treatment, Technologies for Arsenic Removal from Contaminated Water Sources

CONTRIBUTORS Joanna Davies, Syngenta, Bracknell, Berkshire, United Kingdom, The Control of Algal Populations in Eutrophic Water Bodies Maria B. Davoren, Dublin Institute of Technology, Dublin, Ireland, Luminescent Bacterial Biosensors for the Rapid Detection of Toxicants T.A. Delvalls, Facultad de Ciencias del Mar y Ambientales, Cadiz, ´ Spain, Biomarkers and Bioaccumulation: Two Lines of Evidence to Assess Sediment Quality, A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: An Example For Spanish Coasts Nicolina Dias, Centro de Engenharia Biol´ogica, Braga, Portugal, Ciliated Protists as Test Organisms in Toxicity Assessment Galina Dimitrieva-Moats, University of Idaho, Moscow, Idaho, Microbial Detection of Various Pollutants as an Early Warning System for Monitoring of Water Quality and Ecological Integrity of Natural Resources, in Russia Halanaik Diwakara, University of South Australia, Adelaide, Australia, Water Markets in India: Economic and Institutional Aspects Francis G. Doherty, AquaTox Research, Inc., Syracuse, New York, The Submitochondrial Particle Assay as a Biological Monitoring Tool Antonia A. Donta, University of Munster, ¨ Centre for Environmental Research, Munster, ¨ Germany, Sustainable Water Management On Mediterranean Islands: Research and Education Timothy J. Downs, Clark University, Worcester, Massachusetts, Field Sampling and Monitoring of Contaminants, State and Regional Water Supply, Water Resource Sustainability: Concepts and Practices Hiep N. Duc, Environment Protection Authority, NSW, Bankstown, New South Wales Australia, Urban Water Resource and Management in Asia: Ho Chi Minh City Suzanne Du Vall Knorr, Ventura County Environmental Health Division, Ventura, California, Regulatory and Security Requirements for Potable Water Sandra Dunbar, Napier University, Edinburgh, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Diane Dupont, Brock University, St. Catharines, Ontario, Canada, Valuing Water Resources Michael P. Dziewatkoski, Mettler-Toledo Process Analytical, Woburn, Massachusetts, pH Energy Information Administration—Department of Energy, Hydropower—Energy from Moving Water Environment Canada, Water—Here, There, and Everywhere in Canada, Water Conservation—Every Drop Counts in Canada Environmental Protection Agency, Water Recycling and Reuse: The Environmental Benefits M. Eric Benbow, Michigan State University, East Lansing, Michigan, Road Salt Teresa W.-M. Fan, University of Louisville, Louisville, Kentucky, Remediation and Bioremediation of Selenium-Contaminated Waters Federal Emergency Management Agency, Food and Water in an Emergency Huan Feng, Montclair State University, Montclair, New Jersey, Classification and Environmental Quality Assessment in Aquatic Environments ´ N. Buceta Fernandez, Centro de Estudios de Puertos y Costas, Madrid, Spain, A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: An Example For Spanish Coasts Peter D. Franzmann, CSIRO Land and Water, Floreat, Australia, Microbial Activities Management Christian D. Frazar, Silver Spring, Maryland, Biodegradation Rajiv Gandhi Chair, Jawaharlal Nehru University, New Delhi, India, Oil Pollution Suduan Gao, USDA–ARS, Parlier, California, Eh Horst Geckeis, Institut fur ¨ Nukleare Entsorgung, Karlsruhe, Germany, Metal Ion Humic Colloid Interaction Robert Gensemer, Parametrix, Corvallis, Oregon, Effluent Water Regulations in Arid Lands ´ Mario Abel Goncalves, ¸ Faculdade de Ciˆencias da Universidade de Lisoba, Lisoba, Portugal, Background Concentration of Pollutants Neil S. Grigg, Colorado State University, Fort Collins, Colorado, Planning and Managing Water Infrastructure, Drought and Water Supply

Absar Alum, Arizona State University, Tempe, Arizona, Water Quality Management in the U.S.: History of Water Regulation Mohammad N. Almasri, An-Najah National University, Nablus, Palestine, Best Management Practices for Water Resources Linda S. Andrews, Mississippi State University, Biloxi, Mississippi, Shellfish Growing Water Classification, Chlorine Residual Hannah Aoyagi, University of California, Irvine, California, Cytochrome P450 Monooxygenase as an Indicator of PCB/Dioxin-Like Compounds in Fish ¨ Industrie Service GmbH, Munchen, Robert Artinger, TUV ¨ Germany, Column Experiments in Saturated Porous Media Studying Contaminant Transport Mukand Singh Babel, Asian Institute of Technology, Pathumthani, Thailand, Conservation of Water, Integrated Water Resources Management (IWRM) Mark Bailey, Centre for Ecology and Hydrology–Oxford, Oxford, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Shimshon Balanson, Cleveland State University, Cleveland, Ohio, Macrophytes as Biomonitors of Trace Metals Christine L. Bean, University of New Hampshire, Durham, New Hampshire, Protozoa in Water Jennifer Bell, Napier University, Edinburgh, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Lieven Bervoets, University of Antwerp, Antwerp, Belgium, Active Biomonitoring (ABM) by Translocation of Bivalve Molluscs J.M. Blasco, Instituto de Ciencias Marinas de Andaluc´ıa, Cadiz, ´ Spain, A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: An Example For Spanish Coasts Ronny Blust, University of Antwerp, Antwerp, Belgium, Active Biomonitoring (ABM) by Translocation of Bivalve Molluscs Marta Bryce, CEPIS/PAHO, Delft, The Netherlands, Flood of Portals on Water Mario O. Buenfil-Rodriguez, National University of Mexico, Cuernavaca, Morelos, Mexico, Water Use Conservation and Efficiency Jacques Buffle, University of Geneva, Geneva, Switzerland, Colloids and Dissolved Organics: Role in Membrane and Depth Filtration Zia Bukhari, American Water, Belleville, Illinois, Understanding Escherichia Coli O157:H7 and the Need for Rapid Detection in Water John Cairns, Jr., Virginia Polytechnic Institute and State University, Blacksburg, Virginia, Microscale Test Relationships to Responses to Toxicants in Natural Systems Michael J. Carvan III, University of Wisconsin–Milwaukee, Milwaukee, Wisconsin, Genomic Technologies in Biomonitoring M.C. Casado-Mart´ınez, Facultad de Ciencias del Mar y Ambientales, Cadiz, ´ Spain, A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: An Example For Spanish Coasts ´ ´ Angel Tomas Del Valls Casillas, Universidad de Cadiz, Cadiz, Spain, Amphipod Sediment Toxicity Tests, Development and Application of Sediment Toxicity Test for Regulatory Purposes Teresa A. Cassel, University of California, Davis, California, Remediation and Bioremediation of Selenium-Contaminated Waters Augusto Cesar, Universidad de Cadiz, Cadiz, Spain, Amphipod Sediment Toxicity Tests K.W. Chau, The Hong Kong Polytechnic University, Hung Hom, Kowloon, Hong Kong, Water Quality Models: Mathematical Framework Paulo Chaves, Water Resources Research Center, Kyoto University, Japan, Quality of Water in Storage Shankar Chellam, University of Houston, Houston, Texas, Bromide Influence on Trihalomethane and Haloacetic Acid Formation X. Chris Le, University of Alberta, Edmonton, Alberta, Canada, Arsenic Compounds in Water Russell N. Clayshulte, Aurora, Colorado, Water Quality Management in an Urban Landscape Gail E. Cordy, U.S. Geological Survey, A Primer on Water Quality Rupali Datta, University of Texas, San Antonio, Texas, Lead and its Health Effects

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CONTRIBUTORS

Management, Drought Management Planning, Water Infrastructure and Systems, Water Resources Management ˚ ˚ Hakan Hakanson, University of Lund, Lund, Sweden, Dishwashing Water Quality Properties Carol J. Haley, Virginia Water Resources Research Center, Management of Water Resources for Drought Conditions M.G.J Hartl, Environmental Research Institute, University College Cork, Ireland, Development and Application of Sediment Toxicity Tests for Regulatory Purposes Roy C. Haught, U.S. Environmental Protection Agency, Water Supply and Water Resources: Distribution System Research Joanne M. Hay, Lincoln Ventures, Ltd., Lincoln, New Zealand, Biochemical Oxygen Demand and Other Organic Pollution Measures Richard M. Higashi, University of California, Davis, California, Remediation and Bioremediation of Selenium-Contaminated Waters A.Y. Hoekstra, UNESCO–IHE Institute for Water Education, Delft, The Netherlands, Globalization of Water Charles D.D. Howard, Water Resources, Victoria, British Columbia, Canada, River Basin Decisions Support Systems Margaret S. Hrezo, Radford University, Virginia, Management of Water Resources for Drought Conditions Enos C. Inniss, University of Texas, San Antonio, Texas, Perchloroethylene (PCE) Removal James A. Jacobs, Environmental Bio-Systems, Inc., Mill Valley, California, Emerging and Recalcitrant Compounds in Groundwater Chakresh K. Jain, National Institute of Hydrology, Roorkee, India, Water Quality Management, Trace Element Contamination in Groundwater of District Hardwar, Uttaranchal, India, Ground Water Quality in Areas Adjoining River Yamuna at Delhi, India, Irrigation Water Quality in Areas Adjoining River Yamuna At Delhi, India Sanjay Kumar Jain, National Institute of Hydrology, Roorkee, India, Remote Sensing and GIS Application in Water Resources Sharad K. Jain, National Institute of Hydrology, Roorkee, Uttranchal, India, Water Resources of India H.A. Jenner, KEMA Power Generation and Sustainables, Arnhem, The Netherlands, Dose-Response of Mussels to Chlorine Y. Jiang, Hong Kong Baptist University, Kowloon, Hong Kong, Algal Toxins in Water B. Ji, Hong Kong Baptist University, Kowloon, Hong Kong, Algal Toxins in Water ´ N. Jimenez-Tenorio, Facultad de Ciencias del Mar y Ambientales, Cadiz, ´ Spain, Biomarkers and Bioaccumulation: Two Lines of Evidence to Assess Sediment Quality Zhen-Gang Ji, Minerals Management Service, Herndon, Virginia, Water Quality Modeling—Case Studies, Water Quality Models: Chemical Principles Erik Johansson, GS Development AB, Malm¨o, Sweden, Dishwashing Water Quality Properties B. Thomas Johnson, USGS—Columbia Environmental Research Center, Columbia, Missouri, Monitoring Lipophilic Contaminants in the Aquatic Environment using the SPMD-TOX Paradigm Anne Jones-Lee, G. Fred Lee & Associates, El Macero, California, Water Quality Aspects of Dredged Sediment Management, Municipal Solid Waste Landfills—Water Quality Issues Dick de Jong, IRC International Water and Sanitation Centre, Delft, The Netherlands, Flood of Portals on Water Jagath J. Kaluarachchi, Utah State University, Logan, Utah, Best Management Practices for Water Resources Atya Kapley, National Environmental Engineering Research Institute, CSIR, Nehru Marg, Nagpur, India, Salmonella: Monitoring and Detection in Drinking Water I. Katsoyiannis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Arsenic Health Effects Absar A. Kazmi, Nishihara Environment Technology, Tokyo, Japan, Activated Carbon—Powdered, Chlorination Keith O. Keplinger, Texas Institute for Applied Environmental Research, Stephenville, Texas, The Economics of Water Quality Kusum W. Ketkar, Jawaharlal Nehru University, New Delhi, India, Oil Pollution Ganesh B. Keremane, University of South Australia, Adelaide, Australia, Harvesting Rainwater

Rebecca D. Klaper, University of Wisconsin–Milwaukee, Milwaukee, Wisconsin, Genomic Technologies in Biomonitoring Toshiharu Kojiri, Water Resources Research Center, Kyoto University, Japan, Quality of Water in Storage Ken’ichirou Kosugi, Kyoto University, Kyoto, Japan, Lysimeter Soil Water Sampling Manfred A. Lange, University of Munster, ¨ Centre for Environmental Research, Munster, ¨ Germany, Sustainable Water Management On Mediterranean Islands: Research and Education ´ eric ´ Fred Lasserre, Universit´e Laval, Ste-Foy, Qu´ebec, Canada, Water Use in the United States N.K. Lazaridis, Aristotle University, Thessaloniki, Greece, Sorptive Filtration Jamie R. Lead, University of Birmingham, Birmingham, United Kingdom, Trace Metal Speciation G. Fred Lee, G. Fred Lee & Associates, El Macero, California, Water Quality Aspects of Dredged Sediment Management, Municipal Solid Waste Landfills—Water Quality Issues Terence R. Lee, Santiago, Chile, Water Markets: Transaction Costs and Institutional Options, The Provision of Drinking Water and Sanitation in Developing Countries, Spot Prices, Option Prices, and Water Markets, Meeting Water Needs in Developing Countries with Tradable Rights Markku J. Lehtola, National Public Health Institute, Kuopio, Finland, Microbiological Quality Control in Distribution Systems Gary G. Leppard, National Water Research Institute, Burlington, Ontario, Canada, Colloids and Dissolved Organics: Role in Membrane and Depth Filtration Mark LeChevallier, American Water, Voorhees, New Jersey, Understanding Escherichia Coli O157:H7 and the Need for Rapid Detection in Water Nelson Lima, Centro de Engenharia Biol´ogica, Braga, Portugal, Ciliated Protists as Test Organisms in Toxicity Assessment Maria Giulia Lionetto, Universita` di Lecce, Lecce, Italy, Metallothioneins as Indicators of Trace Metal Pollution Jody W. Lipford, PERC, Bozeman, Montana, and Presbyterian College, Clinton, South Carolina, Averting Water Disputes Baikun Li, Pennsylvania State University, Harrisburg, Pennsylvania, Iron Bacteria, Microbial Dynamics of Biofilms, Microbial Forms in Biofouling Events Rongchao Li, Delft University of Technology, Delft, The Netherlands, Transboundary Water Conflicts in the Nile Basin, Institutional Aspects of Water Management in China, Flood Control History in the Netherlands Bryan Lohmar, Economic Research Service, U.S. Department of Agriculture, Will Water Scarcity Limit China’s Agricultural Potential? ´ Inmaculada Riba Lopez, Universidad de Cadiz, Cadiz, Spain, Amphipod Sediment Toxicity Tests M.X. Loukidou, Aristotle University of Thessaloniki, Thessaloniki, Greece, Biosorption of Toxic Metals Scott A. Lowe, Manhattan College, Riverdale, New York, Eutrophication and Organic Loading G. Lyberatos, University of Ioannina, Agrinio, Greece, Cartridge Filters for Iron Removal Kenneth M. Mackenthun, Arlington, Virginia, Water Quality Tarun K. Mal, Cleveland State University, Cleveland, Ohio, Macrophytes as Biomonitors of Trace Metals Philip J. Markle, Whittier, California, Toxicity Identification Evaluation, Whole Effluent Toxicity Controls James T. Markweise, Neptune and Company, Inc., Los Alamos, New Mexico, Assessment of Ecological Effects in Water-Limited Environments, Effluent Water Regulations in Arid Lands Pertti J. Martikainen, University of Kuopio, Kuopio, Finland, Microbiological Quality Control in Distribution Systems M.L. Mart´ın-D´ıaz, Instituto de Ciencias Marinas de Andaluc´ıa, Cadiz, ´ Spain, A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: An Example For Spanish Coasts, Biomarkers and Bioaccumulation: Two Lines of Evidence to Assess Sediment Quality Maria del Carmen Casado Mart´ınez, Universidad de Cadiz, Cadiz, Spain, Amphipod Sediment Toxicity Tests K.A. Matis, Aristotle University, Thessaloniki, Greece, Sorptive Filtration Lindsay Renick Mayer, Goddard Space Flight Center, Greenbelt, Maryland, NASA Helping to Understand Water Flow in the West

CONTRIBUTORS Mark C. Meckes, U.S. Environmental Protection Agency, Water Supply and Water Resources: Distribution System Research Richard W. Merritt, Michigan State University, East Lansing, Michigan, Road Salt Richard Meyerhoff, CDM, Denver, Colorado, Effluent Water Regulations in Arid Lands J. Michael Wright, Harvard School of Public Health, Boston, Massachusetts, Chlorination By-Products Cornelis J.H. Miermans, Institute for Inland Water Management and Waste Water Treatment–RIZA, Lelystad, The Netherlands, SOFIE: An Optimized Approach for Exposure Tests and Sediment Assays Ilkka T. Miettinen, National Public Health Institute, Kuopio, Finland, Microbiological Quality Control in Distribution Systems Dusan P. Miskovic, Northwood University, West Palm Beach, Florida, Oil-Field Brine Diana Mitsova-Boneva, University of Cincinnati, Cincinnati, Ohio, Quality of Water Supplies Tom Mohr, Santa Clara Valley Water District, San Jose, California, Emerging and Recalcitrant Compounds in Groundwater M.C. Morales-Caselles, Facultad de Ciencias del Mar y Ambientales, Cadiz, ´ Spain, Biomarkers and Bioaccumulation: Two Lines of Evidence to Assess Sediment Quality National Drought Mitigation Center, Drought in the Dust Bowl Years National Water-Quality Assessment (NAWQA) Program—U.S. Geological Survey, Source-Water Protection Jennifer Nelson, The Groundwater Foundation, Lincoln, Nebraska, Reaching Out: Public Education and Community Involvement in Groundwater Protection Anne Ng, Swinburne University of Technology, Hawthorne, Victoria, Australia, River Water Quality Calibration, Review of River Water Quality Modeling Software Tools Jacques Nicolas, University of Liege, Arlon, Belgium, Interest in the Use of an Electronic Nose for Field Monitoring of Odors in the Environment Ana Nicolau, Centro de Engenharia Biol´ogica, Braga, Portugal, Ciliated Protists as Test Organisms in Toxicity Assessment Diana J. Oakes, University of Sydney, Lidcombe, Australia, Environmental Applications with Submitochondrial Particles Oladele A. Ogunseitan, University of California, Irvine, California, Microbial Enzyme Assays for Detecting Heavy Metal Toxicity, Cytochrome P450 Monooxygenase as an Indicator of PCB/Dioxin-Like Compounds in Fish J. O’Halloran, Environmental Research Institute, University College Cork, Ireland, Development and Application of Sediment Toxicity Tests for Regulatory Purposes Victor Onwueme, Montclair State University, Montclair, New Jersey, Classification and Environmental Quality Assessment in Aquatic Environments Alper Ozkan, Selcuk University, Konya, Turkey, Coagulation and Flocculation in Practice Neil F. Pasco, Lincoln Ventures, Ltd., Lincoln, New Zealand, Biochemical Oxygen Demand and Other Organic Pollution Measures B.J.C. Perera, Swinburne University of Technology, Hawthorne, Victoria, Australia, River Water Quality Calibration, Review of River Water Quality Modeling Software Tools Jim Philip, Napier University, Edinburgh, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Laurel Phoenix, Green Bay, Wisconsin, Source Water Quality Management, Water Managed in the Public Trust Randy T. Piper, Dillon, Montana, Overview and Trends in the International Water Market John K. Pollak, University of Sydney, Lidcombe, Australia, Environmental Applications with Submitochondrial Particles ¨ Dorte Poszig, University of Munster, ¨ Centre for Environmental Research, Munster, ¨ Germany, Sustainable Water Management On Mediterranean Islands: Research and Education Hemant J. Purohit, National Environmental Engineering Research Institute, CSIR, Nehru Marg, Nagpur, India, Salmonella: Monitoring and Detection in Drinking Water Shahida Quazi, University of Texas, San Antonio, Texas, Lead and its Health Effects S. Rajagopal, Radboud University Nijmegen, Toernooiveld, Nijmegen, The Netherlands, Dose-Response of Mussels to Chlorine

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Krishna Ramanujan, Goddard Space Flight Center, Greenbelt, Maryland, NASA Helping to Understand Water Flow in the West Lucas Reijnders, University of Amsterdam, Amsterdam, The Netherlands, Sustainable Management of Natural Resources Steven J. Renzetti, Brock University, St. Catharines, Ontario, Canada, Water Demand Forecasting, Water Pricing, Valuing Water Resources Martin Reuss, Office of History Headquarters U.S. Army Corps of Engineers, The Development of American Water Resources: Planners, Politicians, and Constitutional Interpretation, The Expansion of Federal Water Projects I. Riba, Facultad de Ciencias del Mar y Ambientales, Cadiz, ´ Spain, Biomarkers and Bioaccumulation: Two Lines of Evidence to Assess Sediment Quality, A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: An Example For Spanish Coasts Matthew L. Rise, University of Wisconsin–Milwaukee, Milwaukee, Wisconsin, Genomic Technologies in Biomonitoring Arthur W. Rose, Pennsylvania State University, University Park, Pennsylvania, Acid Mine Drainage—Extent and Character, Passive Treatment of Acid Mine Drainage (Wetlands) Barry H. Rosen, US Fish & Wildlife Service, Vero Beach, Florida, Waterborne Bacteria Serge Rotteveel, Institute for Inland Water Management and Waste Water Treatment–RIZA, Lelystad, The Netherlands, SOFIE: An Optimized Approach for Exposure Tests and Sediment Assays Timothy J. Ryan, Ohio University, Athens, Ohio, Water Sampling and Laboratory Safety Randall T. Ryti, Neptune and Company, Inc., Los Alamos, New Mexico, Assessment of Ecological Effects in Water-Limited Environments Masaki Sagehashi, University of Tokyo, Tokyo, Japan, Biomanipulation Basu Saha, Loughborough University, Loughborough, United Kingdom, Activated Carbon: Ion Exchange and Adsorption Properties Md. Salequzzaman, Khulna University, Khulna, Bangladesh, Ecoregions: A Spatial Framework for Environmental Management Dibyendu Sarkar, University of Texas, San Antonio, Texas, Lead and its Health Effects Peter M. Scarlett, Winfrith Technology Centre, Dorchester, Dorset, United Kingdom, The Control of Algal Populations in Eutrophic Water Bodies Trifone Schettino, Universita` di Lecce, Lecce, Italy, Metallothioneins as Indicators of Trace Metal Pollution Lewis Schneider, North Jersey District Water Supply Commission, Wanaque, New Jersey, Classification and Environmental Quality Assessment in Aquatic Environments Wolfram Schuessler, Institut fur ¨ Nukleare Entsorgung, Karlsruhe, Germany, Column Experiments in Saturated Porous Media Studying Contaminant Transport K.D. Sharma, National Institute of Hydrology, Roorkee, India, Water Quality Management Mukesh K. Sharma, National Institute of Hydrology, Roorkee, India, Ground Water Quality in Areas Adjoining River Yamuna at Delhi, India, Irrigation Water Quality in Areas Adjoining River Yamuna At Delhi, India Daniel Shindler, UMDNJ, New Brunswick, New Jersey, Methemoglobinemia Slobodan P. Simonovic, The University of Western Ontario, London, Ontario, Canada, Water Resources Systems Analysis, Fuzzy Criteria for Water Resources Systems Performance Evaluation, Participatory Multicriteria Flood Management Shahnawaz Sinha, Malcolm Pirnie Inc., Phoenix, Arizona, Disinfection By-Product Precursor Removal from Natural Waters Joseph P. Skorupa, U.S. Fish and Wildlife Service, Remediation and Bioremediation of Selenium-Contaminated Waters Roel Smolders, University of Antwerp, Antwerp, Belgium, Active Biomonitoring (ABM) by Translocation of Bivalve Molluscs Jinsik Sohn, Kookmin University, Seoul, Korea, Disinfection By-Product Precursor Removal from Natural Waters Fiona Stainsby, Napier University, Edinburgh, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Ross A. Steenson, Geomatrix, Oakland, California, Land Use Effects on Water Quality Leonard I. Sweet, Engineering Labs Inc., Canton, Michigan, Application of the Precautionary Principle to Water Science

xiv

CONTRIBUTORS

Kenneth K. Tanji, University of California, Davis, California, Eh Ralph J. Tella, Lord Associates, Inc., Norwood, Massachusetts, Overview of Analytical Methods of Water Analyses With Specific Reference to EPA Methods for Priority Pollutant Analysis William E. Templin, U.S. Geological Survey, Sacramento, California, California—Continually the Nation’s Leader in Water Use Rita Triebskorn, Steinbeis-Transfer Center for Ecotoxicology and Ecophysiology, Rottenburg, Germany, Biomarkers, Bioindicators, and the Trondheim Biomonitoring System Nirit Ulitzur, Checklight Ltd., Tivon, Israel, Use of Luminescent Bacteria and the Lux Genes For Determination of Water Quality Shimon Ulitzur, Technion Institute of Technology, Haifa, Israel, Use of Luminescent Bacteria and the Lux Genes For Determination of Water Quality U.S. Environmental Protection Agency, How We Use Water in These United States U.S. Agency for International Development (USAID), Water and Coastal Resources U.S. Geological Survey, Water Quality, Water Science Glossary of Terms G. van der velde, Radboud University Nijmegen, Toernooiveld, Nijmegen, The Netherlands, Dose-Response of Mussels to Chlorine F.N.A.M. van pelt, Environmental Research Institute, University College Cork, Ireland, Development and Application of Sediment Toxicity Tests for Regulatory Purposes D.V. Vayenas, University of Ioannina, Agrinio, Greece, Cartridge Filters for Iron Removal Raghuraman Venkatapathy, Oak Ridge Institute for Science and Education, Cincinnati, Ohio, Alternative Disinfection Practices and Future Directions for Disinfection By-Product Minimization, Chlorination ByProducts V.P. Venugopalan, BARC Facilities, Kalpakkam, India, Dose-Response of Mussels to Chlorine Jos P.M. Vink, Institute for Inland Water Management and Waste Water Treatment–RIZA, Lelystad, The Netherlands, Heavy Metal Uptake Rates Among Sediment Dwelling Organisms, SOFIE: An Optimized Approach for Exposure Tests and Sediment Assays Judith Voets, University of Antwerp, Antwerp, Belgium, Active Biomonitoring (ABM) by Translocation of Bivalve Molluscs Mark J. Walker, University of Nevada, Reno, Nevada, Water Related Diseases William R. Walker, Virginia Water Resources Research Center, Management of Water Resources for Drought Conditions Xinhao Wang, University of Cincinnati, Cincinnati, Ohio, Quality of Water Supplies Corinna Watt, University of Alberta, Edmonton, Alberta, Canada, Arsenic Compounds in Water

Janice Weihe, American Water, Belleville, Illinois, Understanding Escherichia Coli O157:H7 and the Need for Rapid Detection in Water June M. Weintraub, City and County of San Francisco Department of Public Health, San Francisco, California, Chlorination By-Products, Alternative Disinfection Practices and Future Directions for Disinfection By-Product Minimization Victor Wepener, Rand Afrikaans University, Auckland Park, South Africa, Active Biomonitoring (ABM) by Translocation of Bivalve Molluscs ˚ Wernersson, GS Development AB, Malm¨o, Sweden, DishwashEva Stahl ing Water Quality Properties Andrew Whiteley, Centre for Ecology and Hydrology–Oxford, Oxford, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Siouxsie Wiles, Imperial College London, London, United Kingdom, Bioluminescent Biosensors for Toxicity Testing Thomas M. Williams, Baruch Institute of Coastal Ecology and Forest Science, Georgetown, South Carolina, Water Quality Management in a Forested Landscape Parley V. Winger, University of Georgia, Atlanta, Georgia, Water Assessment and Criteria M.H. Wong, Hong Kong Baptist University, Kowloon, Hong Kong, Algal Toxins in Water R.N.S. Wong, Hong Kong Baptist University, Kowloon, Hong Kong, Algal Toxins in Water J. Michael Wright, Harvard School of Public Health, Boston, Massachusetts, Alternative Disinfection Practices and Future Directions for Disinfection By-Product Minimization Gary P. Yakub, Kathleen Stadterman-Knauer Allegheny County Sanitary Authority, Pittsburgh, Pennsylvania, Indicator Organisms Yeomin Yoon, Northwestern University, Evanston, Illinois, Disinfection By-Product Precursor Removal from Natural Waters M.E. Young, Conwy, United Kingdom, Water Resources Challenges in the Arab World Mehmet Ali Yurdusev, Celal Bayar University, Manisa, Turkey, Integration of Environmental Impacts into Water Resources Planning Karl Erik Zachariassen, Norwegian University of Science and Technology, Trondheim, Norway, Physiological Biomarkers and the Trondheim Biomonitoring System Luke R. Zappia, CSIRO Land and Water, Floreat, Australia, Microbial Activities Management Harry X. Zhang, Parsons Corporation, Fairfax, Virginia, Water Quality Management and Nonpoint Source Control, Water Quality Models for Developing Soil Management Practices Igor S. Zonn, Lessons from the Rising Caspian A.I. Zouboulis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Biosorption of Toxic Metals, Arsenic Health Effects

CONTRIBUTORS ´ de Moreno, Universidad Nacional de Mar del Plata, Julia E. Aizpun Mar del Plata, Argentina, Occurrence of Organochlorine Pesticides in Vegetables Grown on Untreated Soils from an Agricultural Watershed, Pesticide Chemistry in the Environment Mahbub Alam, Kansas State University, Garden City, Kansas, Vacuum Gauge Tensiometer Absar Alum, Arizona State University, Tempe, Arizona, Microbial Quality of Reclaimed Irrigation: International Perspective Peyman Daneshkar Arasteh, Soil Conservation and Watershed Management Research Institute (SCWMRI), Tehran, Iran, Minimum Environmental Flow Regimes, Large Area Surface Energy Balance Estimation Using Satellite Imagery, Water Spreading Muhammad Nadeem Asghar, International Water Management Institute (IWMI), Lahore, Pakistan, Irrigation Wells, Skimmed Groundwater, Sprinkler Irrigation Mukand Singh Babel, Asian Institute of Technology, Pathumthani, Thailand, Soil Moisture Measurement—Neutron Frank Balon, Buffalo District, U.S. Army Corps of Engineers, Corps Turned Niagara Falls Off, On Again Joseph D. Bankston, Louisiana State University, Baton Rouge, Louisiana, Pumping Stations Nathalie Barrette, Laval University, Qu´ebec, Canada, Greenhouse Gas Emissions From Hydroelectric Reservoirs Luis Berga, ETSIn Caminos, Barcelona, Spain, Floods as a Natural Hazard Asbjørn Bergheim, RF-Rogaland Research, Stavanger, Norway, Water Pollution From Fish Farms K.K.S. Bhatia, National Institute of Hydrology, Roorkee, Uttaranchal, India, Dilution-Mixing Zones and Design Flows, Surface Water Pollution, Assessment of Pollution Outflow From Large Agricultural Areas Sandra Bird, U.S. Environmental Protection Agency, Impervious Cover—Paving Paradise Peter E. Black, State University of New York College of Environmental Science and Forestry, Syracuse, New York, Watershed Hydrology Robert W. Black, National Water Quality Assessment Program, U.S. Geological Survey, Organic Compounds and Trace Elements in Freshwater Streambed Sediment and Fish from the Puget Sound Basin Alexander Brenning, Humboldt–Universitat ¨ zu Berlin, Berlin, Germany, Rock Glacier Karen D. Brettschneider, Houston-Galveston Area Council, Houston, Texas, Urban Water Studies, Watershed Emera Bridger, SUNY-ESF, Syracuse, New York, Forests and Wetlands Alexander Brinker, Fischereiforschungsstelle des Landes BadenWurttemberg, ¨ Langenargen, Germany, Water Pollution From Fish Farms Gary A. Buchanan, Division of Science, Research, and Technology, Trenton, New Jersey, Fish Consumption Advisories Yong Cai, Florida International University, Miami, Florida, Metal Tolerance in Plants: The Roles of Thiol-Containing Peptides Rene´ Canuel, University of Qu´ebec in Montr´eal, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases Virginia Carter, U.S. Department of the Interior, Classification of Wetlands and Deepwater Habitats of the United States Harenda Singh Chauhan, G.B. Pant University of Agriculture and Technology, Pantnagar, India, Microirrigation: An Approach to Efficient Irrigation Shulin Chen, Washington State University, Pullman, Washington, Aquaculture Technology for Producers Pietro Chiavaccini, Universita di Pisa, Pisa, Italy, Flood Control Structures, Hydraulics Brent C. Christopher, Montana State University, Bozeman, Montana, Subglacial Lake Vostok Xuefeng Chu, Grand Valley State University, Muskegon, Michigan, Pesticide Occurrence and Distribution in Relation to Use Michelle Clarke, Cranfield University, North Wyke, Devon, United Kingdom, Water Quality Management in an Agricultural Landscape, Soil Erosion and Control Practices Thomas R. Clarke, USDA ARS U.S. Water Conservation Laboratory, Phoenix, Arizona, Crop Water Stress Detection Using Remote Sensing

B.D. Clinton, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Measuring and Modeling Tree and Stand Level Transpiration Steve Colman, U.S. Geological Survey, Lake Baikal—A Touchstone for Global Change and Rift Studies Charles M. Cooper, USDA Agricultural Research Service National Sedimentation Laboratory, Oxford, Mississippi, Drainage Ditches Dennis L. Corwin, Salinity Laboratory, Riverside, California, Soil Salinity Lewis M. Cowardin, U.S. Department of the Interior, Classification of Wetlands and Deepwater Habitats of the United States Christophe Cudennec, Ecole Nationale Sup´erieure Agronomique, Rennes, France, Unit Hydrograph, Rivers and Streams: One-Way Flow System Rupali Datta, University of Texas, San Antonio, Texas, Phytoremediation By Constructed Wetlands, Soil N Management Impact on The Quality of Surface and Subsurface Water Francesca Dellacasa, Universita` di Pisa, Pisa, Italy, River Basins Richard Dowling, Pittsburgh District, U.S. Army Corps of Engineers, Innovative Pens Hatch Thousands of Trout ´ Eric Duchemin, DREX Environment, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases Julia Duzant, Cranfield University, North Wyke, Devon, United Kingdom, Water Quality Management in an Agricultural Landscape, Soil Erosion and Control Practices J. Gordon Edwards, San Jose, California, Effects of DDT in Surface Water on Bird Abundance and Reproduction—A History Environment Canada, Water—The Canadian Transporter K.J. Elliott, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Measuring and Modeling Tree and Stand Level Transpiration Theodore A. Endreny, SUNY-ESF, Syracuse, New York, Forests and Wetlands, Riparian Systems, Great Lakes Wayne D. Erskine, State Forests of New South Wales, Beecroft, New South Wales, Australia, Stream Classification, Gully Erosion, Sediment Load Measurements Xing Fang, Lamar University, Beaumont, Texas, Culvert Design, Water Turbine, Hydraulics of Pressurized Flow, Open Channel Design, Storage and Detention Facilities, Streamflow Jerry L. Farris, Arkansas State University, State University, Arkansas, Drainage Ditches Douglas H. Fender, International Turf Producers Foundation, Meadows, Illinois, Landscape Water-Conservation Techniques ´ Ferreyra, Ag Connections, Inc., Murray, Kentucky, Stomates R. Andres Markus Flury, Washington State University, Pullman, Washington, Dyes As Hydrological Tracers Julia Freedgood, American Farmland Trust, Washington, District of Columbia, Agriculture and Land Use Planning Annette Geller, UFZ Center for Environmental Research Leipzig-Halle Ltd., Magdeburg, Germany, Limnology Walter Geller, UFZ Center for Environmental Research, Magdeburg, Germany, Limnology, Lakes Walter H. Geller, UFZ, Dept. of Inland Water Research, Magdeburg, Germany, Acidification of Freshwater Resources Robert J. Gilliom, U.S. Geological Survey, Classification and Mapping of Agricultural Land for National Water-Quality Assessment Francis C. Golet, U.S. Department of the Interior, Classification of Wetlands and Deepwater Habitats of the United States Mariana Gonzalez, Universidad Nacional de Mar del Plata, Mar del Plata, Argentina and Consejo Nacional de Investigaciones Cient´ıficas y T´ecnicas (CONICET), Buenos Aires, Argentina, Pesticide Chemistry in the Environment, Occurrence of Organochlorine Pesticides in Vegetables Grown on Untreated Soils from an Agricultural Watershed Eve Gruntfest, (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003), Floods

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CONTRIBUTORS

Suresh K. Gupta, CSSRI, Karnal, Haryana, India, Salt Tolerance, Waterlogging John Hall, New Orleans District, U.S. Army Corps of Engineers, Lasers Scan Levees from the Air Dorota Z. Haman, University of Florida, Gainesville, Florida, Screen Filters for Microirrigation, Media Filters for Microirrigation, Microirrigation Blaine Hanson, LAWR, Davis, California, Deep-Well Turbine Pumps David R. Hargis, Hargis + Associates, Inc., Tucson, Arizona, Forensic Hydrogeology Thomas Harter, University of California, Davis, California, Animal Farming Operations: Groundwater Quality Issues G.J. Harvey, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Measuring and Modeling Tree and Stand Level Transpiration Karl E. Havens, South Florida Water Mgmt. District, West Palm Beach, Florida, Submerged Aquatic Plants Affect Water Quality in Lakes Peder Hjorth, Lund University, Lund, Sweden, River Basin Planning and Coordination Joseph Holden, University of Leeds, Leeds, United Kingdom, Surface Runoff and Subsurface Drainage Robert M. Hordon, Rutgers University, Piscataway, New Jersey, Flowduration Curves ´ Stephane Houel, University of Qu´ebec in Montr´eal, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases Xiangjiang Huang, Colorado State University, Fort Collins, Colorado, Drainage Networks Fen C. Hunt, U.S. Department of Agriculture, Washington, District of Columbia, Agriculture and Land Use Planning Stephen W. Hussey, Dabane Trust, Burnside, Bulawayo, Zimbabwe, Water from Saturated River Sediment—Sand Abstraction Deborah Hutchinson, U.S. Geological Survey, Lake Baikal—A Touchstone for Global Change and Rift Studies James A. Jacobs, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater Assessment Using Soil Sampling Techniques Chakresh K. Jain, National Institute of Hydrology, Roorkee, India, Adsorption of Metal Ions On Bed Sediments Sharad K. Jain, National Institute of Hydrology, Roorkee, Uttranchal, India, Hydroelectric Power, Reservoir Sedimentation, ReservoirsMultipurpose, Base Flow, Ganga River, India Steven Jennings, (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003), Floods Ramakar Jha, National Institute of Hydrology, Roorkee, Uttranchal, India, Hydrological Processes and Measured Pollutant Loads, DilutionMixing Zones and Design Flows, Surface Water Pollution, Assessment of Pollution Outflow From Large Agricultural Areas ¨ Klaus Johnk, University of Amsterdam, Amsterdam, The Netherlands, Heat Balance of Open Waterbodies Bill Johnson, Washington State University, Pullman, Washington, Aquaculture Technology for Producers Brit Johnson, Washington State University, Pullman, Washington, Aquaculture Technology for Producers Anne Jones-Lee, G. Fred Lee & Associates, El Macero, California, Eutrophication (Excessive Fertilization), Unrecognized Pollutants, Urban Stormwater Runoff Water Quality Issues Pierre Y. Julien, Colorado State University, Fort Collins, Colorado, Sedimentation, Rivers Alpana Khairom, University of Texas, San Antonio, Texas, Phytoremediation By Constructed Wetlands Jacob W. Kijne, Herts, United Kingdom, World’s Major Irrigation Areas, Maintaining Salt Balance on Irrigated Land Charles W. Kirby, Paciulli, Simmons & Assoc., Fairfax, Virginia, Water Quality in Suburban Watersheds Matthias Koschorreck, Center of Environmental Research (UFZ), Magdeburg, Germany, Microorganisms in Their Natural Environment Demetris Koutsoyiannis, National Technical University, Athens, Greece, Stochastic Simulation of Hydrosystems, Reliability Concepts in Reservoir Design, Hydrologic Persistence and The Hurst Phenomenon V. Kulik, CYPUM PTY LTD, Canberra City, Australia, Combustible Watersheds

Richard Lanyon, Metropolitan Water Reclamation District of Greater Chicago, Chicago, Illinois, Reversal of the Chicago River Edward T. Laroe, U.S. Department of the Interior, Classification of Wetlands and Deepwater Habitats of the United States Catherine Larose, University of Qu´ebec in Montr´eal, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases Frederic Lasserre, Laval University, Quebec City, Canada, The Aral Sea Disaster: Environment Issues and Nationalist Tensions, Irrigation in The United States Jamie R. Lead, University of Birmingham, Birmingham, United Kingdom, Freshwater Colloids G. Fred Lee, G. Fred Lee & Associates, El Macero, California, Eutrophication (Excessive Fertilization), Unrecognized Pollutants, Urban Stormwater Runoff Water Quality Issues Leo S. Leonhart, Hargis + Associates, Inc., Tucson, Arizona, Forensic Hydrogeology, Cienega Xu Liang, University of California, Berkeley, California, Land Surface Modeling Srinivasa Lingireddy, University of Kentucky, Lexington, Kentucky, Calibration of Hydraulic Network Models Zhi-Qing Lin, Southern Illinois University at Edwardsville, Edwardsville, Illinois, Bioaccumulation Rongchao Li, Delft University of Technology, Delft, The Netherlands, Flood Control in the Yellow River Basin in China G.V. Loganathan, Virginia Polytechnic Institute and State University, Blacksburg, Virginia, Instream Flow Methods Marc Lucotte, University of Qu´ebec in Montr´eal, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases Dorene E. Maccoy, National Water Quality Assessment Program, U.S. Geological Survey, Organic Compounds and Trace Elements in Freshwater Streambed Sediment and Fish from the Puget Sound Basin Chandra Madramootoo, Macdonald Campus of McGill University, SteAnne de Bellevue, Quebec, Canada, Tile Drainage: Impacts, Plant Growth, and Water Table Levels, Water Logging: Topographic and Agricultural Impacts, Water Table Contribution to Crop Evapotranspiration Ole Mark, Asian Institute of Technology, Hørsholm, Denmark, Modeling of Urban Drainage and Stormwater, Modeling of Water Quality in Sewers, Urban Flooding Brane Maticic, Ljubljana, Slovenia, Nitrate Pollution Prevention Matthew P. McCartney, International Water Management Institute, Pretoria, South Africa, Wetlands: Uses, Functions, and Values Marianne McHugh, Cranfield University, Silsoe, United Kingdom, Water Quality Management in an Agricultural Landscape, Soil Erosion and Control Practices, Soil Conservation T.C. Mcintyre, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Plant and Microorganism Selection for Phytoremediation of Hydrocarbons and Metals Karina S.B. Miglioranza, Universidad Nacional de Mar del Plata, Mar del Plata, Argentina and Consejo Nacional de Investigaciones Cient´ıficas y T´ecnicas (CONICET), Buenos Aires, Argentina, Pesticide Chemistry in the Environment, Occurrence of Organochlorine Pesticides in Vegetables Grown on Untreated Soils from an Agricultural Watershed Matjaz Mikos, University of Ljubljana, Ljubljana, Slovenia, Sediment Transport Myron J. Mitchell, State University of New York, Syracuse, New York, Episodic Acidification Jarai Mon, Washington State University, Pullman, Washington, Dyes As Hydrological Tracers Matthew T. Moore, USDA Agricultural Research Service National Sedimentation Laboratory, Oxford, Mississippi, Drainage Ditches M. Susan Moran, USDA ARS Southwest Watershed Research Center, Tucson, Arizona, Crop Water Stress Detection Using Remote Sensing V´ıctor J. Moreno, Universidad Nacional de Mar del Plata, Mar del Plata, Argentina, Pesticide Chemistry in the Environment, Occurrence of Organochlorine Pesticides in Vegetables Grown on Untreated Soils from an Agricultural Watershed R.P.C. Morgan, Cranfield University, Silsoe, United Kingdom, Soil Conservation Roy Morgan, Cranfield University, North Wyke, Devon, United Kingdom, Soil Erosion and Control Practices

CONTRIBUTORS

xiii

National Wild and Scenic Rivers System—National Park Service, River and Water Facts

K.D. Sharma, National Institute of Hydrology, Roorkee, Uttranchal, India, Hydrological Processes and Measured Pollutant Loads

J.R. Newman, IACR-Centre for Aquatic Plant Management, Reading Berkshire, United Kingdom, Weed Control Strategies

G. S. Shrivastava, University of the West Indies, St. Augustine, Trinidad, West Indies, Watershed Management for Environmental Quality and Food Security

Jeffrey D. Niemann, Colorado State University, Fort Collins, Colorado, Drainage Networks NOAA Great Lakes Environmental Research Lab, NOAA Lake Level Forecast for Lake Michigan Right on Target C.S.P. Ojha, Indian Institute of Technology, Roorkee, Uttaranchal, India, Surface Water Pollution, Dilution-Mixing Zones and Design Flows, Assessment of Pollution Outflow From Large Agricultural Areas Lindell E. Ormsbee, University of Kentucky, Lexington, Kentucky, Calibration of Hydraulic Network Models Stefano Pagliara, Universita di Pisa, Pisa, Italy, Flood Control Structures, Hydraulics, River Basins, Levees for Flood Protection Qiang Pan, Washington State University, Pullman, Washington, Aquaculture Technology for Producers

Paul K. Sibley, University of Guelph, Guelph, Ontario, Canada, Chironomids in Sediment Toxicity Testing Slobodan P. Simonovic, The University of Western Ontario, London, Ontario, Canada, Flood Prevention Vijay P. Singh, Louisiana State University, Baton Rouge, Louisiana, Base Flow, Kinematic Wave Flow Routing, Kinematic Shock, Kinematic Wave and Diffusion Wave Theories, Surface Water Pollution, Hydrological Processes and Measured Pollutant Loads Bellie Sivakumar, University of California, Davis, California, Hydrologic Thresholds Nicolas Soumis, University of Qu´ebec in Montr´eal, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases

Laurel Phoenix, Green Bay, Wisconsin, Lakes—Discharges To

Susan-Marie Stedman, NMFS F/HC, Silver Spring, Maryland, Coastal Wetlands Pasquale Steduto, Mediterranean Agronomic Institute, Valenzano, Italy, Agricultural Water Use Efficiency (WUE) and Productivity (WP)

Gloria Post, Division of Science, Research, and Technology, Trenton, New Jersey, Fish Consumption Advisories

Alan Stern, Division of Science, Research, and Technology, Trenton, New Jersey, Fish Consumption Advisories

Simone Pozzolini, Universita` di Pisa, Pisa, Italy, Levees for Flood Protection

Eric A. Strauss, Upper Midwest Environmental Sciences Center, La Crosse, Wisconsin, Microbiology of Lotic Aggregates and Biofilms

John C. Priscu, Montana State University, Bozeman, Montana, Subglacial Lake Vostok

Tim Sullivan, E & S Environmental, Corvallis, Oregon, Acidification—Chronic

Geoffrey Petts, University of Birmingham, Birmingham, United Kingdom, Regulated Rivers, Environmental Flows

Nitish Priyadarshi, Ranchi University, Ranchi, Jharkhand, India, Pollution of Surface Waters, Trace Elements in Water, Sediment, and Aquatic Biota—Effects of Biology and Human Activity, Cultural Eutrophication, Water Quality in Ponds Rudi Rajar, University of Ljubljana, Ljubljana, Slovenia, Numerical Modeling of Currents Tara Reed, University of Wisconsin-Green Bay, Green Bay, Wisconsin, Tile Drainage Lucas Reijnders, Universiteit van Amsterdam, Amsterdam, The Netherlands, Biofuel Alternatives to Fossil Fuels Jane Rickson, Cranfield University, North Wyke, Devon, United Kingdom, Soil Erosion and Control Practices R.J. Rickson, Cranfield University, Silsoe, United Kingdom, Soil Conservation Eliot C. Roberts, Rosehall Associates, Sparta, Tennessee, Soil Water Issues T. Lackey Robert, United States Environmental Protection Agency, Corvallis, Oregon, Fisheries: History, Science, and Management Barry H. Rosen, U.S. Fish and Wildlife Service, Vero Beach, Florida, Potential Health Issues Associated With Blue-Green Algae Blooms in Impoundments, Ponds and Lakes William R. Roy, Illinois State Geological Survey, Champaign, Illinois, Remediation of Pesticide-Contaminated Soil at Agrichemical Facilities Bahram Saghafian, Soil Conservation and Watershed Management Research Institute, Tehran, Iran, Time-Area and The Clark RainfallRunoff Transformation, Time of Concentration and Travel Time in Watersheds, Flood Source Mapping in Watersheds, Floodwater Spreading Dibyendu Sarkar, University of Texas at San Antonio, San Antonio, Texas, Phytoremediation By Constructed Wetlands, Soil N Management Impact on The Quality of Surface and Subsurface Water, Soil Phosphorus Availability and Its Impact on Surface Water Quality Hubert H.G. Savenije, Delft University of Technology, The Netherlands, Interception Miklas Scholz, The University of Edinburgh, Edinburgh, United Kingdom, Urban Runoff

Leonard I. Sweet, Environmental Energy Group, Engineering Labs Inc., Canton, Michigan, Fish Cells in the Toxicological Evaluation of Environmental Contaminants Maria A. Szumiec, Polish Academy of Sciences, Chybie, Poland, Pond Aquaculture—Modeling and Decision Support Systems, An Outline of the History of Fishpond Culture in Silesia, the Western Part of Poland David D. Tarkalson, University of Nebraska-Lincoln, North Platte, Nebraska, Nitrification Gail P. Thelin, U.S. Geological Survey, Classification and Mapping of Agricultural Land for National Water-Quality Assessment Chacharee Therapong, University of Texas, San Antonio, Texas, Phytoremediation By Constructed Wetlands Mladen Todorovic, Mediterranean Agronomic Institute of Bari, Valenzano, Bari, Italy, Crop Water Requirements, Crop Evapotranspiration Mark R. Tompkins, University of California, Berkeley, California, Fish Passage Facilities, Floodplain, Fishing Waters U.S. Environmental Protection Agency—Office of Water, Office of Wetlands, Oceans and Watersheds, Wetlands Overview Vandana Vandanapu, University of Texas, San Antonio, Texas, Soil N Management Impact on The Quality of Surface and Subsurface Water Roger C. Viadero, West Virginia University, Morgantown, West Virginia, Sedimentation and Flotation, Factors Affecting Fish Growth and Production, Geochemistry of Acid Mine Drainage J.M. Vose, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Measuring and Modeling Tree and Stand Level Transpiration Thorsten Wagener, The Pennsylvania State University, University Park, Pennsylvania, Uncertainty Analysis in Watershed Modeling, Watershed Modeling, Modeling Ungauged Watersheds Lizhu Wang, Wisconsin Department of Natural Resources, Monona, Wisconsin, Biotic Integrity Index to Evaluate Water Resource Integrity in Freshwater Systems Sutat Weesakul, DHI Water & Environment, Bangkok, Thailand, Modeling of Urban Drainage and Stormwater Brian B. Weigel, Wisconsin Department of Natural Resources, Monona, Wisconsin, Biotic Integrity Index to Evaluate Water Resource Integrity in Freshwater Systems

Harold L. Schramm, Jr, U.S. Geological Survey, Mississippi State, Mississippi, Water Needs for Freshwater Fisheries Management

Sebastian Weissenberger, University of Qu´ebec in Montr´eal, Montr´eal, Canada, Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases

Andy Seidl, Colorado State University, Fort Collins, Colorado, Agriculture and Land Use Planning

Douglas F. Welsh, Texas A&M University, Collage Station, Texas, Xeriscape

xiv

CONTRIBUTORS

Adrian E. Williams, APEM Ltd., Manchester, United Kingdom, Water Hyacinth—The World’s Most Problematic Weed Tommy S.W. Wong, Nanyang Technological University, Singapore, Kinematic Wave Method For Storm Drainage Design Don J. Wood, University of Kentucky, Lexington, Kentucky, Calibration of Hydraulic Network Models

Chih Ted Yang, Colorado State University, Fort Collins, Colorado, Sedimentation, Dam Removal as River Restoration, Rivers Pablo J. Zarco-Tejada, Instituto de Agricultura Sostenible (IAS-CSIC), C´ordoba, Spain, Crop Water Stress Detection Using Remote Sensing Weihua Zhang, Florida International University, Miami, Florida, Metal Tolerance in Plants: The Roles of Thiol-Containing Peptides

CONTRIBUTORS Eli Dahi, Environmental Development Corporation, Søborg, Denmark, Defluroidation Carl W. David, University of Connecticut, Storrs, Connecticut, The Hydronium Ion Jana Davis, Smithsonian Environmental Research Center, Edgewater, Maryland, Marine Stock Enhancement Techniques Christine Dickenson, Florida Institute of Technology, Melbourne, Florida, An Analysis of the International Maritime Organization–London Convention Annual Ocean Dumping Reports Robert E. Dickinson, (from The Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003), Overview: The Climate System Priyanka K. Dissanayake, Moratuwa, Sri Lanka, Estuarian Waters Iver W. Duedall, Florida Institute of Technology, Melbourne, Florida, An Analysis of the International Maritime Organization–London Convention Annual Ocean Dumping Reports Sandra Dunbar, Napier University, Edinburgh, Scotland, United Kingdom, Partitioning and Bioavailability ´ Anton´ın Dvoˇrak, University of Economics, Prague, Czech Republic, Negotiating between Authority and Polluters: An Approach to Managing Water Quality David R. Easterling, (from Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003), Observations of Climate and Global Change from Real-Time Measurements Theodore A. Endreny, SUNY-ESF, Syracuse, New York, Remote Sensing of Applications in Hydrology, Evaporation, Radar Use in Rainfall Measurements, Rainfall and Runoff Environment Canada, Water—Nature’s Magician Timothy Erickson, National Weather Service New Orleans/Baton Rouge Forecast Office, Slidell, Louisiana, Coastal Fog Along the Northern Gulf of Mexico Kirsten Exall, National Water Research Institute, Burlington, Ontario, Canada, Coagulation and Flocculation David L. Feldman, University of Tennessee, Knoxville, Tennessee, Water Supply Planning—Federal, Water Transfers Montserrat Filella, University of Geneva, Geneva, Switzerland, Antimony in Aquatic Systems Sylwester Furmaniak, N. Copernicus University, Torun, ´ Poland, Mechanisms of Water Adsorption on Carbons Piotr A. Gauden, N. Copernicus University, Torun, ´ Poland, The Effect of Carbon Surface Chemical Composition on the Mechanism of Phenol Adsorption from Aqueous Solutions, Mechanisms of Water Adsorption on Carbons ˜ Geophysical Fluid Dynamics Laboratory—NOAA, El Nino: The Interannual Prediction Problem, Hurricanes: Modeling Nature’s Fury, Global Climate Change Andrea K. Gerlak, Columbia University, New York, New York, Wetlands Policy in the United States: From Drainage to Restoration Michael H. Glantz, (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003), Climate and Society Global Change Research Program—U.S. Geological Survey, Flattop Mountain Snotel Snowpack: Water Year 2004 Jyotsna Goel, Centre for Fire, Explosives, and Environmental Safety, Timarpur, India, Ion Exchange and Inorganic Adsorption Kelly Goodwin, University of Miami, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific David M. Gray, Mettler-Toledo Thornton Inc., Bedford, Massachusetts, Conductivity-Electric Great Lakes Environmental Research Laboratory (NOAA), Technology Development: Hardware Development—Marine Instrumentation Laboratory (MIL) Seth I. Gutman, NOAA Forecast Systems Laboratory, Ground-Based GPS Meteorology at FSL Janusz Guziur, University of Warmia and Mazury in Olsztyn, Olsztyn, Poland, History of Pond Fisheries in Poland David Haley, Picacadilly, Ulverston, Cumbria, United Kingdom, Water and Well-Being, Evolution

M. Abad, Universidad de Huelva. Avda. de las Fuerzas Armadas, Huelva, A Statistical Approach to Critical Storm Period Analysis Edinara Adelaide Boss, Artur-Nogueira-SP, Brazil, Sublimation Joseph H. Aldstadt, III, Genetic Technologies, Inc. Testing Institute, Waukesha, Wisconsin, In Situ Chemical Monitoring Alaska Biological Science Center—U.S. Geological Survey, Oceanographic Environment of Glacier Bay Sergio Alonso, University of the Balearic, Palma de Mallorca, Spain, Relative Humidity Marie de Angelis, Humboldt State University, Arcata, California, Major Ions in Seawater Carolyn Ann Koh, Colorado School of Mines, Golden, Colorado, Clathrate Hydrates, Hydration Yoseph Negusse Araya, Open University, Milton Keynes, United Kingdom, Hydrosphere Ann Azadpour-Keeley, National Risk Management Research Laboratory, ORD, U.S.EPA, Ada, Oklahoma, Water and the History of Man W.D. Bach, Jr., (from The Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003), Basic Research for Military Applications Patrick L. Barry, NASA, Water on the Space Station F. Batmanghelidj, Global Health Solutions, Falls Church, Virginia, The Myth of Bad Cholesterol: Why Water Is A Better Cholesterol-Lowering Medication, Water: The Key to Natural Health And Healing, The Mistake of Waiting to Get Thirsty M. Eric Benbow, Michigan State University, East Lansing, Michigan, A History of Hawaiian Freshwater Resources Mike Bettwy, Goddard Space Flight Center, NASA, African Monsoons David Biggs, University of Washington, Seattle, Washington, Canals in the Mekong Delta: A Historical Overview from 200 C.E. to the Present Jessica Black, NOAA/CREST CCNY, New York, New York, Remote Sensing of Applications in Hydrology Peter E. Black, SUNY ESF, Syracuse, New York, Water Resource Organizations ´ Blasco, Institute of Marine Sciences of Analucia, Cadiz, Spain, Julian Marine and Estuarine Microalgal Sediment Toxicity Tests ¨ M. Bostrom, Link¨oping University, Link¨oping, Sweden, Hofmeister Effects, Dissolved Gases Kristofor R. Brye, University of Arkansas, Fayetteville, Arkansas, Nitrogen Bureau of Labor Statistics, U.S. Department of Labor, Water Transportation Occupations Bureau of Reclamation—U.S. Department of the Interior, Hoover Dam History Robert M. Burgess, U.S. Environmental Protection Agency, Narragansett, Rhode Island, Ammonia Albert J. Burky, University of Dayton, Dayton, Ohio, A History of Hawaiian Freshwater Resources James Butler, Climate Monitoring and Diagnostics Laboratory, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Bradford Butman, U.S. Geological Survey, Mapping the Sea Floor of the Historic Area Remediation Site (HARS) Offshore of New York City Pietro Chiavaccini, Universita´ di Pisa, Pisa, Italy, Shallow Water Waves, Breakwaters G.G. Clarke, University of Hawaii, Hilo, Hawaii, Climate and Water Balance on the Island of Hawaii Jay Clausen, AMEC Earth and Environmental, Inc., Westford, Massachusetts, Technetium in Water, Beryllium in Water Aldo Conti, Frascati (RM), Italy, Floating Ice, Condensation, Desertification, Dew, Distilled Water, Isotopes, Rain Forests, Dew Point Giuseppe Cortese, Alfred Wegener Institute for Polar and Marine Research (AWI), Bremerhaven, Germany, Seawater Temperature Estimates in Paleoceanography Sara Cotton, University of Miami, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific ´ L.M. Caceres, Universidad de Huelva. Avda. de las Fuerzas Armadas, Huelva, A Statistical Approach to Critical Storm Period Analysis

xi

xii

CONTRIBUTORS

Jon Hare, NOAA Center for Coastal Fisheries and Habitat Research, NOS/NMFS Cooperative Research on Coastal Fisheries and Habitats at the Beaufort Laboratory Michael J. Hayes, Climate Impacts Specialist—National Drought Mitigation Center, Drought Indices William Henderson, University of Waikato, Hamilton, New Zealand, Analysis of Aqueous Solutions Using Electrospray Ionization Mass Spectrometry (esi ms) Warwick Hillier, The Australian National University, Canberra, Australia, Isotope Exchange in Gas-Water Reactions Joseph Holden, University of Leeds, Leeds, United Kingdom, Mariotte Bottle—Use in Hydrology, Rain Simulators S. Holgate, Proudman Oceanographic Laboratory, Birkenhead, United Kingdom, The Permanent Service for Mean Sea Level Arthur M. Holst, Philadelphia Water Department, Philadelphia, Pennsylvania, Frost Damage, Droughts, Water Cycle, Monsoon, Permanent Frost, United States Weather Bureau, Waterspout, Cloud Seeding, Cyclones, Fog, Frost, Chinook, Flood Control Act of 1944, Jacob’s Well, Gordon and Franklin Rivers and the Tasmanian Wilderness World Heritage Area Kirk L. Holub, NOAA Forecast Systems Laboratory, Ground-Based GPS Meteorology at FSL Paul R. Houser, (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003), Infiltration and Soil Moisture Processes S.A. Hsu, Louisiana State University, Baton Rouge, Louisiana, Air–Sea Interaction Jason A. Hubbart, University of Idaho, Moscow, Idaho, Hydrologic Cycle, Water Resources, and Society, Hydrologic History, Problems, and Perspectives Basia Irland, University of New Mexico, Albuquerque, New Mexico, A Concise Glimpse of Water in the History of Photography Mohammed Riajul Islam, University of Idaho, Moscow, Idaho, Rain and Rocks: The Recipe for River Water Chemistry James A. Jacobs, Environmental Bio-Systems, Inc., Mill Valley, California, Fenton’s Reaction and Groundwater Remediation, Regulatory Issues and Remediation: Risk, Costs, and Benefits, Water, Bacteria, Life on Mars, and Microbial Diversity Sharad K. Jain, National Institute of Hydrology, Roorkee, India, Isohyetal Method, Hydrologic Cycle Purnima Jalihal, National Institute of Ocean Technology, Chennai, India, Renewable Energies from the Ocean S. Jevrejeva, Proudman Oceanographic Laboratory, Birkenhead, United Kingdom, The Permanent Service for Mean Sea Level Alicia Jimenez, Michigan State University, East Lansing, Michigan, Transboundary Waters in Latin America: Conflicts and Collaboration Anne Jones-Lee, G. Fred Lee & Associates, El Macero, California, Clean Water Act, Water Quality Criteria/Standards, TMDLs, and Weight-ofEvidence Approach for Regulating Water Quality Andrew Juhl, Lamont–Doherty Earth Observatory of Columbia University, Palisades, New York, Food Chain/Foodweb/Food Cycle James O. Juvik, University of Hawaii, Hilo, Hawaii, Climate and Water Balance on the Island of Hawaii K. Kadirvelu, Centre for Fire, Explosives, and Environmental Safety, Timarpur, India, Ion Exchange and Inorganic Adsorption Th.D. Karapantsios, Aristotle University, Thessaloniki, Greece, Sorption Kinetics Thomas R. Karl, (from Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003), Observations of Climate and Global Change from Real-Time Measurements Melinda R. Kassen, Trout Unlimited, Colorado Water Project, Boulder, Colorado, Legal Protection for In-stream Flow, Interface between Federal Water Quality Regulation and State Allocation of Water Quantity Gholam A. Kazemi, Shahrood University of Technology, Shahrood, Iran, Age Dating Old Groundwater, Chlorine-36 and Very Old Groundwaters, Deuterium, Freshwater, Hard Water, Heavy Water, Hydrogen Ion, Isotope Fractionation, Soft Water, Strontium Isotopes in Water and Rock, Chlorofluorocarbons (CFCs) Jack W. Keeley, Environmental Engineer, Ada, Oklahoma, Water and the History of Man Giora J. Kidron, The Hebrew University of Jerusalem, Jerusalem, Israel, Dew Deserts

Daniel King, University of Colorado, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Piotr Kowalczyk, Faculty of Science, Chiba University, Chiba, Japan, The Effect of Carbon Surface Chemical Composition on the Mechanism of Phenol Adsorption from Aqueous Solutions, Mechanisms of Water Adsorption on Carbons Upadhyayula V. K. Kumar, Choa Chu kang Ave-4, Singapore, Coastal Water Pollutants W. Kunz, University of Regensburg, Regensburg, Germany, Hofmeister Effects Frederic Lasserre, Laval University, Quebec City, Quebec, Canada, Great Lakes Governors’ Agreement Marshall Lawson, Land Conservation Legal Services, LLC, Columbia, South Carolina, Quantitative GroundWater Law Jamie R. Lead, University of Birmingham, Birmingham, United Kingdom, Dissolved Organic Carbon G. Fred Lee, G. Fred Lee & Associates, El Macero, California, Clean Water Act, Water Quality Criteria/Standards, TMDLs, and Weight-of-Evidence Approach for Regulating Water Quality Maggie Lee, Santa Fe, New Mexico, Free Flowing Water: A Source of Wisdom Nai Kuang Liang, National Taiwan University, Taipei, Taiwan, Water Waves Sharon L. Lien, The Groundwater Foundation, Lincoln, Nebraska, Effective Water Education Strategies in a Nontraditional Setting Clive D. Lipchin, Ann Arbor, Michigan, Water Between Arabs and Israelis: Researching Twice-Promised Resources, Conflict and Water Use in the Middle East Eileen Loiseau, Bigelow Laboratory for Ocean Sciences, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific M.X. Loukidou, Aristotle University, Thessaloniki, Greece, Sorption Kinetics Helen H. Lou, Lamar University, Beaumont, Texas, Solubility of Hydrocarbons and Sulfur Compounds in Water, Solubility of Chemicals in Water Nancy A. Lowery, San Diego, California, United States-Mexico Border Waters: Conventions, Treaties, and Institutions ´ Institute of Marine Sciences of Analucia, Cadiz, Spain, Lu´ıs M. Lubian, Marine and Estuarine Microalgal Sediment Toxicity Tests Paolo Magni, IMC—International Marine Centre, Torregrande-Oristano, Italy, Laboratory Experiments On Bivalve Excretion Rates of Nutrients, Temporal Scaling of Benthic Nutrient Regeneration in BivalveDominated Tidal Flat, Physical and Chemical Variability of Tidal Streams, Tidally Mediated Changes in Nutrient Concentrations Babs A. Makinde-Odusola, Riverside, California, Radon in Water William E. Marks, Water Consciousness, Inc., Martha’s Vineyard, Massachusetts, Benjamin Franklin: From Kite to Lightning Rod, Benjamin Franklin’s Armonica: A Water Music Instrument, Our Evolving Water Consciousness, Water Clocks, Ben Franklin’s Gulf Stream Weather and Swim Fins D.L. Marrin, Hanalei, Hawaii, Cosmic Water, Molecular Network Dynamics, Sound in Water, Water Symbolism K.A. Matis, Aristotle University, Thessaloniki, Greece, Sorption Kinetics Patricia Matrai, Bigelow Laboratory for Ocean Sciences, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Mollie D. McIntosh, Michigan State University, East Lansing, Michigan, A History of Hawaiian Freshwater Resources Kevin S. McLeary, Physical Properties Lucius O. Mendis, Colombo, Sri Lanka, Ancient Water and Soil Conservation Ecosystems of Sri Lanka Carlos D. Messina, University of Florida, Gainesville, Florida, Degree Day Method Paulette Middleton, (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003), Acid Rain and Society Elsie F. Millano, ERM, Inc., Vernon Hills, Illinois, Great Lakes Water Quality Initiative A. Trent Millet, Newark, Vermont, The Medicinal Properties of the Waters of Saratoga Springs Koichi Miyashita, Okayama University, Kurashiki, Japan, Transpiration

CONTRIBUTORS Eric P. Mollard, Institut de Recherche pour le D´eveloppement, France, Curious Uses of Agricultural Water in the World, Jaubert de Passa: The First World History of Irrigation in 1846 Shigeru Montani, Hokkaido University, Hakodate, Japan, Physical and Chemical Variability of Tidal Streams, Laboratory Experiments On Bivalve Excretion Rates of Nutrients, Tidally Mediated Changes in Nutrient Concentrations, Temporal Scaling of Benthic Nutrient Regeneration in Bivalve-Dominated Tidal Flat, Seasonal Coupling Between Intertidal Macrofauna and Sediment Column Porewater Nutrient Concentrations Robert M. Moore, Dalhousie University, Halifax, Nova Scotia, Canada, Marine Sources of Halocarbons Ignatio Moreno-Garrido, Institute of Marine Sciences of Analucia, Cadiz, Spain, Marine and Estuarine Microalgal Sediment Toxicity Tests Suparna Mukherji, IIT Bombay, Powai, Mumbai, India, Adsorption of Organic Compounds ˜ J.M. Munoz, Universidad de Sevilla. Avda. Reina Mercedes, Sevilla, A Statistical Approach to Critical Storm Period Analysis Prasad K. Narasimhan, Lamar University, Beaumont, Texas, Adsorption Capacity of Activated Carbon for Water Purification, Solubility of Hydrocarbons and Sulfur Compounds in Water, Solubility of Chemicals in Water, Solubility of Hydrocarbons in Salt Water NASA—Goddard Space Flight Center, Black Water Turns the Tide on Florida Coral, The Water Cycle, Weather Forecasting Through The Ages, The Earth Observing System: Aqua NASA—Langley Research Center, CERES: Understanding The Earth’s Clouds and Climate NASA Marshall Space Flight Center, Where Water Floats National Drought Mitigation Center, What is Climatology? National Oceanographic and Atmospheric Administration (NOAA), NOAA’s Atlantic Oceanographic and Meteorological Laboratory Natural Resources Conservation Service, Conservation and the Water Cycle John W. Nielsen-Gammon, (from Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003), Overview of Weather Systems Robert Y. Ning, King Lee Technologies, San Diego, California, Silica in Natural Waters, Carbonate in Natural Waters B.W. Ninham, University of Regensburg, Regensburg, Germany and Australian National University, Canberra, Australia, Hofmeister Effects, Dissolved Gases NOAA Coral Reef Information System (CORIS), Deep Water Corals NOAA National Ocean Service, NOS Sanctuaries Protect Nation’s Maritime History NOAA—Pacific Marine Environmental Laboratory, Pacific Marine Environmental Laboratory—30 Years of Observing The Ocean Northeast Fisheries Science Center—NOAA, Woods Hole: The Early Years ¨ Gertrud K. Nurnberg, Freshwater Research, Baysville, Ontario, Canada, Quantification of Anoxia and Hypoxia in Water Bodies Pacific Fisheries Environmental Laboratory, NOAA, Physical Oceanography Pacific Northwest National Laboratory—Shrub-Steppe Ecology Series, What About Meteorology? Stefano Pagliara, Universita´ di Pisa, Pisa, Italy, Shallow Water Waves, Breakwaters Magni Paolo, IMC-International Marine Centre, Torregrande-Oristano, Italy, Seasonal Coupling Between Intertidal Macrofauna and Sediment Column Porewater Nutrient Concentrations Jose O. Payero, University of Nebraska-Lincoln, North Platte, Nebraska, Evapotranspiration Mauricio Peredo, NASA Goddard Space Flight Center, Electricity as a Fluid Naraine Persaud, Virginia Polytechnic Institute and State University, Blacksburg, Virginia, Adiabatic Cooling, Humidity—Absolute, Heat of Vaporization, Vapor Pressure Geoff Petts, University of Birmingham, Birmingham, United Kingdom, Water in History Tony Phillips, NASA, Water on the Space Station Jim Philp, Napier University, Edinburgh, Scotland, United Kingdom, Partitioning and Bioavailability

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Laurel Phoenix, Green Bay, Wisconsin, Water as a Human Right Physics Laboratory, National Institute of Standards and Technology, Early Clocks Ralph W. Pike, Louisiana State University, Baton Rouge, Louisiana, Solubility of Hydrocarbons and Sulfur Compounds in Water, Solubility of Chemicals in Water R. Pino, Universidad de Sevilla. Avda. Reina Mercedes, Sevilla, A Statistical Approach to Critical Storm Period Analysis Richard Z. Poore, U.S. Geological Survey, Reston, Virginia, Sea Level and Climate Bobby J. Presley, Texas A&M University, College Station, Texas, Trace Element Pollution Nitish Priyadarshi, Ranchi University, Ranchi, Jharkhand, India, Sodium in Natural Waters Jian-Wen Qiu, Hong Kong Baptist University, Kowloon Tong, Kowloon, Hong Kong, Larvae and Small Species of Polychaetes in Marine Toxicological Testing Climent Ramis, University of the Balearic, Palma de Mallorca, Spain, Relative Humidity Jorge Ram´ırez-Vallejo, Harvard University, Cambridge, Massachusetts, Economic Value of Water: Estimation Todd Rasmussen, The University of Georgia, Athens, Georgia, Barometric Efficiency Howard Reminick, Ohno Institute on Water and Health, Willoughby, Ohio, An Analysis of The Impact of Water on Health and Aging: Is All Water The Same? Martin Reuss, Ph.D, Office of History Headquarters, U.S. Army Corps of Engineers, The Constitution and Early Attempts at Rational Water Planning J. Rodr´ıguez Vidal, Universidad de Huelva. Avda. de las Fuerzas Armadas, Huelva, A Statistical Approach to Critical Storm Period Analysis A. Rodr´ıguez-Ram´ırez, Universidad de Huelva. Avda. de las Fuerzas Armadas, Huelva, A Statistical Approach to Critical Storm Period Analysis Romualdo Romero, University of the Balearic, Palma de Mallorca, Spain, Relative Humidity William R. Roy, Illinois State Geological Survey, Champaign, Illinois, Iron F. Ruiz, Universidad de Huelva. Avda. de las Fuerzas Armadas, Huelva, A Statistical Approach to Critical Storm Period Analysis David L. Russell, Global Environmental Operations, Inc., Lilburn, Georgia, A Brief History of the Water Pollution Control Act in the U.S. Timothy J. Ryan, Ohio University, Athens, Ohio, The Clean Water Act Gerhard Rychlicki, N. Copernicus University, Torun, ´ Poland, The Effect of Carbon Surface Chemical Composition on the Mechanism of Phenol Adsorption from Aqueous Solutions, Mechanisms of Water Adsorption on Carbons S. Sabri, University of Malaya, Petaling Jaya Selangor, Malaysia, Islamic Water Law Basu Saha, Loughborough University, Loughborough, United Kingdom, Removal of Organic Micropollutants and Metal Ions from Aqueous Solutions by Activated Carbons Eric Saltzman, University of Miami, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Peter H. Santschi, Texas A&M University, Galveston, Texas, Marine Colloids Edward S. Sarachik, (from The Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003), The Ocean in Climate Eva Saroch, Panjab University, Chandigarh, India, Representing Geopolitics of (Hydro) Borders in South Asia ˇ Petr Sauer, University of Economics, Prague, Czech Republic, Negotiating between Authority and Polluters: An Approach to Managing Water Quality Donald Savage, NASA Headquarters, Washington, Mars Exploration Rover Mission Reuel Shinnar, The City College of the CUNY, New York, The Mirage of The H2 Economy

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CONTRIBUTORS

Melissa A. Singer Pressman, Genetic Technologies, Inc. Testing Institute, Waukesha, Wisconsin, In Situ Chemical Monitoring Pratap Singh, National Institute of Hydrology, Roorkee, India, Snow Density, Snow and Snowmelt, Snow Surveys Vijay P. Singh, Louisiana State University, Baton Rouge, Louisiana, Entropy Theory For Hydrologic Modeling, Unit Hydrograph Theory, Hydrologic Cycle, Isohyetal Method D.C. Singleton, University of Hawaii, Hilo, Hawaii, Climate and Water Balance on the Island of Hawaii Bellie Sivakumar, University of California, Davis, California, Hydropsychology E. Dendy Sloan, Colorado School of Mines, Golden, Colorado, Clathrate Hydrates James A. Smith, (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003), Rainfall Fiona Stainsby, Napier University, Edinburgh, Scotland, United Kingdom, Partitioning and Bioavailability Susan-Marie Stedman, Silver Spring, Maryland, Coastal Waters Kenneth F. Steele, University of Arkansas, Fayetteville, Arkansas, Nitrogen, Carbonate Geochemistry David P. Stern, Goddard Space Flight Center, NASA, Electricity as a Fluid, Weather and the Atmosphere Bradley A. Striebig, Gonzaga University, Spokane, Washington, Chemical Precipitation Georgina Sturrock, Commonwealth Scientific and Industrial Research Organization—Australia, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Artur P. Terzyk, N. Copernicus University, Torun, ´ Poland, The Effect of Carbon Surface Chemical Composition on the Mechanism of Phenol Adsorption from Aqueous Solutions, Mechanisms of Water Adsorption on Carbons Ryszard Tokarczyk, University of Miami, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Marta E. Torres, Oregon State University, Corvallis, Oregon, Distribution and Dynamics of Gas Hydrates in the Marine Environment Christopher Tracey, U.S. Geological Survey, Reston, Virginia, Sea Level and Climate Anne M. Trehu, Oregon State University, Corvallis, Oregon, Distribution and Dynamics of Gas Hydrates in the Marine Environment Robert C. Upstill-Goddard, University of Newcastle upon Tyne, Newcastle upon Tyne, United Kingdom, The Role of Oceans in the Global Cycles of Climatically-Active Trace-Gases David B. Vance, ARCADIS, Midland, Texas, Regulatory Issues and Remediation: Risk, Costs, and Benefits, Water, Bacteria, Life on Mars, and Microbial Diversity Javier Velez-Arocho, U.S. Environmental Protection Agency, Washington, District of Columbia, Marine Debris Abatement Roger C. Viadero Jr., West Virginia University, Morgantown, West Virginia, Henry’s Law

Rolf R. von Oppenfeld, The TESTLaw Practice Group, Phoenix, Arizona, The National Pollution Discharge Elimination System, The Safe Drinking Water Act Linda Voss, U.S. Centennial of Flight Commission, Ballooning and Meteorology in the Twentieth Century U.S. Department of Labor, Bureau of Labor Statistics, Atmospheric Scientists U.S. Environmental Protection Agency—Oceans and Coastal Protection Division, Coral Reefs and Your Coastal Watershed U.S. Geological Survey, Water Quality U.S. Global Change Research Program, The Global Water Cycle Guy Webster, Jet Propulsion Laboratory, Pasadena, California, Mars Exploration Rover Mission Radosław P. Wesołowski, N. Copernicus University, Torun, ´ Poland, Mechanisms of Water Adsorption on Carbons Scott Whiteford, Michigan State University, East Lansing, Michigan, Transboundary Waters in Latin America: Conflicts and Collaboration Patrick Willems, Hydraulics Laboratory, Leuven, Belgium, Uncertainties in Rainfall–Runoff Modeling Richard S. Williams, Jr., U.S. Geological Survey, Woods Hole, Massachusetts, Sea Level and Climate Marek Wi´sniewski, N. Copernicus University, Torun, ´ Poland, Mechanisms of Water Adsorption on Carbons Ming Hung Wong, Hong Kong Baptist University, Kowloon Tong, Kowloon, Hong Kong, Larvae and Small Species of Polychaetes in Marine Toxicological Testing Eve Woods, Denver, Colorado, Reserved Water Rights for Indian and Federal Lands P.L. Woodworth, Proudman Oceanographic Laboratory, Birkenhead, United Kingdom, The Permanent Service for Mean Sea Level Tom Wydrzynski, The Australian National University, Canberra, Australia, Isotope Exchange in Gas-Water Reactions Carl L. Yaws, Lamar University, Beaumont, Texas, Adsorption Capacity of Activated Carbon for Water Purification, Solubility of Hydrocarbons and Sulfur Compounds in Water, Solubility of Chemicals in Water, Solubility of Hydrocarbons in Salt Water Brian Yocis, Bigelow Laboratory for Ocean Sciences, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific David W. Yoskowitz, Texas A&M University—Corpus Christi, Laredo, Texas, U.S./Canadian Boundary Waters Treaty and the Great Lakes Water Quality Agreement Mark A. Young, University of Iowa, Iowa City, Iowa, Environmental Photochemistry in Surface Waters Shari Yvon-Lewis, Atlantic Oceanographic and Meteorological Laboratory, NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific Karl Erik Zachariassen, Norwegian University of Science and Technology, Trondheim, Norway, Freezing and Supercooling of Water, OsmosisDiffusion of Solvent or Caused by Diffusion of Solutes? A. Zaharuddin, University of Malaya, Petaling Jaya Selangor, Malaysia, Islamic Water Law

CONTRIBUTORS Flow, Combined Free and Porous Flow in the Subsurface, Modeling Techniques for Solute Transport in Groundwater Rupali Datta, University of Texas at San Antonio, San Antonio, Texas, Remediation of Contaminated Soils, Genetics of Metal Tolerance and Accumulation in Higher Plants, Phytoextraction of Zinc and Cadmium from Soils Using Hyperaccumulator Plants, Phytoremediation of Selenium-Laden Soils, Phytoextraction and Phytostabilization: Technical, Economic and Regulatory Considerations of the Soil-Lead Issue Ali H. Davani, University of Texas at San Antonio, San Antonio, Texas, Remediation of Contaminated Soils L.C. Davis, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Phytoremediation of Methyl Tertiary-Butyl Ether Melissa R. Dawe, University of New Brunswick, Fredericton, New Brunswick, Canada, River-Connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry Steven A. Dielman, ENVIRON International Corporation, Arlington, Virginia, Hydraulic Conductivity/Transmissibility Craig E. Divine, Colorado School of Mines, Golden, Colorado, Groundwater Sampling with Passive Diffusion Samplers, Detecting Modern Groundwaters with 85 Kr, Groundwater Dating with H–He Shonel Dwyer, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Perchlorate: Chemical Behavior and Treatment Aly I. El-Kadi, University of Hawaii at Manoa, Honolulu, Hawaii, Unconfined Groundwater Environment Canada, Groundwater—Nature’s Hidden Treasure L.E. Erickson, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Phytoremediation of Methyl Tertiary-Butyl Ether Thomas R. Fisher, Horn Point Laboratory—UMCES, Solomons, Maryland, What is a Hydrochemical Model? Craig Foreman, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Cadmium: Chemical Behavior and Treatment Devin L. Galloway, U.S. Geological Survey, Denver, Colorado, Earthquakes—Rattling the Earth’s Plumbing System Lorraine Geddes-McDonald, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Nitrate: Chemical Behavior and Treatment ´ Mario Abel Goncalves, ¸ Faculdade de Ciˆencias da Universidade de Lisoba, Lisoba, Portugal, Metal Organic Interactions in Subtitle D Landfill Leachates and Associated Ground Waters, Geochemical ModelingComputer Codes, Geochemical Models Jason J. Gurdak, U.S. Geological Survey, Lakewood, Colorado and Colorado School of Mines, Golden, Colorado, Groundwater Vulnerability to Pesticides: Statistical Approaches Navraj S. Hanspal, Loughborough University, Loughborough, United Kingdom, Modeling Techniques for Solute Transport in Groundwater, Viscous Flow, Laminar Flow, Finite Element Modeling of Coupled Free and Porous Flow Thomas Harter, University of California, Davis, California, Specific Yield Storage Equation, Vulnerability Mapping of Groundwater Resources, Aquifers Blayne Hartman, H&P Mobile Geochemistry, Solana Beach, California, Applications of Soil Vapor Data to Groundwater Investigations Joseph Holden, University of Leeds, Leeds, United Kingdom, Infiltrometers, Soil Pipes and Pipe Flow, Infiltration and Soil Water Processes, Darcy’s Law, Infiltration/Capacity/Rates Ekkehard Holzbecher, Humboldt Universitat ¨ Berlin, Berlin, Germany, Groundwater Modeling, Ghijben–Herzberg Equilibrium Paul F. Hudak, University of North Texas, Denton, Texas, Mass Transport in Saturated Media John D. Humphrey, Colorado School of Mines, Golden, Colorado, Groundwater Dating with H–He S.L. Hutchinson, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Hydrologic Feasibility Assessment and Design in Phytoremediation Th.A. Ioannidis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Phytoremediation of Lead-Contaminated Soils

Segun Adelana, University of Ilorin, Ilorin, Nigeria, Summary of Isotopes in Contaminant Hydrogeology, Environmental Isotopes in Hydrogeology Mohammad N. Almasri, An-Najah National University, Nablus, Palestine, Groundwater Flow and Transport Process Tom A. Al, University of New Brunswick, Fredericton, New Brunswick, Canada, River-Connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry Larry Amskold, University of New Brunswick, Fredericton, New Brunswick, Canada, River-Connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry Ann Azadpour-Keeley, National Risk Management Research Laboratory, ORD, U.S. EPA, Ada, Oklahoma, Microbial Processes Affecting Monitored Natural Attenuation of Contaminants in the Subsurface, Nitrate Contamination of Groundwater Mukand Singh Babel, Asian Institute of Technology, Pathumthani, Thailand, Groundwater Velocities, Groundwater Flow Properties, Water in The Unsaturated Zone Philip B. Bedient, Rice University, Houston, Texas, Transport of Reactive Solute in Soil and Groundwater David M. Bednar, Jr., Michael Baker, Jr. Inc., Shreveport, Louisiana, Karst Hydrology, Karst Topography, Groundwater Dye Tracing in Karst Milovan Beljin, Cincinnati, Ohio, Horizontal Wells Craig H. Benson, University of Wisconsin-Madison, Madison, Wisconsin, Reactive Transport in The Saturated Zone: Case Histories for Permeable Reactive Barriers Robert A. Bisson, Alexandria, Virginia, Megawatersheds William J. Blanford, Louisiana State University, Baton Rouge, Louisiana, Vadose Zone Monitoring Techniques Thomas B. Boving, University of Rhode Island, Kingston, Rhode Island, Organic Compounds in Ground Water, Innovative Contaminated Groundwater Remediation Technologies Richard C. Brody, UC Berkeley, Berkeley, California, Connate Water Kristofor R. Brye, University of Arkansas, Fayetteville, Arkansas, Lysimeters, Soil and Water Contamination by Heavy Metals Gunnar Buckau, Institut fur ¨ Nukleare Entsorgung, Karlsruhe, Germany, Mobility of Humic Substances in Groundwater Bureau of Indian Affairs and Arizona Department of Water Resources—U.S. Geological Survey, Black Mesa Monitoring Program Karl E. Butler, University of New Brunswick, Fredericton, New Brunswick, Canada, River-Connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry Herbert T. Buxton, United States Geological Survey, Pharmaceuticals, Hormones, and Other Organic Wastewater Contaminants in U.S. Streams Natalie L. Capiro, Rice University, Houston, Texas, Transport of Reactive Solute in Soil and Groundwater Harendra S. Chauhan, G.B. Pant University of Agriculture and Technology, Uttar Pradesh, India, Steady-State Flow Aquifer Tests, Subsurface Drainage Bernard L. Cohen, University of Pittsburgh, Pittsburgh, Pennsylvania, Risk Analysis of Buried Wastes From Electricity Generation David P. Commander, Water and Rivers Commission, East Perth, Australia, Water Dowsing (Witching), Artesian Water Dennis L. Corwin, USDA-ARS George E. Brown, Jr., Salinity Laboratory, Riverside, California, Characterizing Soil Spatial Variability, Modeling Non-Point Source Pollutants in the Vadose Zone Using GIS, Groundwater Vulnerability to Pesticides: An Overview of Approaches and Methods of Evaluation Colin C. Cunningham, The University of Edinburgh, Edinburgh, Scotland, United Kingdom, In Situ Bioremediation of Contaminated Groundwater William L. Cunningham, U.S. Geological Survey, Denver, Colorado, Earthquakes—Rattling the Earth’s Plumbing System Uwe Dannwolf, URS Australia Pty Ltd., Turner, Australia, Groundwater and Vadose Zone Hydrology Diganta Bhusan Das, Oxford University, Oxford, United Kingdom, Viscous Flow, Finite Element Modeling of Coupled Free and Porous

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CONTRIBUTORS

Jahangir Islam, City Design Limited, Auckland, New Zealand, SubSurface Redox Chemistry: A Comparison of Equilibrium and ReactionBased Approaches Irena B. Ivshina, Institute of Ecology and Genetics of Microorganisms of the RAS, Perm, Russia, In Situ Bioremediation of Contaminated Groundwater James A. Jacobs, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Cobalt: Chemical Behavior and Treatment, Limiting Geochemical Factors in Remediation Using Monitored Natural Attenuation and Enhanced Bioremediation, The Role of Heat in Groundwater Systems, Horizontal Wells in Groundwater Remediation, Groundwater Flow in Heterogenetic Sediments and Fractured Rock Systems, Groundwater and Cadmium: Chemical Behavior and Treatment, Groundwater and Benzene: Chemical Behavior and Treatment, Groundwater and Lead: Chemical Behavior and Treatment, Groundwater and Nitrate: Chemical Behavior and Treatment, Groundwater and Uranium: Chemical Behavior and Treatment, Groundwater and Mercury: Chemical Behavior and Treatment, The Environmental Impact of Iron in Groundwater, Water Well Drilling Techniques, Water-Jetting Drilling Technologies for Well Installation And In Situ Remediation of Hydrocarbons, Solvents, and Metals, Source, Mobility, and Remediation of Metals, Particulate Transport in Groundwater—Bacteria and Colloids, Groundwater and Arsenic: Chemical Behavior and Treatment, In Situ Groundwater Remediation for Heavy Metal Contamination, Groundwater Remediation by In Situ Aeration and Volatilization, MTBE, Phytoremediation Enhancement of Natural Attenuation Processes, Groundwater Remediation by Injection and Problem Prevention, Chemical Oxidation Technologies for Groundwater Remediation, Physical Properties of DNAPLs and Groundwater Contamination, Process Limitations of In Situ Bioremediation of Groundwater, Water Contamination by Low Level Organic Waste Compounds in the Hydrologic System, Applications of Soil Vapor Data to Groundwater Investigations, Groundwater Remediation Project Life Cycle, Groundwater Remediation: In Situ Passive Methods, Groundwater and Vinyl Chloride: Chemical Behavior and Treatment, Groundwater and Perchlorate: Chemical Behavior and Treatment, Groundwater Sampling Techniques for Environmental Projects, Low Flow Groundwater Purging and Surging Hamid R. Jahani, Water Research Institute, Hakimieh, Tehran, Iran, Groundwater Tracing, Resistivity Methods Chakresh K. Jain, National Institute of Hydrology, Roorkee, India, Assessment of Groundwater Quality in District Hardwar, Uttaranchal, India, Nonpoint Sources, Fluoride Contamination in Ground Water, Irrigation Water Quality in District Hardwar, Uttaranchal, India John R. Jansen, Aquifer Science & Technology, Waukesha, Wisconsin, Geophysics and Remote Sensing Anthea Johnson, University of Auckland, Auckland, New Zealand, Bacteria Role in the Phytoremediation of Heavy Metals Silvia Johnson, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Mercury: Chemical Behavior and Treatment Tracey Johnston, University of Texas at San Antonio, San Antonio, Texas, Phytoextraction and Phytostabilization: Technical, Economic and Regulatory Considerations of the Soil-Lead Issue Jagath J. Kaluarachchi, Utah State University, Logan, Utah, Groundwater Flow and Transport Process A. Katsoyiannis, Aristotle University of Thessaloniki, Thessaloniki, Greece, The Use of Semipermeable Membrane Devices (SPMDs) for Monitoring, Exposure, and Toxicity Assessment Jack Keeley, Environmental Engineer, Ada, Oklahoma, Nitrate Contamination of Groundwater David W. Kelley, University of St. Thomas, St. Paul, Minnesota, Leaching Lisa Kirkland, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Lead: Chemical Behavior and Treatment Dana W. Kolpin, United States Geological Survey, Pharmaceuticals, Hormones, and Other Organic Wastewater Contaminants in U.S. Streams C.P. Kumar, National Institute of Hydrology, Roorkee, India, Groundwater Balance Maria S. Kuyukina, Institute of Ecology and Genetics of Microorganisms of the RAS, Perm, Russia, In Situ Bioremediation of Contaminated Groundwater Kung-Yao Lee, Horn Point Laboratory—UMCES, Solomons, Maryland, What is a Hydrochemical Model?

Leo S. Leonhart, Hargis Associates, Inc., Tucson, Arizona, Recharge in Arid Regions, Perched Groundwater Scott M. Lesch, USDA-ARS George E. Brown, Jr., Salinity Laboratory, Riverside, California, Characterizing Soil Spatial Variability Len Li, University of Wisconsin-Madison, Madison, Wisconsin, Reactive Transport in The Saturated Zone: Case Histories for Permeable Reactive Barriers Keith Loague, Stanford University, Stanford, California, Groundwater Vulnerability to Pesticides: An Overview of Approaches and Methods of Evaluation, Modeling Non-Point Source Pollutants in the Vadose Zone Using GIS Walter W. Loo, Environmental & Technology Services, Oakland, California, Treatment for Nitrates in Groundwater, Treatment of Arsenic, Chromium, and Biofouling in Water Supply Wells, In Situ Electrokinetic Treatment of MtBE, Benzene, and Chlorinated Solvents, Hydraulic Properties Characterization Kerry T. Macquarrie, University of New Brunswick, Fredericton, New Brunswick, Canada, River-Connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry Mini Mathew, Colorado School of Mines, Golden, Colorado, Modeling of DNAPL Migration in Saturated Porous Media S.C. Mccutcheon, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Hydrologic Feasibility Assessment and Design in Phytoremediation John E. McCray, Colorado School of Mines, Golden, Colorado, Groundwater Vulnerability to Pesticides: Statistical Approaches M.S. Mohan Kumar, Indian Institute of Science, Bangalore, India, Modeling of DNAPL Migration in Saturated Porous Media John E. Moore, USGS (Retired), Denver, Colorado, Well Hydraulics and Aquifer Tests, Drawdown, Groundwater Quality, Hot Springs, Overdraft, Saline Seep, Geological Occurrence of Groundwater Angela Munroe, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Vinyl Chloride: Chemical Behavior and Treatment Jean-Christophe Nadeau, University of New Brunswick, Fredericton, New Brunswick, Canada, River-Connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry NASA Earth Science Enterprise Data and Services, Squeezing Water from Rock Vahid Nassehi, Loughborough University, Loughborough, United Kingdom, Combined Free and Porous Flow in the Subsurface, Viscous Flow Sascha E. Oswald, UFZ Centre for Environmental Research, LeipzigHalle, Germany, Modeling Contaminant Transport and Biodegradation in Groundwater Timothy K. Parker, Groundwater Resources of California, Sacramento, California, Water Contamination by Low Level Organic Waste Compounds in the Hydrologic System Jim C. Philp, Napier University, Edinburgh, Scotland, United Kingdom, In Situ Bioremediation of Contaminated Groundwater Laurel Phoenix, Green Bay, Wisconsin, Fossil Aquifers Nitish Priyadarshi, Ranchi University, Ranchi, Jharkhand, India, Geothermal Water, Rock Fracture, Consolidated Water Bearing Rocks, Groundwater Contamination from Runoff, Groundwater Dating with Radiocarbon, Methane in Groundwater, Permeability S.N. Rai, National Geophysical Research Institute, Hyderabad, India, Artificial Recharge of Unconfined Aquifer Todd Rasmussen, The University of Georgia, Athens, Georgia, Head, Deep Soil-Water Movement, Soil Water, Specific Gravity, Tidal Efficiency Hugh H. Russell, CHR2 Environmental Services, Inc.,, Oilton, Oklahoma, Microbial Processes Affecting Monitored Natural Attenuation of Contaminants in the Subsurface Philip R. Rykwalder, University of Texas at San Antonio, San Antonio, Texas, Vadose Zone Monitoring Techniques Bahram Saghafian, Soil Conservation and Watershed Management Research Institute, Tehran, Iran, Qanats: An Ingenious Sustainable Groundwater Resource System C. Samara, Aristotle University of Thessaloniki, Thessaloniki, Greece, The Use of Semipermeable Membrane Devices (SPMDs) for Monitoring, Exposure, and Toxicity Assessment Dibyendu Sarkar, University of Texas at San Antonio, San Antonio, Texas, Remediation of Contaminated Soils, Genetics of Metal Tolerance and Accumulation in Higher Plants, Phytoextraction of Zinc and Cadmium from Soils Using Hyperaccumulator Plants, Phytoremediation

CONTRIBUTORS of Selenium-Laden Soils, Phytoextraction and Phytostabilization: Technical, Economic and Regulatory Considerations of the Soil-Lead Issue J.L. Schnoor, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Phytoremediation of Methyl Tertiary-Butyl Ether Guy W. Sewell, National Risk Management Research Laboratory, ORD, U.S. EPA, Ada, Oklahoma (formerly with Dynamac Corporation), Microbial Processes Affecting Monitored Natural Attenuation of Contaminants in the Subsurface Raj Sharma, University of KwaZulu-Natal, Durban, South Africa, Laminar Flow Caijun Shi, CJS Technology, Inc., Burlington, Ontario, Canada, High pH Groundwater—The Effect of The Dissolution of Hardened Cement Pastes Naresh Singhal, University of Auckland, Auckland, New Zealand, Bacteria Role in the Phytoremediation of Heavy Metals, Sub-Surface Redox Chemistry: A Comparison of Equilibrium and Reaction-Based Approaches V.P. Singh, Louisiana State University, Baton Rouge, Louisiana, Artificial Recharge of Unconfined Aquifer Joseph Skopp, University of Nebraska, Lincoln, Nebraska, Field Capacity Jeffrey G. Skousen, West Virginia University, Morgantown, West Virginia, Acid Mine Drainage: Sources and Treatment in the United States Ricardo Smalling, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Uranium: Chemical Behavior and Treatment James A. Smith, University of Virginia, Charlottesville, Virginia, Vapor Transport in the Unsaturated Zone Stuart A. Smith, Smith-Comeskey GroundWater Science LLC, Upper Sandusky, Ohio, Well Maintenance, Biofouling in Water Wells, Soil and Groundwater Geochemistry and Microbiology Michelle Sneed, U.S. Geological Survey, Denver, Colorado, Earthquakes—Rattling the Earth’s Plumbing System Roger Spence, Oak Ridge National Laboratory, Oak Ridge, Tennessee, High pH Groundwater—The Effect of The Dissolution of Hardened Cement Pastes Kenneth F. Steele, University of Arkansas, Fayetteville, Arkansas, Soil and Water Contamination by Heavy Metals Mark D. Steele, MDC Systems, Inc., Berwyn, Pennsylvania, Water Level Drawdown P. Takis Elefsiniotis, University of Auckland, Auckland, New Zealand, Bacteria Role in the Phytoremediation of Heavy Metals Henry Teng, The George Washington University, Washington, DC, Water/Rocks Interaction Stephen M. Testa, Mokelumne Hill, California, Dating Groundwaters with Tritium, Brine Deposits Geoffrey Thyne, Colorado School of Mines, Golden, Colorado, Detecting Modern Groundwaters with 85 Kr, Geochemical Modeling—Computer Code Concepts Fred D. Tillman, U.S. Environmental Protection Agency, Athens, Georgia, Vapor Transport in the Unsaturated Zone

xiii

David J. Tonjes, Cashin Associates PC, Hauppauge, New York, Groundwater Contamination From Municipal Landfills in the USA Douglas C. Towne, Phoenix, Arizona, Ambient Groundwater Monitoring Network Strategies and Design Michael D. Trojan, Minnesota Pollution Control Agency, St. Paul, Minnesota, Land Use Impacts on Groundwater Quality, Sensitivity of Groundwater to Contamination Kristine Uhlman, University of Arizona, Tucson, Arizona, Recharge in Desert Regions Around The World Matthew M. Uliana, Texas State University—San Marcos, San Marcos, Texas, Regional Flow Systems, Hydraulic Head, Storage Coefficient David B. Vance, ARCADIS G&M, Inc., Midland, Texas, Groundwater Remediation by In Situ Aeration and Volatilization, Source, Mobility, and Remediation of Metals, Particulate Transport in Groundwater—Bacteria and Colloids, The Environmental Impact of Iron in Groundwater, Groundwater Remediation by Injection and Problem Prevention, Chemical Oxidation Technologies for Groundwater Remediation, Physical Properties of DNAPLs and Groundwater Contamination, Process Limitations of In Situ Bioremediation of Groundwater, Phytoremediation Enhancement of Natural Attenuation Processes, Groundwater and Arsenic: Chemical Behavior and Treatment, Low Flow Groundwater Purging and Surging, The Role of Heat in Groundwater Systems, Horizontal Wells in Groundwater Remediation, Groundwater Flow in Heterogenetic Sediments and Fractured Rock Systems, Limiting Geochemical Factors in Remediation Using Monitored Natural Attenuation and Enhanced Bioremediation Keith Villiers, Environmental Bio-Systems, Inc., Mill Valley, California, Groundwater and Benzene: Chemical Behavior and Treatment Nikolay Voutchkov, Poseidon Resources Corporation, Stamford, Connecticut, Well Design and Construction Atul N. Waghode, Loughborough University, Leicestershire, United Kingdom, Finite Element Modeling of Coupled Free and Porous Flow Roger M. Waller, U.S. Geological Survey,, Ground Water: Wells Lise Walter, Environmental Bio-Systems, Mill Valley, California, Groundwater and Cobalt: Chemical Behavior and Treatment J.W. Weaver, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Hydrologic Feasibility Assessment and Design in Phytoremediation Jason J. Wen, City of Downey, Downey, California, Treatment for Nitrates in Groundwater, Treatment of Arsenic, Chromium, and Biofouling in Water Supply Wells Dennis E. Williams, Geoscience Support Services, Claremont, California, Well TEST, Radial Wells, Well Screens Eric S. Wilson, E. L. Montgomery & Associates, Inc., Tucson, Arizona, Safe Yield of an Aquifer, Specific Capacity S.K. Winnike-McMillan, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Phytoremediation of Methyl Tertiary-Butyl Ether Q. Zhang, (from Phytoremediation: Transformation and Control of Contaminants, Wiley 2003), Phytoremediation of Methyl Tertiary-Butyl Ether A.I. Zouboulis, Aristotle University of Thessaloniki, Thessaloniki, Greece, Phytoremediation of Lead-Contaminated Soils

Contents

Prefaces Preface (Volume 1) ...............................................................................................................

xi

Preface (Volume 2) ...............................................................................................................

xii

Preface (Volume 3) ...............................................................................................................

xiii

Preface (Volume 4) ...............................................................................................................

xiv

Preface (Volume 5) ...............................................................................................................

xv

Acknowledgments ...............................................................................................................

xvi

Contributors Contributors (Volume 1) ........................................................................................................

xvii

Contributors (Volume 2) ........................................................................................................

xxi

Contributors (Volume 3) ........................................................................................................

xxv

Contributors (Volume 4) ........................................................................................................

xxix

Contributors (Volume 5) ........................................................................................................

xxxiii

Volume 1. Domestic, Municipal, and Industrial Water Supply and Waste Disposal 1.1

Domestic Water Supply .........................................................................................................

1:1

1.1.1

The Arsenic Drinking Water Crisis in Bangladesh ...............................................

1:1

1.1.2

Bottled Water .......................................................................................................

1:3

1.1.3

Corrosion Control in Drinking Water Systems .....................................................

1:5

1.1.4

Economics of Residential Water Demands ..........................................................

1:12

1.1.5

Gray Water Reuse in Households .......................................................................

1:16

1.1.6

Water and Human Health ....................................................................................

1:19

1.1.7

Nitrate Health Effects ...........................................................................................

1:30

1.1.8

Domestic Water Supply – Public-private Partnership ..........................................

1:42

1.1.9

Methods of Reducing Radon in Drinking Water ...................................................

1:51

1.1.10

Water Reuse ........................................................................................................

1:53

1.1.11

Roof Drainage Hydraulics ....................................................................................

1:54

1.1.12

Septic Tank Systems ...........................................................................................

1:61

1.1.13

Domestic Solar Water Heaters ............................................................................

1:63

1.1.14

Household Drinking Water Treatment and Safe Storage .....................................

1:67

1.1.15

Virus Transport in the Subsurface .......................................................................

1:70

1.1.16

Windmills .............................................................................................................

1:73

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xxxvii

xxxviii 1.2

Contents

Municipal Water Supply .........................................................................................................

1:76

1.2.1

Mixing and Agitation in Water Treatment Systems ..............................................

1:76

1.2.2

Arsenic in Natural Waters ....................................................................................

1:81

1.2.3

Evaluation of Microbial Components of Biofouling ...............................................

1:83

1.2.4

Threat Agents and Water Biosecurity ..................................................................

1:87

1.2.5

Granular Activated Carbon ..................................................................................

1:92

1.2.6

Competitive Adsorption of Several Organics and Heavy Metals on Activated Carbon in Water ................................................................................................... 1:107

1.2.7

A Real-time Hydrological Information System for Cities ...................................... 1:121

1.2.8

Chlorine and Chlorine Residuals ......................................................................... 1:127

1.2.9

Modeling Chlorine Residuals in Urban Water Distribution Systems .................... 1:131

1.2.10

Particulate Matter Removal by Coagulation ......................................................... 1:137

1.2.11

Selective Coagulant Recovery from Water Treatment Plant Residuals Using the Domain Membrane Process .......................................................................... 1:139

1.2.12

Physical Water Conditioning ................................................................................ 1:141

1.2.13

Consumer Confidence Reports ............................................................................ 1:145

1.2.14

Water Conservation Measures ............................................................................ 1:146

1.2.15

Preventing Well Contamination ............................................................................ 1:149

1.2.16

Corrosion Control ................................................................................................. 1:152

1.2.17

Cross Connection and Backflow Prevention ........................................................ 1:155

1.2.18

Molecular-based Detection of Cryptosporidium Parvum in Water ....................... 1:158

1.2.19

Cryptosporidium ................................................................................................... 1:162

1.2.20

Measuring Cryptosporidium Parvum Oocyst Inactivation Following Disinfection with Ultraviolet Light ......................................................................... 1:165

1.2.21

Dechlorination ...................................................................................................... 1:169

1.2.22

Desalination ......................................................................................................... 1:170

1.2.23

Diatomaceous Earth Filtration for Drinking Water ................................................ 1:174

1.2.24

Emerging Waterborne Infectious Diseases .......................................................... 1:177

1.2.25

Improving Waterborne Disease Surveillance ....................................................... 1:183

1.2.26

Disinfectants ........................................................................................................ 1:192

1.2.27

Disinfection .......................................................................................................... 1:196

1.2.28

Water Distribution System Operation ................................................................... 1:200

1.2.29

Water Quality in Distribution Systems .................................................................. 1:204

1.2.30

Design of Water Distribution Systems ................................................................. 1:207

1.2.31

What Is in Our Drinking Water? ........................................................................... 1:213

1.2.32

The Economics of Water Resources Allocation ................................................... 1:215

1.2.33

Answering the Challenge ..................................................................................... 1:218

1.2.34

Key Causes of Drinking Water Quality Failure in a Rural Small Water Supply of South Africa ..................................................................................................... 1:221

1.2.35

Filtration ............................................................................................................... 1:227 This page has been reformatted by Knovel to provide easier navigation.

Contents

xxxix

1.2.36

Water Filtration .................................................................................................... 1:230

1.2.37

Filtration with Granular Media .............................................................................. 1:233

1.2.38

Slow Sand Filtration and the Impact of Schmutzdecke ........................................ 1:235

1.2.39

Multistage Drinking Water Filtration ..................................................................... 1:237

1.2.40

Multistage Filtration: an Innovative Water Treatment Technology ....................... 1:238

1.2.41

Particulate Matter Removal by Filtration and Sedimentation ............................... 1:243

1.2.42

Filtration Water Treatment ................................................................................... 1:245

1.2.43

Synthetic and Natural Organic Removal by Biological Filtration .......................... 1:248

1.2.44

Granular Bed and Precoat Filtration .................................................................... 1:249

1.2.45

Flocculation .......................................................................................................... 1:252

1.2.46

Fluoridation .......................................................................................................... 1:254

1.2.47

Giardiasis ............................................................................................................. 1:257

1.2.48

Gravity Separation/Sedimentation ....................................................................... 1:259

1.2.49

Water Hammer .................................................................................................... 1:261

1.2.50

Health Effects of Commonly Occurring Disinfection Byproducts in Municipal Water Supplies .................................................................................................... 1:264

1.2.51

Health Effects of Microbial Contaminants and Biotoxins in Drinking Water ......... 1:277

1.2.52

Drinking Water and Public Health Protection ....................................................... 1:281

1.2.53

1962 U.S. Public Health Service Standards ........................................................ 1:292

1.2.54

Ion Exchange and Demineralization .................................................................... 1:297

1.2.55

The State of the Water Industry – 2004 ............................................................... 1:301

1.2.56

Iron and Manganese Removal ............................................................................. 1:312

1.2.57

Extraterritorial Land Use Control to Protect Water Supplies ................................ 1:315

1.2.58

Leak Detection and Water Loss Control .............................................................. 1:317

1.2.59

Lime-soda Ash Processes ................................................................................... 1:320

1.2.60

Lime Softening ..................................................................................................... 1:322

1.2.61

Ion Exchange – Use of Magnetic Ion Exchange Resin for DOC Removal .......... 1:325

1.2.62

Membrane Filtration ............................................................................................. 1:331

1.2.63

Water Meters ....................................................................................................... 1:337

1.2.64

Microbiological Concerns of Drinking Water Distribution Systems ...................... 1:341

1.2.65

Nitrification of Potable Water Using Trickling Filters ............................................ 1:346

1.2.66

Organic Removal ................................................................................................. 1:350

1.2.67

Ozone .................................................................................................................. 1:354

1.2.68

Ozone with Activated Carbon for Drinking Water Treatment ............................... 1:357

1.2.69

Ozone-bromide Interactions ................................................................................. 1:357

1.2.70

Municipal Water Supply: Ozonation ..................................................................... 1:362

1.2.71

Review of Parasite Fate and Transport in Karstic Aquifers ................................. 1:365

1.2.72

Particulate Removal ............................................................................................. 1:370

1.2.73

Pharmaceuticals in Water Systems ..................................................................... 1:372

1.2.74

Point-of-use/Point-of-entry Systems (POU/POE) ................................................ 1:378 This page has been reformatted by Knovel to provide easier navigation.

xl

Contents 1.2.75

Assessing the Bactericidal Efficiency of Polydex for the Disinfection of Drinking Water in Rural Areas of South Africa ..................................................... 1:382

1.2.76

Private Sector Participation, Marketing and Corporate Strategies in Municipal Water Supply and Sewerage ............................................................... 1:387

1.2.77

Pumps .................................................................................................................. 1:391

1.2.78

Radionuclides ...................................................................................................... 1:395

1.2.79

Use of Redox Potentials in Wastewater Treatment ............................................. 1:399

1.2.80

Repairing Distribution Line Breaks ....................................................................... 1:400

1.2.81

Role of Small Water Reservoirs in Environment .................................................. 1:403

1.2.82

Reservoirs, Towers, and Tanks Drinking Water Storage Facilities ...................... 1:408

1.2.83

Water Treatment Plant Residuals Management .................................................. 1:411

1.2.84

Reverse Osmosis, Process Chemistry ................................................................ 1:414

1.2.85

Reverse Osmosis, Membrane Foulants ............................................................... 1:416

1.2.86

Reverse Osmosis, Membrane Cleaning .............................................................. 1:419

1.2.87

Application of Risk Assessments in Crafting Drinking Water Regulations ........... 1:422

1.2.88

Potential Risks of Waterborne Transmission of Escherichia Coli O157:H7 ......... 1:429

1.2.89

Slow Sand Filtration ............................................................................................. 1:431

1.2.90

Approaches for Securing a Water Distribution System ........................................ 1:434

1.2.91

Water Security: an Emerging Issue ..................................................................... 1:437

1.2.92

Guide to Selection of Water Treatment Processes .............................................. 1:439

1.2.93

Source Water Assessment .................................................................................. 1:444

1.2.94

Hydraulic Design of Water Distribution Storage Tanks ........................................ 1:448

1.2.95

System Control and Data Acquisition (SCADA) ................................................... 1:449

1.2.96

Settling Tanks ...................................................................................................... 1:452

1.2.97

Treatment for Technologies for Small Drinking Water Systems .......................... 1:457

1.2.98

Ultraviolet Disinfection ......................................................................................... 1:466

1.2.99

Ultraviolet Irradiation ............................................................................................ 1:469

1.2.100

Water Disinfection Using UV Radiation – a Sri Lankan Experience .................... 1:471

1.2.101

Drinking Water Quality Standards (DWQS) – United States ............................... 1:476

1.2.102

Valves .................................................................................................................. 1:482

1.2.103

Removal of Pathogenic Bacteria, Viruses, and Protozoa .................................... 1:485

1.2.104

Water Meter ......................................................................................................... 1:489

1.2.105

Municipal Watersheds .......................................................................................... 1:495

1.2.106

Public Water Supply World .................................................................................. 1:500

1.2.107

Ten Key Trends That Will Shape the Future of the World Water Industry ........... 1:508

1.2.108

Zebra Mussel Control without Chemicals ............................................................ 1:510

1.2.109

Package Plants .................................................................................................... 1:514

1.2.110

Anaerobic Sewage Treatment ............................................................................. 1:517

1.2.111

Persistence of Pathogens in Water ...................................................................... 1:521

1.2.112

Well Head Protection ........................................................................................... 1:524 This page has been reformatted by Knovel to provide easier navigation.

Contents 1.2.113 1.3

1.4

xli

Chemical Drinking Water Standards, Past, Present, and Future ......................... 1:529

Industrial Water ..................................................................................................................... 1:534 1.3.1

Magnetic Water Conditioning ............................................................................... 1:534

1.3.2

Water Impacts from Construction Sites ............................................................... 1:537

1.3.3

Industrial Cooling Water – Biofouling ................................................................... 1:538

1.3.4

Industrial Cooling Water – Corrosion ................................................................... 1:542

1.3.5

Industrial Cooling Water – Scale Formation ........................................................ 1:545

1.3.6

Economics of Industrial Water Demands ............................................................. 1:549

1.3.7

Electric Generating Plants – Effects of Contaminants ......................................... 1:553

1.3.8

Energy Dissipation ............................................................................................... 1:558

1.3.9

Water Use in Energy Production .......................................................................... 1:560

1.3.10

Evaluation of Toxic Properties of Industrial Effluents by on-line Respirometry ....................................................................................................... 1:565

1.3.11

Polycyclic Aromatic Hydrocarbons ....................................................................... 1:571

1.3.12

Hydrocarbon Treatment Techniques ................................................................... 1:575

1.3.13

Use of Anaerobic-aerobic Treatment Systems for Maize Processing Plants ....... 1:581

1.3.14

Bonding of Toxic Metal Ions ................................................................................. 1:586

1.3.15

Application of Microfiltration to Industrial Wastewaters ........................................ 1:591

1.3.16

Water Treatment in Spent Nuclear Fuel Storage ................................................. 1:595

1.3.17

Industrial Mine Use: Mine Waste ......................................................................... 1:609

1.3.18

Sugarcane Industry Wastewaters Treatment ....................................................... 1:614

1.3.19

Estimated Use of Water in the United States in 1990 Industrial Water Use ........ 1:620

Waste Water Treatment ........................................................................................................ 1:623 1.4.1

Aeration ............................................................................................................... 1:623

1.4.2

Fine Bubble Diffused Air Aeration Systems ......................................................... 1:626

1.4.3

Air Stripping ......................................................................................................... 1:631

1.4.4

Land Applications of Wastewater in Arid Lands: Theory and Case Studies ........ 1:632

1.4.5

Technologies for Arsenic Removal from Contaminated Water Sources .............. 1:636

1.4.6

Biochemical Oxygen Demand .............................................................................. 1:639

1.4.7

Molecular Biology Tools for Monitoring Biodiversity in Wastewater Treatment Plants ................................................................................................................... 1:642

1.4.8

Biosolids .............................................................................................................. 1:646

1.4.9

Integrated Capacity Building Needs for Water Supply and Wastewater Sanitation ............................................................................................................. 1:651

1.4.10

Wastewater Characterization ............................................................................... 1:656

1.4.11

Chemically Enhanced Primary Treatment of Wastewater ................................... 1:659

1.4.12

Getting Our Clean Water Act Together ................................................................ 1:660

1.4.13

Inadequate Treatment of Wastewater: a Source of Coliform Bacteria in Receiving Surface Water Bodies in Developing Countries – Case Study: Eastern Cape Province of South Africa ............................................................... 1:661 This page has been reformatted by Knovel to provide easier navigation.

xlii

Contents 1.4.14

Denitrification in the Activated Sludge Process ................................................... 1:667

1.4.15

Detergents ........................................................................................................... 1:669

1.4.16

Ecological Wastewater Management ................................................................... 1:675

1.4.17

Waste Treatment in Fish Farms ........................................................................... 1:681

1.4.18

Flotation as a Separation Process ....................................................................... 1:684

1.4.19

Degradation of Chloro-organics and Hydrocarbons ............................................. 1:688

1.4.20

Landfill ................................................................................................................. 1:695

1.4.21

Landfill Leachates, Part 1: Origin and Characterization ....................................... 1:699

1.4.22

Landfill Leachates: Part 2: Treatment .................................................................. 1:702

1.4.23

Macrophytes as Biomonitors of Polychlorinated Biphenyls ................................. 1:714

1.4.24

Wastewater Management for Developing Countries ........................................... 1:718

1.4.25

Mercury Removal from Complex Waste Waters .................................................. 1:722

1.4.26

Metal Speciation and Mobility as Influenced by Landfill Disposal Practices ........ 1:723

1.4.27

Microbial Foaming in the Activated Sludge Process ............................................ 1:728

1.4.28

Introduction to Wastewater Modeling and Treatment Plant Design ..................... 1:730

1.4.29

Practical Applications of Wastewater Modeling and Treatment Plant Design ..... 1:738

1.4.30

New York City Harbor Survey .............................................................................. 1:745

1.4.31

Nitrification in the Activated Sludge Process ....................................................... 1:751

1.4.32

Effluent Limitations and the NPDES Permit ......................................................... 1:755

1.4.33

Odor Abatement in Wastewater Treatment Plants .............................................. 1:760

1.4.34

Aqueous Reactions of Specific Organic Compounds with Ozone ....................... 1:765

1.4.35

The Fate of Persistent Organic Pollutants (POPs) in the Wastewater Treatment Process .............................................................................................. 1:766

1.4.36

The Role of Organoclay in Water Cleanup .......................................................... 1:771

1.4.37

Combined Sewer Overflow Treatment ................................................................. 1:782

1.4.38

Biological Phosphorus Removal in the Activated Sludge Process ...................... 1:788

1.4.39

Photocatalytic Membrane Reactors in Water Purification .................................... 1:791

1.4.40

EPA’s National Pretreatment Program, 1973–2003: Thirty Years of Protecting the Environment .................................................................................. 1:798

1.4.41

Problems Encountered during Pipe Repair and Renewal .................................... 1:801

1.4.42

Radioactive Waste ............................................................................................... 1:802

1.4.43

Reclaimed Water ................................................................................................. 1:805

1.4.44

Wastewater Treatment and Recycling Technologies ........................................... 1:808

1.4.45

Wastewater Treatment Processes and Water Reuse .......................................... 1:814

1.4.46

Wastewater Reclamation and Reuse Research .................................................. 1:819

1.4.47

Wastewater Reclamation and Reuse ................................................................... 1:825

1.4.48

Wastewater Reclamation and Reuse Treatment Technology .............................. 1:826

1.4.49

Sewage ................................................................................................................ 1:828

1.4.50

Domestic Sewage ................................................................................................ 1:830

1.4.51

Solidification/Stabilization of Hazardous Solid Wastes ........................................ 1:835 This page has been reformatted by Knovel to provide easier navigation.

Contents

xliii

1.4.52

Wastewater Treatment – Small Scale .................................................................. 1:840

1.4.53

Microbial Foaming and Bulking in Activated Sludge Plants ................................. 1:844

1.4.54

Aqueous Behavior of Elements in a Flue Gas Desulfurization Sludge Disposal Site ........................................................................................................ 1:848

1.4.55

Sludge Treatment and Disposal .......................................................................... 1:853

1.4.56

Wastewater Sludge .............................................................................................. 1:861

1.4.57

Processing of Sludge ........................................................................................... 1:864

1.4.58

Municipal Storm Water Management ................................................................... 1:866

1.4.59

What Wastewater Utilities Can Do Now to Guard Against Terrorist and Security Threats ................................................................................................... 1:870

1.4.60

Wastewater Treatment Techniques – Advanced ................................................. 1:871

1.4.61

Trenchless Repair and Rehabilitation Techniques .............................................. 1:876

1.4.62

Basics of Underground Water and Sewer Pipeline Assessment, Repair, and Rehabilitation ....................................................................................................... 1:883

1.4.63

Water Hammer: Quantitative Causes and Effects ............................................... 1:891

1.4.64

Constructed Wetlands .......................................................................................... 1:892

1.4.65

Using Ecosystem Processes in a Constructed Wetland to Treat Mine Wastewater in Ireland .......................................................................................... 1:897

1.4.66

Water and Wastewater Properties and Characteristics ....................................... 1:900

1.4.67

Anaerobic Wastewater Treatment ....................................................................... 1:904

1.4.68

Sewerage Odors – How to Control ...................................................................... 1:910

1.4.69

Ultrafiltration – Complexation in Wastewater Treatment ...................................... 1:916

Volume 2. Water Quality and Resource Development 2.5

Water Quality Control ............................................................................................................

2:1

2.5.1

Acid Mine Drainage – Extent and Character ........................................................

2:1

2.5.2

The Control of Algal Populations in Eutrophic Water Bodies ...............................

2:2

2.5.3

Arsenic Compounds in Water ..............................................................................

2:7

2.5.4

Arsenic Health Effects .........................................................................................

2:15

2.5.5

Background Concentration of Pollutants ..............................................................

2:18

2.5.6

Waterborne Bacteria ............................................................................................

2:20

2.5.7

Water Assessment and Criteria ...........................................................................

2:24

2.5.8

Physiological Biomarkers and the Trondheim Biomonitoring System ..................

2:28

2.5.9

Biomarkers, Bioindicators, and the Trondheim Biomonitoring System ................

2:29

2.5.10

Active Biomonitoring (ABM) by Translocation of Bivalve Molluscs ......................

2:33

2.5.11

Biochemical Oxygen Demand and Other Organic Pollution Measures ...............

2:37

2.5.12

Biodegradation .....................................................................................................

2:41

2.5.13

Bioluminescent Biosensors for Toxicity Testing ...................................................

2:45

2.5.14

Biomanipulation ...................................................................................................

2:50

2.5.15

Genomic Technologies in Biomonitoring .............................................................

2:58

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xliv

Contents 2.5.16

Macrophytes as Biomonitors of Trace Metals ......................................................

2:64

2.5.17

Biosorption of Toxic Metals ..................................................................................

2:68

2.5.18

Bromide Influence on Trihalomethane and Haloacetic Acid Formation ...............

2:74

2.5.19

Activated Carbon: Ion Exchange and Adsorption Properties ...............................

2:79

2.5.20

Activated Carbon – Powdered .............................................................................

2:86

2.5.21

Chlorination ..........................................................................................................

2:88

2.5.22

Chlorination by-products ......................................................................................

2:91

2.5.23

Classification and Environmental Quality Assessment in Aquatic Environments .......................................................................................................

2:94

2.5.24

Coagulation and Flocculation in Practice .............................................................

2:98

2.5.25

Colloids and Dissolved Organics: Role in Membrane and Depth Filtration .........

2:99

2.5.26

Column Experiments in Saturated Porous Media Studying Contaminant Transport ............................................................................................................. 2:103

2.5.27

Cytochrome P450 Monooxygenase as an Indicator of PCB/Dioxin-like Compounds in Fish .............................................................................................. 2:106

2.5.28

Water Related Diseases ...................................................................................... 2:111

2.5.29

Dishwashing Water Quality Properties ................................................................ 2:112

2.5.30

Disinfection by-product Precursor Removal from Natural Waters ....................... 2:115

2.5.31

Alternative Disinfection Practices and Future Directions for Disinfection byproduct Minimization ............................................................................................ 2:118

2.5.32

Water Quality Aspects of Dredged Sediment Management ................................ 2:122

2.5.33

The Economics of Water Quality ......................................................................... 2:127

2.5.34

Understanding Escherichia Coli O157:H7 and the Need for Rapid Detection in Water ............................................................................................................... 2:136

2.5.35

Eutrophication and Organic Loading .................................................................... 2:142

2.5.36

Trace Element Contamination in Groundwater of District Hardwar, Uttaranchal, India ................................................................................................. 2:143

2.5.37

Iron Bacteria ........................................................................................................ 2:149

2.5.38

Cartridge Filters for Iron Removal ........................................................................ 2:152

2.5.39

Irrigation Water Quality in Areas Adjoining River Yamuna at Delhi, India ........... 2:155

2.5.40

Water Sampling and Laboratory Safety ............................................................... 2:161

2.5.41

Municipal Solid Waste Landfills – Water Quality Issues ...................................... 2:163

2.5.42

Land Use Effects on Water Quality ...................................................................... 2:169

2.5.43

Monitoring Lipophilic Contaminants in the Aquatic Environment Using the SPMD-TOX Paradigm .......................................................................................... 2:170

2.5.44

Use of Luminescent Bacteria and the Lux Genes for Determination of Water Quality .................................................................................................................. 2:172

2.5.45

Water Quality Management ................................................................................. 2:176

2.5.46

Water Quality Management and Nonpoint Source Control .................................. 2:184

2.5.47

Water Quality Management in an Urban Landscape ........................................... 2:189

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Contents

xlv

2.5.48

Water Quality Management in the U.S.: History of Water Regulation ................. 2:193

2.5.49

Water Quality Management in a Forested Landscape ......................................... 2:199

2.5.50

Trace Metal Speciation ........................................................................................ 2:202

2.5.51

Metal Ion Humic Colloid Interaction ..................................................................... 2:205

2.5.52

Heavy Metal Uptake Rates Among Sediment Dwelling Organisms ..................... 2:211

2.5.53

Methemoglobinemia ............................................................................................ 2:219

2.5.54

Microbial Activities Management ......................................................................... 2:223

2.5.55

Microbial Dynamics of Biofilms ............................................................................ 2:228

2.5.56

Microbial Enzyme Assays for Detecting Heavy Metal Toxicity ............................ 2:233

2.5.57

Microbial Forms in Biofouling Events ................................................................... 2:239

2.5.58

Microbiological Quality Control in Distribution Systems ....................................... 2:243

2.5.59

Water Quality Models for Developing Soil Management Practices ...................... 2:248

2.5.60

Water Quality Modeling – Case Studies .............................................................. 2:255

2.5.61

Field Sampling and Monitoring of Contaminants ................................................. 2:263

2.5.62

Water Quality Models: Chemical Principles ......................................................... 2:269

2.5.63

Water Quality Models: Mathematical Framework ................................................ 2:273

2.5.64

Environmental Applications with Submitochondrial Particles ............................... 2:278

2.5.65

Interest in the Use of an Electronic Nose for Field Monitoring of Odors in the Environment ......................................................................................................... 2:281

2.5.66

Oil-field Brine ....................................................................................................... 2:284

2.5.67

Oil Pollution .......................................................................................................... 2:290

2.5.68

Indicator Organisms ............................................................................................. 2:292

2.5.69

pH ........................................................................................................................ 2:294

2.5.70

Perchloroethylene (PCE) Removal ...................................................................... 2:299

2.5.71

A Primer on Water Quality ................................................................................... 2:301

2.5.72

Overview of Analytical Methods of Water Analyses with Specific Reference to EPA Methods for Priority Pollutant Analysis .................................................... 2:304

2.5.73

Source-water Protection ...................................................................................... 2:311

2.5.74

Protozoa in Water ................................................................................................ 2:313

2.5.75

Water Quality ....................................................................................................... 2:314

2.5.76

Water Quality ....................................................................................................... 2:316

2.5.77

Emerging and Recalcitrant Compounds in Groundwater ..................................... 2:316

2.5.78

Road Salt ............................................................................................................. 2:319

2.5.79

Review of River Water Quality Modeling Software Tools .................................... 2:325

2.5.80

River Water Quality Calibration ............................................................................ 2:331

2.5.81

Salmonella: Monitoring and Detection in Drinking Water ..................................... 2:337

2.5.82

Lysimeter Soil Water Sampling ............................................................................ 2:340

2.5.83

Regulatory and Security Requirements for Potable Water .................................. 2:343

2.5.84

A Weight of Evidence Approach to Characterize Sediment Quality Using Laboratory and Field Assays: an Example for Spanish Coasts ........................... 2:350 This page has been reformatted by Knovel to provide easier navigation.

xlvi

2.6

Contents 2.5.85

Remediation and Bioremediation of Selenium-contaminated Waters .................. 2:355

2.5.86

Shellfish Growing Water Classification ................................................................ 2:360

2.5.87

Sorptive Filtration ................................................................................................. 2:362

2.5.88

Quality of Water in Storage .................................................................................. 2:367

2.5.89

Quality of Water Supplies .................................................................................... 2:370

2.5.90

The Submitochondrial Particle Assay as a Biological Monitoring Tool ................ 2:376

2.5.91

Microscale Test Relationships to Responses to Toxicants in Natural Systems ............................................................................................................... 2:379

2.5.92

Toxicity Identification Evaluation .......................................................................... 2:380

2.5.93

Whole Effluent Toxicity Controls .......................................................................... 2:382

2.5.94

Development and Application of Sediment Toxicity Tests for Regulatory Purposes .............................................................................................................. 2:383

2.5.95

Algal Toxins in Water ........................................................................................... 2:387

2.5.96

Ground Water Quality in Areas Adjoining River Yamuna at Delhi, India ............. 2:392

2.5.97

Chlorine Residual ................................................................................................ 2:398

2.5.98

Source Water Quality Management ..................................................................... 2:399

2.5.99

Dose-response of Mussels to Chlorine ................................................................ 2:401

2.5.100

Metallothioneins as Indicators of Trace Metal Pollution ....................................... 2:406

2.5.101

Amphipod Sediment Toxicity Tests ...................................................................... 2:408

2.5.102

Ciliated Protists as Test Organisms in Toxicity Assessment ............................... 2:413

2.5.103

SOFIE: an Optimized Approach for Exposure Tests and Sediment Assays ........ 2:418

2.5.104

Passive Treatment of Acid Mine Drainage (Wetlands) ........................................ 2:423

2.5.105

Biomarkers and Bioaccumulation: Two Lines of Evidence to Assess Sediment Quality ................................................................................................. 2:426

2.5.106

Lead and Its Health Effects .................................................................................. 2:432

2.5.107

Microbial Detection of Various Pollutants as an Early Warning System for Monitoring of Water Quality and Ecological Integrity of Natural Resources, in Russia .................................................................................................................. 2:440

2.5.108

Luminescent Bacterial Biosensors for the Rapid Detection of Toxicants ............. 2:453

2.5.109

Development and Application of Sediment Toxicity Test for Regulatory Purposes .............................................................................................................. 2:458

2.5.110

Eh ......................................................................................................................... 2:464

Water Resource Development and Management ................................................................. 2:470 2.6.1

Water Resources Challenges in the Arab World ................................................. 2:470

2.6.2

Effluent Water Regulations in Arid Lands ............................................................ 2:475

2.6.3

California – Continually the Nation’s Leader in Water Use .................................. 2:478

2.6.4

Lessons from the Rising Caspian ........................................................................ 2:480

2.6.5

Institutional Aspects of Water Management in China .......................................... 2:484

2.6.6

Will Water Scarcity Limit China’s Agricultural Potential? ..................................... 2:488

2.6.7

Water and Coastal Resources ............................................................................. 2:489

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Contents

xlvii

2.6.8

Water Use Conservation and Efficiency .............................................................. 2:489

2.6.9

Conservation of Water ......................................................................................... 2:495

2.6.10

The Development of American Water Resources: Planners, Politicians, and Constitutional Interpretation ................................................................................. 2:498

2.6.11

Water Markets: Transaction Costs and Institutional Options ............................... 2:499

2.6.12

Averting Water Disputes ...................................................................................... 2:501

2.6.13

Water Supply and Water Resources: Distribution System Research .................. 2:509

2.6.14

Drought in the Dust Bowl Years ........................................................................... 2:511

2.6.15

Drought Management Planning ........................................................................... 2:514

2.6.16

Drought and Water Supply Management ............................................................. 2:515

2.6.17

Assessment of Ecological Effects in Water-limited Environments ....................... 2:516

2.6.18

Reaching Out: Public Education and Community Involvement in Groundwater Protection ....................................................................................... 2:518

2.6.19

Integration of Environmental Impacts into Water Resources Planning ................ 2:520

2.6.20

The Expansion of Federal Water Projects ........................................................... 2:522

2.6.21

Flood Control History in the Netherlands ............................................................. 2:524

2.6.22

Food and Water in an Emergency ....................................................................... 2:526

2.6.23

Water Demand Forecasting ................................................................................. 2:529

2.6.24

Remote Sensing and GIS Application in Water Resources ................................. 2:531

2.6.25

Globalization of Water ......................................................................................... 2:536

2.6.26

Water Science Glossary of Terms ....................................................................... 2:541

2.6.27

Harvesting Rainwater ........................................................................................... 2:548

2.6.28

Urban Water Resource and Management in Asia: Ho Chi Minh City .................. 2:552

2.6.29

Hydropower – Energy from Moving Water ........................................................... 2:554

2.6.30

Water Markets in India: Economic and Institutional Aspects ............................... 2:555

2.6.31

Water Resources of India .................................................................................... 2:559

2.6.32

Water Infrastructure and Systems ....................................................................... 2:567

2.6.33

Overview and Trends in the International Water Market ...................................... 2:568

2.6.34

Best Management Practices for Water Resources .............................................. 2:570

2.6.35

Integrated Water Resources Management (IWRM) ............................................. 2:574

2.6.36

Management of Water Resources for Drought Conditions .................................. 2:576

2.6.37

Water Resources Management ........................................................................... 2:586

2.6.38

NASA Helping to Understand Water Flow in the West ........................................ 2:587

2.6.39

Transboundary Water Conflicts in the Nile Basin ................................................ 2:590

2.6.40

Planning and Managing Water Infrastructure ...................................................... 2:594

2.6.41

Application of the Precautionary Principle to Water Science ............................... 2:595

2.6.42

Water Pricing ....................................................................................................... 2:603

2.6.43

Spot Prices, Option Prices, and Water Markets ................................................... 2:606

2.6.44

Water Managed in the Public Trust ...................................................................... 2:608

2.6.45

Water Recycling and Reuse: the Environmental Benefits ................................... 2:610 This page has been reformatted by Knovel to provide easier navigation.

xlviii

Contents 2.6.46

State and Regional Water Supply ........................................................................ 2:613

2.6.47

River Basin Decisions Support Systems .............................................................. 2:619

2.6.48

Water Resource Sustainability: Concepts and Practices ..................................... 2:624

2.6.49

The Provision of Drinking Water and Sanitation in Developing Countries ........... 2:630

2.6.50

Sustainable Management of Natural Resources ................................................. 2:633

2.6.51

Sustainable Water Management on Mediterranean Islands: Research and Education ............................................................................................................. 2:638

2.6.52

Meeting Water Needs in Developing Countries with Tradable Rights ................. 2:643

2.6.53

Water Use in the United States ........................................................................... 2:645

2.6.54

How We Use Water in These United States ........................................................ 2:650

2.6.55

Valuing Water Resources .................................................................................... 2:653

2.6.56

Water – Here, There, and Everywhere in Canada ............................................... 2:656

2.6.57

Water Conservation – Every Drop Counts in Canada ......................................... 2:660

2.6.58

Ecoregions: a Spatial Framework for Environmental Management ..................... 2:667

2.6.59

Flood of Portals on Water .................................................................................... 2:668

2.6.60

Fuzzy Criteria for Water Resources Systems Performance Evaluation ............... 2:674

2.6.61

Participatory Multicriteria Flood Management ...................................................... 2:678

2.6.62

Water Resources Systems Analysis .................................................................... 2:683

Volume 3. Surface and Agricultural Water 3.7

Surface Water Hydrology ......................................................................................................

3:1

3.7.1

Acidification – Chronic .........................................................................................

3:1

3.7.2

Episodic Acidification ...........................................................................................

3:5

3.7.3

Acidification of Freshwater Resources .................................................................

3:7

3.7.4

Geochemistry of Acid Mine Drainage ..................................................................

3:13

3.7.5

The Aral Sea Disaster: Environment Issues and Nationalist Tensions ................

3:15

3.7.6

Lake Baikal – a Touchstone for Global Change and Rift Studies ........................

3:20

3.7.7

Base Flow ............................................................................................................

3:22

3.7.8

River Basins .........................................................................................................

3:28

3.7.9

River Basin Planning and Coordination ...............................................................

3:33

3.7.10

Bioaccumulation ..................................................................................................

3:34

3.7.11

Biotic Integrity Index to Evaluate Water Resource Integrity in Freshwater Systems ...............................................................................................................

3:36

3.7.12

Reversal of the Chicago River .............................................................................

3:41

3.7.13

Flood Control in the Yellow River Basin in China ................................................

3:45

3.7.14

Chironomids in Sediment Toxicity Testing ...........................................................

3:50

3.7.15

Cienega ...............................................................................................................

3:57

3.7.16

Time-area and the Clark Rainfall-runoff Transformation ......................................

3:60

3.7.17

Stream Classification ...........................................................................................

3:65

3.7.18

Coastal Wetlands .................................................................................................

3:71

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xlix

3.7.19

Freshwater Colloids .............................................................................................

3:74

3.7.20

Culvert Design .....................................................................................................

3:75

3.7.21

Dilution-mixing Zones and Design Flows .............................................................

3:78

3.7.22

Drainage Ditches .................................................................................................

3:87

3.7.23

Drainage Networks ..............................................................................................

3:93

3.7.24

Dyes as Hydrological Tracers ..............................................................................

3:95

3.7.25

Flow-duration Curves ........................................................................................... 3:102

3.7.26

Environmental Flows ............................................................................................ 3:106

3.7.27

Eutrophication (Excessive Fertilization) ............................................................... 3:107

3.7.28

Cultural Eutrophication ........................................................................................ 3:114

3.7.29

Fish Cells in the Toxicological Evaluation of Environmental Contaminants ......... 3:115

3.7.30

Fish Consumption Advisories .............................................................................. 3:118

3.7.31

Fisheries: History, Science, and Management .................................................... 3:121

3.7.32

Factors Affecting Fish Growth and Production .................................................... 3:129

3.7.33

Water Needs for Freshwater Fisheries Management .......................................... 3:133

3.7.34

An Outline of the History of Fishpond Culture in Silesia, the Western Part of Poland .................................................................................................................. 3:135

3.7.35

Floods .................................................................................................................. 3:142

3.7.36

Flood Control Structures ...................................................................................... 3:150

3.7.37

Floods as a Natural Hazard ................................................................................. 3:153

3.7.38

Flood Source Mapping in Watersheds ................................................................. 3:155

3.7.39

Urban Flooding .................................................................................................... 3:159

3.7.40

Floodwater Spreading .......................................................................................... 3:163

3.7.41

Minimum Environmental Flow Regimes ............................................................... 3:166

3.7.42

Forensic Hydrogeology ........................................................................................ 3:168

3.7.43

Forests and Wetlands .......................................................................................... 3:170

3.7.44

Rock Glacier ........................................................................................................ 3:174

3.7.45

Great Lakes ......................................................................................................... 3:175

3.7.46

Greenhouse Gas Emissions from Hydroelectric Reservoirs ................................ 3:180

3.7.47

Gully Erosion ....................................................................................................... 3:183

3.7.48

Potential Health Issues Associated with Blue-green Algae Blooms in Impoundments, Ponds and Lakes ....................................................................... 3:188

3.7.49

Heat Balance of Open Waterbodies .................................................................... 3:190

3.7.50

Hydraulics ............................................................................................................ 3:194

3.7.51

Hydraulics of Pressurized Flow ............................................................................ 3:196

3.7.52

Hydroelectric Power ............................................................................................. 3:199

3.7.53

Hydroelectric Reservoirs as Anthropogenic Sources of Greenhouse Gases ...... 3:203

3.7.54

Hydrologic Persistence and the Hurst Phenomenon ........................................... 3:210

3.7.55

Unit Hydrograph ................................................................................................... 3:221

3.7.56

Hydrological Processes and Measured Pollutant Loads ...................................... 3:222 This page has been reformatted by Knovel to provide easier navigation.

l

Contents 3.7.57

Hydrologic Thresholds ......................................................................................... 3:229

3.7.58

Ganga River, India ............................................................................................... 3:232

3.7.59

Interception .......................................................................................................... 3:235

3.7.60

Kinematic Shock .................................................................................................. 3:239

3.7.61

Kinematic Wave Method for Storm Drainage Design .......................................... 3:242

3.7.62

Kinematic Wave and Diffusion Wave Theories .................................................... 3:246

3.7.63

Kinematic Wave Flow Routing ............................................................................. 3:253

3.7.64

Reliability Concepts in Reservoir Design ............................................................. 3:259

3.7.65

Lakes ................................................................................................................... 3:265

3.7.66

The Theory of Alternative Stable States in Shallow Lake Ecosystems ................ 3:272

3.7.67

NOAA Lake Level Forecast for Lake Michigan Right on Target .......................... 3:274

3.7.68

Submerged Aquatic Plants Affect Water Quality in Lakes ................................... 3:275

3.7.69

Lakes – Discharges to ......................................................................................... 3:281

3.7.70

Lasers Scan Levees from the Air ......................................................................... 3:284

3.7.71

Levees for Flood Protection ................................................................................. 3:286

3.7.72

Limnology ............................................................................................................ 3:291

3.7.73

Adsorption of Metal Ions on Bed Sediments ........................................................ 3:295

3.7.74

Microbiology of Lotic Aggregates and Biofilms .................................................... 3:305

3.7.75

Microorganisms in Their Natural Environment ..................................................... 3:309

3.7.76

Calibration of Hydraulic Network Models ............................................................. 3:313

3.7.77

Numerical Modeling of Currents .......................................................................... 3:320

3.7.78

Uncertainty Analysis in Watershed Modeling ....................................................... 3:325

3.7.79

Watershed Modeling ............................................................................................ 3:327

3.7.80

Modeling of Water Quality in Sewers ................................................................... 3:331

3.7.81

Modeling of Urban Drainage and Stormwater ...................................................... 3:337

3.7.82

Modeling Ungauged Watersheds ......................................................................... 3:342

3.7.83

Corps Turned Niagara Falls Off, on Again ........................................................... 3:345

3.7.84

Open Channel Design ......................................................................................... 3:346

3.7.85

Organic Compounds and Trace Elements in Freshwater Streambed Sediment and Fish from the Puget Sound Basin ................................................. 3:349

3.7.86

Impervious Cover – Paving Paradise ................................................................... 3:363

3.7.87

Phytoremediation by Constructed Wetlands ........................................................ 3:364

3.7.88

Unrecognized Pollutants ...................................................................................... 3:371

3.7.89

Pollution of Surface Waters ................................................................................. 3:373

3.7.90

Pond Aquaculture – Modeling and Decision Support Systems ............................ 3:375

3.7.91

Pumping Stations ................................................................................................. 3:379

3.7.92

Regulated Rivers ................................................................................................. 3:381

3.7.93

Reservoirs-multipurpose ...................................................................................... 3:382

3.7.94

Dam Removal as River Restoration ..................................................................... 3:387

3.7.95

Riparian Systems ................................................................................................. 3:390 This page has been reformatted by Knovel to provide easier navigation.

Contents

li

3.7.96

Rivers ................................................................................................................... 3:392

3.7.97

River and Water Facts ......................................................................................... 3:394

3.7.98

Sediment Load Measurements ............................................................................ 3:397

3.7.99

Sedimentation ...................................................................................................... 3:401

3.7.100

Sedimentation and Flotation ................................................................................ 3:404

3.7.101

Reservoir Sedimentation ..................................................................................... 3:408

3.7.102

Water from Saturated River Sediment – Sand Abstraction .................................. 3:412

3.7.103

Sediment Transport ............................................................................................. 3:417

3.7.104

Stochastic Simulation of Hydrosystems ............................................................... 3:421

3.7.105

Storage and Detention Facilities .......................................................................... 3:430

3.7.106

Urban Stormwater Runoff Water Quality Issues .................................................. 3:432

3.7.107

Rivers and Streams: One-way Flow System ....................................................... 3:437

3.7.108

Streamflow ........................................................................................................... 3:439

3.7.109

Water Quality in Suburban Watersheds ............................................................... 3:441

3.7.110

Surface Water Pollution ....................................................................................... 3:444

3.7.111

Surface Runoff and Subsurface Drainage ........................................................... 3:451

3.7.112

Trace Elements in Water, Sediment, and Aquatic Biota – Effects of Biology and Human Activity .............................................................................................. 3:454

3.7.113

Innovative Pens Hatch Thousands of Trout ......................................................... 3:458

3.7.114

Watershed ........................................................................................................... 3:460

3.7.115

Combustible Watersheds ..................................................................................... 3:461

3.7.116

Time of Concentration and Travel Time in Watersheds ....................................... 3:469

3.7.117

Watershed Hydrology .......................................................................................... 3:472

3.7.118

Watershed Management for Environmental Quality and Food Security .............. 3:479

3.7.119

Water Hyacinth – The World’s Most Problematic Weed ...................................... 3:479

3.7.120

Water Quality in Ponds ........................................................................................ 3:484

3.7.121

Water Turbine ...................................................................................................... 3:487

3.7.122

Wetlands: Uses, Functions, and Values .............................................................. 3:489

3.7.123

Wetlands Overview .............................................................................................. 3:493

3.7.124

Classification of Wetlands and Deepwater Habitats of the United States ............ 3:496

3.7.125

Urban Runoff ....................................................................................................... 3:498

3.7.126

Urban Water Studies ............................................................................................ 3:501

3.7.127

Subglacial Lake Vostok ....................................................................................... 3:503

3.7.128

Water – the Canadian Transporter ...................................................................... 3:507

3.7.129

Flood Prevention .................................................................................................. 3:510

3.7.130

Effects of DDT in Surface Water on Bird Abundance and Reproduction – a History .................................................................................................................. 3:513

3.7.131

Instream Flow Methods ........................................................................................ 3:526

3.7.132

Floodplain ............................................................................................................ 3:527

3.7.133

Fish Passage Facilities ........................................................................................ 3:529 This page has been reformatted by Knovel to provide easier navigation.

lii

3.8

Contents 3.7.134

Fishing Waters ..................................................................................................... 3:532

3.7.135

Land Surface Modeling ........................................................................................ 3:533

Agricultural Water .................................................................................................................. 3:538 3.8.1

Animal Farming Operations: Groundwater Quality Issues ................................... 3:538

3.8.2

Aquaculture Technology for Producers ................................................................ 3:540

3.8.3

Biofuel Alternatives to Fossil Fuels ...................................................................... 3:545

3.8.4

Soil Conservation ................................................................................................. 3:549

3.8.5

Landscape Water-conservation Techniques ........................................................ 3:553

3.8.6

Crop Water Requirements ................................................................................... 3:557

3.8.7

Agricultural Water Use Efficiency (WUE) and Productivity (WP) ......................... 3:558

3.8.8

Large Area Surface Energy Balance Estimation Using Satellite Imagery ............ 3:560

3.8.9

Soil Erosion and Control Practices ...................................................................... 3:565

3.8.10

Water Table Contribution to Crop Evapotranspiration ......................................... 3:570

3.8.11

Crop Evapotranspiration ...................................................................................... 3:571

3.8.12

Water Pollution from Fish Farms ......................................................................... 3:579

3.8.13

World’s Major Irrigation Areas .............................................................................. 3:581

3.8.14

Irrigation in the United States ............................................................................... 3:586

3.8.15

Irrigation Wells ..................................................................................................... 3:594

3.8.16

Agriculture and Land Use Planning ..................................................................... 3:595

3.8.17

Waterlogging ........................................................................................................ 3:599

3.8.18

Water Quality Management in an Agricultural Landscape ................................... 3:604

3.8.19

Classification and Mapping of Agricultural Land for National Water-quality Assessment ......................................................................................................... 3:608

3.8.20

Metal Tolerance in Plants: the Roles of Thiol-containing Peptides ...................... 3:609

3.8.21

Microirrigation ...................................................................................................... 3:615

3.8.22

Microirrigation: an Approach to Efficient Irrigation ............................................... 3:620

3.8.23

Plant and Microorganism Selection for Phytoremediation of Hydrocarbons and Metals ........................................................................................................... 3:628

3.8.24

Nitrate Pollution Prevention ................................................................................. 3:637

3.8.25

Nitrification ........................................................................................................... 3:640

3.8.26

Occurrence of Organochlorine Pesticides in Vegetables Grown on Untreated Soils from an Agricultural Watershed ................................................................... 3:643

3.8.27

Pesticide Chemistry in the Environment .............................................................. 3:647

3.8.28

Remediation of Pesticide-contaminated Soil at Agrichemical Facilities ............... 3:651

3.8.29

Pesticide Occurrence and Distribution in Relation to Use ................................... 3:655

3.8.30

Assessment of Pollution Outflow from Large Agricultural Areas .......................... 3:657

3.8.31

Deep-well Turbine Pumps .................................................................................... 3:664

3.8.32

Microbial Quality of Reclaimed Irrigation: International Perspective .................... 3:667

3.8.33

Soil Salinity .......................................................................................................... 3:673

3.8.34

Maintaining Salt Balance on Irrigated Land ......................................................... 3:677 This page has been reformatted by Knovel to provide easier navigation.

Contents

liii

3.8.35

Salt Tolerance ...................................................................................................... 3:681

3.8.36

Groundwater Assessment Using Soil Sampling Techniques ............................... 3:688

3.8.37

Skimmed Groundwater ........................................................................................ 3:691

3.8.38

Soil Moisture Measurement – Neutron ................................................................ 3:692

3.8.39

Soil N Management Impact on the Quality of Surface and Subsurface Water ................................................................................................................... 3:694

3.8.40

Soil Phosphorus Availability and Its Impact on Surface Water Quality ................ 3:701

3.8.41

Soil Water Issues ................................................................................................. 3:706

3.8.42

Water Spreading .................................................................................................. 3:708

3.8.43

Sprinkler Irrigation ................................................................................................ 3:712

3.8.44

Stomates .............................................................................................................. 3:714

3.8.45

Crop Water Stress Detection Using Remote Sensing ......................................... 3:719

3.8.46

Vacuum Gauge Tensiometer ............................................................................... 3:724

3.8.47

Tile Drainage ....................................................................................................... 3:729

3.8.48

Tile Drainage: Impacts, Plant Growth, and Water Table Levels .......................... 3:731

3.8.49

Measuring and Modeling Tree and Stand Level Transpiration ............................ 3:732

3.8.50

Water Logging: Topographic and Agricultural Impacts ........................................ 3:741

3.8.51

Weed Control Strategies ...................................................................................... 3:742

3.8.52

Screen Filters for Microirrigation .......................................................................... 3:748

3.8.53

Xeriscape ............................................................................................................. 3:750

3.8.54

Media Filters for Microirrigation ............................................................................ 3:752

Volume 4. Oceanography; Meteorology; Physics and Chemistry; Water Law; and Water 4.9

Oceanography .......................................................................................................................

4:1

4.9.1

Air-sea Interaction ................................................................................................

4:1

4.9.2

NOAA’s Atlantic Oceanographic and Meteorological Laboratory ........................

4:4

4.9.3

Laboratory Experiments on Bivalve Excretion Rates of Nutrients .......................

4:6

4.9.4

Temporal Scaling of Benthic Nutrient Regeneration in Bivalve-dominated Tidal Flat ..............................................................................................................

4:11

4.9.5

Breakwaters .........................................................................................................

4:14

4.9.6

The Ocean in Climate ..........................................................................................

4:21

4.9.7

Coastal Waters ....................................................................................................

4:23

4.9.8

Marine Colloids ....................................................................................................

4:27

4.9.9

Deep Water Corals ..............................................................................................

4:32

4.9.10

Marine Debris Abatement ....................................................................................

4:38

4.9.11

Larvae and Small Species of Polychaetes in Marine Toxicological Testing ........

4:42

4.9.12

El Niño: the Interannual Prediction Problem ........................................................

4:43

4.9.13

Renewable Energies from the Ocean ..................................................................

4:44

4.9.14

Estuarian Waters .................................................................................................

4:49

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liv

Contents 4.9.15

NOS/NMFS Cooperative Research on Coastal Fisheries and Habitats at the Beaufort Laboratory .............................................................................................

4:55

4.9.16

Distribution and Dynamics of Gas Hydrates in the Marine Environment .............

4:57

4.9.17

Oceanographic Environment of Glacier Bay ........................................................

4:61

4.9.18

NOS Sanctuaries Protect Nation’s Maritime History ............................................

4:62

4.9.19

Quantification of Anoxia and Hypoxia in Water Bodies ........................................

4:64

4.9.20

Floating Ice ..........................................................................................................

4:69

4.9.21

Technology Development: Hardware Development – Marine Instrumentation Laboratory (MIL) ..................................................................................................

4:70

Seasonal Coupling between Intertidal Macrofauna and Sediment Column Porewater Nutrient Concentrations ......................................................................

4:73

Mapping the Sea Floor of the Historic Area Remediation Site (HARS) Offshore of New York City ...................................................................................

4:77

NOAA and University Scientists Study Methyl Bromide Cycling in the North Pacific ..................................................................................................................

4:80

4.9.25

Tidally Mediated Changes in Nutrient Concentrations .........................................

4:81

4.9.26

The Role of Oceans in the Global Cycles of Climatically-active Trace-gases .....

4:85

4.9.27

Pacific Marine Environmental Laboratory– 30 Years of Observing the Ocean ..................................................................................................................

4:89

4.9.28

Seawater Temperature Estimates in Paleoceanography .....................................

4:92

4.9.29

Physical Oceanography .......................................................................................

4:95

4.9.30

Coastal Water Pollutants .....................................................................................

4:96

4.9.31

Trace Element Pollution ....................................................................................... 4:109

4.9.32

Coral Reefs and Your Coastal Watershed ........................................................... 4:113

4.9.33

Sea Level and Climate ......................................................................................... 4:117

4.9.34

The Permanent Service for Mean Sea Level ....................................................... 4:118

4.9.35

Marine and Estuarine Microalgal Sediment Toxicity Tests .................................. 4:120

4.9.36

Marine Stock Enhancement Techniques ............................................................. 4:124

4.9.37

Physical and Chemical Variability of Tidal Streams ............................................. 4:128

4.9.38

Black Water Turns the Tide on Florida Coral ....................................................... 4:133

4.9.39

Shallow Water Waves .......................................................................................... 4:135

4.9.40

Water Waves ....................................................................................................... 4:138

4.9.41

Woods Hole: the Early Years ............................................................................... 4:139

4.9.42

An Analysis of the International Maritime Organization – London Convention Annual Ocean Dumping Reports ......................................................................... 4:144

4.9.43

Marine Sources of Halocarbons ........................................................................... 4:149

4.9.44

Food Chain/Foodweb/Food Cycle ....................................................................... 4:151

4.9.45

Plankton ............................................................................................................... 4:154

4.9.46

Major Ions in Seawater ........................................................................................ 4:159

4.9.47

Tsunami ............................................................................................................... 4:160

4.9.22 4.9.23 4.9.24

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Contents

lv

4.10 Meteorology ........................................................................................................................... 4:164 4.10.1

Ballooning and Meteorology in the Twentieth Century ........................................ 4:164

4.10.2

Barometric Efficiency ........................................................................................... 4:166

4.10.3

CERES: Understanding the Earth’s Clouds and Climate ..................................... 4:169

4.10.4

Chinook ................................................................................................................ 4:170

4.10.5

Global Climate Change ........................................................................................ 4:171

4.10.6

Observations of Climate and Global Change from Real-time Measurements ..... 4:172

4.10.7

Overview: the Climate System ............................................................................. 4:179

4.10.8

Climate and Society ............................................................................................. 4:183

4.10.9

What Is Climatology? ........................................................................................... 4:186

4.10.10

Cloud Seeding ..................................................................................................... 4:187

4.10.11

Condensation ....................................................................................................... 4:188

4.10.12

Cosmic Water ...................................................................................................... 4:189

4.10.13

The Water Cycle .................................................................................................. 4:191

4.10.14

Cyclones .............................................................................................................. 4:194

4.10.15

Water Cycle ......................................................................................................... 4:196

4.10.16

Degree Day Method ............................................................................................. 4:197

4.10.17

Desertification ...................................................................................................... 4:199

4.10.18

Dew ...................................................................................................................... 4:200

4.10.19

Dew Deserts ........................................................................................................ 4:201

4.10.20

Dew Point ............................................................................................................ 4:207

4.10.21

Droughts .............................................................................................................. 4:208

4.10.22

Drought Indices .................................................................................................... 4:209

4.10.23

The Earth Observing System: Aqua .................................................................... 4:214

4.10.24

Entropy Theory for Hydrologic Modeling .............................................................. 4:217

4.10.25

Evaporation .......................................................................................................... 4:223

4.10.26

Evapotranspiration ............................................................................................... 4:226

4.10.27

Fog ....................................................................................................................... 4:229

4.10.28

Coastal Fog along the Northern Gulf of Mexico ................................................... 4:230

4.10.29

Rain Forests ........................................................................................................ 4:239

4.10.30

Frost ..................................................................................................................... 4:240

4.10.31

Frost Damage ...................................................................................................... 4:241

4.10.32

The Global Water Cycle ....................................................................................... 4:242

4.10.33

Ground-based GPS Meteorology at FSL ............................................................. 4:244

4.10.34

Climate and Water Balance on the Island of Hawaii ............................................ 4:255

4.10.35

Heat of Vaporization ............................................................................................ 4:263

4.10.36

Hydrologic History, Problems, and Perspectives ................................................. 4:265

4.10.37

Humidity – Absolute ............................................................................................. 4:269

4.10.38

Relative Humidity ................................................................................................. 4:270

4.10.39

Hurricanes: Modeling Nature’s Fury .................................................................... 4:274 This page has been reformatted by Knovel to provide easier navigation.

lvi

Contents 4.10.40

Hydrologic Cycle .................................................................................................. 4:275

4.10.41

Hydrosphere ........................................................................................................ 4:283

4.10.42

Hydrologic Cycle, Water Resources, and Society ............................................... 4:287

4.10.43

Isohyetal Method ................................................................................................. 4:290

4.10.44

What about Meteorology? .................................................................................... 4:292

4.10.45

Basic Research for Military Applications .............................................................. 4:295

4.10.46

Uncertainties in Rainfall-runoff Modeling ............................................................. 4:297

4.10.47

Monsoon .............................................................................................................. 4:304

4.10.48

African Monsoons ................................................................................................ 4:304

4.10.49

Permanent Frost .................................................................................................. 4:305

4.10.50

Radar Use in Rainfall Measurements .................................................................. 4:306

4.10.51

Rainfall ................................................................................................................. 4:309

4.10.52

Rainfall and Runoff .............................................................................................. 4:315

4.10.53

Remote Sensing of Applications in Hydrology ..................................................... 4:319

4.10.54

Atmospheric Scientists ......................................................................................... 4:328

4.10.55

Rain Simulators .................................................................................................... 4:330

4.10.56

Snow Density ....................................................................................................... 4:333

4.10.57

Flattop Mountain Snotel Snowpack: Water Year 2004 ........................................ 4:334

4.10.58

Snow and Snowmelt ............................................................................................ 4:336

4.10.59

Snow Surveys ...................................................................................................... 4:337

4.10.60

A Statistical Approach to Critical Storm Period Analysis ..................................... 4:338

4.10.61

Sublimation .......................................................................................................... 4:343

4.10.62

Transpiration ........................................................................................................ 4:345

4.10.63

Waterspout .......................................................................................................... 4:347

4.10.64

United States Weather Bureau ............................................................................ 4:347

4.10.65

Weather Forecasting through the Ages ............................................................... 4:348

4.10.66

Overview of Weather Systems ............................................................................. 4:352

4.10.67

Unit Hydrograph Theory ....................................................................................... 4:355

4.10.68

Weather and the Atmosphere .............................................................................. 4:360

4.10.69

Vapor Pressure .................................................................................................... 4:362

4.10.70

Adiabatic Cooling ................................................................................................. 4:366

4.10.71

Rain and Rocks: the Recipe for River Water Chemistry ...................................... 4:371

4.11 Physics and Chemistry of Water ........................................................................................... 4:377 4.11.1

Acid Rain and Society .......................................................................................... 4:377

4.11.2

Adsorption Capacity of Activated Carbon for Water Purification .......................... 4:381

4.11.3

Adsorption of Organic Compounds ...................................................................... 4:384

4.11.4

Age Dating Old Groundwater ............................................................................... 4:388

4.11.5

Ammonia .............................................................................................................. 4:390

4.11.6

Beryllium in Water ................................................................................................ 4:394

4.11.7

Dissolved Organic Carbon ................................................................................... 4:399 This page has been reformatted by Knovel to provide easier navigation.

Contents

lvii

4.11.8

Mechanisms of Water Adsorption on Carbons .................................................... 4:400

4.11.9

The Effect of Carbon Surface Chemical Composition on the Mechanism of Phenol Adsorption from Aqueous Solutions ........................................................ 4:404

4.11.10

Carbonate Geochemistry ..................................................................................... 4:408

4.11.11

Carbonate in Natural Waters ............................................................................... 4:413

4.11.12

Chlorine-36 and Very Old Groundwaters ............................................................. 4:416

4.11.13

Chlorofluorocarbons (CFCs) ................................................................................ 4:420

4.11.14

Coagulation and Flocculation ............................................................................... 4:424

4.11.15

Conductivity-electric ............................................................................................. 4:429

4.11.16

Conservation and the Water Cycle ...................................................................... 4:433

4.11.17

Defluroidation ....................................................................................................... 4:434

4.11.18

Deuterium ............................................................................................................ 4:438

4.11.19

Distilled Water ...................................................................................................... 4:441

4.11.20

Electricity as a Fluid ............................................................................................. 4:442

4.11.21

Analysis of Aqueous Solutions Using Electrospray Ionization Mass Spectrometry (ESI MS) ........................................................................................ 4:443

4.11.22

Fenton’s Reaction and Groundwater Remediation .............................................. 4:445

4.11.23

Where Water Floats ............................................................................................. 4:448

4.11.24

Freshwater ........................................................................................................... 4:449

4.11.25

Dissolved Gases .................................................................................................. 4:450

4.11.26

Hard Water .......................................................................................................... 4:452

4.11.27

An Analysis of the Impact of Water on Health and Aging: Is All Water the Same? ................................................................................................................. 4:455

4.11.28

Heavy Water ........................................................................................................ 4:462

4.11.29

Henry’s Law ......................................................................................................... 4:466

4.11.30

Hofmeister Effects ................................................................................................ 4:468

4.11.31

Clathrate Hydrates ............................................................................................... 4:471

4.11.32

Hydration ............................................................................................................. 4:475

4.11.33

The Mirage of the H2 Economy ............................................................................ 4:477

4.11.34

Hydrogen Ion ....................................................................................................... 4:480

4.11.35

The Hydronium Ion .............................................................................................. 4:482

4.11.36

Infiltration and Soil Moisture Processes ............................................................... 4:484

4.11.37

Ion Exchange and Inorganic Adsorption .............................................................. 4:490

4.11.38

Iron ....................................................................................................................... 4:496

4.11.39

Isotopes ............................................................................................................... 4:499

4.11.40

Isotope Fractionation ........................................................................................... 4:500

4.11.41

Mariotte Bottle – Use in Hydrology ...................................................................... 4:503

4.11.42

Mars Exploration Rover Mission .......................................................................... 4:504

4.11.43

Removal of Organic Micropollutants and Metal Ions from Aqueous Solutions by Activated Carbons ........................................................................................... 4:506 This page has been reformatted by Knovel to provide easier navigation.

lviii

Contents 4.11.44

Molecular Network Dynamics .............................................................................. 4:511

4.11.45

In situ Chemical Monitoring .................................................................................. 4:514

4.11.46

Nitrogen ............................................................................................................... 4:517

4.11.47

Osmosis-diffusion of Solvent or Caused by Diffusion of Solutes? ....................... 4:520

4.11.48

Partitioning and Bioavailability ............................................................................. 4:521

4.11.49

Physical Properties .............................................................................................. 4:527

4.11.50

Environmental Photochemistry in Surface Waters ............................................... 4:529

4.11.51

Isotope Exchange in Gas-water Reactions .......................................................... 4:535

4.11.52

Radon in Water .................................................................................................... 4:541

4.11.53

Silica in Natural Waters ........................................................................................ 4:548

4.11.54

Sodium in Natural Waters .................................................................................... 4:551

4.11.55

Soft Water ............................................................................................................ 4:553

4.11.56

Solubility of Chemicals in Water .......................................................................... 4:555

4.11.57

Solubility of Hydrocarbons in Salt Water .............................................................. 4:559

4.11.58

Solubility of Hydrocarbons and Sulfur Compounds in Water ............................... 4:561

4.11.59

Sorption Kinetics .................................................................................................. 4:564

4.11.60

Sound in Water .................................................................................................... 4:569

4.11.61

Water on the Space Station ................................................................................. 4:572

4.11.62

Strontium Isotopes in Water and Rock ................................................................ 4:574

4.11.63

Technetium in Water ............................................................................................ 4:578

4.11.64

Water – Nature’s Magician ................................................................................... 4:583

4.11.65

Freezing and Supercooling of Water ................................................................... 4:585

4.11.66

Chemical Precipitation ......................................................................................... 4:586

4.11.67

Antimony in Aquatic Systems .............................................................................. 4:589

4.12 Water Law and Economics .................................................................................................... 4:595 4.12.1

The Clean Water Act ........................................................................................... 4:595

4.12.2

Clean Water Act, Water Quality Criteria/Standards, TMDLs, and Weight-ofevidence Approach for Regulating Water Quality ................................................ 4:598

4.12.3

The Constitution and Early Attempts at Rational Water Planning ........................ 4:604

4.12.4

Economic Value of Water: Estimation .................................................................. 4:605

4.12.5

Water Supply Planning – Federal ........................................................................ 4:612

4.12.6

Flood Control Act of 1944 .................................................................................... 4:616

4.12.7

Great Lakes Governors’ Agreement .................................................................... 4:617

4.12.8

U.S./Canadian Boundary Waters Treaty and the Great Lakes Water Quality Agreement ........................................................................................................... 4:620

4.12.9

Great Lakes Water Quality Initiative .................................................................... 4:621

4.12.10

Quantitative Groundwater Law ............................................................................ 4:627

4.12.11

Islamic Water Law ................................................................................................ 4:634

4.12.12

Transboundary Waters in Latin America: Conflicts and Collaboration ................. 4:636

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Contents

lix

4.12.13

United States-Mexico Border Waters: Conventions, Treaties, and Institutions ............................................................................................................ 4:643

4.12.14

Negotiating between Authority and Polluters: an Approach to Managing Water Quality ....................................................................................................... 4:647

4.12.15

Water Resource Organizations ............................................................................ 4:648

4.12.16

A Brief History of the Water Pollution Control Act in the U.S. .............................. 4:651

4.12.17

The National Pollution Discharge Elimination System ......................................... 4:655

4.12.18

Legal Protection for in-stream Flow ..................................................................... 4:659

4.12.19

Water Quality ....................................................................................................... 4:664

4.12.20

Interface between Federal Water Quality Regulation and State Allocation of Water Quantity ..................................................................................................... 4:671

4.12.21

Regulatory Issues and Remediation: Risk, Costs, and Benefits .......................... 4:674

4.12.22

The Safe Drinking Water Act ............................................................................... 4:676

4.12.23

Representing Geopolitics of (Hydro) Borders in South Asia ................................ 4:680

4.12.24

Water Transfers ................................................................................................... 4:685

4.12.25

Reserved Water Rights for Indian and Federal Lands ......................................... 4:689

4.12.26

Wetlands Policy in the United States: from Drainage to Restoration ................... 4:690

4.13 Water History, Art, and Culture .............................................................................................. 4:695 4.13.1

Curious Uses of Agricultural Water in the World .................................................. 4:695

4.13.2

Water between Arabs and Israelis: Researching Twice-promised Resources ............................................................................................................ 4:699

4.13.3

The Myth of Bad Cholesterol: Why Water Is a Better Cholesterol-lowering Medication ........................................................................................................... 4:701

4.13.4

Early Clocks ......................................................................................................... 4:702

4.13.5

Water Clocks ....................................................................................................... 4:704

4.13.6

Our Evolving Water Consciousness .................................................................... 4:707

4.13.7

Effective Water Education Strategies in a Nontraditional Setting ........................ 4:713

4.13.8

Evolution .............................................................................................................. 4:715

4.13.9

History of Pond Fisheries in Poland ..................................................................... 4:718

4.13.10

Water: the Key to Natural Health and Healing ..................................................... 4:722

4.13.11

Water in History ................................................................................................... 4:726

4.13.12

Hoover Dam History ............................................................................................. 4:732

4.13.13

Hydropsychology ................................................................................................. 4:733

4.13.14

Jacob’s Well ......................................................................................................... 4:735

4.13.15

Jaubert de Passa: the First World History of Irrigation in 1846 ............................ 4:736

4.13.16

Benjamin Franklin: from Kite to Lightning Rod ..................................................... 4:741

4.13.17

Water and the History of Man .............................................................................. 4:745

4.13.18

Water, Bacteria, Life on Mars, and Microbial Diversity ........................................ 4:746

4.13.19

Canals in the Mekong Delta: a Historical Overview from 200 C.E. to the Present ................................................................................................................ 4:748

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lx

Contents 4.13.20

Conflict and Water Use in the Middle East .......................................................... 4:753

4.13.21

Benjamin Franklin’s Armonica: a Water Music Instrument .................................. 4:758

4.13.22

Water Transportation Occupations ...................................................................... 4:762

4.13.23

A Concise Glimpse of Water in the History of Photography ................................ 4:766

4.13.24

Water as a Human Right ...................................................................................... 4:769

4.13.25

Ancient Water and Soil Conservation Ecosystems of Sri Lanka .......................... 4:772

4.13.26

Ben Franklin’s Gulf Stream Weather and Swim Fins ........................................... 4:780

4.13.27

Water Symbolism ................................................................................................. 4:785

4.13.28

Gordon and Franklin Rivers and the Tasmanian Wilderness World Heritage Area ..................................................................................................................... 4:788

4.13.29

The Mistake of Waiting to Get Thirsty .................................................................. 4:789

4.13.30

Water and Well-being .......................................................................................... 4:791

4.13.31

Free Flowing Water: a Source of Wisdom ........................................................... 4:793

4.13.32

The Medicinal Properties of the Waters of Saratoga Springs .............................. 4:797

4.13.33

A History of Hawaiian Freshwater Resources ..................................................... 4:801

Volume 5. Ground Water 5.14 Ground Water ........................................................................................................................

5:1

5.14.1

Acid Mine Drainage: Sources and Treatment in the United States ......................

5:1

5.14.2

Aquifers ................................................................................................................

5:9

5.14.3

Artificial Recharge of Unconfined Aquifer ............................................................

5:11

5.14.4

Groundwater and Arsenic: Chemical Behavior and Treatment ............................

5:17

5.14.5

Treatment of Arsenic, Chromium, and Biofouling in Water Supply Wells ............

5:22

5.14.6

Artesian Water .....................................................................................................

5:29

5.14.7

Modeling Contaminant Transport and Biodegradation in Groundwater ...............

5:30

5.14.8

Biofouling in Water Wells .....................................................................................

5:35

5.14.9

In situ Bioremediation of Contaminated Groundwater .........................................

5:38

5.14.10

Process Limitations of in situ Bioremediation of Groundwater .............................

5:42

5.14.11

Black Mesa Monitoring Program ..........................................................................

5:48

5.14.12

Brine Deposits .....................................................................................................

5:51

5.14.13

Connate Water .....................................................................................................

5:54

5.14.14

Consolidated Water Bearing Rocks .....................................................................

5:55

5.14.15

Sensitivity of Groundwater to Contamination .......................................................

5:56

5.14.16

Water Contamination by Low Level Organic Waste Compounds in the Hydrologic System ...............................................................................................

5:60

5.14.17

Darcy’s Law .........................................................................................................

5:63

5.14.18

Groundwater Dating with Radiocarbon ................................................................

5:64

5.14.19

Groundwater Dating with H–He ...........................................................................

5:65

5.14.20

Dating Groundwaters with Tritium ........................................................................

5:69

5.14.21

Recharge in Desert Regions around the World ...................................................

5:72

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Contents

lxi

5.14.22

Hydrologic Feasibility Assessment and Design in Phytoremediation ..................

5:76

5.14.23

Well Design and Construction ..............................................................................

5:87

5.14.24

Physical Properties of DNAPLs and Groundwater Contamination ......................

5:91

5.14.25

Water Dowsing (Witching) ...................................................................................

5:92

5.14.26

Subsurface Drainage ...........................................................................................

5:94

5.14.27

Drawdown ............................................................................................................ 5:101

5.14.28

Water Level Drawdown ........................................................................................ 5:102

5.14.29

Water Well Drilling Techniques ............................................................................ 5:105

5.14.30

Groundwater Dye Tracing in Karst ....................................................................... 5:107

5.14.31

Earthquakes – Rattling the Earth’s Plumbing System ......................................... 5:111

5.14.32

In situ Electrokinetic Treatment of MtBE, Benzene, and Chlorinated Solvents ............................................................................................................... 5:116

5.14.33

Field Capacity ...................................................................................................... 5:124

5.14.34

Groundwater Flow Properties .............................................................................. 5:128

5.14.35

Fluoride Contamination in Ground Water ............................................................ 5:130

5.14.36

Rock Fracture ...................................................................................................... 5:136

5.14.37

Geochemical Models ........................................................................................... 5:138

5.14.38

Geochemical Modeling – Computer Codes ......................................................... 5:140

5.14.39

Geochemical Modeling – Computer Code Concepts ........................................... 5:142

5.14.40

Geological Occurrence of Groundwater ............................................................... 5:145

5.14.41

Geophysics and Remote Sensing ........................................................................ 5:145

5.14.42

Geothermal Water ............................................................................................... 5:156

5.14.43

Ghijben-Herzberg Equilibrium .............................................................................. 5:158

5.14.44

Groundwater Balance .......................................................................................... 5:162

5.14.45

Hydraulic Head .................................................................................................... 5:169

5.14.46

The Role of Heat in Groundwater Systems ......................................................... 5:172

5.14.47

Groundwater Flow in Heterogenetic Sediments and Fractured Rock Systems ............................................................................................................... 5:175

5.14.48

Horizontal Wells ................................................................................................... 5:177

5.14.49

Horizontal Wells in Groundwater Remediation .................................................... 5:178

5.14.50

Head .................................................................................................................... 5:180

5.14.51

Well Hydraulics and Aquifer Tests ....................................................................... 5:182

5.14.52

Hydraulic Properties Characterization ................................................................. 5:184

5.14.53

Mobility of Humic Substances in Groundwater .................................................... 5:188

5.14.54

Assessment of Groundwater Quality in District Hardwar, Uttaranchal, India ....... 5:192

5.14.55

Irrigation Water Quality in District Hardwar, Uttaranchal, India ............................ 5:204

5.14.56

Infiltration and Soil Water Processes ................................................................... 5:210

5.14.57

Infiltration/Capacity/Rates .................................................................................... 5:212

5.14.58

Infiltrometers ........................................................................................................ 5:214

5.14.59

Summary of Isotopes in Contaminant Hydrogeology ........................................... 5:216 This page has been reformatted by Knovel to provide easier navigation.

lxii

Contents 5.14.60

Environmental Isotopes in Hydrogeology ............................................................ 5:227

5.14.61

Water-jetting Drilling Technologies for Well Installation and in situ Remediation of Hydrocarbons, Solvents, and Metals .......................................... 5:234

5.14.62

Karst Hydrology ................................................................................................... 5:235

5.14.63

Karst Topography ................................................................................................ 5:243

5.14.64

Detecting Modern Groundwaters with 85Kr .......................................................... 5:248

5.14.65

Land Use Impacts on Groundwater Quality ......................................................... 5:250

5.14.66

Groundwater Contamination from Municipal Landfills in the USA ....................... 5:253

5.14.67

Metal Organic Interactions in Subtitle D Landfill Leachates and Associated Ground Waters .................................................................................................... 5:258

5.14.68

Leaching .............................................................................................................. 5:260

5.14.69

Well Maintenance ................................................................................................ 5:263

5.14.70

Megawatersheds ................................................................................................. 5:266

5.14.71

Mass Transport in Saturated Media ..................................................................... 5:273

5.14.72

Soil and Water Contamination by Heavy Metals .................................................. 5:275

5.14.73

Source, Mobility, and Remediation of Metals ....................................................... 5:280

5.14.74

Genetics of Metal Tolerance and Accumulation in Higher Plants ........................ 5:284

5.14.75

In situ Groundwater Remediation for Heavy Metal Contamination ...................... 5:290

5.14.76

Methane in Groundwater ..................................................................................... 5:293

5.14.77

Fossil Aquifers ..................................................................................................... 5:294

5.14.78

What is a Hydrochemical Model? ........................................................................ 5:295

5.14.79

Modeling Non-point Source Pollutants in the Vadose Zone Using GIS ............... 5:299

5.14.80

Modeling Techniques for Solute Transport in Groundwater ................................ 5:305

5.14.81

Ambient Groundwater Monitoring Network Strategies and Design ...................... 5:313

5.14.82

MTBE ................................................................................................................... 5:318

5.14.83

Limiting Geochemical Factors in Remediation Using Monitored Natural Attenuation and Enhanced Bioremediation .......................................................... 5:319

5.14.84

Nitrate Contamination of Groundwater ................................................................ 5:322

5.14.85

Treatment for Nitrates in Groundwater ................................................................ 5:323

5.14.86

Nonpoint Sources ................................................................................................ 5:331

5.14.87

Organic Compounds in Ground Water ................................................................. 5:337

5.14.88

Overdraft .............................................................................................................. 5:340

5.14.89

Chemical Oxidation Technologies for Groundwater Remediation ....................... 5:344

5.14.90

Particulate Transport in Groundwater – Bacteria and Colloids ............................ 5:349

5.14.91

Perched Groundwater .......................................................................................... 5:352

5.14.92

Permeability ......................................................................................................... 5:355

5.14.93

Groundwater Vulnerability to Pesticides: an Overview of Approaches and Methods of Evaluation ......................................................................................... 5:357

5.14.94

High pH Groundwater – the Effect of the Dissolution of Hardened Cement Pastes .................................................................................................................. 5:362 This page has been reformatted by Knovel to provide easier navigation.

Contents

lxiii

5.14.95

Phytoextraction and Phytostabilization: Technical, Economic and Regulatory Considerations of the Soil-lead Issue .................................................................. 5:365

5.14.96

Phytoextraction of Zinc and Cadmium from Soils Using Hyperaccumulator Plants ................................................................................................................... 5:369

5.14.97

Phytoremediation Enhancement of Natural Attenuation Processes .................... 5:374

5.14.98

Bacteria Role in the Phytoremediation of Heavy Metals ...................................... 5:376

5.14.99

Phytoremediation of Lead-contaminated Soils .................................................... 5:381

5.14.100

Phytoremediation of Methyl Tertiary-butyl Ether .................................................. 5:385

5.14.101

Phytoremediation of Selenium-laden Soils .......................................................... 5:397

5.14.102

Soil Pipes and Pipe Flow ..................................................................................... 5:401

5.14.103

Low Flow Groundwater Purging and Surging ...................................................... 5:404

5.14.104

Groundwater Quality ............................................................................................ 5:406

5.14.105

Radial Wells ......................................................................................................... 5:407

5.14.106

Recharge in Arid Regions .................................................................................... 5:408

5.14.107

Sub-surface Redox Chemistry: a Comparison of Equilibrium and Reactionbased Approaches ............................................................................................... 5:413

5.14.108

Regional Flow Systems ....................................................................................... 5:417

5.14.109

Groundwater Remediation by Injection and Problem Prevention ........................ 5:421

5.14.110

Groundwater Remediation: in situ Passive Methods ........................................... 5:423

5.14.111

Groundwater Remediation by in situ Aeration and Volatilization ......................... 5:426

5.14.112

Remediation of Contaminated Soils ..................................................................... 5:432

5.14.113

Groundwater Remediation Project Life Cycle ...................................................... 5:436

5.14.114

Innovative Contaminated Groundwater Remediation Technologies .................... 5:438

5.14.115

Resistivity Methods .............................................................................................. 5:443

5.14.116

Risk Analysis of Buried Wastes from Electricity Generation ................................ 5:448

5.14.117

Groundwater Contamination from Runoff ............................................................ 5:451

5.14.118

Saline Seep ......................................................................................................... 5:453

5.14.119

Groundwater Sampling Techniques for Environmental Projects ......................... 5:454

5.14.120

Groundwater Sampling with Passive Diffusion Samplers .................................... 5:456

5.14.121

Specific Capacity ................................................................................................. 5:460

5.14.122

Soil Water ............................................................................................................ 5:461

5.14.123

Soil and Groundwater Geochemistry and Microbiology ....................................... 5:463

5.14.124

Characterizing Soil Spatial Variability .................................................................. 5:465

5.14.125

Deep Soil-water Movement .................................................................................. 5:471

5.14.126

Specific Gravity .................................................................................................... 5:473

5.14.127

Hot Springs .......................................................................................................... 5:475

5.14.128

Squeezing Water from Rock ................................................................................ 5:477

5.14.129

Storage Coefficient .............................................................................................. 5:480

5.14.130

Qanats: an Ingenious Sustainable Groundwater Resource System .................... 5:483

5.14.131

Lysimeters ........................................................................................................... 5:487 This page has been reformatted by Knovel to provide easier navigation.

lxiv

Contents 5.14.132

Steady-state Flow Aquifer Tests .......................................................................... 5:491

5.14.133

Tidal Efficiency ..................................................................................................... 5:497

5.14.134

Combined Free and Porous Flow in the Subsurface ........................................... 5:498

5.14.135

Groundwater Tracing ........................................................................................... 5:501

5.14.136

Hydraulic Conductivity/Transmissibility ................................................................ 5:507

5.14.137

Groundwater Flow and Transport Process .......................................................... 5:514

5.14.138

Reactive Transport in the Saturated Zone: Case Histories for Permeable Reactive Barriers ................................................................................................. 5:518

5.14.139

Transport of Reactive Solute in Soil and Groundwater ........................................ 5:524

5.14.140

Water in the Unsaturated Zone ............................................................................ 5:531

5.14.141

Groundwater and Vadose Zone Hydrology .......................................................... 5:533

5.14.142

Vadose Zone Monitoring Techniques .................................................................. 5:538

5.14.143

Vapor Transport in the Unsaturated Zone ........................................................... 5:543

5.14.144

Applications of Soil Vapor Data to Groundwater Investigations ........................... 5:548

5.14.145

Groundwater Velocities ........................................................................................ 5:554

5.14.146

Viscous Flow ........................................................................................................ 5:555

5.14.147

Vulnerability Mapping of Groundwater Resources ............................................... 5:561

5.14.148

Water/Rocks Interaction ...................................................................................... 5:566

5.14.149

Ground Water: Wells ............................................................................................ 5:571

5.14.150

Well Screens ........................................................................................................ 5:572

5.14.151

Well TEST ............................................................................................................ 5:574

5.14.152

Safe Yield of an Aquifer ....................................................................................... 5:575

5.14.153

Specific Yield Storage Equation ........................................................................... 5:576

5.14.154

Microbial Processes Affecting Monitored Natural Attenuation of Contaminants in the Subsurface .......................................................................... 5:578

5.14.155

Groundwater Vulnerability to Pesticides: Statistical Approaches ......................... 5:594

5.14.156

Groundwater – Nature’s Hidden Treasure ........................................................... 5:599

5.14.157

Pharmaceuticals, Hormones, and Other Organic Wastewater Contaminants in U.S. Streams .................................................................................................... 5:605

5.14.158

The Environmental Impact of Iron in Groundwater .............................................. 5:608

5.14.159

Groundwater and Cobalt: Chemical Behavior and Treatment ............................. 5:610

5.14.160

Groundwater and Cadmium: Chemical Behavior and Treatment ........................ 5:613

5.14.161

Groundwater Modeling ........................................................................................ 5:619

5.14.162

Groundwater and Benzene: Chemical Behavior and Treatment ......................... 5:626

5.14.163

Groundwater and Nitrate: Chemical Behavior and Treatment ............................. 5:628

5.14.164

Groundwater and Perchlorate: Chemical Behavior and Treatment ..................... 5:631

5.14.165

Groundwater and Vinyl Chloride: Chemical Behavior and Treatment ................. 5:634

5.14.166

Groundwater and Uranium: Chemical Behavior and Treatment .......................... 5:640

5.14.167

Groundwater and Mercury: Chemical Behavior and Treatment ........................... 5:642

5.14.168

Groundwater and Lead: Chemical Behavior and Treatment ................................ 5:645 This page has been reformatted by Knovel to provide easier navigation.

Contents

lxv

5.14.169

Laminar Flow ....................................................................................................... 5:649

5.14.170

Finite Element Modeling of Coupled Free and Porous Flow ................................ 5:655

5.14.171

Unconfined Groundwater ..................................................................................... 5:662

5.14.172

Modeling of DNAPL Migration in Saturated Porous Media .................................. 5:668

5.14.173

The Use of Semipermeable Membrane Devices (SPMDs) for Monitoring, Exposure, and Toxicity Assessment .................................................................... 5:672

5.14.174

River-connected Aquifers: Geophysics, Stratigraphy, Hydrogeology, and Geochemistry ....................................................................................................... 5:677

Index ..................................................................................................................................

This page has been reformatted by Knovel to provide easier navigation.

I:1

DOMESTIC WATER SUPPLY THE ARSENIC DRINKING WATER CRISIS IN BANGLADESH

NEPAL88 26

90 Rangpur

CHARLIE BRYCE JIM PHILP

0

50 100 km

0

50

100 km

26

Brahma p u t r a INDIA

Ja m u na

Napier University Edinburgh, Scotland, United Kingdom

Sylhet

G

Rajshahi an ge s

Mymensingh

24

INTRODUCTION

24

Chittagong

22

s

22

a

Jessore Khulna Barisal Mungla

Megh n

DHAKA Comilla Narayanganj

INDIA

Although the incidence of arsenic poisoning in groundwater is worldwide and includes Bangladesh and India, Taiwan, Vietnam, Chile, China, North America, and Finland, the area of the highest demand for a resolution of the problem is Bangladesh. The source seems to be geological, for arsenic has been found in tube well water used for drinking and irrigation, although the geochemistry is not completely understood. As many thousands of boreholes have been produced to support modern irrigation systems, the underground aquifers are aerated, which causes transformation of anaerobic conditions to aerobic conditions. The presence of oxygen in this way decomposes arsenopyrite-releasing arsenic acid. At low pH, this arsenic dissolves in water and hence leads to water contamination. The arsenic content of sediments is high relative to crustal concentrations. The biogeochemical cycling of arsenic and iron are coupled in deltaic systems; iron oxyhydroxides act as a carrier for the deposition of arsenic in sediments. From there, it can be mobilized by bicarbonate, which can extract arsenic from sediments under both aerobic and anaerobic conditions. Arsenic also becomes a pollutant as a result of various industrial uses and activities. Arsenic is a metalloid, and its primary usage has been in agriculture, in formulating herbicides, especially for controlling weeds in cotton fields. Sodium arsenite has been used as an insecticidal ingredient in sheep-dips. In industry, arsenic has found use in glass manufacture and a new role in the semiconductor industry. Copper smelting releases significant amounts into soils.

Tungi

a he G

New Moore Island 88

of t M o uths Bay of Bengal 90

ng

e

Keokradong Cox's Bazar

BURMA

92

Figure 1. The geography of Bangladesh.

Bangladeshi districts that have the same geographical continuity and aquifers as the West Bengal districts; this yielded the result that slightly more than 20% of the samples contained arsenic at levels ranging from 0.01 to 0.4 mg/L. Ten million people populate these areas and hence are at risk of arsenic toxicity. Since that time, it has been shown that there is groundwater contamination in more than 40 districts that endanger in excess of 50 million people. The problem has been described in The Lancet as the world’s worst episode of arsenic poisoning; more than 220,000 people reportedly suffer from arsenic-related diseases. In a recent study of 27 districts in Bangladesh, 58% of the water samples were unsuitable for drinking. The worst case was in Nawabganj district, where one well contained 60 times the WHO maximum permissible level. TOXICITY AND DISEASE

ARSENIC AND THE GEOGRAPHY OF BANGLADESH Arsenic occurs principally in the forms of organic arsenic (methyl arsonic acid, dimethyl arsonic acid, arsenobetaine, and arsenocholine) and inorganic arsenic (trivalent and pentavalent arsenic). Of these, the trivalent form is the most toxic to humans (20 times more so than the pentavalent form) and is the most difficult to remove chemically from water. Arsenic is a suspected carcinogen and has many acute effects on human health. But at the concentrations present in drinking water, it has no immediate side affects. The latency (i.e., the time from first exposure to manifestation of disease) for arsenic-caused skin lesions, in particular keratoses, is typically of the order of 10 years, and so a major increase in the number of cases of arsenic-caused diseases can be projected into the future (Fig. 2).

Bangladesh, 85% of which is deltaic and alluvial plain, is situated in the lower end of three large river systems, the Ganges, the Brahmaputra, and the Meghna, whose catchment area is about 600,000 square miles (Fig. 1). The sediments produced in the catchment areas are very high and expose the underlying rocks, including arsenicbearing rocks. Arsenic pollution became a live issue in Bangladesh as recently as 1993, following a warning by the World Health Organisation (WHO) that levels of arsenic in groundwater above the permissible limit of 0.05 mg/L had been reported in seven districts of adjoining West Bengal in India. The Department of Public Health Engineering of Bangladesh was invited to test water samples from the adjoining eight 1

2

THE ARSENIC DRINKING WATER CRISIS IN BANGLADESH

test kit is being used (Fig. 3). The disadvantage of this approach is that the sensitivity of the chemistry (poor below 100 µg/L) is not compatible with the levels of contamination that need to be detected (50 µg/L and less). Often, at best, the presence or absence of arsenic can be inferred, but not the level of contamination. In addition, such testing is slow and can take about 6 months to cover some 2000 villages in a district. Bioavailability Biosensors for Detecting and Quantifying Arsenic

Figure 2. Common manifestations of long-term, chronic arsenic poisoning.

Exposure to arsenic in this way can lead to latent or manifest clinical symptoms through even low-level exposure over a period of time. This can result in an accumulation of this toxicant in various organs and systems, affecting their normal functioning, including the kidney and nervous system. Arsenic causes skin cancers and internal cancers such as lung and bladder cancer. The most common manifestations in afflicted people in Bangladesh are melanosis (93.5%), keratoses (68.3%), hyperkeratosis (37.6%), and dipigmentation (leucomelanosis) (39.1%). Cancers are found in 0.8% of the afflicted population. Preliminary work indicates that there may be several factors triggering arsenic-related diseases, but experts generally feel that poor nutrition may be a primary cause. Studies in Taiwan have shown that there is an increased occurrence of diabetes in the population exposed to arsenic via drinking water. Recent studies have shown that arsenic is also a teratogen. Further, at the 5th International Conference on Arsenic held in Dhaka, 2004, one of the key messages and cause for increased concern is that there is very good evidence that the environmental contaminant is getting into the food chain, thus putting even more lives at risk. DRINKING WATER STANDARDS The World Health Organisation has set 10 µg/L as the allowable level for arsenic in drinking water. On January 22, 2001, the U.S. EPA adopted this standard, and public water systems must comply by January 23, 2006.

A few strains of bacteria are resistant to arsenate, arsenite, and antimonite through the action of the gene products of the ars operon. The ars operon consists of five genes that code for three structural and two regulatory proteins. Two structural genes in the ars operon, arsA and arsB, code for proteins that form an efflux pump that transports arsenite and antimonite out of cells. A means of measuring available arsenicals would be to construct a gene fusion plasmid in which part of the ars operon is fused upstream of a reporter gene system, such as the bacterial lux operon, which results in the production of light. A transcriptional gene fusion has been done (1) that consists of E. coli arsB :: luxAB. The detection limit of arsenic is of the order of 10 µg/L. Moreover, bioluminescence may be inducible in a concentration-dependent manner (Fig. 4) by arsenic salts; high concentrations result in higher bioluminescence, so that such biosensors may be able to • Quantify arsenic within the required range of drinking water in Bangladesh • Provide a measure of the bioavailability of the arsenic for risk assessment REMOVAL OF ARSENIC FROM DRINKING WATER Coprecipitation of arsenate with ferric (Fe3+ ) ion is currently the most effective and practical method of arsenic removal. Optimum stability of the FeAsO4 precipitate occurs at Fe/As molar ratios of >4; this ratio increases significantly, in practice, depending on water turbidity, slime levels, dissolved solids, and the presence of

DETECTING ARSENIC IN DRINKING WATER In the modern analytical laboratory, arsenic is quantified by soluble arsenic assaying, preferable with GF-AAS (graphite furnace-atomic absorption spectrometry) for detection levels of less than 50 µg/L. However, given the highly dispersed nature of tube wells in Bangladesh, the transport of the many samples to central laboratories is logistically impossible. Field techniques are more important, so that samples can be processed as they are taken. In Bangladesh, this requires inexpensive and completely portable techniques. At present, a chemical

Figure 3. Testing for arsenic in drinking water using test kits at a village.

BOTTLED WATER

Relative light units

400 350

5.0 µM 10 µM 2.5 µM

300

1.25 µM

3

250 200

0.625 µM

150 100 50 0

0

20

40

60

80

100

120

140

160

180

200

0 220

Time, minutes

iron-consuming species. However, Fe3+ ion coprecipitation of arsenite (AsO3 3− ) is moderately effective at pH ∼7.0. The trivalent As(III) species must be oxidized to As(V) for complete precipitation with Fe3+ ion. Oxidation may be achieved through aeration or by adding oxidizers such as hypochlorite, permanganate, peroxide, and ozone. The application of other technologies, including alum and lime precipitation together with activated alumina adsorption, are not fully effective. In Bangladesh, for geographical and financial reasons, there is likely to be a preference for local treatment rather than large-scale treatment plants. The ideal solution would be to modify each tube well at low cost for arsenic removal by, for example, ion exchange.

SOME OTHER ARSENIC LINKS Harvard University Arsenic Project Website Natural Resources Defense Council—FAQs: Arsenic in drinking water U.S. Agency for Toxic Substances and Disease Registry— ToxFAQs: Arsenic U.S. Environmental Protection Agency—Arsenic Standard pages and Q & A’s: Occurrence of Arsenic in Ground Water West Bengal and Bangladesh Arsenic Crisis Information Centre World Health Organisation http://www.who.int/mediacentre/ factsheets/fs210/en/

BOTTLED WATER

BIBLIOGRAPHY 1. Cai, J. and DuBow, M.S. (1997). Use of a bioluminescent bacterial biosensor for biomonitoring and characterization of arsenic toxicity of chromated copper arsenate (CCA). Biodegradation 8(2): 105–111.

READING LIST Akhtar, S.A. et al. (1997). Arsenic contamination in ground water and arsenicosis in Bangladesh. Int. J. Environ. Health Res. 7: 71–276. Anawar, H.M., Akai, J., and Sakugawa, H. (2004). Mobilization of arsenic from subsurface sediments by effect of biocarbonate ions in groundwater. Chemosphere 54: 753–762. Dave, J.M. (1997). Arsenic contamination of drinking water in Bangladesh. WHO SEA/EH/Meet.3/6.17. Dipankar, D. et al. (1996). Arsenic in groundwater in six districts of West Bengal, India. Environ. Geochem. Health 18: 5–15. Karim, M.D.M. (2000). Arsenic in groundwater and health problems in Bangladesh. Water Res. 34: 304–310. Tareq, S.M., Safiullah, S., Anawar, H.M., Majibur Rahman, M., and Ishizuka, T. (2003). Arsenic pollution in groundwater: a self-organizing complex geochemical process in the deltaic sedimentary environment, Bangladesh. Sci. Total Environ. 313: 213–226. Wadud Khan, A. et al. (1997). Arsenic contamination in ground water and its effect on human health with particular reference to Bangladesh. J. Preventative Soc. Med. 16: 65–73.

Figure 4. The response of a bacterial whole-cell bioavailability biosensor to increasing concentrations of arsenic.

ROBERT M. HORDON Rutgers University Piscataway, New Jersey

INTRODUCTION Bottled water sales in the United States have increased dramatically during the past decade. Total domestic and imported sales rose 142% from almost 2.5 billion gallons (9.4 billion liters) in 1992 to more than 6 billion gallons (22.8 billion liters) in 2002. Bottled water revenues rose nearly 190% from $2.66 billion in 1992 to $7.7 billion in 2002. Using the same 11-year period from 1992 to 2002 (1), per capita consumption in the U.S. increased more than 119% from 9.8 to 21.5 gallons (37.1 to 81.4 liters). The global water market also shows comparable increases. For example, the world total consumption of bottled water rose nearly 63% from 21.3 billion gallons (80.6 billion liters) in 1997 to 34.7 billion gallons (131.3 billion liters) in 2002. Using the same 6-year period from 1997 to 2002, global per capita consumption increased 107% from 5.7 to 11.8 gallons (21.6 to 44.7 liters). Italy, Mexico, and France were the top three countries in per capita consumption in 2002 at 44.2, 37.7, and 37.1 gallons (167.3, 142.7, and 140.4 liters), respectively. In per capita consumption, the United States rose from a rank of 15 in 1997 to a rank of 11 in 2002 (1).

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BOTTLED WATER

The rising consumption in the United States is attributed to an increasingly effective advertising campaign that touts bottled water as a safer and better tasting alternative to tap water. Packaging labels that show massive glaciers and springs in alpine settings have also helped. HISTORICAL USES OF BOTTLED OR MINERAL WATER The presumed reason for drinking mineral water and more recently bottled water is for the purported therapeutic effects and associated health benefits. Using mineral waters as a form of therapy was discussed by the ancient Egyptians and Greeks. Based on legend, the thermal springs of Bath, England, date to 800 B.C. Hannibal’s army (and elephants) was reputed to have refreshed itself in the pools of Vergeze in southern France in 218 B.C. while enroute to attack Rome. Later on, the spring waters at Vergeze became known as the source of Perrier Water. The mineral waters from 12 springs at Vichy, France, date back to Roman times. Spring water from Fiuggi near Rome, Italy, was used by Michaelangelo. Another Italian luminary from the fifteenth century, Leonardo da Vinci, used the mineral waters from the three deep springs (396 m; 1300 ft) at San Pellegrino. Famous spas and watering holes, such as Hot Springs in Arkansas and Saratoga Springs in New York and many others in Europe, developed near mineral springs during the late 1800s and early 1900s (2). Poland Spring water from Maine began to be distributed during the mid-1800s. Bottled water from Mountain Valley near Hot Springs, Arkansas, goes back to 1871. Commercial bottling of San Pellegrino water from Italy began in 1899, and export of Evian water from France to the United States began in 1905 (2). TYPES OF BOTTLED WATER Bottled water can be grouped into the following several categories depending upon the nature of the water and its source. Nonsparkling Water This includes spring water, artesian water, mineral water, and purified water. Domestic production of this type of water in the United States made up more than 95% of the bottled water market in 2002 (1). 1. Spring water is ground water. It comes from a waterbearing subsurface geologic formation known as an aquifer from which water flows naturally to the earth’s surface. Water of this type can be collected only at the spring or from a well that taps the aquifer that feeds the spring. 2. Artesian water is derived from a well in an aquifer that is under pressure due to overlying confining layers. Artesian or confined well water can be collected with external pumps that supplement the natural underground pressure. The word ‘‘artesian’’ was derived from the first deep wells that were

drilled into confined aquifers in the province of Artois in northern France from about 1750 (3). 3. Mineral water naturally contains at least 250 ppm of mineral salts such as calcium, chloride, sulfate, carbonate, and bicarbonate. No minerals can be added artificially, and it cannot come from a municipal source. 4. Purified (or demineralized) water may come from a municipal source and is treated by one or more of the following water treatment processes: a) Distillation: heating of water to produce water vapor which is then condensed and collected; b) reverse osmosis, where water is filtered by passing it though a membrane; c) deionization: a process where minerals are drawn to particles of the opposite electrical charge and then removed. Sparkling Water Sparkling water may include any type of naturally carbonated water. In addition, if the water is treated, CO2 can be added to the product as long as the water has the same amount of CO2 as it had when it emerged from its source. Domestic and imported sparkling water made up about 2.6% and 2.1% of the U.S. bottled water market in 2002, respectively (1). Beverages that contain certain ingredients or additives, such as sugar, fall into a separate category called soft drinks. Thus, tonic water, soda water, and seltzer are not considered bottled waters and are regulated differently. REGULATORY AGENCIES Public potable water supplies in the United States are regulated by the Environmental Protection Agency (EPA) under the federal Safe Drinking Water Act (SDWA). This means that all public water systems that operate under either public or private investor-owned companies that serve 25 or more people must be tested regularly for up to 118 chemicals and bacteria that are specified by the SDWA. In contrast, bottled water in the United States is regulated as a packaged food product by the federal Food and Drug Administration (FDA). By law, the FDA must follow the same water quality standards as outlined in the Safe Drinking Water Act. In addition, bottled water companies are required to comply with FDA’s quality standards, labeling rules, and good manufacturing practices. Finally, bottlers that are members of the International Bottled Water Association (IBWA) may opt to receive random, unannounced site inspections annually by a third-party organization. However, not all bottled water companies comply with the standards of the IBWA. To compound the regulatory issues associated with bottled water, the standards of the IBWA are not legally enforceable (4). Also, the results of any water quality tests that are made by the bottled water companies need not be released to the public (5). This stands in sharp contrast with the water quality reports that all United States

CORROSION CONTROL IN DRINKING WATER SYSTEMS

public water supply purveyors must furnish each year to their customers. The EPA states that bottled water is not inherently safer than tap water distributed by public water systems (6). Although tap water and bottled water must meet the standards set by their respective regulatory agencies, the FDA requires testing only once a year for bottled water, whereas the EPA requires much more frequent, often daily, testing. Another factor worth noting is that the FDA rules apply only to water sold in interstate commerce. The Natural Resources Defense Council (7) estimates that about 60% of all bottled water is sold in the same state where it is bottled; thus, FDA rules do not apply. State regulations and industry standards affect bottled water at the state level. Some states abide by the FDA standards; others are even more stringent. For example, the State of New Jersey requires that the labels of all bottled water products must contain an expiration date of 2 years from the date of bottling (8). However, about 20% of the states have either very limited enforcement powers or no regulations at all. Fluorides It is well known that fluorides are a key factor in reducing tooth decay, particularly in children. Drinking water that has been adjusted to optimal fluoride levels prevents cavities and thereby improves dental health. The consumption of fluoridated water is more effective than the use of fluoridated toothpastes or mouth rinses as the latter is only on the teeth for a short time, whereas fluoridated water can be delivered to the teeth continuously through the bloodstream and saliva (5). Most bottled water does not have an optimal fluoride level. Although some bottlers provide fluoride information on their labels, they are not required to do so. Safety of Plastic Bottles For reasons of convenience and nonbreakability, most bottled waters are sold in plastic containers that contain phthalates. Water is a universal solvent, so phthalates can be leached from a plastic bottle. It is still not known if there are any negative health effects from human exposure to phthalates. However, rodents have experienced adverse effects from the chemical in some studies. It is apparent that additional research is needed on this issue (5). COST TO THE CONSUMER One thing is very clear about the differences between bottled water and tap water: the former is orders of magnitude more expensive (240–10,000 times greater). Bottled water packaged in convenient sizes of plastic containers costs about $6.60/gallon ($1.74/liter) compared to average costs of 1/100 of a cent/gallon for municipal tap water. Is the difference warranted by taste, quality, and convenience in carrying, or does the explosive growth in bottled water sales reflect the success of mass marketing appeals to a more affluent generation that follows current fashion trends in beverage types? Time will tell, but for the

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forseeable future, bottled water sales continue to increase, even though the EPA states on its website that bottled water is not necessarily safer than regular tap water. BIBLIOGRAPHY 1. Beverage Marketing Corp. (2003). Bottled Water. New York. 2. LaMoreaux, P.E. and Tanner, J.T. (Eds.). (2001). Springs and Bottled Waters of the World: Ancient History, Source, Occurrence, Quality and Use. Springer-Verlag, New York. 3. Todd, D.K. (1980). Groundwater Hydrology, 2nd Edn. John Wiley & Sons, New York. 4. International Bottled Water Association. www.bottledwater. org. 5. Minnesota Department of Health. (2003). Bottled Water: Questions and Answers. St. Paul, MN. 6. U.S. EPA (Environmental Protection Agency). Ground Water and Drinking Water. www.epa.gov/safewater. 7. Natural Resources Defense Council. (1999). Bottled Water: Pure Drink or Pure Hype? New York, NY. 8. New Jersey Department of Health and Senior Services. (2002). Report to the New Jersey Legislature Summarizing Laboratory Test Results on the Quality of Bottled Drinking Water for the period January 1, 2001 Through December 31, 2001. Trenton, NJ.

CORROSION CONTROL IN DRINKING WATER SYSTEMS CHRISTIAN J. VOLK Indiana-American Water Richmond, Indiana

CORROSION IN DISTRIBUTION SYSTEM Corrosion Process and Corrosion Cell Corrosion can be defined as the wearing away or deterioration of a material because of chemical reaction with its environment. When iron or steel is exposed to water, rust (oxidized iron) forms (1). Water that promotes corrosion is defined as aggressive or corrosive. The corrosion processes consist of a series of electrochemical reactions occurring at the metal surface in contact with water and its constituents. Corrosion is an extremely complex chemical and electrochemical phenomenon. During oxidative reactions, local galvanic couples form on the surface of the metal, in which the metal is oxidized, while the oxidant is reduced. Each couple is a microbattery where the corrosion reaction proceeds with a flow of electric current between anodic and cathodic sites on the metal (Fig. 1). The electrochemical corrosion corresponds to the destruction of a metal by electron transfer reaction. All the components of an electrochemical cell must be present for this type of corrosion to occur. The components include an anode and a cathode (which are sites that have a different electrical potential on the metal), an electrical path between the anode and cathode for electron transport (internal circuit), an electrolyte solution that will conduct ions between the anode and cathode (external circuit), and an oxidizing agent to

6

CORROSION CONTROL IN DRINKING WATER SYSTEMS Water

Me+

Me+

OH−

e− + 1/4 O2 + 1/2 H2O Cathode

e− Pipe

Anode

OH−

Figure 1. Pipe corrosion in water.

be reduced at the cathode (2). As metallic plumbing materials are not completely homogeneous, anodic and cathodic sites occur on the pipe surface. Oxidation and dissolution of the metal occur at the anode (Fig. 1). Electrons generated at the anode migrate to the cathode, where they are accepted by an electron acceptor, such as oxygen, after chemical reduction. Consequently, the positive ions generated at the anode migrate through the solution to the cathode and negative ions generated at the cathode tend to migrate to the anode. The mechanisms of corrosion are extremely complex and depend on the interactions of physical and chemical factors and the material itself. Table 1 shows different materials, their use in distribution systems, and corrosion-associated problems. Generally, inert and nonmetallic materials like concrete or plastic are more corrosion resistant than metallic pipes. Scale Formation The formation of a scale on the pipe surface protects the pipe from corrosion by separating the corrodible metal from the water. Scale is formed when the divalent metallic cations associated with hardness (calcium and magnesium) combine with other minerals contained in water and precipitate to coat the pipe wall. Scale generally includes calcium carbonate (CaCO3 ), but also

magnesium carbonate (MgCO3 ), calcium sulfate (CaSO4 ), or magnesium chloride (MgCl2 ). Water can hold a certain amount of a given chemical in solution. If more is added, it will precipitate instead of dissolve. The point at which no more chemical can be dissolved is the point of saturation. The saturation point depends on water quality, including pH, temperature, and total dissolved solids (TDS) (1). Types of Corrosion Different types of corrosion exist, which can be divided into two broad classes: uniform or localized, depending on the material to be corroded, system construction, and water characteristics. Localized corrosion resulting in pitting is produced after galvanic corrosion and concentration cell corrosion. Uniform corrosion takes place in an equal rate over the entire surface (3). The different types of corrosion are summarized in Table 2. Microbial Corrosion Bacteria adhere to the pipe surface and form a biofilm (4). Microbes can promote corrosion by creating areas with different concentrations in oxygen, hydrogen ions, minerals, and metals. These concentration differentials promote corrosion. Some microorganisms also catalyze reactions associated with the corrosion process. Iron precipitating bacteria (such as Gallionella) can convert Fe (II) to Fe (III) and influence the structure of Fe (III) precipitates (2,5). Organisms involved in the sulfur cycle in water also affect the corrosion process. Sulfate reducers have been found in tubercles under anoxic conditions. Bacteria involved in the nitrogen reaction affect the concentration of oxygen, leading to oxygen concentration cells that produce localized corrosion and pitting. Corrosion also protects bacteria from disinfection. Corrosion products offer a large surface area for microbial attachment. One corrosion product

Table 1. Different Material Types and Corrosion-Associated Problems in the Distribution Network Material

Sources

Corrosion Problems

Iron and steel

Most common material in water systems

–Buildup of corrosion products on pipe walls and release or iron oxide products

Galvanized pipe

Oldest and most common plumbing material, quality varies

–Better service in hard water –Subject to galvanic corrosion –Leaching of zinc, iron, cadmium, and lead (impurities)

Lead

Lead service lines, lead pipe joints, gaskets

–Corrodes in soft water with pHaluminum>iron>cast iron>lead>brass>copper: cathode)

Pitting corrosion

–Localized attack, pitting may occur if imperfections in the metal or regions of high stress exist –Imperfection or high-stress area is anodic and surrounded by cathodic area –Corrosion occurs rapidly at point of failure –Chloride is associated with pitting

Tuberculation

–Occurs when pitting corrosion leads to a product buildup at the anode next to a pit

Crevice corrosion

–Form of localized corrosion caused by changes in acidity, decrease in oxygen, or dissolved iron

Erosion corrosion

–Removal of protective films and the pipe metal as well –Results from high flow velocities and turbulence

Cavitation corrosion

–Type of erosion corrosion –Water pressure drops, causing water to form water vapor bubbles that collapse with an explosive effect –Removal of protective coating on metal –Occur in pump impellers, partially closed valves, and reducers

Selective leaching

–Preferential removal of one metal

Stay-current corrosion

–Localized corrosion caused by grounding of home appliances to water pipes

Concentration cell corrosion

Occurs when concentrations of aqueous species (like oxygen) differ between two parts of the metal

(goethite, α–Fe–O–OH) quickly reacts and consumes chlorine. Chlorine is consumed before it can diffuse to the core of corrosion tubercles. The microorganisms within the tubercles are not exposed to lethal concentrations of disinfectant and thus are allowed to grow in the distribution systems. Corrosion products including iron oxides are also capable of adsorbing natural organic matter (NOM) from the bulk fluid (6). NOM accumulates on the surface of corrosion products. It was found that heterotrophic microorganisms found in drinking water supply could readily extract the NOM from the corrosion products for cell growth (7,8). In highly corroded environments, biofilm microorganisms can cause many problems such as bad taste and odor, slime formation, or coliform occurrences. Consequences of Corrosion Corrosion is one of the main problems in the drinking water industry. It can affect public health, water aesthetics, and operations. Corrosive water can leach toxic metals from distribution and household plumbing systems. Lead and cadmium may occur in tap water. US EPA promulgated the lead and copper rule in 1991 in order to reduce lead and copper levels in drinking water (9). The methods to reduce lead and copper are the following: removal of these metals from the source water (if present), implementation of a corrosion control program, replacement of lead service lines, and public education. The lead and copper rule defines an action level for the tap concentration of lead and copper higher than 15 ppb and 1.3 ppm, respectively (for the 90th percentile of the samples). Periodic sampling is required to

monitor lead and copper concentrations at the customer’s tap after leaving the water stagnant in the service lines for 6–8 hours. Moreover, optimal corrosion control water parameters (pH, calcium, alkalinity, temperature, inhibitor level) are defined for the plant effluent and distributed waters. When the lead or copper concentration is above the action level, the water utility has to implement a corrosion control program (9). Copper in water can cause blue stains and a metallic taste, whereas zinc leads to a metallic taste. Corrosion of cast iron pipes can cause the formation of iron deposits called tubercles in the mains. Red water problems occur when iron is dissolved from cast iron by corrosive water. Iron stains plumbing fixtures, laundry, and makes water appear unappealing for drinking. Responding to customer complaints of colored water or bad taste is expensive in terms of money and public relations. Corrosioncaused problems that add to the cost of water include increased pumping costs after a buildup of corrosion products (tuberculation), uncontrolled scale deposit that can seriously reduce pipeline capacity, and increased resistance to flow. Aggressive water reduces the life of valves and can shorten the service life of plumbing fixtures and hot water heaters. Water leaks lead to loss of water and water pressure (1,3,10). FACTORS IMPACTING CORROSION The corrosion rates depend on many site-specific conditions and their interactions, including water characteristics and pipe conditions (2,3,10). The following section describes major factors impacting corrosion. Table 3 shows the effects of some chemicals present in water.

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CORROSION CONTROL IN DRINKING WATER SYSTEMS Table 3. Effects of Chemicals on Corrosion

Parameter

Effect on Corrosion

Hardness (measures calcium and magnesium)

Hard water is less corrosive than soft water because of the formation of a film of calcium carbonate (CaCO3 ) on the pipe wall

Chloride (Cl− )

Leads to pitting by causing the metal to stay in solution

Sulfate (SO4 2− )

Leads to pitting by causing the metal to stay in solution

Hydrogen sulfide (H2 S)

Accelerates corrosion

Ammonia (NH3 )

Increases corrosion rates

Organic Matter

Can increase or decrease corrosion rates

household electrical systems or electric railway systems leads to electrical current in water pipes. Dissolved Oxygen (DO) Corrosion rate increases with increasing DO concentrations. Oxygen is the molecule accepting the electrons given up by the corroding metal. Oxygen also reacts with soluble ferrous iron ions to form ferric iron, which precipitates and forms a tubercle. Total Dissolved Solids (TDS) TDS levels are critical because electrical flow is necessary for corrosion to occur. Corrosion rates increase with increasing concentrations of TDS because water becomes a better conductor. pH

Temperature Corrosion generally increases with temperature as temperature accelerates chemical reactions. Temperature changes the solubility constants and can favor the precipitation of different substances or transform the identities of corrosion products. These changes result in either more or less protection of the pipe surface, depending on the conditions. Temperature also affects the dissolving of CaCO3 . CaCO3 tends to precipitate and form a protective coating more readily at higher temperatures. Temperature can affect the nature of corrosion. Corrosion that can be of pitting type at cold temperatures can become uniform at hot temperatures (2,3,10). Flow/Velocity The velocity of water increases or decreases corrosion rates depending on the conditions. When water is corrosive, higher flow velocities bring dissolved oxygen to the corroding surface more rapidly. For water with protective properties or containing corrosion inhibitors, high flow velocities aid in the formation of a protective film. At low velocities, the slow movement does not aid the effective diffusion of protective ingredients to the metal surface. High velocities can lead to the erosion of pipes, especially in copper lines. Stagnant waters in water main and house plumbing have been shown to promote tuberculation accompanied with biological growth (4).

pH is a measure of hydrogen ions (H+ ). H+ is a substance accepting the electrons given up by the metal. Generally, corrosion rate decreases as pH increases. Alkalinity and Dissolved Inorganic Carbon (DIC) Alkalinity measures the ability of water to neutralize acids or bases. The corrosion rates decreases as alkalinity increases. Chlorine Residual Gaseous chlorine lowers the pH of water by reacting with water to form hypochlorous acid (HOCl), hydrogen ion, and chloride ion. This reaction makes the water more corrosive. For low alkalinity water, the problem is greater because water has less ability to resist pH changes. MEASUREMENT OF CORROSION A comprehensive corrosion control program should include several techniques to monitor corrosion, because no single technique provides all the information on corrosion. The corrosion rates are expressed in mils (1/1000 inch) per year (MPY). Corrosion rates can be determined by weight loss method or electrochemical methods. Physical observations of a pipe exposed to water can be conducted, and corrosion indexes can be determined for given water (2,3,10). Table 4 presents different methods to monitor corrosion.

Metal and Manufacturing Process

Coupon Weight Loss Method

A metal that easily gives up electrons will corrode easily. When dissimilar metals are connected together, the metal corroding easier becomes an anode, whereas the metal resistant to corrosion becomes a cathode (defined as galvanic corrosion). The anode metal will corrode rapidly, whereas the cathode is protected. Manufacturing process can also impact the durability of the pipe, especially for galvanized piping.

A metallic coupon is inserted inside a main. The coupon method determines the average corrosion rate over a period of exposure, which is accomplished by weighing the coupon before and after exposure. The corrosion rate is calculated from the weight loss, initial surface area of the coupon, and time of exposition.

Electrical Current Corrosion is accelerated when an electrical current is passed through the metal. Improper grounding of

Loop System Weight Loss This method uses a pipe loop or sections of a pipe for determining the effect of water quality on pipe material. Water flows through the loop under a continuous or intermittent flow to simulate the flow patterns of a

CORROSION CONTROL IN DRINKING WATER SYSTEMS

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Table 4. Different Methods to Assess Corrosion Levels Method Coupon weight loss

Advantages –Provides corrosion over a specific period –Economical –Coupons placed in system

Disadvantages –Takes a long time –Does not show variations occurring during test –Analytical error because of weighing –Coupon may not be representative of system material

Loop system weight loss –Actual pipe sections used –Loops placed in the system –Loop can be used to test different chemicals

–Takes a long time –Does not show variations occurring during test

Electrochemical rate measurements

–Simple and reliable –Instantaneous readings –Continuous monitoring possible –Gives uniform and pitting corrosion

–Relatively expensive –Gives corrosion rates for a particular material

Microscopic techniques

–Examination of particles in water and films on pipes –Require equipment –Determination of the morphology of corrosion products

X-ray analysis and diffraction

–Identification of the elements or class of compounds –Require equipment present in corrosion products and films with possible quantification depending of the technique

Corrosion indices

–Various indices available

household. Pipe sections are removed for weight loss measurement and inspection. Electrochemical Methods Several instruments are based on the electrochemical nature of metal corrosion in water. They are based on electrical resistance, linear polarization, or galvanic current. Electrical resistance probes measure the resistance of a thin metal probe; as corrosion causes metal to be removed from the probe, its resistance increases. Linear polarization resistance (LPR) is an electrochemical technique that measures the DC current through the metal/fluid interface, which results from polarization of one or two electrodes of a material after application of a small electrical potential. The corrosion current density, which corresponds to the current flowing in a corrosion cell per unit area, is related to the DC current. The galvanic current method measures corrosion of dissimilar alloys of metals. Radiography Methods X-ray emission spectroscopy and X-ray diffraction help characterize the corrosion scales. Microscopic Methods The deterioration of pipe surface can be evaluated using optical or scanning electron microscopes. Chemical Analysis of Water and Corrosion Indexes Water quality data can be used to calculate stability indexes or indicators of water corosivity (Table 5) (10–14). The Langelier Saturation Index (LSI) indicates whether a given water is likely to form or dissolve a protective film of calcium carbonate (11). The calcium carbonate precipitation potential (CCPP) estimates the quantity of

–Have some limitations, can be misapplied –Useful after the fact but not to predict corrosion problems

CaCO3 that can be precipitated from oversaturated water and the amount that can be dissolved by unsaturated water (12). Larson and Skold (13) studied the effects of chloride and sulfate ions on iron and proposed two indices. In general, it is very difficult to find a relationship between these indices and potential corrosion problems in the system. METHODS OF CORROSION CONTROL The complete elimination of corrosion is almost impossible. However, it is possible to reduce corrosion. As corrosion depends on both water quality and pipe characteristics, optimal corrosion control methods are site specific. Three basic approaches to control corrosion exist: (a) modify water quality so that it is less corrosive to the metal, (b) place a protective barrier between the water and pipe, and (c) use pipe material and design the system so that it is not corroded by water. Methods used to achieve corrosion control involve modifying water quality (changing pH and alkalinity), forming a calcium carbonate coating, using corrosion inhibitors, providing cathodic protection, and using a corrosion-resistant coating (3,10). pH Adjustment Adjusting the pH is one of the most common methods of corrosion control. As most metals used in the distribution system tend to dissolve more readily at lower pH (presence of H+ ), an increase in pH and alkalinity levels can reduce corrosion by reducing the solubility of metals. The optimum pH is related to water and system characteristics; it is generally above 7.0. Various chemicals can be used in corrosion control treatment (Table 6). For example, lime is commonly used to increase both pH and alkalinity. It is less expensive than the other chemical products. However, lime softening can cause severe scale problems

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CORROSION CONTROL IN DRINKING WATER SYSTEMS Table 5. Indexes of Water Corrosion

Index

Formula

Langelier Saturation Index (LSI)

LSI = pHactual − pHs pHactual : pH of water sample pHs: Theorical pH at which water is saturated with calcium carbonate LSI = 0: stable water, LSI < 0: corrosive water, calcium carbonate dissolves, LSI > 0: calcium carbonate precipitates

Calcium carbonate precipitation potential (CCPP)

CCPP = 50,000 (Talki − Talkeq ) (in mg of CaCO3 /L) Talki : initial total alkalinity Talkeq : Total alkalinity at equilibrium

Larson Index

LI1 =

Ryznar Stability Index (RSI)

2[SO4 2− ] + [Cl− ] The brackets indicate the concentration in mole/L [HCO3 − ] [Cl− ] LI2 = [HCO3 − ] RSI = pHs-pH RSI = 6.5–7: water at saturation RSI < 6.5: Scale forming water RSI > 7: Undersaturated water

Table 6. Various Chemicals Used to Adjust pH Chemical

Effect

Addition Equipment

Lime, Ca(OH2 )

Increases pH, alkalinity, and calcium levels

Caustic soda, NaOH Soda Ash, Na2 CO3 Sodium bicarbonate, NaHCO3 Sulfuric acid, H2 SO4 Carbon dioxide, CO2

Raises pH and converts CO2 to alkalinity species Increases alkalinity with moderate pH increase Increases alkalinity Lowers pH Lowers pH, converts hydroxyls to bicarbonate and carbonate

when it is not stabilized. Stabilization after softening can be accomplished by feeding carbon dioxide or sulfuric acid to decrease pH so that calcium carbonate does not precipitate in the distribution system (1). Formation of a Calcium Carbonate Coating Protective coating can be applied by controlling the chemistry of water. A common protective coating technique is to adjust the pH of water to a level just above the saturation of calcium carbonate. This process has to be closely controlled because a pH that is too low may result in corrosion and a pH too high may lead to precipitation and cause plugging of the lines. Corrosion Inhibitors Some waters do not contain enough calcium or alkalinity to lead to the formation of a coating. Corrosion inhibitors used in potable water act by forming a protective scale over anodic or cathodic sites. These films are commonly inorganic precipitates containing the ions added as inhibitors. They provide a barrier between the water and the pipe. Chemical inhibitors include phosphates and silicates. The success of an inhibitor in controlling corrosion depends on three requirements. First, when treatment is initiated by adding two to three times the normal concentration of inhibitor to build up a protective film rapidly. Several weeks are required for the film to develop. Second, the inhibitor must be fed continuously at a proper concentration. An interruption in chemical addition can lead to

Dry storage, gravimetric or volumetric dry feeders, slurry feed 50% solution, metering pumps Dry storage with solution feed Dry storage with solution feed Metering pumps Pressurized gas feeder

the dissolution of the film. Third, water flow rates must be adequate to transport the inhibitor to all parts of the system, otherwise the protection film will not form (3). Several types of phosphates are used for corrosion control, including linear and cyclic polyphosphates, orthophosphates, glassy polyphosphates, and bimetallic phosphates. It is also possible to use zinc along with polyphosphates and orthophosphates, or blends of ortho and polyphosphates (15). The choice of a particular phosphate product depends on the water and the distribution system characteristics as well as the utility treatment goals. Some phosphates work better than others for a given system (3). It is recommended to conduct laboratory and pilot testing to evaluate the effectiveness of different products. Phosphate inhibitors need particular pH, alkalinity, and concentrations to be effective; the balance required is poorly understood. Orthophosphates seem to be very effective for a wide range of plumbing materials (15). It can be used alone (phosphoric acid, H3 PO4 ) or in combination with zinc. Commercial formulations of orthophosphates exist that contain various levels of zinc. Zinc orthophosphate products are used primary in water where the potential for formation of CaCO3 scale is low (soft and slightly acidic water). When using zinc phosphates, limitations of metal levels in wastewaters can limit the use of products with a high zinc content. Zinc orthophosphate leads to the formation of a zinc phosphate scale (16). When zinc is not available with phosphates, a protective scale still forms, such as ferrous or ferric phosphate scale on iron pipes.

CORROSION CONTROL IN DRINKING WATER SYSTEMS

Phosphate is most effective in the absence of prior scale formation because phosphates must be part of the precipitate to be effective. When fed into the system, phosphates soften the previously formed scales, causing red water and bacteriological problems as the scale washes out of the system. Orthophosphates can reduce corrosion rate of iron, lead, zinc, and galvanized pipes; although it is not considered to be very effective in preventing copper corrosion (15,17). A distinction should be made between orthophosphates and polyphosphates (10). Polyphosphates have been used to control corrosion in cast iron pipes; however, little evidence exists of any beneficial effects of polyphosphates on lead corrosion. Some studies (18,19) showed that lead levels could increase in water after solubilization of protective films on the pipes. Some polyphosphates can be used as sequestering agents to keep in solution scale-forming ions (calcium and magnesium) and iron. These polyphosphate products tie up with iron and prevent red water (3). Inorganic silicates have also been used to reduce corrositivity; they lead to the formation of a protective film onto various metal surfaces. Silicates inhibit corrosion in asbestos cement pipes, where its effectiveness may be attributed to a surface catalyzed conversion to quartz (20). Soluble silicates are adsorbed onto the metal pipe surface at the anodic area and form a thin film. High dosages (20 mg/L) are required during the first 30–60 days of treatment. Then, doses of 4–8 mg/L are added in the system. Silicates have not been widely used because their effects on corrosion are debatable (21,22). Cathodic Protection This process limits corrosion of metallic structures. It is used to prevent internal corrosion in water storage tanks. It consists of using an inert electrode (such as high silicon cast iron or graphite) powered by an external source of current. The current forces the inert electrode to act as an anode, preventing the metal to be protected to become an anode and corrode. Another method involves a magnesium or zinc electrode acting as a galvanic anode. The electrode produces a galvanic action with iron; it is sacrificed and corrodes, whereas the iron is protected from corroding. Lining, Coating, and Paints

11

when the plant was feeding a corrosion inhibitor (constant doses of 0.86 mg PO4 /l over the year). Higher phosphate levels in the test reactor resulted in low corrosion rates, especially in the summer. Corrosion rates were maintained below 3 mpy when phosphate dosages were slightly increased (between 1.5 and 2 PO4 /l), especially during warm periods. For this site, a seasonal corrosion control strategy was developed that would require slightly higher corrosion inhibitor concentrations during the summer and possibly lower dosages during winter months, rather than using a constant concentration over the entire year (23). CONCLUSION As distribution system pipes are in place for long periods of time (50 years), corrosion control is critical to maintain microbial and aesthetic water quality and pipe integrity. Water utilities should set corrosion control goals and monitor corrosion rates on a regular basis to determine seasonal changes in corrosivity and adjust corrosion control programs to prevent excessive corrosion. BIBLIOGRAPHY 1. American Water Works Association (AWWA). (1995). Water Treatment, 2nd Edn. AWWA, Denver, CO. 2. AWWA Research Foundation. (1996). Internal Corrosion of Water Distribution Systems. AWWARF-DVGW-TZW Cooperative Research Report. Denver, CO, p. 586. 3. Schock, M. (1990). Internal corrosion and deposition control. In: Water Quality and Treatment, 4th Edn. McGraw-Hill, New York. 4. Geildreich, E.E. (1996). Microbial Quality of Water Supply in Distribution Systems. CRC Press, Boca Raton, FL. 5. Kolle, W. and Roshe, H. (1980). Von Wasser. 55: 159–167. 6. Tipping, E. and Cooke, D. (1982). Geochimica et Cosmochimica Acta. 46: 75–80. 7. Chang, Y., Li, C., and Benjamin, M.M. (1997). J. Amer. Water Works Assoc. 89: 100–113. 8. Abernathy, C.G. and Camper, A.K. (1997). Interactions between pipe materials, disinfectants, corrosion inhibitors, organics and distribution biofilms. Proc. AWWA WQTC. Denver, CO.

A protective coating can keep corrosive water away from the pipe surface and storage tanks. Some linings include coal tar enamels, epoxy paint, cement mortar, polypropylene, or vinyl. These linings are applied when pipes are manufactured or in the field.

10. Faust, S.D. and Aly, O.M. (1999). Chemistry of Water Treatment, 2nd Edn. Lewis Pubilsher, Boca Raton, FL.

UTILITY EXPERIENCE WITH CORROSION CONTROL

11. Langelier, W.F. (1936). J. Amer. Water Works Assoc. 28(10): 1500–1505.

Figure 2 shows corrosion rates as a function of temperature and corrosion inhibitor levels in two pilot systems (23). The first system received a constant dose of phosphate over the entire study (system fed after conventional treatment of the Mississippi water, IL-American Water, E. St. Louis), whereas the second system received changing inhibitor levels (test system to optimize treatment). Corrosion rates were strongly related to water temperature (and/or other seasonal factors). For the plant condition system, corrosion rates could vary up to 7 mpy, even

9. U.S. EPA. (1991). Maximum Contaminant Level Goals and NPDWR for Lead and Copper. Federal Register.

12. Rossum, J.R. and Merrill, D.T. (1984). J. Amer. Water Works Assoc. 76(8): 72–82. 13. Larson, T.E. and Skold, R.V. (1957). J. Amer. Water Works Assoc. 49(10): 94–97. 14. Ryznar, J.W. (1944). J. Amer. Water Works Assoc. 36: 472–479. 15. Benjamin, M.M., Reiber, S.H., Fergusson, J.F., Van der Werff, E.A., and Miller, M.W. (1990). Chemistry of Corrosion Inhibitors in Potable Water. AWWA Research Foundation, Denver, CO. 16. Swayze, J. (1983). J. Amer. Water Works Assoc. 75: 101–102.

12

ECONOMICS OF RESIDENTIAL WATER DEMANDS 30

10 Plant conditions

Test conditions

Temperature

9 25

7 20 6 15

5 4

Temperature ( C)

Corrosion rate (mils/year)

8

10 3 2 5 1 0

3/27 5/5 5/17 5/31 6/20 7/10 7/22 8/16 10/4 10/1610/2811/1212/2012/30 1/24 2/10 2/21 3/5 3/26 4/16 5/7 5/29 6/26 96 97

0

Date 3

Phosphate concentration (mg/l)

Plant conditions

Test conditions

2.5 2 1.5 1 0.5 0 3/27 5/5 5/17 5/31 6/13 6/25 7/10 7/22 8/6 9/30 10/9 10/22 11/5 12/1112/27 1/7 2/10 2/21 3/5 3/26 4/16 5/7 5/29 6/26 96 97

Date

Figure 2. Corrosion rates as a function of corrosion inhibitor levels and temperature (the arrows show the changes in inhibitor levels to optimize corrosion control).

17. Ryder, R. and Wagner, I. (1985). Corrosion inhibitors. In: Internal Corrosion of Distribution Systems. AWWA Research Foundation, Denver, CO. 18. Schock, M.R. (1989). J. Amer. Water Works Assoc. 81(7): 88–95. 19. Holin, T.R. and Shock, M.M. (1991). J. Amer. Water Works Assoc. 83(7): 76–83. 20. Schock, M.R. and Buelow, R.W. (1981). J. Amer. Water Works Assoc. 73: 636–751.

23. Volk, C., Dundore, E., Schiermann, J., and LEChevallier, M. (2000). Water Res. 34: 1967–1974.

ECONOMICS OF RESIDENTIAL WATER DEMANDS STEVEN J. RENZETTI Brock University St. Catharines, Ontario, Canada

21. Sheiham, I. and Jackson, P.J. (1981). J. Inst. Water Engrs. Scientists. 35(6): 491–496. 22. Rompre, A., Prevost, M., Brisebois, P., Lavois, J., and Lafrance, P. (1997). Comparison of corrosion control strategy efficiency and their impacts on biofilm growth. Proc. AWWA Water Quality Tech. Conf., Denver, CO.

This entry examines what is known regarding the economic characteristics of residential water use and how this type of information can be applied in managing water resources. Understanding the economic dimension

ECONOMICS OF RESIDENTIAL WATER DEMANDS

of household water use is important for several reasons. First and most importantly, all households on the planet share the need for access to potable water supplies. On the other hand, households have very different degrees of access to safe water. For most of North America and Europe, access to clean water is almost taken for granted by many households. In contrast, a great number of households in low-income countries do not have access to reliable supplies of potable water. As the World Bank contends, ‘‘The challenge is enormous: one billion people still lack access to safe water, two billion lack safe sanitation. Slow progress is not acceptable, as more than three million children still die every year from avoidable water-related disease’’ (World Bank Water Supply and Sanitation web-site, September 1, 2001 http://www.worldbank.org/html/fpd/water/). Second, the agricultural sector dominates global water use, but residential and industrial demands have grown much more rapidly during the twentieth century. Third, empirical evidence indicates that the amount of water used by a household is a complex function of a number of influences, including the price of water (both in monetary terms and in time costs), household income, and household characteristics (for example, the number of residents and water-using appliances).

Table 1. Residential Water Use in 1995a Annual Residential Water Use (m3 /person)

Country Chad Nigeria China India Greece Israel South Africa Argentina Uruguay France Japan Canada U.S.A. a

4 9 21 25 45 45 49 67 73 94 137 157 203

Ref. 1.

3500

Global water use (km3/year) 3,250

Households use water for personal hygiene, waste removal, cooking, cleaning and a number of outdoor applications (lawn and garden watering, etc.) Table 1, Fig. 1. Except in some arid portions of the United States, the bulk of residential water use occurs indoors and is related to personal hygiene and waste removal (4). In North America, for example, these two applications account for 60% to 70% of the average household’s daily indoor water use. Another important feature of residential water use is its cyclical character. On a daily basis, household water use typically displays two peak periods of use: early morning and early evening. On an annual basis, residential water use in summer months is usually substantially higher, than during the rest of the year due to the increase in outdoor water use. Hanemann (5), for example, indicates that for households in the Western states of the United States, total summer water use exceeds total winter water use by 50–60%. The multiplicity of water’s uses and the cyclical nature of water use suggest that potentially a number of factors influence a household’s water use decisions. Economic theory provides a useful framework within which the nature of residential water use may be examined. Economists usually assume that a household’s demand for any good such as potable water is, in the most general case, a function of all of the prices facing a household as well as its income and demographic characteristics. The household’s estimated demand for water may be used to predict household consumption levels and to predict how households will respond to changes in the price of water. This degree of responsiveness is captured by a variable known as ‘‘elasticity.’’ This variable

1900 1995

3000 2500 2000 1500

1,250

1000

THE ECONOMIC PERSPECTIVE ON RESIDENTIAL WATER DEMANDS

13

500 0

500

425 25 Agricultural

Residential

50 Industrial

Figure 1. Global water use (1,2).

is measured as the ratio of the percent change in the demand for a good to the percent change in a price or income. A household’s demand is said to be ‘inelastic’ if the estimated elasticity is less than one. For example, if the price elasticity of demand equals −0.5, then a 10% increase in price, it is predicted, with to induce a 5% decrease in water use. The empirical water demand literature seeks to estimate households’ price and income elasticities of water demand as well as those households’ valuation of access to potable water supplies. There are several ways in which the general model of household demand has been altered to reflect conditions that are important in the consumption of water. These include situations when households’ choice of water use is constrained by their stock of water-using appliances and the size of the house and property (6), when households must choose not only the quantity of water to consume but also the source of its potable water (7), and when households face water prices that are complex functions of the quantity of water used. The last of these extensions is quite important because of the growing frequency of complex price schedules facing households and because of the implications of this type of price structure for the statistical estimation of residential water demands (5,8). EMPIRICAL MODELS AND RESULTS Residential water use has received a substantial amount of attention from economic researchers. There are a

14

ECONOMICS OF RESIDENTIAL WATER DEMANDS

number of surveys that summarize the theoretical and empirical research (4,9–11). The main goal of statistical models of water demand is to estimate the relationship between the observed quantity of residential water use and the explanatory variables (such as the price of water) suggested by economic theory. During the evolution of the residential water demand literature, a number of issues have been at the forefront, including the appropriate definition of the price of water, the choice of statistical technique, and the role of other explanatory variables (such as climate and household characteristics). The modern residential water demand literature begins with the work of Howe and Linaweaver (12). The authors estimate demand models for domestic (indoor) and sprinkling (outdoor) water demands. The demand models are further divided according to climatic zones. The authors assume that the quantity of water demanded by the average household is a function of water and sewage prices, age and value of property, the number of people in the household, and climate conditions. The estimation models indicate that indoor water use is responsive to the price of water to a limited degree (price elasticities ranging from −0.214 to −0.231) and outdoor water use is more responsive to prices (price elasticities ranging from −0.438 to −1.57). Income elasticities follow the same pattern: 0.314 to 0.378 for indoor water use and 0.447 to 1.45 for outdoor. Since the work of Howe and Linaweaver, statistical models of residential water demand have become more sophisticated, and data on household water use and characteristics have become more detailed and comprehensive. A particularly important issue has been the manner in which the price of water is represented. When prices are set out as complex functions in which prices can increase or decrease with the amount purchased, researchers must take more care in measuring the influence of prices on water use (the variety of statistical methods used to do this is reviewed in Ref. 11). A number of other factors, it has been found, are influential in determining household water demand. Researchers have found positive relationships between water use and family size, property value, and household income (4). In Hanemann’s (5) comprehensive listing of estimated water demand elasticities, the average of the income elasticities reported is 0.52. Another important set of factors that, it has been found, influences residential water demand is related to climate. There is a consensus in the literature that increases in temperature or evaporation rates lead to higher residential water demand, whereas increases in precipitation have the opposite effect. However, indoor residential water use appears to be relatively insensitive to weather conditions, and most of the influence of climate on residential water use works through outdoor water uses. LOW-INCOME COUNTRIES Households in many low-income countries face a variety of challenges in their efforts to secure potable water. In large urban centers, water supply systems are often unreliable and fail to provide service to many of the

city’s poorest residents (13). Residents in rural areas frequently confront even greater risks from contaminated and distant water supplies (14). Households in low-income countries also differ from their high-income counterparts in that they spend a larger share of their income on water. They also may face a different set of circumstances regarding their supply of potable water. For example, it is common for households in low-income countries to have several possible sources of potable water. Mu, Whittington and Briscoe (14) report that the members of a small town in Kenya choose among private pipe, communal wells and pumps, kiosks, and water vendors for their potable water. These sources differ in their relative cost, convenience, reliability, and quality. It is important to note that the cost of each alternative is a combination of the time spent obtaining water, installation charges, continuing connection charges, and prices. Households face the challenge of deciding which supply source(s) to use as well as how much water to obtain from each source, and researchers face the challenge of understanding and modeling these decisions (15). There have been a number of attempts to model, first, household decision-making regarding its choice of the source of its drinking water and, second, households’ valuation of improved access to reliable water supplies in low-income countries. Researchers have conducted surveys that collect information on households’ supply–source choices and characteristics. For example, Madanat and Humplick (16) examine the behavior of 900 households in Faisalabad, Pakistan. The authors conclude ‘‘the more expensive the in-house pipe connection relative to the other sources, the less likely the household is to connect’’ (p. 1337). In addition, the connection decision is strongly influenced by household expectations regarding the relative reliability and quality of alternative sources as well as their past experiences with alternative supply sources. With respect to the value they assign to access to safe water, many households in low-income countries find themselves in what some researchers have termed a ‘‘lowlevel equilibrium trap’’ (17, p. 1931). By this, the authors mean that the existing water supply system has few connections, low prices, low revenues, low maintenance, poor reliability, and low usage by households. One of the ways to improve this situation is to demonstrate that the value of improved service to households (as expressed in their willingness to pay through higher prices) exceeds the costs of improving service. The World Bank Water Research Team (15) is the most extensive effort to assess the preferences of households in low-income countries for connection to water supply systems. The authors find that household income is positively linked to the demand for improved services, but the link between the two is not strong. Households with higher levels of education are also willing to pay more for improved access. Gender is often an important factor in explaining willingness to pay, but ‘‘the direction of the influence depends on the specific cultural context’’ (p. 53). In Tanzania and Haiti, female respondents’ willingness to pay exceeds that of males, but in Nigeria and India, the reverse is true. As expected, economic variables play

ECONOMICS OF RESIDENTIAL WATER DEMANDS

an important role, and higher connection charges and monthly prices reduce demand. Finally, households are willing to pay more for private connections and for sources higher expected quality levels and reliability. RESIDENTIAL WATER DEMAND MANAGEMENT Historically, the challenge of managing water resources and providing potable water has concentrated on the engineering task of conveying and treating water so that it would be available for household use. However, as the construction and operation of water distribution and treatment systems becomes increasingly expensive (in monetary and environmental terms), attention has turned to the idea of balancing the costs of developing water supplies with the benefits derived from consumption. From this adoption of a more balanced perspective has come a renewed interest in using the information contained in estimated residential water demand to assist in planning and operating water delivery systems. Estimated residential water demand equations provide information that can be used in a variety of ways to promote water conservation and in the development and operation of water supply systems. First, the structure of residential water demand can provide water utility operators with information regarding the relative efficacy of alternative policy instruments aimed at promoting water conservation. For example, Renwick and Archibald (18) examine the factors influencing Californian households’ adoption of water conserving technologies such as low-flow toilets and showerheads. The results of the authors’ empirical model suggest that both price and nonprice measures reduce household water use, although their impact varies across households. Forecasting future water use is a second way in which demand information is used. Dziegielewski (19) provides a brief review of the history of urban water demand forecasting. In the ‘‘traditional’’ method of forecasting that has dominated historically, total future demand is predicted as the product of expected population growth and a fixed per capita water use coefficient. This method was not very accurate as it neglected other influences (such as prices) on water use. The development of the municipal and industrial needs (MAIN) model represents a major change in forecasting methods. The MAIN model disaggregates total water use into a large number (approximately 400) of categories and locations. The factors that influence water demand for each category are determined through statistical analysis. Changes in these explanatory variables (such as income, climate, and energy prices) translate into anticipated changes in water use and, in turn, form the basis for water demand forecasts. A third way in which water demand information is used in assessing the construction and operation of water delivery systems. For example, the costs of improving the reliability of a water supply system can be compared with households’ valuation of that increase in reliability (20). CONCLUSIONS All households share the need for access to potable water, but actual consumption levels vary significantly due to

15

differences in income, water prices, proximity to reliable water supplies, climate, and a variety of other factors. Economic models of household decision-making regarding water use indicate that households are influenced by these factors and that water prices, income levels, and climate play particularly influential roles. A different line of research highlights the challenges faced by households in low-income countries in their efforts to acquire potable water. One of the facets of this situation that has received attention recently concerns households’ decision-making when confronted with more than one source of potable water. As predicted by economic theory, most households consider the relative quality, reliability, and cost of alternative sources when making their choices. The last topic considered here is the use of information regarding the economic features of residential water demand to encourage water conservation. Research indicates that both price and nonprice based conservation programs are effective in curtailing demands. BIBLIOGRAPHY 1. Gleick, P. (2000). The World’s Water 2000–2001, Washington, DC, Island Press. 2. Abramovitz, J. (1996). Imperiled Waters, Impoverished Future: The Decline of Freshwater Ecosystems, WorldWatch paper 128, Washington, DC, Worldwatch Institute. 3. Biswas, A.K. (1997). Water development and environment. In: Water Resources: Environmental Planning, Management and Development. A.K. Biswas (Ed.). McGraw-Hill, New York, pp. 1–37. 4. Baumann, D.J., Boland, and Hanemann, W.M. (1998). Urban Water Demand Management and Planning. McGraw-Hill, New York. 5. Hanemann, W.M. (1998). Determinants of urban water use. In: Urban Water Demand Management and Planning. D.J. Baumann, Boland, and W.M. Hanemann. McGraw-Hill, New York, pp. 31–75. 6. Lyman, R.A. (1992). Peak and off-peak residential water demand. Water Resour. Res. 28(9): 2159–2167. 7. Asthana, A. (1997). Where the water is free but the buckets are empty: Demand analysis of drinking water in rural India. Open Economies Review 8(2): 137–149. 8. Shin, J. (1985). Perception of price when information is costly: Evidence from residential electricity demand. Rev. Econ. Stat. 67(4): 591–598. 9. Hanke, S. and de Mar´e, L. (1984). Municipal water demands. In: Modeling Water Demands. J. Kindler and C. Russell, (Eds.). Academic Press, London, pp. 149–170. 10. Gibbons, D.C. (1986). The Economic Value of Water, Resources for the Future, Washington, DC. 11. Renzetti, S. (2002). The Economics of Water Demands, Kluwer Academic, Norwell, MA. 12. Howe, C. and Linaweaver, F. (1967). The impact of price on residential water demand and its relation to system design and price structure. Water Resour. Res. 3(1): 13–31. 13. Rivera, D. (1996). Private Sector Participation in the Water Supply and Wastewater Sector: Lessons from Six Developing Countries. The World Bank, Washington, DC. 14. Mu, X., Whittington, D., and Briscoe, J. (1990). Modeling village water demand behavior: a discrete choice approach. Water Resour. Res. 26(4): 521–529.

16

GRAY WATER REUSE IN HOUSEHOLDS

15. World Bank Water Research Team. (1993). The demand for water in rural areas: Determinants and policy implications. World Bank Res. Obs. 8(1): 47–70.

Outside supply 4% Shower 5%

Kitchen sink 15%

16. Madanat, S. and Humplick, F. (1993). A model of household choice of water supply systems in developing countries, Water Resour. Res. 29(5): 1353–1358.

Basin 9%

17. Singh, B., Ramasubban, R., Bhatia, R., Briscoe, J., Griffin, C., and Kim, C. (1993). Rural water supply in Kerala, India: How to emerge from a low-level equilibrium trap. Water Resour. Res. 29(7): 1931–1942.

Bath 15%

18. Renwick, M. and Archibald, S. (1998). Demand side management policies for residential water use: Who bears the conservation burden? Land Econ. 74(3): 343–59.

WC 31%

19. Dziegielewski, B. (1996). Long-term forecasting of urban water demands. In: Marginal Cost Rate Design and Wholesale Water Markets, Advances in the Economics of Environmental Resources, Vol. 1, D. Hall (Ed.). JAI Press, Greenwich, CT, pp. 33–47.

Washing machine 20%

20. Howe, C. and Griffin Smith, M. (1993). Incorporating public preferences in planning urban water supply reliability. Water Resour. Res. 29(10): 3363–3369.

Figure 1. Components of domestic water use (2).

Table 1. Gray Water Quality from Various Sourcesa

GRAY WATER REUSE IN HOUSEHOLDS FAYYAZ A. MEMON D. BUTLER Imperial College London, United Kingdom

Gray water is a loosely defined term representing discharges from wash basins, baths, showers, dishwashers and washing machines. This generally excludes wastewater from kitchen sinks and toilets, commonly known as black water. Gray water accounts for about 50% of the total household water consumption activities (Fig. 1). The quality of gray water depends on factors, including the habits and affluence of the water users, the types of products used for clothes and personal washing, and the nature of the substances disposed of through sinks and other appliances. Substances found in gray water include detergents, shaving foam, toothpaste, soap, hair, body oils, and dried skin residues. Small amounts of fecal material arising from washing of baby diapers and traces of urine are also present in gray water. These pollutants exert oxygen demand and contain some disease causing microorganisms. Typical gray water pollutant concentrations from different sources are shown in Table 1. The average pollutant concentration measured in the effluent from different appliances housed in a residence hall (for example) is shown in Table 2. Freshly produced gray water usually does not have any objectionable odor. Compared to black water, gray water has a relatively higher temperature and readily degradable pollutants. Therefore, it requires immediate treatment after collection. If stored untreated for long periods, oxygen deficient conditions will develop, and scum will be formed that can float or sink in the collection tank. Experiments have shown that bacterial population also increases with increased storage time (5).

Dishwasher 1%

P Total BOD COD Turbidity NH3 (mg/l) (mg/l) (NTU) (mg/l) (mg/l) Coliforms Single person Single family Block of flats College Large college a

110 – 33 80 96

256 – 40 146 168

14 76.5 20 59 57

– 0.74 10 10 0.8

– 9.3 0.4 – 2.4

– – 1 × 106 – 5.2 × 106

Ref. 3.

Before gray water is reused, a certain level of treatment is required to minimize aesthetic concerns and the potential for health risk. Table 3 shows a summary of gray water quality criteria for toilet flushing followed in different countries. The level of treatment required depends on the scale and purpose of use. On a small scale (domestic level), a two-stage treatment consisting of filtration of coarse pollutants (hair and suspended impurities) followed by disinfection with chlorine, bromine, or UV is probably sufficient. On larger scales (hotels, commercial buildings), more complex and expensive methods of treatment could be employed. Domestic gray water recycling systems, normally employed, produce water for toilet flushing. A recycling system (Fig. 2), typically, consists of an underground collection tank and an overhead distribution tank to supply toilet cisterns. The collection tank is designed to prevent groundwater contamination and ingress and is sized to accommodate water volumes intended for reuse. The optimal size of the collection tank has been modeled by Dixon (6), and systems storing 100–200 liters are considered sufficient for a family of five persons (7). Any excess gray water is diverted to the sanitary drain (i.e., the drain going out of the household). Devices are installed to prevent back-flow from the foul drain to the tank. Filtration is typically carried out at the tank inlet. The clogged filters are either replaced or cleaned using water jets. A submersible pump fitted with a float switch

GRAY WATER REUSE IN HOUSEHOLDS

17

Table 2. Average Pollutant Concentration in Gray Water Measured in a Residence Halla Parameter

Bath/ Shower

Washbasin

Washing Machine

Laundry and Dishwashing

BOD (mg/L) COD (mg/L) Ammonia as N (mg/L) Phosphate as P (mg/L) Total coliforms (cfu/100 mL) Faecal coliforms (cfu/100 mL) Turbidity (NTU) Inorganic carbon (mg/L) TOC (mg/L) Total solids (mg/L) Suspended solids (mg/L) Dissolved solids (mg/L) Volatile solids (mg/L) pH Copper (mg/L) Lead (mg/L) Zinc (mg/L) Cadmium (mg/L)

216 424 1.56 1.63 6 × 106 600 92 26 104 631 76 559 318 7.6 111 3 59 0.54

252 433 0.53 45.5 5 × 104 32 102 20 40 558 40 520 240 8.1 – – – –

472 725 10.7 101 7 × 105 728 108 25 110 658 68 590 330 8.1 322 33 308 0.63

110 – – – 5×106 462 148 20 84 538 90 449 277 7.8 – – – –

a

Ref. 4.

Table 3. International Water Quality Criteria for Toilet Flushinga

US EPA (g) Florida (m) Texas (m) Germany (g) Japan (m) South Africa (g) WHO lawn irrigation EC bathing water

UK (BSRIA) Proposed (g)

Fecal coliforms (cfu/100 mL)

Total coliforms (cfu/100 mL)

BOD (mg/L)

Turbidity (NTU)

TSS

DO% (% saturation)

pH

Cl2 residual (mg/L)

14 for any sample 0 for 90% samples 25 for any sample 0 for 75% samples 75 (m) 100 (g) 10 for any sample 0 (g) 200 (g) 1000 (m) 100 (g) 2000 (m) – – 14 for any sample 0 for 90%

– – – – – 500 (g) 10 – – – 500 (g) 10000 (m) – – – –

10 – 20 – 5 20 (g) 10 – – – 2 (m) (g) 1 (m) (m) – –

2 – – – 3 1–2 (m) 5 – – – – – – – – –

– – 5 – – 30 – – – – – – – – – –

– – – – – 80–120 – – – – 80–120 – – – – –

6–9 – – – – 6–9 6–9 – – – 6–9 – – – – –

1 – 1 – – – – – – – – – – – – –

a Ref. 4. g = guideline. m = mandatory.

is normally used in the tank to transfer filtered water to the overhead tank. This then contains the disinfectant feeding arrangement and switches to control the water level. When the water volume in the overhead tank drops below a certain level, the pump turns on and stops when the water level in the tank reaches the design level. The overhead tank normally has an inlet to provide a top up supply of mains water when the treated gray water is not sufficient to meet the demand or the recycling system is inoperative. An air gap is typically provided between the inlet pipes of gray water and mains water. The pipework carrying gray water needs to be clearly marked or colored differently to avoid cross connections

and contamination of potable water. The underground and overhead tanks will be designed to drain down fully to avoid problems of prolonged gray water storage. There are several packaged recycling systems available on the market. Each offers a different degree of treatment and safety controls. Experience has shown that they are not ‘fit and forget’ systems but require monitoring to ensure their smooth operation. Therefore, a clear warning mechanism that can show the failure of system components should be installed within the household, and clear maintenance instructions provided. Gray water recycling is not a problem-free option, and particularly, issues related to health risk must

18

GRAY WATER REUSE IN HOUSEHOLDS

water stagnation for a prolonged period, the recycling system should be kept free from ‘dead legs.’ Past case studies on gray water recycling systems suggest that the relative health risk from gray water reuse is not high if it is properly treated and does not come in direct contact with users (8). For additional safety, the use of treated gray water should be discontinued when any of the users living in the household is ill. Spray irrigation with gray water is not advised because it will increase the bacterial exposure potential. There is also concern about the use of chlorine as a disinfectant. It has been found that chlorine can corrode metal switches and fittings in the overhead tank and toilet cistern. Excessive buildup of chlorine gas in the overhead tank (if located in a loft) may produce an unpleasant smell in the household and may be linked to asthma (9). Other disinfectants, such as ultraviolet radiation, are also available. UV lamps are expensive, and their germicidal efficiency reduces with time. They are most effective in waters of low turbidity. Therefore, fine filtration will be needed to achieve improved pathogen removal. The residual effect of UV as a disinfectant is not stable and microbial regrowth is possible, so UV treated waters should not be stored for long periods. Bromine is also used for pathogen removal from gray water. Some forms of bromine disinfectants such as hypobromous acid are considered harmful to plants, and the treated water would not be suitable for irrigation. The health risks from bromine use are yet to be quantified (7). At present, gray water recycling systems on a single household scale are hardly financially viable. Although there are some savings from reduced consumption of mains water, the capital and operating expenditures incurred for these systems are relatively high, and the pay-back period is 20–25 years (10). The payback period reduces with increase in occupancy.

Disinfection dosing Potable water top up

Gray water from baths, showers, and washbasins

To flush WCs Filtration To sewer

Pump Figure 2. Water flows and components of a typical gray water recycling system.

be addressed. Although most waterborne pathogens are killed by conventional disinfectants, there are certain species (e.g., Legionella spp.) that are resistant to normal modes of treatment. Legionella pneumophila, a naturally occurring bacterium in domestic hot water supplies, showerheads, cooling waters, and other water services in buildings, has been linked to outbreaks of Legionnaires’ disease. Surface fouling, biofilm formation, slow moving or stagnant waters, and increased temperatures are favorable conditions for Legionella growth. Therefore, recycling systems could provide an ideal environment for their growth. Legionella has an infection route through inhalation, and it is suspected that some bacteria may be inhaled through water vapors during toilet flushing (8). Fortunately, however, research shows that the Legionella count in gray water is typically low (3). To avoid in-pipe

100 BOD Suspended solids

Removal efficiency, %

80

Total coliforms

60

40

20

0

ed

BR

am

g er

m

b Su

Rs

s

s

BR

M

d Si

tre es

M

M

AB

Fs

es

an br m e

BA

c

flo

M

g

sin

O2

do

Ti

i

at

ul

ag Co

+ on

n

io

at

l cu

Figure 3. Technological performance in removing pollutants from gray water (3).

WATER AND HUMAN HEALTH

Gray water recycling on a medium to large scale (e.g., hotels, blocks of flats, commercial buildings,) may be more viable. A stepwise complex treatment sequence, instead of relying just on simple filtration and disinfection, produces water that has relatively low potential for health risk. A wide selection of gray water treatment technologies is currently available. These include biological aerated filters (BAFs), membranes, sidestream and submerged membrane bioreactors (MBRs), UV treatment, titanium dioxide (TiO2 ) dosing, membrane aeration bioreactors (MABRs), and coagulation/flocculation with alum and ferric. Trials with these technologies have shown efficient and reliable removal of pollutants from gray water (3). The comparative efficiency of these technologies in removing BOD, suspended solids, and total coliforms is shown in Fig. 3. Large recycling units perform well, and their use in large buildings in the developed world and particularly in Japan is well established. The main barrier to wider uptake of gray water recycling systems is lack of adequate consolidated legislation, high capital and maintenance costs, and potential health risks due to technology failure. Studies carried out to gauge public perception have shown that individuals have a positive attitude toward using treated gray water produced within their own households for toilet flushing, as long as safety is guaranteed and it is costeffective (3). BIBLIOGRAPHY 1. Butler, D. and Dixon, A. (1998). Assessing the viability of the household grey water and rainwater re-use. 21AD: Water, Architectural Digest for the 21st Century. Oxford Brookes University, Oxford, pp. 20–23. 2. POST. (2000) Water Efficiency in the Home. Parliamentary Office of Science and Technology Note 135, London. 3. Jeffrey, P. and Jefferson, B. (2001). Water recycling: How feasible is it? Filtration + Separation 38(4): 26–29. 4. Surendran, S. and Wheatley, A.D. (1998). Gray water reclamation for non-potable reuse. J. Chartered Inst. Water Environ. Manage. 12(6): 406–413. 5. Dixon, A., Butler, D., Fewkes, A., and Robinson, M. (2000). Measurement and modelling of quality changes in stored untreated gray water. Urban Water 1(4): 293–306. 6. Dixon, A. (2000). Simulation of Domestic Water Reuse Systems: Gray Water and Rainwater in Combination. PhD Thesis, Imperial College, London. 7. CIRIA. (2000). Rainwater and Gray Water Use in Buildings: Best Practice Guidance. Construction Industry Research Information Association Report, London. 8. Dixon, A., Butler, D., and Fewkes, A. (1999). Guidelines for gray water reuse—health issues. J. Chartered Inst. Water Environ. Manage. 13(5): 322–326. 9. HAZOP. (2000). HAZOP-Single house gray water recycling system. Report No. WROCS HAZOP1, The School of Water Science, Cranfield University. http://www.cranfield.ac.uk/ sims/water/gray waterhazopstudy.pdf. 10. WROCS. (2000). Water Recycling Opportunities For City Sustainability. Final Report prepared by Cranfield University, Imperial College and Nottingham Trent University in collaboration with industrial partners. http://www.cranfield.ac.uk/ sims/water/finalreport-wrocs.pdf.

19

WATER AND HUMAN HEALTH S.M.A ADELANA University of Ilorin Ilorin, Kwara State, Nigeria

INTRODUCTION Water in the literal sense is the source of life on the earth. Research has shown that the human body is 70% water. Generally, human beings begin to feel thirst after a loss of only 1% of body fluids and risk death if fluid loss exceeds 10%. It has been proved that human beings can survive for only a few days without fresh water. Although it is true that life depends on water, society does not usually act as though water has value equal to life itself. The reason is that the supply of water in many parts of the world far exceeds what is required to sustain life. Estimates revealed that about 9000 cubic meters (9.0× 106 liters) of water is available for use per person per year. Based on projected population growth, this amount will drop to 5100 cubic meters per person by the year 2025 because another 2 billion people are expected to join the world’s population by the year 2025. Despite this sharp drop (by nearly 50% in 35 years), the amount of water available would be sufficient to meet human needs if it were distributed equally among the world’s population and less polluted by human activities. Present estimates give a false picture of freshwater available for human use because the distribution of the world’s available freshwater is uneven throughout the seasons and from year to year. According to Falkenmark (1), water is not always where we want it. Sometimes, it is not available in sufficient quantities where we want it or at another time too much water is in the wrong place. Yet, in many parts of the world, people are withdrawing water from surface and ground sources at a rate faster than they can be recharged. In the last century, world population has tripled, but water withdrawals have increased by more than six times (2,3). For example, since 1940, annual global water withdrawals have increased by an average of 2.5% to 3% a year compared with annual population growth of 1.5% to 2% (4,5). In the past decade, however, water withdrawal has increased from 4% to 8% a year, especially in developing countries (6). If the present consumption patterns continue, by year 2025, about two billion people will be living in areas where it will be difficult or impossible to meet all their needs for fresh water. Half of them will face severe shortages (3,7,8). Apart from the pressure of population growth on water resources, the supply of freshwater available to humanity is shrinking, in effect, because of increasing pollution. Population growth, urbanization, and industrialization with little regard for the environment are polluting and decreasing the quantity of freshwater available for human consumption (or use). Farming is said to be responsible for a great deal of water pollution in the United States (9). Similarly, in India (where there is heavy dependence on irrigation farming for food supplies), more than 4

20

WATER AND HUMAN HEALTH

million hectares of high-quality land have been abandoned as a result of salinization and waterlogging caused by excessive irrigation (10,11). More than 90% of the rivers in Europe have high nitrate concentrations, mostly from agrochemicals, and 5% of these rivers have concentrations at least 200 times higher than nitrate levels naturally occurring in unpolluted rivers (12–14). Moreover, in the Czech Republic, 70% of all surface waters are heavily polluted mostly with municipal and industrial wastes (15). Havas-Szilagyl (16) reported that 600 out of the 1600 well fields tapping groundwater in Hungary are already contaminated, mostly by agrochemicals. In developing countries, an average of 90% of all domestic sewage and 75% of industrial wastes are discharged into surface waters without any kind of treatment (17,18). Generally, oil and salts are washed off city streets, and heavy metals are leached from municipal and industrial dump sites. There is also the possibility that pollutants, such as sulfur dioxide and oxygen or nitrogen, combine in the atmosphere to form acid rain which has devastating effects on both surface water and land ecosystems (19) and accompanying health implications (3). On the whole, both water scarcity and water pollution pose serious health problems. Unclean water is by far the largest environmental killer around the world; it claims millions of lives every year. According to the World Health Organization (WHO), a large percentage of urban population in developing countries do not have access to proper sanitation facilities, and about half lacks a regular supply of potable water (12). In the year 2000, an estimated 1.1 billion people remained without access to improved drinking water (7), and the number of persons drinking water contaminated by human sewage was

much higher (21). Obviously, scarce and unclean water supplies are critical public health problems in many parts of the world and are likely to be one of the major factors that will limit economic development in the near future (10,11,13,14). WATER SCARCITY AND STRESS: IMPLICATIONS FOR HUMAN HEALTH The world’s population of nearly 6 billion is growing by about 80 million people each year (3). This rapid population growth coupled with increasing demands for water for irrigation agriculture, domestic (municipal), and industrial uses puts tremendous pressure on the world’s freshwater resources. As population grows, water use per person rises and freshwater withdrawal becomes faster than it can be recharged, resulting in water mining (7,22). If this continues and water needs consistently outpace available supplies, a level will be reached when depletion of surface and groundwater resources results in chronic water shortages (23), as illustrated in Fig. 1. Investigations have revealed that up to 31 countries, which represent nearly 8% of the world population, face chronic water shortages (3). It has also been estimated that by the year 2025, the number of countries facing water shortages is expected to be near 48, affecting more than 2.8 billion people—35% of the world’s projected population (24–27). Figure 2 shows the population in water-scarce and water-stressed countries. It is obvious from this figure that, as population grows, many more countries will face water shortages. Accordingly, a more optimistic outlook predicts that 2.8 billion people in 48 countries will be struggling with water scarcity by the year 2025, whereas worst case scenarios for water shortages

Population dynamics Growth-Migration-DensityDistribution-UrbanizationMorbidity-Mortality

Human outcomes Food shortage-Water-related illnessSocial & political instability-Conflicts over water-Slowed economic growthPopulation displacement

Water use Agriculture-IndustryHousehold use-Sanitation & waste disposal-HydroelectricityFish farming

Environmental outcomes Depletion of surface & groundwaterWater pollution-Land degradationEcosystem degradation-Declining fisheriesDisruptions to the hydrologic cycle

Figure 1. Links between world population and freshwater (3).

WATER AND HUMAN HEALTH

foresee 4 billion people in 54 countries facing water shortages in 2050. This critical trend is supported by the water scarcity and population projections of the United Nations (28–30). From the growing consensus among hydrologists, a country is said to experience water stress when its annual water supply is between 1000 and 1700 cubic meters per person. Such a country can expect to experience temporary or limited water shortages. But when the annual supply of renewable freshwater drops below 1000 cubic meters per person, the country faces water scarcity (1,20,25). In such cases, chronic and widespread shortages of water that hinder development result, and this could lead to severe health problems. Among countries likely to run short of water in the next 25 years are Ethiopia, Kenya, Nigeria, India, and Peru, whereas most parts of China and

Population in water-scarce and water-stressed countries, 1995–2050 4.0

Population, in billions

4

54 countries 2.8

3

48 countries 2

1 0.46 31 countries 0

1995

2025

2050

Figure 2. The rising trend in water scarcity and stress, 1995—2025 (3).

Pakistan are already approaching water stress. The list of water-stressed countries and those already suffering from water scarcity has been tabulated in GardnerOutlaw and Engleman (25) and PR (3) based on the 1995 world population and water per capita and a projection for the year 2025. Gardner-Outlaw and Engelman (25) based their calculations on United Nations Population Division population estimates and growth rate and total fertility rate (TFR) data from the Population Reference Bureau (24), World Population Data Sheet. According to the estimates, it is obvious that in this century, water crises in more and more countries will present obstacles to better living standards and better health and may even bring risks of outright conflict over access to scarce freshwater supplies. Available statistics show that more than half of the world’s population suffers from water services that are inferior to those of the ancient Greeks and Romans (8). According to Gleick, this has long been recognized as a serious global water problem that even generated attention at the World Water Conference organized by the United Nations at Mar del Plata in 1977, where strong commitments and resolution were made to finding lasting solutions. Since this initial attempt, considerable efforts have been geared toward providing access to safe drinking water and adequate sanitation services. The United Nations in its Millennium Developmental Goal (during the World Summit on Sustainable Development held in Johannesburg in September 2002) planned to reduce by half, by the year 2015, the proportion of people without sustainable access to safe drinking water and basic sanitation services. Yet, a high percentage of the world’s population is still without access to adequate water supply and sanitation (see Fig. 3a, b). The 1990 world population without access to clean drinking water was estimated at 1300 million, and close to 2600 million people have no access to basic sanitation

Distribution of unserved populations by region (source: Ref. 31). (a)

(b)

Europe 2%

Europe 2%

Latin America and the Caribbean 5%

Latin America and the Caribbean 6%

Africa 27%

Africa 13%

Asia 65%

Total unserved: 1.1 billion

21

Asia 80%

Total unserved: 2.4 billion

Figure 3. Distribution of unserved populations by region.

22

WATER AND HUMAN HEALTH

services (32). In Africa alone, 54% of the 1994 total estimated population of 707 million have no access to clean drinking water (33). Table 1 shows access to safe drinking water in developing countries by region; the percentage of population that has access to sanitation services is tabulated in Table 2. Illustrative world maps showing estimates of population without access to clean drinking water and adequate sanitation services are included in Gleick (8). It is important to point out that the failure to provide basic clean water and sanitation services takes a serious toll on human health and results in economic loss in many countries of the world. Reports have shown that water, shortages, polluted water, and unsanitary living conditions claim millions of lives annually (7,20,34). The World Health Organisation reports that 80% of diseases are overtly or covertly waterborne (35) and/or consequent to freshwater shortages. Moreover, in much of the world, polluted water, improper waste disposal, and poor water management cause serious public health problems. For example, diarrheal diseases leave millions of children underweight, mentally as well as physically handicapped, and vulnerable to other diseases.

basic human needs should be an obligation of governments and nongovernmental organizations. The postconference reports on the International Symposium and Technology Expo on small drinking water and wastewater systems held at the Hyatt Regency in Phoenix, Arizona (from January 12–15, 2000), demonstrated that the provision of safe drinking water and effective wastewater system managements are key elements that ensure safe and healthful living linked to social and economic development. The basic water requirements for humans depend on the purpose for which water is used in the different sectors of our society. Among these are drinking (and other domestic use), removing and diluting waste (including disposing of human waste), growing food, producing manufactured goods, producing and using energy, and so on. The water requirement for each of these activities varies with domestic conditions, lifestyle, culture, tradition, diet, technology, and wealth (36). It has been argued that water requirements for humans should also include any water necessary for disposing of human wastes (33). For example, in regions where absolute water quantity is a major problem, waste disposal options that require no water are available, in most developing nations, preference is given to alternatives that use at least some water (8). However, there are societies that use enormous amounts of fresh water to dispose of wastes. Based on the research carried out so far, a recommendation for a basic water requirement has been made. In 1996, Gleick proposed the overall minimum water required per person per day as 50 liters. This covers

BASIC WATER, REQUIREMENTS FOR HUMAN HEALTH Generally, in developing and using water resources, priority has to be given to the satisfaction of basic human needs (35). Therefore, providing water sufficient to meet

Table 1. Population that has Access to Safe Drinking Water in Developing Countries, by Region, 1980 and 1994a

Region and Country

1980 Population, Millions

Percent With Access

Number Unserved, Millions

1994 Population, Millions

Percent with Access

Number Unserved, Millions

239 468 707

64 37 46

86 295 381

348 125 473

88 56 80

42 55 97

955 2167 3122

84 78 80

150 477 627

1 11 12

52 29 81

98 69 88

1 9 10

215 1612 1827

1594 2789 4383

82 70 74

279 836 1115

Africa Urban Rural Total

120 333 453

83 33 46

20 223 243

Latin America & the Caribbean Urban Rural Total

237 125 362

82 47 70

43 66 109

Asia & the Pacific Urban Rural Total

549 1823 2373

73 28 38

Urban Rural Total

28 22 49

95 51 75

148 1,313 1,461

Western Asia

Total Urban Rural Grand total a

Reference 8.

933 2303 3236

77 30 44

WATER AND HUMAN HEALTH

23

Table 2. Population that has Access to Basic Sanitation Services in Developing Countries, by Region, 1980 and 1994a

Region and Country

1980 Population, Millions

Percent with Access

Number Unserved, Millions

1994 Population, Millions

Percent with Access

Number Unserved, Millions

239 468 707

55 24 34

108 356 464

148 125 473

73 34 63

94 83 176

955 2167 3,122

61 15 29

371 1835 2206

6 14 20

52 29 81

69 66 68

16 10 26

289 1451 1740

1594 2789 4383

63 18 34

589 2284 2873

Africa Urban Rural Total

120 333 453

65 18 30

42 273 315

Latin American & the Caribbean Urban Rural Total

237 125 362

Urban Rural Total

549 1823 2373

78 22 59

52 97 150

Asia & the Pacific 65 42 47

192 1058 1250

Western Asia Urban Rural Total

28 22 49

79 34 59 Total

Urban Rural Grand total a

933 2303 3236

69 37 46

Reference 8.

the minimum standards for drinking, sanitation, domestic (bathing and washing), and cooking. Out of this overall water requirement, 25 liters/person/day is required for basic hygiene (washing, showering, and bathing) and for cooking (33). A minimum of 20 liters/person/day offers the maximum benefits of combining waste disposal and related hygiene, thereby meeting cultural and societal preferences for water-based disposal (8). In other words, the minimum amount of water needed for drinking, cooking, bathing, and sanitation is 50 liters. The average person needs a minimum of 5 liters of water per day to survive in a moderate climate at an average activity level. However, average people in the United States uses between 250 to 300 liters of water per day for drinking, cooking, bathing, and watering their yards, whereas the average person in the Netherlands uses only 104 liters per day for the same tasks (33,37). This amount is slightly above the minimum target of 20–40 liters/person/day set by the United States Agency for International Development, the World Bank, and the World Health Organisation. Many people in the poorest nations survive on far less than the recommended amount. For example, the average person in Somalia uses only 8.9 liters of water per person per day (7,37). Although different sources use different figures for total water consumption and for water use by sector of the economy (1,25,33,34,37–39), yet from drinking water and sanitation needs, it became obvious that a basic requirement of 25 liters/person/day of clean water must be provided for drinking and sanitation by water agencies, governments, or community organisations. An estimate

made in 1990 revealed that about 55 countries whose population was nearly 1 billion people did not meet this standard (33). Yet, it is a desirable goal from a health perspective and from a broader objective of meeting a minimum quality of life. Further information on basic water requirements can be obtained in The World’s Water 1998–1999, The World’s Water 2000–2001, and The World’s Water 2002–2003, which are available from Island Press, Washington (http://www.islandpress.com/). WATER-RELATED, DISEASES Water-related diseases that affect human health are relatively widespread and abundant, especially in rural communities of developing nations, although there is evidence that they have been reduced to a greater extent as a serious health problem in industrialized countries. The incidence of these diseases depends on local climate, geography, culture, sanitary habits and facilities, and on the quantity and quality of the local water supply as well as the methods of waste disposal (3). Changes in water supply do affect different groups of diseases in different ways. Some may depend on changes in water quality, others on water availability, and yet others on the indirect effects of standing water. A World Health Organisation (40) estimate of the number of people suffering from water-related diseases is staggering (see Table 3). Generally, in many developing countries, waterborne diseases such as cholera, dysentery,

24

WATER AND HUMAN HEALTH Table 3. Estimates of Global Morbidity and Mortality from Water-Related Diseases (Early 1990s) Culled from Reference 8a Morbidity, Episodes/Year or People Infected

Disease Diarrheal diseases Intestinal helminths Schistosomiasis Dracunculiasis Trachoma Malaria Dengue fever Poliomyelitis Trypanosomiasis Bancroftian filariasis Onchocerciasis a

1,000,000,000 1,500,000,000 (people infected) 200,000,000 (people infected) 150,000 (in 1996) 150,000,000 (active cases) 400,000,000 1,750,000 114,000 275,000 72,800,000 (people infected) 17,700,000 (people infected; 270,000 blind)

Mortality, Deaths/Year 3,300,000 100,000 200,000 — — 1,500,000 20,000 — 130,000 — 40,000 (mortality caused by blindness)

Original Source: Reference 33.

typhoid, malaria, and schistosomiasis are increasing and harm or kill millions of people every year. The Pacific Institute’s recent research indicates that lack of clean drinking water leads to nearly 250 million cases of water-related disease each year and roughly five to ten million result in deaths (7). Earlier estimates have shown much higher numbers of people in the world suffering from diseases that are linked with water (5,41) and resultant death (8,15). The true extent of these waterrelated diseases is unknown, and even the WHO data (40) suggest there may be many more cases of the diseases and resultant death. However, about 60% of all infant mortality is linked to infectious and parasitic diseases; most are waterrelated (42), and a large percentage of these diseases is attributable to inadequate water supply and sanitation. Research has shown that, at any one time, there are probably millions of people who have trachoma, elephantiasis, bilharzias (snail fever), malaria, diarrhea, dracunculiasis (guinea worm disease), and onchocerciasis (river blindness). For example, according to Edungbola (43), at least 15 million Africans suffered from guinea-worm infection; of these, nearly 75,000 people are permanently disabled every year, and about 3 million individuals were irreversibly crippled in Africa. His estimates have further shown that subsistence farmers in Africa lost at least 80 million man-days each year to guinea worm disease. Water-related diseases are generally classed into four categories: waterborne, water-washed, water-based and water-related insect vectors (8,44,45). Waterborne diseases include those caused by both fecal—oral organisms and those caused by toxic substances; water-washed (also referred to as water-scarce) consists of diseases that develop where clean fresh water is scarce (44). Aquatic organisms that spend part of their life cycles in water and other part as parasites of animals cause water-based diseases. Insects that transmit infections, such as mosquitoes and tsetse flies, cause water-related vector diseases. A full description of each class of water-related disease together with their causative agents and routes of transmission as well as the geographical extent and number of reported cases has been compiled in Population Reports (3).

According to Population Reports, diarrheal disease (which belongs to the class of waterborne disease) is prevalent in many countries where sewage treatment is inadequate or where human wastes are disposed of in open latrines, ditches, canals, and watercourses or are spread indiscriminately on farmland. In the mid-1990s, a large number of people drank water contaminated with human sewage (28), and the World Health Organisation reported that drinking contaminated water contributes directly to diarrhea-related deaths (46). An estimated 4 billion cases of diarrheal diseases are reported annually that cause 3–4 million deaths, mainly among children (34,47–49). In Nigeria alone, more than 300,000 children less than 5 years of age die annually from diarrheal diseases (50). For example, in 1996, a large outbreak of severe diarrhea (which was later confirmed as cholera from tests conducted at the Institut Pasteur, Paris, France) struck the commercial city of Kano in northern Nigeria. According to Hutin et al. (51), a total of 5600 cases and 340 deaths (attack rate = 86.3 per 100, 000 inhabitants) were reported to the Kano State Ministry of Health within 5 months of the incident. This incidence was highest among children less than 5-years-old and was linked to drinking streetvended water and failure to wash hands with soap before meals (51). Earlier, the consumption of street-vended water was reportedly associated with a cholera outbreak in Latin America (52–54). A similar cause of a cholera outbreak was also reported in India (55,56) and in Peru (57). Recent research has also shown that childhood mortality from diarrhea in Latin America remains high (58). Gleick (8), using available data on the prevalence of different water-related diseases, presented and discussed two of these diseases extensively—dracunculiasis (guinea worm) and cholera—as case studies. He traced the history and reported the total global cases of these diseases by region and the recurrent deaths as a result of the epidemic from 1990 to 1997 with an update on the complete eradication programs. Reported guinea worm cases, globally, have fallen from an estimated 3.5 million in 1986 to 150, 000 in 1996 (59,60). This is approximately a 97% reduction, and there are hopes that it may have

WATER AND HUMAN HEALTH

been completely eradicated in accordance with the ‘‘New Millennium Plan.’’ On the other hand, little has been achieved in the effort to control the transmission of other parasitic infections such as schistosomiasis, intestinal helminthiasis, and malaria which are related to water supply and sanitation, especially in the developing world and particularly in Africa. For further information and statistics on other water-related diseases, readers are referred to (5,12,14,24,34,47–49,59–68). The issue of water quality or maximum permissible limits of certain elements that can constitute health risks in drinking water should be included here as another source of waterborne disease. For example, increased nitrate concentrations in drinking water add to the variety of water-related health risks. Health problems from nitrate in water sources are generating serious concern in almost all countries of the world, particularly in urban and rural communities where agricultural practises are intensive (69–73). There is increasing evidence that nitrate levels in many aquifers are rising and that the problem of increased exposure of the world population to high nitrate inputs will become more pressing, as speculated earlier by the WHO (74). Agricultural activities such as fertilizer and pesticides applications are frequent sources of contamination in surface and groundwaters. An estimate from Population Reports has shown that in more than 150 countries, nitrate from the application of fertilizers has seeped into water wells and polluted drinking water (75). Increased concentrations of nitrate often cause blood disorders (76). High levels of nitrate and phosphates in drinking water also encourage the growth of blue-green algae, resulting in deoxygenation (eutrophication) and subsequent reduction in metabolic activities of the organisms that purify fecalpolluted water in the human system. Details of nitrate health hazards are discussed in ADELNA (in this volume). Other sources of water pollution include animal wastes, excess nutrients, salinity, pathogens, and sediments that often render water unusable for drinking, unless it is purified (77–81). Even when any of these substances or chemicals occurs in low concentrations, they can accumulate in humans over time to cause serious health problems such as cancer if the water is used for drinking. For maximum permissible and acceptable levels of ions/elements in water, refer to the standards of the World Health Organisation (62) and of most national authorities, which are consistent with standards for the composition of drinking water (82). The average contribution of drinking water to the daily intake of mineral nutrients is important in health considerations. Of note here are those for fluoride (F− ) and arsenic (As). Generally, excessive concentrations of these elements often limit the use of groundwater for drinking. Too high an intake of fluoride is often the general cause of painful skeleton deformations called fluorosis, which is a common disease in East Africa, especially in Kenya and Ethiopia. The occurrence of fluoride in groundwater for human consumption has also been reported in Argentina (83). High concentrations of arsenic in groundwater used for drinking are reported

25

in many countries such as India, Bangladesh, China, Thailand, Vietnam, Taiwan, Hungary, Mexico, and Finland (84–88). Nearly 50 million people are at risk of cancer and other arsenic-related diseases due to consumption of high arsenic groundwater in India and Bangladesh (64,89). About 44% of the population of West Bengal (India) is suffering from arsenic-related diseases such as conjunctivitis, melanosis, hyperkeratosis, and hyperpigmentation (90). In certain areas, gangrene in the limb, malignant neoplasm, and even skin cancer have also been observed. High arsenic concentrations lead to black-foot disease. This is sometimes visible in a blackening of the fingers and toe tops and induces general lethargy in the patient. Arsenic toxicity affects almost all organs of the human body. Ingestion of large doses of arsenic usually results in symptoms within 30 to 60 minutes but may be delayed when taken with food (90). High arsenic concentrations have also been reported in Southeast Asia (91), the United States (92), Argentina, and Chile (93–97); all have consequent health implications. THE IMPACT OF IMPROVED WATER SUPPLY AND SANITATION The direct consequence of water scarcity and failure to meet basic water requirements is the prevalence of most water-related diseases. In the past, this has caused serious economic and social loss to both governments and communities. Estimates in the late 1970s have shown that water-related diseases cost more than $125 billion per year, excluding social costs, the loss of education and other opportunities, lost economic productivity of sick workers, and other hidden costs (8,98). For example, in subSaharan Africa, malaria costs an estimated $1.7 billion US annually in treatment and lost productivity (48). A study in Pakistan (within its capital city, Karachi) has shown that people living in areas without proper sanitation or hygiene education spent six times more on medical care than residents in areas with access to sanitation and basic hygiene (63). In Peru, an economic loss of more than $1 billion dollars in seafood exports and tourist revenues has been reported due to a cholera epidemic (99). However, the huge investments by governments in Asia, Africa, and Latin America in basic water and sanitation services (5,100) have reduced the prevalence of these diseases in the last decade. The World Bank estimates spending for water and sanitation in developing countries at nearly $26 billion per year (101). Not until clean drinking water and improved sanitation services are universally available will millions of people stop dying from preventable water-related illnesses (8). Rogers (101) estimated that completely meeting basic water supply needs up to the year 2020 would require total capital costs of about $24 billion per year. If the additional costs of meeting a higher level of services, such as advanced wastewater treatments, were included, the cost would be up to $50 billion a year. Several studies have reported the high reduction in water-related morbidity and mortality as a result of improvements in water services and sanitation

26

WATER AND HUMAN HEALTH

consequent to these huge financial investments. According to Population Reports, a review in 1991 of more than 100 studies of the effects of clean water and sanitation on human health revealed a medium reduction in deaths from water-related diseases (up to 69%) among residents, who have access to clean drinking water and improved sanitation services, because effective disposal of human wastes controls the spread of infectious agents and interrupts the transmission of water-related diseases. Table 4 shows the impact of improved water infrastructure on reducing water-related diseases. Another approach that has shown improved water quality and reduced the incidence of waterborne disease (for example, diarrhea) is the Center for Disease Control (CDC) safe water system. This system combines locally produced sodium hypochlorite solution (chlorine bleach), a CDC water storage vessel, and a public health campaign to change the behavior of rural dwellers to basic hygiene (102,103). This system has improved water quality and reduced the incidence of diarrhea by 68% in Uzbekistan (104), by 44% in Bolivia (105), and by 48% in Zambia (106). It further serves as an alternative method of disinfecting drinking water in rural Guatemala and prevents excessive morbidity and mortality from waterborne diseases (103). Moreover, according to Population Reports (3), some water development schemes have started disease control programs along with construction of water and sanitation facilities. As a result of such a program in the Philippines, for example, the prevalence of water-related diseases fell from 24% in 1979 to 9% in 1985 (67). There are indications of good progress made so far, and, at this point, some water-related diseases are on the verge of complete eradication. A good example is guinea worm (dracunculiasis) eradication. Pakistan, reportedly the first country to have completely eradicated guinea worm during the new global eradication program, recorded zero cases every month since October 1993 (8). In India, guinea worm was completely eliminated from the Tamil Nadu area in 1984, the Gujarat area in 1989, and Maharashtra in 1991; only nine cases were reported for the entire country in

1996 (60). In the Kwara State of Nigeria (where guinea worm once had devastating effects on the rural populace), any reported case of guinea worm now attracts a monetary prize (107). Globally, only five of the countries that had guinea worm recorded slightly above 100 cases in 1996. The number of cases has generally dropped by nearly 97% during the past decade (8). Although the eradication program has shown impressive progress, guinea worm is still prevalent in nearly 17 developing nations, mostly in Africa as of the end of 1996 (8). In the final analysis, guinea worm has been eradicated most effectively by providing protected clean drinking water in all countries where the disease was prevalent. Successful eradication programs for guinea worm and other water-related diseases are documented in the series of articles in Population Reports, Series M (3), as well as in (5,59,60,100,108–110). Therefore, an improved water supply and sanitation system will consequently generate tremendous improvements in the health, social welfare, and economic development of any nation, especially a developing one. For details of this and several other studies related to reduction in waterborne morbidity and mortality as a result of improvements in sanitation and water supplies, refer to Esrey et al. (111), Alam (112), Aung and Thein (113), Baltazar (114), Cairncross and Cliff (115), Young and Briscoe (110), Esrey and Habicht (116), Henry (117), Rahman (118), Haines and Avery (119), Khan (120), Torun (121), Ankar and Knowles (122), Koopman (123), Misra (124), and White et al. (36). SUMMARY Water is essential for life and health and has cultural and religious significance. Water plays a vital role in transmitting infectious diseases, and 80% of diseases reported are directly or indirectly water-related. Scarce and unclean water supplies are critical public health problems in many parts of the world and are likely to be one of the major factors that will limit economic development in the near future. It has been reported that

Table 4. Impact of Improved Water Infrastructure on Reducing Water-Related Diseasesa Place Teknaf, Bangladesh Northeast Brazil

Khuzestan, Iran Uttar Pradesh, India Peninsular Malaysia Kwara State, Nigeria Cebu, Philippines St. Lucia Lusaka, Zambia a

Type of Facilities or Improvement Hand pumps and health education Latrines, communal taps, laundry facilities, showers, and hand pumps Courtyard latrine and public standpipes Piped water Toilets and running water Boreholes, hand pumps, and health education Private, sanitary latrines Household water and latrines Extension of piped water supply

Source: Selected studies compiled in Reference 3.

Type of Study

Diseases

Difference in Incidence After Improvement

Case-control Case-control

Diarrheal diseases Schistosomiasis

17% difference between groups 77% difference between groups

Case-control

Ascariasis

16% difference between groups

Before and after Case-control

Dysentery Diarrheal diseases

Before and after

Dracunculosis

76% reduction 82% difference in infant mortality between groups 81% reduction

Before and after Case-control Before and after

Diarrheal diseases Ascariasis Typhoid

42% reduction 31% difference between groups 37% reduction

WATER AND HUMAN HEALTH

water shortages, polluted water, and unsanitary living conditions claim millions of lives annually via various water-related diseases. Research has shown that, at any one time, there are probably millions of people who have trachoma, elephantiasis, bilharzias (snail fever), malaria, diarrhea, dracunculiasis (guinea worm disease), and onchocerciasis (river blindness). The incidence of these diseases, it has been shown, depends on local climate, geography, culture, sanitary habits and facilities, and on the quantity and quality of the local water supply as well as methods of waste disposal. Effective disposal of human wastes controls the spread of infectious agents and interrupts the transmission of water-related diseases. The role of good quality drinking water and access to adequate sanitation facilities in reducing water-related diseases has been reviewed in this article. The universal provision of treated pipe-borne water is not currently feasible due to economic and political constraints, and this consequently leaves millions of people without access to safe drinking water. Generally, the failure to provide clean drinking water and adequate sanitation services has serious implications for human health and is consequent to severe economic loss in many countries. Furthermore, the water requirements for each of the basic human activities vary with domestic conditions, lifestyle, culture, tradition, diet, technology, and wealth. In any case, the minimum amount of water needed for drinking, cooking, bathing, and sanitation is set at 50 liters. The average person needs a basic minimum of 5 liters of water per day to survive in a moderate climate at an average activity level. This is the absolute minimum amount of water required to maintain adequate human health, independent of lifestyle and culture. About 60% of all infant mortality is linked to infectious and parasitic diseases; most are water-related, and a large percentage of these diseases is attributable to inadequate water supply and sanitation. Diarrheal diseases, for example, leave millions of children underweight, mentally as well as physically handicapped, and vulnerable to other diseases. However, the huge financial investments by governments in Asia, Africa, and Latin America (as well as by nongovernmental organizations) in basic water and sanitation services have reduced the prevalence of these diseases in the last decade.

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103. Rangel, J.M. et al. (2003). A novel technology to improve drinking water quality: A microbiological evaluation of inhome flocculation and chlorination in rural Guatemala. J. Water Health 1(1): 15–22. 104. Semenza, J. et al. (1998). Water distribution system and diarrheal disease transmission: a case study in Uzbekistan. Am. J. Trop. Med. Hyg. 59(6): 941–946. 105. Quick, R. et al. (1999). Diarrhea prevention in Bolivia through point-of-use water treatment and safe storage: A promising new strategy. Epidemiol. Infect. 122(1): 83–90. 106. Quick, R. et al. (2002). Diarrhea prevention through household-level water disinfection and safe storage in Zambia. Am. J. Trop. Med. Hyg. 66(5): 584–589. 107. Adebisi, S.A. (2003). Medical Consultant & Director, Community-Based Health Program, Chemical Pathology & Immunology Department. University of Ilorin, Ilorin, personal communication, June 14, 2003. 108. Carter Center. Former President Jimmy Carter travels to Africa for guinea worm eradication. News release, Mar. 17, 1995. Online: http://www.cc.emory.edu/Carter Center/ RLS95/GWtrip.htm. 109. Carter Center. USAID joins Carter Center for final assault on guinea worm. Carter Center News, Fall 1994. Online: http://www.cc.emory.edu/Carter Center/CCN-F94/gw.htm. 110. Young, B. and Briscoe, J. (1987). A case-control study of the effect of environmental sanitation on diarrhea morbidity in Malawi. J. Epidemiol. Community Health 42: 83–88. 111. Esrey, S.A., Potash, J.B., Roberts, L., and Shiff, C. (1991). Effects of improved water supply and sanitation on ascariasis, diarrhea, dracunculiasis, hookworm infection, schistosomiasis, and trachoma. Bull. World Health Organ. 69(5): 609–621. 112. Alam, N. (1989). Mother’s personal and domestic hygiene and diarrhea incidences in young children in rural Bangladesh. Int. J. Epidemiol. 18: 242–247. 113. Aung, M.H. and Thein, H. (1989). Prevention of diarrhea and dysentery by hand washing. Trans. R. Soc. Trop. Med. Hyg. 83: 128–131. 114. Baltazar, J. (1988). Can the case-control method be used to assess the impact of water supply and sanitation on diarrhea? A study in the Philippines. Bull. World Health Organ. 66: 627–635. 115. Cairncross, S. and Cliff, J.L. (1987). Water use and health in Mueda, Mozambique. Trans. R. Soc. Trop. Med. Hyg. 81: 51–54. 116. Esrey, S.A. and Habicht, J.P. (1986). Epidemiologic evidence for health benefits from improved water and sanitation in developing countries. Epidemiol. Rev. 8: 117–128. 117. Henry, F.J. (1981). Environmental sanitation infection and nutritional status of infants in rural St. Lucia, West Indies. Trans. R. Soc. Trop. Med. Hyg. 75: 507–513. 118. Rahman, M. (1985). Impact of environmental sanitation and crowding on infant mortality in rural Bangladesh. Lancet 8445(2): 28–31. 119. Haines, M.R. and Avery, R.C. (1982). Differential infant and child mortality in Costa Rica: 1968–1973. Population Stud. 36: 31–43. 120. Khan, M.U. (1982). Interruption of shigellosis by handwashing. Trans. R. Soc. Trop. Med. Hyg. 76: 164–168. 121. Torun, B. (1982). Environmental and educational interventions against diarrhea in Guatemala. In: Diarrhea and Malnutrition: Interactions, Mechanisms, and Interventions. L.C. Chen and N.S. Scrimshaw (Eds.). Plenum Press, New York, pp. 235–266.

122. Anker, R. and Knowles, J.C. (1980). An empirical analysis of morbidity differentials in Kenya at the macro- and microlevels. Econ. Dev. Cul. Change 29: 165–185. 123. Koopman, J.S. (1978). Diarrhea and school toilet hygiene in Cali, Colombia. Am. J. Epidemiol. 107: 412–420. 124. Misra, K.K. (1975). Safe water in rural areas. Int. J. Health Educ. 18: 53–59.

READING LIST Agardy, T. (1997). Marine Protected Areas and Ocean Conservation. Academic Press, Austin. Falkenmark, M. and Lindh, G. (1993). Water and economic development. In: Water in Crisis. P. Gleick (Ed.). Oxford University Press, New York, pp. 80–91. Safe Drinking Water Communication. (1980). Drinking Water and Health, Vol. 3. National Academy Press, Washington, DC, p. 415.

NITRATE HEALTH EFFECTS SEGUN MICHAEL ADE ADELANA University of Ilorin Ilorin, Kwara State, Nigeria

INTRODUCTION Nitrate is a stable nitrogen (N) species under certain natural conditions and forms highly soluble compounds. These are peculiar features that allow nitrate ion to be transported in some groundwater systems to environments where it can be converted into other nitrogen species that either promote surface water eutrophication or are hazardous to humans, livestock, and the environment. Nitrate test results are usually expressed in milligrams per liter as either nitrogen (NO3 N, sometimes written as plain N) or as nitrate (NO3 ). The following conversion factors can be useful for nitrate reporting: 1 milligram (mg) of compound expressed as nitrogen (N) is equivalent to 4.43 mg when expressed as nitrate (NO3 − ). 1 milliequivalent (meq) of compound expressed as N is equivalent to 62 mg when expressed as NO3 − . Therefore the nitrate reporting expressions (mg NO3 N/L, mg NO3 /L) are used interchangeably throughout this article without any special preference for one or the other. Problems of nitrate pollution, particularly in groundwater, are widespread in many countries of the world. In the United States, several illustrations of nitrate pollution in groundwater are available in the literature. For example, more than 30 references dealing with nitrate groundwater pollution studies in 15 of the 50 states in America have been reviewed. Information from all these references have been tabulated and presented in detail in Canter (1). Survey results from a study of 25 pesticides and nitrate in 201 rural wells in eight agricultural areas in Missouri

NITRATE HEALTH EFFECTS

revealed that 22% of the wells exceeded the drinking water standard for nitrate (2). A summary of the nature and geographical extent of nitrate pollution of groundwater in Nebraska is given in Exner and Spalding (3). A total of 27 references dealing with case studies of nitrate groundwater pollution (outside the United States) have been reviewed. Of these references, 22 addressed groundwater nitrate pollution studies in Europe. One of these references (4) is a conference proceeding with 52 papers addressing nitrogen as a surface and groundwater pollutant; Tessendorff (5) and Kraus (6) provided a general discussion of groundwater nitrate problems in the European community countries (EEC). The countries noted with adequate references on groundwater nitrate problems include Czechoslovakia, Germany, The Netherlands, United Kingdom, Denmark, Israel, and Chile. Overgaard (7), in a nationwide investigation of nitrate concentrations in groundwater in Denmark, based on analyses of samples from about 11,000 wells and drinking water from 2800 groundwater works, revealed that the overall mean level of nitrate in groundwater had trebled within the last 20–30 years and is increasing at a rate of about 3.3 mg NO3 /L per year. Consequently, the result showed that 8% of the water produced in Danish Waterworks now has a nitrate concentration above the EEC guide limit of 50 mg NO3 /L (6). It has been estimated that 800,000 people in France, 850,000 in the United Kingdom, and 2.5 million in Germany are drinking water whose nitrate concentrations are above the permissible limit of the European Community (1). A survey conducted in France revealed that 81% of the population had nitrate levels less than 25 mg NO3 /L and 96 to 98% had levels less than 50 mg NO3 /L in their drinking water supplies (8). Out of the 53 million people accounted for in the survey, 280,000 at most had a water supply exceeding 100 mg NO3 /L at least once during a 3-year period. Of 20,000 distribution units surveyed, nearly 1000 had nitrate above 50 mg NO3 /L; however, only 61 units were above 100 mg NO3 /L. Most of the high nitrate levels were found in groundwater supplies. Custodio (9) reported that agricultural nitrate pollution is a widespread problem in irrigated areas in Spain, where nitrate in groundwater often exceeds 50 mg NO3 /L and sometimes reaches 500 mg NO3 /L. A regional survey of nitrate in the Anglian area of the United Kingdom in 1975 indicated that 50 public supply boreholes, wells, and springs had recorded nitrate levels in excess of 11.3 mg/L (as nitrogen), 50 mg NO3 /L (10). Due to increasing agricultural activity after the 1960s, both shallow and deep-water resources in the Czech Republic, including karstic systems, have been contaminated by infiltrating nitrate. The nitrate content of one of the largest springs (yielding up to 19 L/s during minimum discharge) in the Republic now varies from 50 to 60 mg NO3 /L (11). Others reported nitrate pollution studies in Canada (12,13), India (14,15), Israel (16), Chile (17), Portugal (18), Southern Africa (19,20), Nigeria (21–25), Ghana (26), Burkina Faso (27), and Senegal (28). Note the pollution of groundwater from industry and waste dumps is a serious problem, particularly in the more

31

developed countries in the European community. This is also the case in the United States, Canada, and Australia. High concentrations of nitrates are the other main cause of groundwater pollution. Concentrations of nitrate that approach or exceed 10 mg/L as N, equivalent to 44.3 mg/L as NO3 , present health hazards. Thus, the international drinking water quality standard is set at 10 mg/L for NO3 N (or plain nitrogen), and it is approximately 45 mg/L as NO3 (29). Therefore, it is important to consider the health implications of this common pollutant of groundwater with respect to humans, livestock, and the environment, which is the main focus of this paper. OCCURRENCE AND EXTENT OF NITRATE IN WATER AND FOODS The occurrence of nitrate concentrations has been reported in many parts of the world such as Europe, the United States, Australia, Chile, Ghana, South Africa, Nigeria, and Cote d’Ivoire. The World Health Organization (WHO) studied the occurrence of nitrate in water and came to the conclusion that nitrate concentrations in surface waters have increased substantially over the last 30–40 years (29). Many countries, mostly in Western Europe and the United Kingdom, showed a more marked increase in the levels of nitrates in groundwater, especially between 1970 and 1980 (30). The reason for these increases in groundwater nitrate is not unconnected with the vast increase in fertilizer application and other forms of animal manure. In the United States, the occurrence of high nitrate concentrations in groundwater is widespread, particularly as a result of agricultural usage of fertilizers or land disposal of domestic wastewaters. Much research has been conducted to determine the amounts of nitrates in drinking water wells. ‘‘USGS data show that the 20 states with the largest agricultural marketing in 1989 had a notably higher percentage of wells with nitrate concentrations above 10 mg NO3 -N/L than the remaining 30 states, 7.1% compared to 3.0%, respectively’’ (31). Research conducted by private firms also links high nitrogen content in wells with agricultural activity. A survey of 1430 randomly selected and sampled drinking water wells in agricultural areas of 26 states, conducted in 1988 and 1989 by the Monsanto Agricultural Products Company, found nitrate above 10 mg NO3 -N/L in 4.9% of the wells. For wells on farmsteads only, however, the proportion was 10% (31). Monsanto, however, concluded that the frequency of wells with nitrate exceeding 10 mg NO3 -N/L doubled for wells located on farm property. Several case histories have indicated the geographical extent and seriousness of high nitrate in groundwater in the United States. For example, the Metropolitan Water District of Southern California has indicated that annually it loses 4% of its drinking water supply primarily to nitrate pollution, compared to less than 1/2% from toxic organic chemicals (1). About 12% of the wells sampled in the service area exceeded the state maximum contaminant level for nitrate. It was estimated that by the year 2000, the groundwater in most of the water table aquifers in Salinas Valley

32

NITRATE HEALTH EFFECTS Table 1. Summary of Nitrate Data in the United Statesa State Illinois Indiana Kentucky Louisiana New Jersey Ohio Virginia West Virginia a

Counties Tested

Number of Samples

Average Nitrate Concentration, mg/L

Percent Over 10 mg/L

8 33 90 23 5 80 24 13

286 5,685 4,559 997 1,108 18,202 1,054 1,288

5.76 0.92 2.50 1.19 2.60 1.32 2.92 0.83

19.9 3.5 4.6 0.8 6.8 3.0 7.1 0.8

Reference 33.

would have exceeded the state’s drinking water standard of 45 mg/L for nitrate (32). It is expected that the rising trend in groundwater nitrate will continue for many more years, even if nitrate leaching from soils is reduced by changes in agricultural practices. A summary of nitrate data by state is presented in Table 1. This was the result of the multistate groundwater quality testing undertaken for several years at Heidelberg College in Tiffin, Ohio and reported by Swanson (33). The data indicate higher nitrate occurrence in the sampled groundwater from Illinois, Virginia, New Jersey, and Kentucky. The extent of nitrate and nitrite as (both anthropogenic and natural) water pollutants has been widely reported and published in the hydrogeologic and biochemical/pharmacological literature, especially, in developed nations (1,13,34–43). Such information is scanty in developing countries until recently (19–27,39,44–48). It has been reported that nitrate health hazards are also posed greatly by solid foods from agricultural and diary products and preservatives (24,44,49), yet the greatest threats come from polluted water; fluids and foods cooked and washed are contaminated directly or indirectly with such waters. The occurrence of nitrate is known in foods, especially in African foods. Okonkwo et al. (49) researched and reported high concentrations (in parts per million) for various African foods (Table 2). Most of these foods are heavily consumed in West Africa and are now largely exported to Europe, America, Canada, and Australia. If the rate and volume at which these foods are consumed continues, nitrate could become potentially hazardous. A high percentage of the world population, as projected by the WHO (29), would ingest increased nitrates in the near future. Nitrates are easily converted to nitrite (the more poisonous form of nitrogen) by various mechanisms. It has been observed that drying tends to change the nitrite content of food items (Table 3). Drying is a common method of food preservation in the tropics, particularly in developing countries. Drying reduces the ascorbate level in food thereby inhibiting its antagonism to the carcinogenic action of nitrites (50). Thus, dried foods especially vegetables, have high nitrite concentrations and consequently increase the hazard. Ezeonu (51) reported high nitrate concentration in some Nigerian beers with frightening statistics. All ten brands of beer selected for study whose production locations are in different parts of the country showed high nitrate content. Yet large volumes of beer are consumed daily in Nigeria. Most breweries in

Table 2. Nitrate Levels in Typical African Foodsa Beverages

Range, ppm

Mean, ppm

0.1–1.4 0.1–0.3

1.2 0.2

2.0–4.9

3.0

Palm Wine Tap water (pooled urban supply) Cereals Guinea Corn & Rice

Proteins (Animal Source) Crayfish Fish Meat

18.2–30.8 7.9–10.4 0.4–1.7

28.7 8.6 0.7

Proteins (Plant Source) Beans (black) Bean (white) Ground nuts Melon Pigeon pea

3.2–6.6 3.1–6.5 4.4–9.7 9.0–11.6 1.9–2.7

4.9 4.9 6.1 10.3 2.4

Vegetables and Fruits Bitter leaf Flutted pumpkin leaf Garden egg leaf Green amaranth leaf Okra Pumpkin lead

8.0–9.8 0.8–1.2 0.7–1.2 0.5–0.8 2.0–2.2 5.6–15.0

8.9 1.1 1.0 0.7 2.1 11.3

Nitrate Not Detected in these Foods Beer (premier brand) Native Gin a

Corn Cassava (Gari)

Sida Yams

After Reference 49.

Nigeria do not treat their water sources (i.e., boreholes) for nitrate, yet anthropogenic nitrate is widespread in the country (21). WATER QUALITY CONCERNS RELATED TO NITRATE LEVELS The issue of water quality related to nitrate concentrations, especially in drinking water, is of concern worldwide in view of the various health implications of ingesting high doses of nitrates. Usually, drinking water contributes 0 ≥10

85,300,000 2,980,000

78,100,000 1,600,000

98,900,000 4,260,000

a

Reference 59.

be the origin of the nitrate. Nitrate sources other than applied fertilizers such as wastes from livestock, dairy, or poultry and accidents or careless precautionary handling of fertilizers near well sites may be involved (31). In the United States, for example, there are locations with high agricultural production in at least 14 states where nitrate contamination has been associated with the application of nitrogen fertilizers (1). In certain categories, organic and inorganic compounds of nitrogen, phosphorus, and potassium that originate from many commercial fertilizers may be released into groundwater. In some cases, it has been demonstrated that the leaching of nitrate is accelerated by irrigation (1,31,61). Septic tank systems also represent a significant fraction of the nitrogen load to groundwater in the United States (62). About 25% of the population in America is served by individual home sewage disposal systems. Research revealed that effluent from a typical septic tank system has a total nitrogen content of 25 to 60 mg/L (62). In the Netherlands, the reason for the rising nitrate level can be due to the application of nitrogen fertilizer (63,64). Jacks and Sharma (15) reported nitrate levels in excess of 300 mg/L (as N) in wells in Southern India owing to anthropogenic and agricultural influences. In Australia, biological fixation in the soil is considered the principal origin, although point sources such as sewage effluent, animal and industrial waste could be significant locally (65). However, in Nigeria and most parts of West Africa, high nitrate levels in groundwater result mostly from indiscriminate waste disposal (21–25,27) and agricultural activities (26,46,47). Table 7 is a summary of nitrate sources in groundwater. HEALTH EFFECTS OF NITRATE Nitrate and the nitrite form of nitrogen constitute a general public health concern, related especially

to infant methemoglobinemia (infantile cyanosis) and carcinogenesis (66). The concentrations of nitrate and nitrite in foods that include vegetables, crayfish, meat, etc. and drinking water may indicate serious potentials for pollution and also could result in severe health problems. For example, the nitrate levels in Nigerian foods that include drinking water (from surface and subsurface sources) and beverages are reportedly high and generally perceived to be associated with adverse health effects in humans (24,49,51,67). These have resulted in reported cases of water-related diseases such as diarrhea in children or cancerous diseases that claimed lives yearly (24,68,69). According to Population Reports (70), diarrheal disease is a class of waterborne disease, which is prevalent in many countries where sewage treatment is inadequate or where human wastes are disposed of in open latrines, ditches, canals, and watercourses or is spread indiscriminately on farmland. These practices are frequent in developing nations and favor the accumulation of anthropogenic nitrate. Certain vegetables (e.g., lettuce, spinach, beetroot, and celery) contain relatively high levels of nitrate [>3000 mg/kg for lettuce (Ref. 71)] but the nitrite levels are usually very low. Nitrates and nitrites are also added as preservatives in some foods, such as cured meats, consequently exposing consumers to higher health risks. The World Health Organization (29) estimated daily dietary intake of nitrate and nitrite in different countries. In most European countries, the mean nitrate intake is about 10–30 mg/day. Vegetarians usually have a two to fourfold higher intake of nitrates than nonvegetarians. In India, it has been estimated that 20–50% of the wells in areas of high population density produce water whose nitrate level is above 50 mg/L, thus causing severe health hazards (72,73). Terblanche (74) reviewed the health hazards of nitrate in drinking water in many developed countries, including South Africa. Of an estimated 219 million people using public drinking water

NITRATE HEALTH EFFECTS

35

Table 7. Examples of the Various Sources of Nitrate in Groundwatera Natural Sources Geologic nitrogen which can be mobilized and leached to groundwater via irrigation practices Unmanaged (natural) climax forests that are normally nitrogen conserving; however, nitrogen losses to groundwater can occur from human-initiated clear cutting and other forest disturbances Waste Materials Animal manures, which may be concentrated in large commercial poultry, dairy, hog, and beef operations Land application of municipal or industrial sludge or liquid effluent on croplands, forests, parks, golf courses, etc. Disposal of household wastes or small business wastes into septic tank systems (septic tank plus soil absorption field) Leachates from sanitary or industrial landfills or upland dredged material disposal sites Row Crop Agricultureb Nitrogen losses to the subsurface environment can occur as a result of excessive fertilizer application, inefficient uptake of nitrogen by crops, and mineralization of soil nitrogen Nitrogen losses to the subsurface environment can occur as a function of fertilizer application rates, seasonal rainfall and temperature patterns, and tillage practices Irrigated Agriculture Enhanced leaching of nitrogen from excessive fertilizer application rates and inefficient irrigation rates Associated leaching of nitrogen from soils periodically subjected to leaching to remove salts so that the soils do not become saline and unproductive a b

Culled from Reference 1. Refers to annual crops.

supplies in the United States, approximately 1.7 million are exposed to nitrate levels above 10 mg/L. About twothirds of those exposed, 1.1 million, are served by public water systems using groundwater supply sources. Almost 27,000 infants a year are exposed to tap water with nitrate levels exceeding 10 mg/L (31). The resulting health hazards and associated statistics in the United States are documented in the Federal Register (75). The following section describes the details of the various health effects of nitrate. Methemoglobinemia High nitrate levels in water can cause infant methemoglobinemia. Methemoglobinemia is a disease primarily affecting babies and is often described by the lay term ‘‘blue baby syndrome.’’ Infants are the primary concern because they are the most vulnerable. The USEPA standard for nitrate in drinking water is set at 10 mg/L to protect babies under about 3 months of age. Such infants are much more sensitive to nitrate toxicity than the rest of the population for many reasons. For example, bacteria that live in the digestive tracts of such infants convert nitrate into toxic nitrite. Nitrite transforms hemoglobin to methemoglobin, preventing transport of oxygen and producing symptoms of asphyxiation (another term for blue baby syndrome). This methemoglobin is considerably more stable than the oxygen hemoglobin complex that fulfils the oxygen transport function of the blood. Once the concentration of methemoglobin in the blood exceeds 5% the first symptoms of ‘cyanosis’ are generally noticeable; anoxia (death) results at levels of 50% and higher (74) or if the condition is left untreated (66). After babies reach the age of 3 to 6 months, acid in their stomachs increases, thereby creating an unfavorable environment for the bacteria

causing the problem (31). It must be borne in mind that nitrate itself has low primary toxicity, but acute toxicity occurs as nitrate is reduced to nitrite (NO2 ), a process that can occur under specific conditions in the stomach and saliva (66). Consequently, the nitrite ion formed becomes an oxidizing agent, transforming hemoglobin in the blood to methemoglobin (29), thereby preventing transport of oxygen and resulting in methemoglobinemia. Most reported cases of infantile methemoglobinemia have been associated with the use of water containing more than 10 mg/L NO3 -N. The occurrence of infant methemoglobinemia from consumption of water with high nitrate concentrations was first recognized clinically by Comly (72). The infants were both less than 1 month old and had received rural well water containing 90 and 140 mg/L, respectively. Earlier, Comly (72) suggested a recommended limit of 10 mg/L NO3 -N in drinking water and a maximum of 20 mg/L. Later on, Shuval and Gruener (76) studied 1702 infants living in the Israel coastal plain in areas with medium to high nitrate (11.3 to 20.3 mg NO3 -N/L) and compared them with a control group of 759 infants in Jerusalem where only 1.1 mg/L of nitrate is in the water supply. There were no significant differences found between the methemoglobin levels in the 1702 infants in the study areas compared to the 758 infants in the control area. In most countries, methemoglobinemia is not a notifiable disease, making its true incidence unknown. From 1945 until 1970, some 2000 cases of methemoglobinemia have been reported in the world literature (76) with a case fatality of about 8%. The WHO (29) cites literature indicating that 10 cases of methemoglobinemia have been reported in the United Kingdom since 1950 when the first cases of methemoglobinemia were reported in East Anglia. Only one death was reported during this period. In 1986, a 2-month-old infant in South Dakota

36

NITRATE HEALTH EFFECTS

(USA) died of methemoglobinemia (77). The exact nitrate concentration is unknown. In another nonfatal case in Iowa, the water apparently contained 285 mg/L nitrate (as N) but the 5-week-old infant survived (42). Hungary is one of the countries, with exceptions, that possibly has the best statistics on the occurrence of infantile methemoglobinemia. Table 8 shows the occurrence of methemoglobinemia in Hungary between 1976 and 1990 (78). Methemoglobinemia became a notifiable disease in Hungary in 1968 (29), and in the first 5 years after 1968, 883 cases were reported. Of the recorded cases, 92% had a nitrate level in the drinking water exceeding 22.6 mg/L as N; in the remaining 8%, it was between 9 and 22.6 mg/L. The highest number of cases was reported in 1977, and the measures taken to supply the population with drinking water low in nitrate have resulted in a definite decrease in the number of cases each year. For detailed statistics on infant methemoglobinemia, readers are referred to WHO (29) and Csanady (78). In a later publication, Shuval and Gruener (79) confirm a direct relationship between the occurrence of methemoglobinemia in infants and high concentrations (>10 mgNO3 -N/L) of nitrate in water. According to Ross and Desforges (80), other factors important in the pathogenesis of the disease are age, the presence of bacteria in sufficient numbers in the gastrointestinal tract, gastric acidity (a pH >4), gastrointestinal disturbances, the types of powdered milk product used as baby food, high fluid intake, and the effect of nutrition because foods rich in nitrate can increase the severity of illness. The first epidemiological survey in South Africa to assess the effect of well water nitrates on infant health was published (57) after the review of health hazards by Terblanche (74). The survey was undertaken due to the risk of methemoglobinemia in infants in the Rietfontein area as a result of the large number of boreholes where nitrate-nitrogen exceeded 10 mg/L. Unfortunately, no correlation was found between the nitrate content of the groundwater used and the methemoglobin levels in the Table 8. The Occurrence of Methemoglobinemia in Hungarya Year 1976 1977 1978 1979 1980 1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 TOTAL a

Reference 78.

Number of Cases

Number of Fatalities

207 293 239 180 172 166 91 67 33 46 41 30 31 35 22 1653

4 7 3 2 3 1 1 — — 1 — — 2 2 — 26

blood. A clinical health risk assessment was attempted in South Africa due to high ingestion of nitrate water (81). There was an increased risk of methemoglobinemia as a result of increased in bottle-feeding by HIV positive mothers. In South Africa, it has been shown that breastfeeding increases the risk by 12–43% that HIV-positive mothers transmit the virus to their children (82). For this reason, the South African Department of Health advises HIV-positive mothers to bottle-feed infants to reduce the risk of mother to child transmission of the HIV virus via breast milk (81). This assessment attempts to quantify the potential additional total exposed population (PATEP) facing increased risk from nitrates. Table 9 shows the potential additional total exposed population at significant risk of methemoglobinemia as a result of bottle-feeding infants by HIV-positive mothers. The number of infants has been calculated as proportional to the area of the province in which groundwater contains >10 mg/L NO3 N (81). According to Colvin (81), Gauteng has 10% of the PATEP for >10 mg/L NO3 -N groundwater, due mainly to the high infant population density (7.5 per km2 ). This province includes over half of the total area underlain by groundwater with >50 mg/L NO3 -N and has 88% of the potential additional population exposed to very high risk (in South Africa), shown in Table 10. The qualitative banding of degrees of risk following the matrix model developed by Carpenter and Maragos (83) and simplified by Genthe (84) is shown in Fig. 1. Severity is defined according to the potential for exposure to nitrate in groundwater sources. For example, a risk that combines high severity with frequent probability is rated ‘‘high,’’ whereas a risk that combines low severity with occasional or the remote probability of occurrence is rated ‘‘acceptable.’’ The highest risk is for infants 50 mg/L NO3 -N, and confounding factors (such as lack of vitamin C and gastrointestinal infections) are evident (81). In another epidemiological study on a comparable population group in Namibia (56), a correlation was found between the nitrate level in the groundwater and blood methemoglobin levels. The main difference between the two studies was that the level of nitrate in groundwater in the Namibian study was much higher, up to 56 mg/L of nitrate-nitrogen (52). However, the ingestion of nitrate has no apparent short-term effects on adults such as methemoglobinemia. Research has shown that adults on a farm near Otjiwarongo, Namibia, continued drinking water with 268 mg/L of nitrate with no apparent ill effects, even after stock losses occurred on that farm (85). Gastric Cancer Infants are not the only ones at risk; it is possible that high nitrate concentrations can cause cancer in adults. ‘‘Nitrate itself is not directly carcinogenic. However, there is recognition of the fact that nitrate could be converted to nitrite in the human body that can react with secondary and tertiary amines to form nitrosamines—which have been identified as potent carcinogens’’ (31). Several studies have shown that simultaneous ingestion of nitrite (or nitrate with amines) results in cancers of many organ

NITRATE HEALTH EFFECTS

37

Table 9. Potential Additional Total Population Exposed to Significant Risk of Methemoglobinemia as a Result of Bottle-Feeding by HIV-Positive Mothersa >10 mg/L Total Infants %HIV+, Estimated% %Area of 50 mg/L NO3 -N Area (km2 )

%Area of Province

Total Infants 50 mg/L NO3–N

Reasonably Probable Bottle-fed withwater containing 20–50 mg/L NO3–N

Occasional Bottle-fed with water containing 10–20 mg/L NO3–N

Remote Bottle-fed with water containing 12 months old

Key

Risk

Recommendation

High risk

Do not advise consumption of this water — use alternative low nitrate source.

Medium risk

Only use this water mixed with low nitrate water and if infants vitamin C intake is sufficient.

Low risk

Preferably mix the water for feeding with low nitrate water and ensure infant’s vitamin C intake is sufficient.

Acceptable risk

None.

Figure 1. Risk characterization of methemoglobinemia (after References 83 and 84).

chronic exposure to high levels of nitrate in drinking water may have adverse effects on the cardiovascular system. The WHO (29), however, reported that an inverse relationship between cardiovascular mortality and nitrate concentration in water supplies had been demonstrated. Moreover, excessive nitrates in drinking water have also resulted in problems with ruminants (cud-chewing animals with divided stomachs). Sheep and cattle, in particular, can be seriously affected by nitrates from birth through adulthood (1). Infants of monogastric (singlestomach) animals like horses, pigs, and chickens are also susceptible to problems from nitrate ingestion. However, as chickens and pigs mature, they are much

less susceptible to the health effects of nitrate, but horses can be affected through adulthood (90). Other possible effects of nitrates relate to the thyroid function in animals. Some animal studies indicate that chronic exposure to high levels of nitrates can reduce the intrathyroid iodine pool and thus render the gland more sensitive to goitrogens (29). However, whether or not exposure to nitrate is an etiological factor in human goiter remains unclear (52). Symptoms of nitrate-nitrite poisoning in livestock include cyanosis in or about the non pigmented areas (mouth and eyes), shortness of breath, rapid heartbeat, staggered gait, frequent urination, and collapse. In severe cases, convulsions, coma, and death may result within a

Gastric cancer death rate/100,000 population

NITRATE HEALTH EFFECTS

50

40

Japan R = +0.88 Romania

30 Czechoslovakia West Germany Yugoslavia 20 Norway Netherlands UK Sweden Switzerland Denmark 10 USA 0 Daily per capital nitrate intake, mg

Figure 2. Relationship between gastric cancer mortality rates and nitrate ingestion in 12 countries (after Reference 88).

few hours (1). Loss of milk production in cows and aborted calves are also indicative of nitrate poisoning (90). Stock losses due to nitrate poisoning have been reported in Namibia (91), in the dolomitic area of South Africa (92), and in Bophuthatswana (52). Exposure to high doses of nitrate is also associated with adverse effects such as the ‘hot-dog headache’ (52). The ‘hot-dog headache’ has been described in the literature as related to nitrites used in curing meat to give it a uniform color (93). Nitrites are also vasodilators, so that some people find that soon after eating these meat products, they develop flushing of the face and headache. The ‘hotdog’ is the classical example, but other meat products, including bacon, ham, and salami, can also cause these symptoms (52). A farmer in the Springbok Flats regularly complained about hot-dog headaches. The problem was solved when he started using nitrate-free water (94). Finally, detailed information on the health effects of nitrate and nitrite in humans and animals, the mechanism and quantification of toxicological effects of nitrate and nitrite, as well as other health-related information on nitrates are available in literature (29,43,56,74,76,79, 95–97). Obviously, pollution and contamination problems from NO3 -NO2 compounds must be causing environmental public health havoc yet to be fully determined and documented. Therefore, nitrate-nitrite pollution control programs must be established to reduce the health effects of these common and widespread contaminants. Excessive chemical fertilizer and animal manure applications must be controlled and curtailed. Exceptionally high NO3 borehole waters must be abandoned, and unpolluted groundwater exploited. These and possible biodenitrification water treatment (for affected waters before use) could contribute solutions. SUMMARY The occurrence of high nitrate concentrations in groundwater is widespread, particularly from agricultural usage of fertilizers and animal manure or land disposal of domestic waste and wastewaters. Much research has been

39

conducted to determine the amounts of nitrates in drinking water wells as well as in foods. Exposure to high doses of nitrate is generally perceived to be associated with adverse health effects in humans and other species. These range from infant methemoglobinemia, cancers, the ‘hot dog headache,’ and hypertension, to other adverse effects such as birth defects (congenital malformations) and spontaneous abortions. Most reported cases of infantile methemoglobinemia have been associated with the use of water containing more than 10 mg/L NO3 -N. Pollution and contamination problems from NO3 -NO2 compounds may be causing environmental public health havoc yet to be fully determined and documented. The evidence as outlined in this article is overwhelming. Therefore, nitrate-nitrite pollution control programs must be established to reduce the health effects of these common and widespread contaminants. Excessive chemical fertilizer and animal manure applications must be controlled and curtailed. Exceptionally high NO3 borehole waters must be abandoned, and unpolluted groundwater exploited. These and possible biodenitrification water treatment could contribute solutions. BIBLIOGRAPHY 1. Canter, L.W. (1996). Nitrates in Groundwater. Lewis, New York. 2. Sievers, D.M. and Fulhage, C.D. (1992). Survey of rural wells in Missouri for pesticides and nitrates. Groundwater Monitoring Rev. 12(4): 142–150. 3. Exner, M.E. and Spalding, R.F. (1990). Occurrence of Pesticides and Nitrate in Nebraska’s Ground Water. Report WC1, Water Center, Institute of Agriculture and Natural Resources, University of Nebraska, Lincoln, Nebraska, pp. 3–30. 4. Anonymous. (1975). Proc. Conf. Nitrogen Water Pollutant, Int. Assoc. Water Pollut. Res., August, Copenhagen, Denmark. 5. Tessendorff, H. (1985). Nitrates in groundwater: A European problem of growing concern. Aqua 4: 192–193. 6. Kraus, H.H. (1993). The European Parliament and the EC Environment Policy. Working Paper W-2, April, European Parliament, Luxembourg, p. 12. 7. Overgaard, K. (1984). Trends in nitrate pollution of groundwater in Denmark. Nordic Hydrol. 15(4–5): 174–184. 8. Anonymous. (1983). Nitrate in water for human consumption: The situation in France (1979–1981). Aqua 2: 74–78. 9. Custodio, E. (1982). Nitrate build-up in Catalonia coastal aquifers. Memoires, Int. Assoc. Hydrogeol. 16(1): 171–181. 10. Greene, L.A. (1978). Nitrates in water supply abstractions in the Anglian region: Current trends and remedies under investigation. Water Pollut. Control 77(4): 478–491. ´ R., and Zak, ´ K. (1998). Nitrate pollution 11. Buzek, F., Kadlecova, of a karstic groundwater system. In: Isotope Techniques in the Study of Environmental Change. IAEA, Vienna, pp. 453–464. 12. Hill, A.R. (1982). Nitrate distribution in the ground water of the Alliston region of Ontario, Canada. Ground Water 20(6): 696–702. 13. Egboka, B.C.E. (1984). Nitrate contamination of shallow groundwater in Ontario, Canada. Sci. Total Environ. 35: 53–70. 14. Handa, B.K. (1983). Effect of Fertilizer use on Groundwater Quality in India. Proc. Symp. Ground Water Water Resour. Plann., Vol. II, IAHS Publication, No. 142, pp. 1105–1119.

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NITRATE HEALTH EFFECTS 52. Tredoux, G. (1993). Preliminary Investigation of Nitrate Content of Groundwater and Limitation of the Nitrate Input. WRC Report No. 368/1/93, Pretoria. 53. Packham, R.F. (1991). Public health and regulatory aspects of inorganic nitrogen compounds in drinking water. Int. Workshop, Inorg. Nitrogen Comp. Water Supply, Hamburg, November 1991. 54. Bosch, H.M. et al. (1950). Methemoglobinemia and Minnesota well supplies. J. Am. Water Works Assoc. 42: 91–95. 55. Walton, G. (1951). Suvey of literature relating to infant methemoglobinemia due to nitrate contaminated water. Am. J. Public Health 41: 1141–1152. 56. Super, M. et al. (1981). An epidemiological study of well water nitrates in a group of Namibian infants. Water Res. 15: 1265–1270. 57. Hesseling, P.B., Toens, P.D., and Visser, H. (1991). An epidemiological survey to assess the effects of well water nitrates on infants’ health at Rietfontein in the northern Cape Province, South Africa. South Afr. J. Sci. 87: 300–304. 58. United States Environmental Protection Agency. (1990). National Survey of Pesticides in Drinking Water Wells—Phase I Report, EPA 570/09-90-015, Office of Water, Washington, DC. 59. United States Environmental Protection Agency. (1992). Another Look: National Survey of Pesticides in Drinking Water Wells—Phase II Report, EPA 579/09-91-020, January, Office of Water, Washington, DC. 60. Cohen, S., (1992). Results of the National Drinking Water Survey: Pesticides and well characterized. Water Well J. 46(8): 35–38. 61. Keeney, D. (1989). Sources of nitrate to ground water. In: Nitrogen Management and Groundwater Protection. R.F. Follett (Eds.). Elsevier Science, Amsterdam, The Netherlands, Chap. 2, pp. 23–34. 62. United States Environmental Protection Agency. (1994). Nitrogen Control. Technomic, Lancaster, PA, pp. 1–22. 63. Van Duyvenboden, W.V. and Loch, J.P.G. (1983). Nitrate in the Netherlands—A serious threat to groundwater. Aqua 2: 59–60. 64. Van Beek, C.G.E.M. (1985) Nitrate situation. Paper presented at Conf. Nitrates Water, Paris. 65. Lawrence, C.R. (1983). Occurrence and genesis of nitraterich groundwaters of Australia. Int. Conf. Groundwater Man, Sydney 2: 237–247. 66. United States Environmental Protection Agency. (1993). Nitrogen Control. EPA 625/R-93-010, September 1993, Office of Water, Washington, DC. 67. Bassir, O. and Maduagwu, E.N. (1978). Occurrence of nitrate, nitrite, dimethylamine and dimethylnitrosamine in some fermented Nigerian beverages. J. Agric. Food Chem. 26(1): 200–203. 68. United States Agency for International Development. (1990). Strategies for Linking Water and Sanitation Programs to Child Survival. USAID, Washington, DC, pp. 1–62. 69. Warner, D. (1998). Drinking water supply and environmental sanitation for health. Presented at the Int. Conf. Water Sustainable Dev., Paris, March 19–21, pp. 1–10. 70. Population Reports. (1998). Solutions for a Water-short World, Series M, 14, September. 71. Parsons, M.L (1977). Current research suggests the nitrate standard in drinking water is too low. J. Environ. Health 40: 140–142. 72. Comly, H.H. (1945). Cyanosis in infants caused by nitrates in well water. J. Am. Med. Assoc. 129: 112–116.

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73. Pande, S.P., Hassan, M.Z., and Saxena, K.L. (1986). Nitrates and nitrites in the environment. J.I.W.W.A. XVIII(3): 145. 74. Terblanche, A.P.S. (1991). Health hazards of nitrate in drinking water. Water SA 17(1): 77–82. 75. Federal Register. (1985). National Primary Drinking Water Standards. 219: November 13, 46,880–47,022. 76. Shuval, H.I. and Gruener, N. (1972). Epidomiological and toxicological aspects of nitrates and nitrites in the environment. Am. J. Public Health 62: 1045–1052. 77. Groundwater Newsletter. (1986). Blue-baby death from contaminated well in South Dakota stirs investigation. Groundwater Newsl. 15(18): 1. 78. Csanady, M. (1991). Nitrite formation and bacteriological deterioration of water quality in distribution networks. Paper presented at Int. Workshop: Inorg. Nitrogen Comp. Water Supply, Hamburg, IWSA, Pre-Prints, 45–49. 79. Shuval, H.I. and Gruener, N. (1977). Infant methaemoglobinemia and other health effects of nitrates in drinking water. Prog. Water Technol. 8: 183–193. 80. Ross, J.D. and Desforges, J.F. (1959). Reduction of methemoglobin by erythrocytes from cord blood. Pediatrics 23: 718–721. 81. Colvin, C. (1999). Increased risk of methemoglobinemia as a result of bottle-feeding by HIV positive mothers in South Africa. Paper presented at IAH Congr. 1999, Bratislava, Slovakia. 82. McIntyre, J. (1997). Affordable option for the prevention of mother to child transmission of HIV 1 from research to clinical care. National AIDs Programme Perinatal HIV Res. Unit Conf., November 1997, Johannesburg, South Africa. 83. Carpenter, R.A. and Maragos, J.E. (1989). How to assess environmental impact on tropical islands and coastal areas. In: Training Manual for South Pacific Regional Environment Programme, South Pacific Regional Environment Programme (SPREP), Apia, pp. 119–122. 84. Genthe, B. (1998). Specialist Study on the Potential Health Impacts of the Proposed Wastewater Treatment Facility on the West Bank of East London. CSIR report ENV/S—198003. 85. Tredoux, G. (1974). ‘n Intensiewe opname van waterbronne in die omgewing van Otjiwarongo, Project Report No. 12 (unpublished), NIWR-CSIR, Windhoek, Namibia. 86. Tannenbaum, S.R. and Green L.C. (1985). Selected Abstracts on the Role of Dietary Nitrate and Nitrite in Human Carcinogenesis. International Cancer Research Data Bank Program, National Cancer Institute, Washington, DC. 87. Wolff, I.A. and Wasserman, A.E. (1972). Nitrates, nitrites and nitrosamines. Science 177: 15–19. 88. Hartman, P.E. (1983). Review: Putative mutagens and carcinogens in foods. Environ. Mutagenesis 5: 111–121. 89. Kleinjans, J.C.S. et al. (1991). Nitrate contamination of drinking water evaluation of genotoxic risk in human populations. Environ. Health Perspect. 94: 189–193. 90. Chandler, J. (1989). Nitrates in water. Water Well J. 3(5): 45–47. 91. Anonymous. (1974). Nitrate vergif water in noorde: Beeste en skape vrek. Die Suidwester, Windhoek, Namibia, June 26. 92. Marais, S. (1991). Written communication. Department of Water Affairs, Mmabatho, Bophuthatswana. 93. Johnson, G.T. and Goldfinger, S.E. (1982). The Harvard Medical School Health Letter Book. Warner Books, New York. 94. Talma, A.S. (1991). Written Communication, CSIR, Pretoria. 95. Life Systems, Inc. (1987). Drinking Water Criteria Document for Nitrate/Nitrite, TR-832-77. Cleveland, Ohio.

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96. Dahab, M.F. and Bogardi, I. (1990). Risk Management for Nitrate-Contaminated Groundwater Supplies. University of Nebraska, Lincoln, NE, pp. 9–16. 97. Kross, B.C., Ayebo, A.D., and Fourtes, L.J. (1992). Methemoglobinemia: nitrate toxicity in rural America. Am. Family Physician, July. 98. Secretaria de Salud. (1988). Reglamento de la Ley General de Salud en Materia de Control Sanitario de Actividades, Establesimientos, Productos y Servicios (General law for sanitary control of activities, products and services): Official News of the Mexican Government, January 8, Mexico, p. 27.

DOMESTIC WATER SUPPLY—PUBLIC–PRIVATE PARTNERSHIP MUHAMMAD SOHAIL Loughborough University Leicestershire, United Kingdom

INTRODUCTION ‘‘Food and water are basic rights. But we pay for food. Why should we not pay for water?’’ Ismail Serageldin ‘‘Water should not be privatised, commodified, traded or exported in bulk for commercial purposes.’’ Maude Barlow

such sources in many different ways, but these can then be divided into either piped or nonpiped options. The behavior of humans in terms of their consumption of water also has historical, geographical, and cultural dimensions. For example, in some parts of Bolivia people only consume 5 or 6 L per capita each day as compared with the 30 to 250 L consumed each day by a person in developed economies. Water has various uses, including agricultural, recreational, industrial, and domestic. In terms of its domestic use, water is used for sanitation facilities as well as for drinking. Indeed, a high level of consumption—up to 80%—is caused by piped sanitation wherever such facilities are available. Think about how much water a person flushes as compared with how much he or she drinks! With a limited amount of useable water, there is competition, sometimes tension, among various water users. Both market-oriented and hierarchy-based rules are used to distribute water among its various consumers. In the supply chain of water—its production, distribution, management, and consumption—private sector organisations are key players. Before moving on to the main discussion of this paper, it is useful to summarize some of the more important concepts.

CONCEPTS Public and Private Goods

The above two quotes typify the two extremes of the arguments surrounding public–private partnerships (PPPs). Although this ideological debate is intellectually exciting, the more challenging problem can be summarized as follows (1): • How do you provide access to safe water to around I billion people? • How do you provide sanitation services to around 2.6 billion people? Time is ticking by, and the global community must meet the above targets soon. Policymakers and water and sanitation practitioners alike should be prepared to include PPPs as a tool in achieving the above. Before going into detail, a brief summary of the context follows. Water is essential for life. This is not only in terms of its biological utility, for water also has social, economic, health, technical, financial, and political dimensions. Historically, the availability of a domestic water supply has been a significant factor in the development and sustenance of civilizations. Water is one of Earth’s most important natural resources. If naturally occurring freshwater is polluted as a result animal (including human), plant, or other activities, processes are required to convert the then raw water to a quality fit for a particular use, such as drinking. In most cases, because of high levels of pollution by humans, water must be treated before and after its use. Various water sources exist, for example, rainwater, groundwater, spring water, surface water, rivers, lakes, ponds, fog, and even glaciers. Water can be supplied from

Private goods are those for which consumption (or use) by one person prevents consumption (or use) by another. Public goods are those that can be used by one person without diminishing the opportunity for use by others. There is a seemingly unending debate over whether water should be treated as a public or private good, or both. Commodification Commodification is the process of converting a good or service formerly subject to non-market social rules into one that is subject to market rules. Treating water as an economic good implies that the resource will be allocated across competing uses in a way that maximizes its economic value across society. However, it also implies that safety nets will be needed for people who cannot afford to pay. Meanwhile, for some people, considering water to be a commodity at all is sacrilegious! Governance Governance refers to the relationship that can be manifested in various types of partnerships and networks. Water governance is the wider context within which water services procurement plays a key role and under which PPP is a niche tool. What is Partnering? At the very least, partnering should be viewed as the absence of adversarial behavior.

DOMESTIC WATER SUPPLY—PUBLIC–PRIVATE PARTNERSHIP

WHAT ARE PPPS? The involvement of the private sector in partnership with government has long been advocated as a means of improving the development of sustainable water and sanitation systems. The author uses PPPs in this document as a general term to cover a wide range of agreements or partnerships between private sector (nongovernment) concerns or organizations, public sector utilities, government departments, and consumer groups in relation to the delivery of water and sanitation services. The community has a direct role to play in such arrangements as a beneficiary and in expressing the price people would be willing to pay for an acceptable level of service. It also has an indirect role to play in shaping policy for the urban environment. In a small PPP, the community could take the role of the private partner. One of the difficulties in determining the scope of discussion concerning private sector involvement in water and sanitation is the sheer diversity of possible partnership arrangements and potential actors. The three main roles are those of the private sector, users (consumers), and the government (often referred to a the client or sponsor; it also may act as regulator). The possible arrangements include complex concession arrangements operated by multinational corporations lasting perhaps 30 years; shorter duration, simpler forms of management or service contract undertaken by mediumsized private enterprises; and service delivery by smallscale independent providers (local entrepreneurs). An outline of formal contracting arrangements is produced in Appendix A. WHY PPPS FOR DOMESTIC WATER SUPPLY? The role of the private sector in domestic water supply is not new. Many water supply-related activities in many municipalities in the world have started out as small private water firms or informal organizations of people. Considering the potential demand for water sanitation, it is obvious that already budget-constrained governments cannot improve services alone. Some estimates indicate huge capital outlays are needed to meet water and sanitation targets; Camdessus and Winpenny (2), for example, estimate US$90 billion per year is needed globally to meet such targets. On the other hand, there is an argument that unless national governments make environmental sustainability a priority, the shortage of water and sanitation services will remain as they are. Other people believe that improved water management and better use of current assets is what is crucial and will reduce the need for more capital infrastructure significantly. In any case, private finance is only one of the benefits of PPPs. Others include the managerial capacity of the private sector in managing assets for water and sanitation services. The private sector is also more likely to be innovative than the public sector, which leads to more effective and efficient services provision. Governments are turning to PPP arrangements for the provision of services for a variety of reasons; therefore, these may include the following:

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• Lack of capacity of government institutions to deliver a reasonable level of service or to improve service quality. • Financial weakness of public utilities. • The inability of public institutions to respond to an increasing growth in demand for services because of, for example, rapid urbanization. • Problems related to the large numbers of employees in public sector providers. • The requirement that international financing institutions (IFIs) place on certain indebted countries to reduce domestic spending as part of structural adjustment programmes. Such institutions promote the idea that the private sector is more efficient and effective and delivers a better quality service. HOW PUBLIC-PRIVATE PARTNERSHIPS GET STARTED There are various ways in which a partnership can be initiated. In some cases, the initiative comes from the likely partners and in some cases the demand comes from an third party. Public-private partnerships may not always seem to be a desirable solution at first. Most organizations prefer to stay on paths they know well, sharing goals and work practices with other groups that think and act like them—governments working with governments, businesses with businesses, and nonprofit groups with nonprofit groups. Governments and private firms have long worked together under simple arrangements, such as government purchase of products produced by the private sector. However, both parties often hesitate to enter into more complex relationships. Governments are frequently concerned that private businesses will take advantage of them, whereas businesses often consider government approaches to be burdensome and a waste of time. Therefore, it is useful to allow some time for trust to be established among the key partners. Three main conditions favor the formation of a partnership: urgency, the involvement of a champion, and some kind of catalyst. Urgency Generally, it takes a widely acknowledged urgency—for instance, the lack of particular services or the waste of resources—before key stakeholders start looking for partnerships and partners are open to cooperate to resolve the problem. Although it is hoped that progress can be made in the absence of a crisis, in practice, the inertia that keeps many people on familiar paths is usually broken only by a pressing need for change. Entrepreneurs/Champion Sometimes, even in the absence of a urgency, an individual, group, or organization may realize that separate, uncoordinated actions are creating redundancies, missed opportunities, and less-than-optimal use of scarce resources. In reality, even if the crisis is clear and the interest is there, partnership arrangements will not succeed

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without the drive and commitment of a few individuals. Such ‘‘champions’’ (leaders or pioneers) can be government officials, NGOs, business people, or citizens who—through their personal motivation—make partnerships happen. In other cases, champions are service providers who stand to profit from the partnership. Catalyst Frequently, there is a need for some kind of catalyst to bring the partners together. Such a catalyst could be the actions of an external actor, such as one or more international finance institutions or other bodies that are respected and trusted by all partners. WHAT ARE THE KEY FEATURES OF SUCCESSFUL PARTNERSHIPS?1 Characteristics of Successful Partnerships Compatible Goals. Government , businesses, and community leaders must understand and respect one another’s goals. For instance: • Government may initially have difficulty accepting the profit motive of private businesses. • Private companies may be tempted to walk away from the more bureaucratic decision-making processes used in the public sector. • Local communities may not have the patience needed to address issues affecting other areas of the city. To resolve these differences, all parties must focus on the broader, complementary goals that are to be achieved. It is important for them to realize that public and private goals do not necessarily need to be the same for partnerships to work—they must be merely compatible. Enabling Environments. An enabling regulatory, legal, and political environment is the cornerstone of sustainable private sector participation. Legal Framework. Early on, the public sector must establish an appropriate legal framework for contract procurement and private sector investment. It is important that mechanisms be put into place to minimize the likelihood or appearance of corruption in any procurement processes. Unpredictable and unfair procurement processes reduce both political acceptability and the interest of many private investors. Regulatory Framework. The government must also establish a clear regulatory framework, and it must implement appropriate tariff regimes and subsidy mechanisms. The creation of a regulatory framework alone, however, does not necessarily guarantee effective regulation. As all local governments are different, the public and private sectors will face a steep learning curve as they try to define and regulate their relationship with one another and their 1

The following section is based on Sohail and Olena (3).

roles in providing services. In particular, the public sector needs to define a clear allocation of responsibilities between the national and municipal governments and a clear statement of its role as a provider and a regulator. In general, private sector companies prefer that the contract serve as the major regulatory mechanism, and that governments have limited regulatory discretion once the contract is in place. Highly specific contract terms that establish duties, performance targets, rules for changing prices, and dispute resolution procedures allow the private sector to better predict the profitability of the venture and decide whether and what to bid for the contract. Given these preferences, governments will have to make important decisions about the degree of regulatory discretion they are willing to give up, particularly for long-term contracts. Political Environment. In addition to the regulatory climate, a bad political climate caused by the pressure of election cycles, the potential instability of new democracies, the personal agendas of government officials, and the special status of some services (particularly in terms of access to water, for example) can create barriers to starting or maintaining public–private collaborations. Governments must provide assurances whenever possible to private sector partners that such political factors will not disrupt the contractual partnership. Acceptance. The government and business leaders cannot build partnerships alone; political and social acceptance of private sector involvement is essential. The population must see private sector participation as beneficial if the partnership is to last over time. Public support of private involvement over the long term will depend on primarily the delivery of promised services and benefits at reasonable costs. Therefore, it is of the utmost importance that mechanisms be developed to ensure that the organization providing the service, whether it is a public or a private sector organization, be accountable to its customers. Public support will also depend on the ability of the partnership to meet the needs of all stakeholders. For example, public sector workers can be a source of tremendous opposition to increased private involvement in the provision of services. Contracts should ensure the employment or placement of public employees and local residents to the greatest degree possible. Credibility and Transparency. Effective cooperation between local government, businesses, and the community is always difficult to achieve because of the wide range of participants involved, the low level of trust that often exists between potential partners, and the lack of predictability in the process. The credibility of champions and other leaders involved, as well as transparency in the process, are critical determinants of long-term success. Experience suggests that genuine partnerships must include the principles of equity, transparency of operations, and mutual benefit. Trust and confidence in any project is necessary for successful partnerships.

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Factors Contributing to the Durable Partnerships Governments clearly want to establish PPPs that are sustainable over time. Essential ingredients of durable PPPs include those listed below. Commitment of Resources. All partners to the arrangement should be obliged to commit resources (financial, human, capital) to increase their interest in seeing the partnership succeed, which implies shared risks and rewards. Capacity Development. Projects requiring substantial institutional change or large capital investments will require capacity development within all groups of stakeholders. For example, development of: • Consumers, in terms of their knowledge of the service they are to receive and the costs associated with its provision • Service providers, particularly local organizations, in terms of entrepreneurial skills • Governments, in terms of their capacity to adopt the frameworks necessary for, and oversee the provision of, the service Roles and Responsibilities. The delineation of appropriate roles and responsibilities is another element necessary in the development of effective, durable partnerships. It is essential that partnerships be organized in a concerted fashion to make the most effective use of the resources committed by both parties. Individual responsibilities should be clearly outlined from the beginning so that there is no ambiguity in the tasks that each party is expected to perform. Furthermore, these responsibilities need to be defined realistically with a clear understanding of the strengths and weaknesses of each partner. Flexibility. All partnerships are context-based and different locally. Partnerships should draw on other experiences but at the same time should be opportunistic about exploiting the comparative advantage of local resources. Over the long term, changes in investment plans, technology choices, and priority actions will be necessary in response to unforeseen circumstances. Including clear procedures for making such changes over the life of the project will reduce the chance that they will have a negative impact on the partnership. Time. Partnerships take time. The process of understanding the problems to be addressed and the impacts on potential partners, as well as those partners’ needs and aspirations, all takes time. Progress can certainly be made along the way, but the process of achieving and maintaining acceptance among users, providers, and regulators is a continuing one; a cooperative dialogue to address shared needs must be maintained throughout the project. Patience. Projects requiring substantial institutional change or large capital investments require a lot of

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patience. Careful attention must be paid to the balance between responding rapidly to the most pressing crises and developing integrated solutions that will last. Political cycles and the desire for immediate improvement in a crisis situation often lead to the development of time frames that are too short. Such short-term agendas and limited horizons lead to unrealistic expectations and unsustainable solutions. It is not realistic to expect that private sector involvement will overcome public institutional and operational inefficiencies quickly, nor that it will compensate immediately for a history of insufficient public sector resources and funding. Social Responsibility. Public services provide public goods—in other words, goods that should be available to everyone. Improving provision of such services is about making people’s lives better, especially those of the urban poor. Governments should always make sure that the changes they make promote increased access to, and better quality of, services. An emphasis on social responsibility will also increase political gain, as better services will lead to greater political acceptance by the general population. What are the Major Obstacles to Forming a Successful Partnership? A range of possible obstacles or deficiencies in the capacity of both public and private actors could hinder the formation of a successful partnership. Major obstacles in this respect include: • Reciprocal mistrust and lack of understanding of one another’s interests and needs across the public and private sectors • Absence of locally available information on, and experience with, arranging sustainable partnerships • Underlying legal, political, and institutional obstacles to forming effective public–private relationships These obstacles often lead to lengthy negotiations, increased transaction costs, and make smaller projects much less attractive to potential investors. To minimize the harm from such obstacles, PPP arrangements should provide certain safeguards for the public and private sectors and for the community. The public sector usually expects the private sector to contribute in one or all of the following ways: • • • •

To provide agreed services To make agreed investments To meet agreed standards/targets To not exploit any monopoly might exist

situation

that

The private sector expects the public sector to contribute in one or all of the following ways: • To create an enabling environment suitable for the PPP • To pay agreed fees promptly and in full • To implement tariff increases as agreed

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• To prevent unexpected competition from others during operation (exclusivity) The community expects the PPP to: • Provide appropriate levels of services • Be affordable to the community, either through direct charges or indirectly through general taxation THE CURRENT SITUATION Quantitative information in this section is based mainly on the World Bank Private Participation in Infrastructure (PPI) database (4). • At least 203 water and sanitation projects are taking place in 43 developing countries, with a corresponding commitment of some US $40 billion. • A few large projects in, for example, Argentina, Chile, and Manila, Philippines can explain peaks in water and sanitation investment during 1991–2000. • Latin America and the Caribbean are the most active regions in terms of global investment (52%), followed by East Asia and the Pacific (38%) and only then Europe and Central Asia (8%). • Concession contracts provide the largest proportion of such investment (69%) through the largest number of projects (90). • Private investment reached a peak in 1999. • The top five countries by cumulative investment in water and sewerage projects with private participation were Argentina, the Philippines, Malaysia, Chile, and Brazil during 1991–2000. The above figures relate only to formal PPPs and do not take into account many informal private sector and community-based operations; such operations could be serving many users well. CHALLENGES2 Political Acceptance Infrastructure services such as water and sanitation are intrinsically political and that fact, along with the politicians’ individual agendas, should be acknowledged in policy discussions. For example, PPPs are still seen to be a political risk by many politicians, whose principle objectives include acquiring or retaining power. Quick fixes may be favored to attain popularity rather than making hard decisions for the long-term sustainability of water and sanitation services. Nonetheless, socially sensitive PPPs are more likely to be politically acceptable.

services provision is expected to also fulfill the social role of the public sector. However, where this is the case, it must be negotiated as part of the PPP right from the beginning of the procurement process. Knowledge and Understanding of Advisers. Terms of Reference mostly appear to be drafted by advisers, who tend to be distant from the sponsor/client’s organization. Their expertise will be determined by the requirements specified by the sponsor/client and normally focus on technology and finance. A lack of understanding on the part of those responsible for developing and negotiating the contracts can lead to the omission of important existing informal arrangements among public utility staff, water vendors, and low-income customers. Hence the inclusion of expert staff members that have PPP experience and a workable knowledge of how to improve water and sanitation facilities for low-income (poor) areas and communities is important if PPPs are to become socially sensitive. The time scale for the sponsor/client’s advisers to prepare the necessary bidding information is relatively short, and this inevitably results in workloads being prioritized. Consequently, the complex and little understood issues involved in service provision for the poor are unlikely to receive serious consideration at this stage. The ‘‘time factor’’ may therefore have contributed to the historic underrepresentation of services to low-income communities in many PPP contract documents. Governance and Relationships with Consumers. Public sector utilities and private companies are experienced in service provision to regular settlements and housing developments occupied by middle and high-income consumers with individual service connections. However, public sector utilities rarely have much experience in dealing with service delivery to the poor, whereas operators from the formal private sector appear to have even less. A key feature of extending coverage to the urban poor therefore involves working out new relationships between a diverse range of actors who have little if any experience of one another. Direct Links: Partnership does not have to Involve Formal Private Sector. In cases where small local entrepreneurs or civil society groups take on the role of private sector operator, they generally act as an intermediary between the public sector sponsor/client and the consumer. This may, for instance, involve acting as a retailer of water services. In such circumstances, there tends to be more negotiation to address problems and issues at the local level, with local entrepreneurs communicating directly with consumers or their representative groups. Although there is less recourse to legally binding agreements, the key issue remains the same, namely, whether the mechanism works for poor consumers.

Social Acceptance in Developing Countries Policy Issue: The Poor are Seldom Mentioned. In many developing countries, private sector involvement in 2

This section is based on Sohail (5).

User Perceptions In recent years, consumers have been asked to voice their opinions on the overall process of PPP development involving the formal private sector. Concerns that have

DOMESTIC WATER SUPPLY—PUBLIC–PRIVATE PARTNERSHIP

emerged include lack of consultation; concern over lack of public control or safeguards; fear of corruption; high tariffs; unemployment; and the assumption that there will be increased burdens on the ‘‘common people,’’ who in turn have no clear idea of the benefits. Involvement in the processes leading up to change is critical. Although this does not guarantee success, case studies have found that projects in which stakeholder participation was absent or minimal were the least successful in terms of feedback from people on low incomes. In such cases, consumers tended to raise objections to the involvement of private companies in their water supply, especially foreign companies. Commonly, such involvement is precipitated by increases in tariffs, which are perceived to be making profits for private companies at the consumer’s expense. Financial Issues and Tariffs Payment Problems. Some of the problems faced by households include high arrears, high repayment levels, disconnection of the water supply, and inability to pay reconnection charges. The introduction of higher charges through both metering and increased standing charges creates additional pressure on the household budget. Low-income people are clearly under most pressure to economize, and this results in reduced water consumption and less cash being available for other needs. Technology Scope for Innovation. Simple yet innovative technology changes may form part of the implementation of PPP arrangements. Such changes are the result of various factors, including: • Analysis of settlements and their needs • Taking community needs and preferences into account • Knowledge of alternative options, relevant technologies, and proven models—for example, simplified, lower cost designs • Willingness to experiment to find innovative alternatives to standard technological approaches Levels of Service and Service Differentiation. There are several cases where PPP has led to the introduction of more appropriate and flexible levels of service. This has involved lowering the conventionally accepted, high levels of service that predominate in wealthier areas. Information The Need to Communicate Effectively. The introduction of PPPs normally changes existing roles and responsibilities, and these have to be communicated to consumers. If one of the partners (usually the operator) has no working knowledge of the area, then it is crucial for that partner to find out about consumer needs. Regulation, Monitoring and Complaints Formal Regulatory Systems. These should be developed during the preoperational phases of PPP contracts, and

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they are notably absent during the preparation for many PPPs. No provisions are made to check the operation of a monopoly supplier or to impose penalties for substandard performance. In addition, there may be no defined procedures for routine external monitoring of performance. In particular, lack of capacity for monitoring and reporting is a problem for the relatively small municipalities involved in many PPPs. Effectiveness of Regulation. International PPP experience has shown there to be situations where regulation does not necessarily safeguard the different interests of the various parties to the arrangement. For instance, poor consumers are often dissatisfied with services and tariffs in formally regulated environments; billing and bill collection by the public partner may be inefficient; and the private operator may perceive lack of control over customer management to be a significant risk. CREATING TRUST Communication and information alone are not enough. In many PPPs, consumers are unsure of what to expect from privatization because they have not been involved in the development process. If user groups and other stakeholders had been integrated into the process from the start, then there might not be such widespread opposition. In conclusion, then: • There is a pressing need to develop a base of information about low-income groups; this can then be used directly in the development of PPP arrangements. Such action requires a clearer understanding of information needs on the part of both the designers of PPPs and the local institutions and organizations. • There is also a clear need to provide information for consumers, particularly about proposed roles and responsibilities. Lack of understanding and consensus leads to operational problems and is ultimately disempowering. • Communication is a vital component of PPPs, and investment in this will pay dividends in operational terms. Lack of information could result in low-income communities refusing to accept or comply with the partnership, and this in turn could lead to the risk of nonpayment. BIBLIOGRAPHY 1. World Health Organization and UNICEF (2004). Meeting the MDG Drinking Water and Sanitation Target: A Mid-Term Assessment of Progress. WHO, Geneva. 2. Camdessus, M. and Winpenny, J. (2003). Report of the World Panel on Financing Water Infrastructure: Financing Water For All. World Water Council, 3rd World Water Forum, Global Water Partnership. Ahmed, N. and Sohail, M. (2003). Alternate water supply arrangements in peri-urban localities: Awami (People’s) tanks in Organgi Township, Karachi. Environ. Urbanisation 15(2): 33–42. 3. Sohail and Olena (2003).

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DOMESTIC WATER SUPPLY—PUBLIC–PRIVATE PARTNERSHIP

Schusterman, R. and Hardoy, A. (1997). Reconstructing social capital in a poor urban settlement: The integral improvement programme in Barrio San Jorge. Environ. Urbaniz. 9(1). Serageldin, I., Barrett, R., and Martin, J. (Eds.). (1994). The business of sustainable cities: Public-private partnerships for creative technical and institutional solutions: An associated event of the second annual World Bank Conference on Environmentally Sustainable Development. World Bank Conference on Environmentally Sustainable Development. World Bank, Washington, DC. Sohail, M. and Baldwin, A.N. (2001). Partnering with the community—an option for infrastructure procurement. Proc. Institution of Civil Engineers Municipal Engineer. 1454: 293–297. Sohail, M. and Baldwin, A.N. (2003). Urban infrastructure procurement in low-income countries. ICE Proc. Institute of Civil Engineers, Municipal Engineer, Engineering Sustainability 156(ES2): 0–1. Sohail, M. and Baldwin, A.N. (2004). Performance indicators for ‘micro-projects’ in developing countries. Constr. Manag. Econom. 22: 11–23. Sohail, M. and Cavill, S. (2000). Public Private Partnerships and Poor, Part B—Interim Findings and Case Studies. WEDC, Loughborough. Sohail, M. and Cotton, A.P. (2000). Public Private Partnerships and Poor Interim Findings, Part A—Summary and Lessons Learned. WEDC, Loughborough. Sohail, M. and Aslyukivska, M. (Eds.). (2004). Tools for Propoor municipal PPP, United Nations Development programme Margraf Publishers, Frankfurt, Germany. Solo, T.M. (1999). Small-scale entrepreneurs in the urban water and sanitation market. Environ. Urbaniz. 11(1): 117– 131. Solo, T.M. (2003). Independent Water Entrepreneurs in Latin America: The Other Private Sector in Water Services. World Bank, Washington, DC. United Nations (1992). Earth Summit Agenda 21, Chapter 33 (Finance). United Nations, Washington, DC. Waddell, S. (2002). Core competences: A key force in businessgovernment-civil society collaborations. J. Corporate Citizenship 7: 43–56. Webb, M. and Ehrhardt, D. (1998). Improving water services through competition. World Bank Public Policy for the Private Sector Note 164. World Bank (1997). Toolkits for Private Participation in Water and Sanitation. The World Bank, Washington, DC. World Bank (1997). Toolkits for Private Sector Privatization in Water and Sanitation. World Bank, Washington, DC.

APPENDIX A: OVERVIEW OF MODELS OF PRIVATE SECTOR PARTICIPATION IN WATER AND SANITATION PROVISION Full Privatization (Divestiture) Private company not only takes full responsibility for operation, maintenance, and investment, but ownership of infrastructure is transferred from the public to the private sector at an agreed fee. The government is responsible for regulation. Partial Private-Sector Responsibilitys Responsibility for service provision is shared between the private and public sectors, with differing levels of responsibility being delegated to the private partner depending on the contract type. In all of the following models, ultimate ownership of assets remains with the public sector. Service Contract Service contracts are usually short-term agreements whereby specific operations and maintenance activities are contracted to the private sector. The public sector retains overall responsibility for the administration of the service. Management Contract A management contract entails private sector responsibility for utility operation and maintenance but without the obligation of investment or commitment of private investment capital. Lease Contract (Affermage) Under lease contracts, the private firm operates and maintains the utility at its commercial risk, deriving revenue directly from tariffs, but it does not invest in new infrastructure. Concession Contract Under concession contracts, the private company manages the infrastructure facility and operates it at its commercial risk and accepts investment obligations. The role of the government in concession contracts is predominantly regulatory. Build-Own-[Operate]-[Train]-[Transfer]-Type Contracts (BOO/BOT/BOOT/BOTT)

World Bank (2003). Private Participation in Infrastructure data base. Available at http://ppi.worldbank.org/

These are similar to concession contracts, but they are usually used for greenfield projects as the private contractor is also responsible for constructing the infrastructure. At the end of the contract, the assets may either remain with the private company or be transferred back to the government.

World Bank (2004). Global Development Finance 2004. World Bank, Washington, DC.

Cooperative Model

Yepes, G. (1999). Do Cross-subsidies help the poor to benefit from water and wastewater subsidies? Lessons from Guayaquil, Ecuador. UNDP-World Bank Water and Sanitation Program.

The cooperative model is a type of government-owned public-limited company (plc) subject to the rules and regulations of other plcs and of which most shares are publicly owned (either by government or citizens/users).

World Bank (1998). Global Development Finance 1998. World Bank, Washington, DC.

METHODS OF REDUCING RADON IN DRINKING WATER

Informal Sector Provision Provision of water and sanitation services to the poor by ‘‘informal’’ and/or small-scale operators is common in most low- and middle-income countries, especially where the poor lack access to formal service provision. In an increasing number of cases, governments are supporting small-scale private initiatives to increase services provision to the poor. Sources: Blokland et al. (6), Calaguas (7), Kempe and Schreiber (8), Johnstone and Wood (9), Lewis and Miller (10), Nickson (11), and Ramaema (12).

METHODS OF REDUCING RADON IN DRINKING WATER PAUL D. ROBILLARD World Water Watch Cambridge, Massachusetts

WILLIAM E. SHARPE BRYAN R. SWISTOCK Pennsylvania State University University Park, Pennsylvania

Since the late 1980s, radon has become a highly publicized health threat. This naturally occurring radioactive gas is seeping out of the earth’s crust and into the basements of thousands of homes across the nation. Until recently, radon concerns have focused primarily on airborne radon; radon in drinking water was not considered a problem. Now, water tests reveal its presence, and many homeowners are asking water treatment dealers, ‘‘How can I reduce radon in my water supply?’’ WATERBORNE RADON Waterborne radon usually originates in deep wells that tap radon-contaminated groundwater. Radon increases household air levels during showering, laundering, and dishwashing. The EPA estimates that 2–5% of airborne radon comes from household water. They further estimate that even these small percentages increase the incidence of cancer. If radon is discovered in water, it is likely that radon is entering the house through the basement as well. Currently, the EPA has not set official standards for either airborne or waterborne radon. EPA suggests that an airborne level of 4 pCi/L is a point at which remedial action should be taken. Recognize that for every 10,000 pCi/L in water, about 1 pCi/L will be released in the air. The EPA’s proposed limit for radon in water is 300 pCi/L.

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from the sample. Direct water sampling is by far the most accurate testing method. Treatment Radon water treatment should remove radon before it becomes airborne. Methods of Home Aeration Home Aeration Units. Home aeration exposes the water to enough air so that radon can escape to the air before the water reaches the taps. Using new technological advancements in home aeration, these units can have radon removal efficiencies up to 99.9%. They are also ideal for high waterborne radon levels.

Spray Aeration Unit. A spray aeration unit, as shown in Fig. 1, sprays radon contaminated water into the tank using a spray nozzle. The increased surface area of the sprayed water droplets causes the radon to come out of solution, and the air blower carries the radon contaminated air to a vent outside the home. About 50% of the radon will be removed in the initial spraying. The water must be sprayed several times to increase removal efficiencies. To keep a supply of treated water, at least a 100-gallon holding tank must be used. Packed Column Aeration Unit. In a packed column system, water moves through a thin film of inert packing material in a column. The air blower forces radon contaminated air back through the column to an outdoor vent. If the column is high enough, removal efficiencies can be between 90 to 95%. For a 6-foot column (shown in Fig. 2), the removal efficiency is around 95%. Packed columns become impractical if the radon level exceeds 20,000 pCi/L. Shallow Aeration Unit. A final aeration system uses a shallow tray to contact air and water. Water is sprayed into the tray and then flows over the tray as air is sprayed up through tiny holes in the tray bottom (see Fig. 3). The system removes more than 99.9% of the radon and vents

Radon contaminated air to vent

Air blower

Detection and Testing Radon and its daughters are radioactive—continually decaying and emitting radioactive particles called alpha and beta rays. Therefore, testing for radon in water requires special sampling and laboratory analytical techniques that measure its presence before it escapes

Float switch Contaminated inflow from well

Outflow to transfer pump

Figure 1. Radon removal using a home spray aeration system.

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METHODS OF REDUCING RADON IN DRINKING WATER

Radon water with sediment and other contaminants

Purified water

Radon contaminated air to vent Contaminated inflow from well

Granular activated carbon Sediment filter Air blower

Outflow to transfer pump

Figure 2. Radon removal using a packed column.

Radon contaminated air to vent

Air blower

Contaminated inflow from well

Outflow to transfer pump

Figure 3. Radon removal by horizontally extended shallow aeration.

it outside the home. The treated water collects in the tank bottom and is pumped to the water pressure tank. Advantages of this type of aeration include: • Low pressure air blower • No fouling problems in tray holes • Small unit size However, this unit uses 100 cubic feet per minute of air compared to the others, which can depressurize the basement. Point-of-Entry Treatment Another method for removing radon from water is a granular activated carbon (GAC) unit. Figure 4 shows a typical GAC unit. For radon removal, GACs are constructed of a fiberglass tank containing granular activated carbon—a fine material that traps and holds the radon. Because of the carbon’s fine particle size, it easily clogs with sediments or other contaminants in the water. Important points to consider with GAC units are: • Some GAC units come with a special backwashing feature for removing sediment, which eventually reduces the effectiveness of the carbon to remove radon.

Figure 4. Treatment by activated carbon.

• Elimination of the sediment source or a sediment filter placed ahead of the GAC tank is the best protection against clogging. • The maximum radon level at which a GAC unit operates effectively is uncertain. Some estimates show that it should not be used if waterborne radon levels exceed 30,000 pCi/L. Other experts say 5,000 pCi/L. • It is important that the filter size matches the water use and conditions. According to the EPA, a 3-cubicfoot unit can handle as much as 250 gallons of water per day and effectively reduce radon levels. Typical water use in a home ranges from 50 to 100 gallons per person per day. • GAC filter will remove radon indefinitely providing that sediments or organic pollutants have not clogged the filter. • A major drawback to using a GAC filter is that if radon is present, the filter becomes radioactive as it picks up the gas. Lead-210 (a radon daughter) builds up on the carbon filter and then gives off its harmful radioactive rays, as it continues to decay. • A GAC filter may produce a radiation problem when the device is used to remove other contaminants. For example, a GAC unit is installed to remove a pesticide without testing the water for radon. The GAC unit sits under the sink harmlessly removing the problem contaminant. Right? Wrong. Unfortunately, what the homeowner doesn’t know is that the water supply has very high radon levels. So, while the GAC traps the pesticide, it also traps radon, thus producing a radioactive filter and a radiation hazard. • Proper maintenance and handling of the GAC unit can minimize exposure risks. Redevelopment of the well intake or a sediment filter is vital to protecting the fine carbon from fouling and clogging; protected filters won’t need to be changed as often. The water should also be periodically retested to insure that radon is still being removed.

WATER REUSE

WATER REUSE PETER S. CARTWRIGHT, P.E.

53

Undoubtedly, a paradigm shift is taking place with regard to water conservation and reuse. The barriers to reuse activity are

Minneapolis, Minnesota

Water reuse is the general term applied to the act of recovering water from a process and reusing it in the same process, or another one, before discharging it. Some experts attach labels to this activity based on the specific use of the water: Recycle involves redirecting the recovered water back to the same process. Reuse means reusing the recovered water in a different application. Recovery generally refers to the technologies used to accomplish this. Black water is defined as the effluent from toilets and garbage disposers in residences. Gray water in the residential environment typically refers to the effluent from bathing, showering, laundry, and dishwashing and other effluents from normal household activities. Gray water is distinguished from black water in that it contains much less organic loading and is expected to contain much less fecal coliform bacteria and other pathogenic organisms. We all are familiar with the admonitions addressing the finite quality of water on this planet: the fact that only 1% is considered ‘‘fresh’’ (nonseawater), but is used for virtually all human activities and, as a result, is rapidly deteriorating in quality. The mindset is there, as are the technologies. The barrier is the commitment to make the economic investment. From a technical standpoint, there are virtually no barriers to the quality improvement of either black water or gray water, even to the point of drinking water quality! Today it is possible to install a ‘‘black box’’ on the sewer line from residences and drink the treated water coming from it. Presently, the deterrents to this are economics and the ‘‘yuck’’ factor. Drinking water regulations continue to place new standards for water quality, and industrial and commercial activities are requiring higher quality water in many applications. Due to population growth, water quantity requirements are also steadily increasing. For example, today it is estimated that one-fifth of the world’s population does not have access to safe drinking water; by 2025, the global population is expected to be 8.3 billion people (up from about 6.4 billion in 2004), and the usage of industrial water is expected to double by then. Interestingly, even in cities experiencing severe drinking water shortages, there always seems to be enough water to flush toilets. Industries, in general, are very poor stewards of water conservation. Most water brought into the plant is used only once, even though that use may have had very little impact on water quality. In most of the industrialized world, we have labored under the misconception that our water supply was inexhaustible as well as inexpensive. This cost aspect has been exacerbated by the fact that, in many areas, the price of water has been partially subsidized by local government.

1. 2. 3. 4.

the ‘‘yuck’’ factor lack of practical reuse technologies economic factors commitment to reuse

1. The average consumer recoils at the thought of drinking (or even reusing) ‘‘sewer water,’’ not considering the fact that the drinking water for one community is very likely the wastewater discharged from another community on the same river or lake. It is estimated that in major U.S. rivers, water is reused as many as 20 times by the time that river empties into the sea. Most ‘‘groundwater recharge’’ is treated municipal wastewater injected into groundwater supplies to provide sufficient storage capacity or to serve as a barrier to salt water intrusion from the oceans. 2. The key to water reuse is to have an arsenal of technologies available to remove hazardous or undesirable contaminants efficiently from the water supply. There is no single technology that efficiently removes all classes of contaminants; however, for the past century or so, there have been significant developments in treatment technologies that effectively reduce the concentration of virtually any contaminant to acceptable levels for any water use. There is little argument that reducing the huge variety of contaminants that may be encountered in typical wastewater, resulting from the combination of sewage, industrial wastewaters, and perhaps even effluent from surface water runoff, requires a stunning array of advanced technologies. The linchpins are the membrane separation technologies of reverse osmosis, nanofiltration, ultrafiltration, and microfiltration. Whereas reverse osmosis has been around for more than 50 years, the others, most notably nanofiltration, are relatively new developments. By separating the treatment process into its key components, pretreatment, primary treatment, and posttreatment, it is now possible to create optimum technology trains that can purify the stream from virtually any source and condition it for virtually any reuse. 3. In addition to treatment technology costs, it is also necessary to factor in the fact that raw water costs are steadily increasing and will continue to do so, reflecting the diminishing supply and costs of meeting new regulations. 4. It is estimated that as of the year 2000, there were more than 10,000 water reuse systems installed in the United States, almost all of this recovered water was used to irrigate agricultural fields or residential and commercial landscaping. In California alone, it is estimated that 120 billion gallons per year are reused.

54

ROOF DRAINAGE HYDRAULICS

choice of removal technologies consists of membrane technologies of reverse osmosis, nanofiltration, and electrodialysis; adsorptive resin technology known as ion exchange and the hybrid of electrodialysis and ion exchange known as electrodeionization. Distillation is the oldest technology, mirroring the natural water cycle, but, because of its high energy cost, is now used only in specialized applications.

Water reuse, billion gallons/day

14 12 10 8 6 4 2 0 01 02 03 04 05 06 07 08 09 10 11 12 13 14 15 Year: 2001–2015 Figure 1. Projected water reuse in 21st century.

The rate of water reuse in the United States is expected to increase markedly over the next decade, as indicated by Fig. 1. SPECIFIC TECHNOLOGIES To understand the technologies most applicable to water reuse, it is necessary to understand the contaminants prevalent in water supplies. Contaminants can be categorized by their physical and chemical properties. Table 1 provides this classification. Suspended Solids Removing suspended solids from water supplies is probably the oldest water treatment procedure. Throughout history, humankind has used everything from containers of sand to cloth to charcoal to ‘‘clarify’’ water to make it look and taste better. The chronology of water treatment technology development also underscores the improvements to suspended solids removal processes; the latest is microfiltration.

Microorganisms Microorganism contaminants in most water supplies are from one or more of the following categories: bacteria viruses protozoan cysts fungi algae There are a number of disinfection technologies that inactivate or remove microorganisms. These include chemical (chlorine compounds and ozone). Certain heavy metals in solution inactivate microorganisms. The most common of these are silver for bacteria reduction and copper for algae inactivation. The most prevalent nonchemical technology is ultraviolet irradiation, although heat is still used occasionally in specialized applications, such as preparation of ‘‘water for injection’’ in the pharmaceutical industry. Many disinfectants are effective on only certain classes or types of microorganisms or under very specific conditions. The most troublesome class of microorganisms is bacteria. Because bacteria are viable and grow under virtually any condition, they are impossible to eliminate completely. In most applications, the goal is to minimize bacterial growth so as not to interfere with the water use.

Dissolved Organics Removing dissolved organic contaminants requires the greatest variety of technologies, reflecting the diversity of dissolved organic chemicals. For most of these contaminants, there are several choices of technologies to effect removal. Some may involve adding chemicals, such as alum, powdered activated carbon, or an acid; other technologies may include physical separation as with a coalescer or ultrafiltration technology.

CONCLUSION Water reuse is not an abstract concept; it is both a reality and a necessity. For the reasons cited, the requirements and opportunities for water reuse will continue to grow at an increasingly rapid rate.

ROOF DRAINAGE HYDRAULICS

Dissolved Ionics (salts)

SCOTT ARTHUR GRANT WRIGHT

These are contaminants that have ionic charges and are almost all inorganic chemicals. The somewhat limited

Heriot-Watt University Edinburgh, Scotland, United Kingdom

Table 1. Chemical Properties of Contaminants Class Suspended solids Dissolved organics Dissolved ionics (salts) Microorganisms Gases

Typical Example Dirt, clay, colloidal materials Trihalomethanes, synthetic organic chemicals, humic acids, fulvic acids Heavy metals, silica, arsenic, nitrate Bacteria, viruses, protozoan cysts, fungi, algae Hydrogen sulfide, methane, radon

INTRODUCTION Over the past decade, urban drainage systems have moved toward what are now commonly known as ‘‘sustainable urban drainage systems’’ (SUDS) or ‘‘best management practice’’ (BMPs). Fundamental to the implementation of these systems is addressing both runoff quantity and

ROOF DRAINAGE HYDRAULICS

quality at a local level in a manner which may also have the potential to offer benefits to stakeholders. This has led to a change in the way new developments now look and interact within catchments. However, despite the availability of such tools to reduce, attenuate, and treat urban runoff, substantial areas of the urban environment are still 100% impermeable and drain rapidly, namely, roof surfaces. Normally, roof drainage systems do not always receive the attention they deserve in design, construction, and maintenance. Although the cost of a system is usually only a small proportion of a building’s total cost, it can be far outweighed by the costs of the damage and disruption resulting from a failure of the system to provide the degree of protection required. There are basically two different types of roof drainage system; conventional and siphonic (refer to Fig. 1). Conventional systems operate at atmospheric pressure, and the driving head is thus limited to the gutter flow depths. Consequently, conventional roof drainage systems normally require a considerable number of relatively large diameter vertical downpipes, all of which have to connect to some form of underground collection network before discharging to a surface drain. In contrast, siphonic roof drainage systems are designed to run full bore, resulting in subatmospheric pressures, higher driving heads, and higher flow velocities. Turbulent gutter conditions mean that there will always be a small percentage of entrained air within the system (typically 5%). Hence, siphonic systems normally require far fewer downpipes, and the depressurized conditions also mean that much of the collection pipework can be routed at high level, thus reducing the extent of any underground pipework. Both types of drainage system comprise three basic interacting components: • the roof surface • the rainwater collection gutters (including outlets) • the system pipework Each of these components can alter the runoff hydrograph substantially as it is routed through the system. This text focuses on the role and performance of each of these components. As the principles of siphonic drainage are generally less well understood and certainly less well documented, particular emphasis placed on the performance of siphonic roof drainage systems.

(a) Free surface outlet conditions

55

ROOF SURFACE The design of the roof surface is usually within the remit of the architect rather than the drainage designer. Notionally, there are three types of roof surfaces: Flat Roofs Flat roofs are normally associated with domestic properties in climates with low rainfall and with industrial buildings in developed countries. Such roofs are seldom truly ‘‘flat’’ but simply fall below the minimum gradient associated with sloped roofs in the jurisdiction under consideration; for example, in the United Kingdom, a flat roof is one whose gradient is less than 10◦ (1). Minimum gradients are usually specified to avoid any unwanted ponding (BS EN 6229:2003 specifies a 1 in 80 minimum gradient) and to help prevent the development of any adverse gradient due to differential settlement (2). Although flat roofs can be problematic if not maintained properly, they are often preferred; they reduce the amount of dead space within the building and they attenuate flows more than sloped surfaces. Sloped Roofs Most residential and many commercial properties have sloped roofs. Such roofs are generally favored because their ability to drain naturally means that there is less risk of leakage. In temperate climates, their specification also means that snow loading is less of an issue. Once a rainfall is underway, the rate at which the runoff flows across a roof is a function of roof slope and roughness. Where rainfall data are available, runoff rates from roof surfaces may be readily assessed using kinematic wave theory (3). Green/Brown Roofs (Sloped or Flat) The oldest type of permanent roof is a green roof. These involve planting roof areas to attenuate and/or dissipate rainfall and can take the form of a rooftop garden with trees and shrubs (termed intensive) or a lightweight carpet of growth media and flora (termed extensive). The latter technology is already employed widely (e.g., the Rolls Royce plant at Goodwood; purportedly Europe’s largest green roof). Many of these applications tend to focus on the aesthetic benefits such systems offer to high profile developments and are often installed to ‘‘green’’

(b) Submerged outlet conditions Driving heads

Annular flow in vertical downpipes

Extensive underground network

Full-bore flow in all pipework

Minimal underground network

Figure 1. Schematics of a typical conventional (a) and (b) siphonic roof drainage system (at normal design condition).

56

ROOF DRAINAGE HYDRAULICS

a development and thus help secure planning consent in sensitive areas (4). However, as well as being aesthetically pleasing and hydraulically beneficial, green roofs may also offer thermal insulation (5), reduce the heat island effect, the phenomenon whereby absorption of solar radiation by urban surfaces causes a marked increase in ambient air temperature (6), provide acoustic damping, and extend the service life of the roof membrane (7–10). Green roof systems are used extensively in Germany and to a lesser extent in North America, but again their specification is primarily due to a desire for a reduced aesthetic impact associated with a particular development. Germany probably has the most experience to date, a direct result of their use in the 1800s as a low fire risk alternative to tarred roofs in deprived urban areas (11). Currently, German research is focused predominantly on planting issues, and there is only a limited understanding of how the systems may be used to mitigate the impact of urban runoff. One research project, which ran from 1987–1989 in Neubrandenburg (8), found that an installed green roof with 70 mm of substrate could reduce annual runoff from a roof by 60–80%. Work in Vancouver (Canada), based on an uncalibrated computer model, suggests that for catchments where the roof area comprises 70% of the total surface, installing an extensive system could reduce total runoff to approximately 60% over 12 months (12). The same model was also used to assess specific synthetic rainfall; these results indicated that the catchment experienced increased runoff during longer rainfalls. Neither of these studies detail how green roofs could be expected to perform during a particular rainfall or where efficiencies may be gained in the design of collection pipework. Limited testing in the United States (13), where green roofs are often irrigated, has indicated that runoff can be reduced by 65% during a single rain. The most authoritative design guidance for green roofs in the United States is produced by The New Jersey Department of Environmental Protection (14). This is focused on lightweight structures and gives guidance on how to ensure ‘‘rapid draining’’ where the rainfall return period exceeds 2 years. Rainfall return periods are normally set within the context of failure probability and consequence. Conventional systems are usually designed assuming 100% runoff for a 2-minute storm; the 2-minute duration is selected because it is the typical time of concentration for conventional systems. Although advice is given in codes for setting higher runoff rates, there is little guidance on setting runoff rates below 100%. These observations mean that inadequacies are encountered if conventional codes are used to design green roofs: • Runoff coefficients should be expected to be below that used for conventional roofs; 100% is used by BS EN 12056-3:2000 and 98.7% was recorded by Pratt and Parkar (15). • Peak runoff rates are reduced; even where there is no infiltration, the surface roughness has a significant impact.

• Time of concentration is expected to be greater than 2 minutes; particularly relevant when designing collection pipework for large roof areas for public sector, commercial, and industrial properties. • As with other elements of urban drainage design, it is not efficient for a complex system such as a green roof to be matched to a single rainfall. It is probable that the duration of runoff hydrographs will be orders of magnitude longer compared with conventional systems, and runoff interactions between independent rainfalls are probable; this may make a time-series approach more appropriate. RAINWATER COLLECTION GUTTERS The basic requirement for rainwater collection gutters is that they have sufficient flow capacity to accommodate flows from the design storm (16). Although it is common practice to install gutters at a slight gradient to prevent ponding, the nature of the construction industry and the process of settlement means that it is normal to assume that gutters laid at slack gradients are actually flat; for example, BS EN 12056-3:2000 stipulates that gutters at gradients less than 0.3% shall be treated as flat (17). In a level gutter, the water surface profile will slope toward the outlet, and it is the difference in hydrostatic pressure along the gutter that gives the incoming water the required momentum to flow toward the outlet (18). Gutter Outlet Depths Key to ensuring whether or not collection gutters have sufficient capacity are the conditions that occur at the gutter outlets. As well as affecting the flow rates entering the drainage system pipework, the outlet depths also affect upstream gutter depths (via the backwater surface profile). Hence, although the depth at a gutter outlet may not cause any particular problems, the greater depths at the upstream end of the gutter may result in overtopping. Extensive experimental studies in the 1980s determined that the flow conditions in the vicinity of a gutter outlet in a conventional roof drainage system could be categorized as either ‘‘weir’’ type or ‘‘orifice’’ type, depending on the depth of water relative to the size of the outlet (19). At depths below that equivalent to half of the outlet diameter, the flow conditions are ‘‘weir’’ type, and outlet conditions are calculated using an appropriate sharp-edged weir equation (18). At higher flow depths, the flow effectively ‘‘chokes,’’ and the flow regime changes to ‘‘orifice’’ type; the outlet conditions are calculated by an appropriate sharp-edged orifice equation (18). Although conventional roof drainage systems are usually designed to ensure free discharge at gutter outlets, design restrictions may mean that the outlets cannot discharge freely; in such circumstances, additional gutter capacity (storage) is normally required to accommodate the resulting higher flow depths. In siphonic roof drainage systems, the outlets are designed to become submerged to allow full-bore flow to develop and be sustained; if this is the case, the determination of outlet depth is complicated as the gutter

ROOF DRAINAGE HYDRAULICS

conditions depend on downstream conditions (within the connected pipework) as well as gutter inflows. Recent experimental work has also indicated that conventional roof drainage systems incorporating ‘‘nonstandard’’ gutter sections, whose base width and height are significantly greater than the diameter of the outlet, can result in the development of full-bore flow in the vertical downpipe and siphonic action (20); for a given gutter section, the onset and extent of such conditions depend on the diameter of the downpipe. Similar phenomena have also been observed in ‘‘standard’’ gutter sections (semicircular and elliptical); in these cases, limited siphonic action occurs for only a short distance below the outlet (18).

friction losses are minor and may be ignored (18). If a gutter outlet allows free discharge and frictional effects are neglected, the backwater profile may be determined by applying Equation 1 to determine the horizontal distance (L) between any given upstream depth (h1 ) and downstream depth (h2 ).
h2 L = h1

In terms of flow division between multiple outlets in a gutter under free discharge, it can be seen from Fig. 2a that the flow splits evenly in any given gutter section (between two outlets or between an end wall and an outlet), whether or not the gutter inflow is uniform or nonuniform. Figure 2b, c indicates the effect of outlet placement within a gutter; evenly spaced outlets require far less gutter capacity than those placed at gutter extremities. Where outlets are not freely discharging, the flow division between multiple outlets in a gutter may not be as described, as the individual gutter sections may ‘‘hydraulically merge’’ to form one continuous channel and/or downstream system conditions may become significant. For example, the pipework in a siphonic system runs full bore when operating at or near its design point, and the flow division between outlets depends on the relative losses for each branch of the system.

0.5 Q 2

(1)

Current Design Methods The foregoing discussion has highlighted the key elements that should be considered when designing a rainwater gutter. However, without recourse to some form of numerical modeling, it is not feasible to calculate backwater surface profiles, and hence gutter capacities, for roof drainage systems; this is particularly the case for large commercial or manufacturing developments which may incorporate many kilometers of different types of guttering. Consequently, current gutter design methods for gutters installed in conventional drainage systems are based primarily on empirical relationships (19) and the assumption of free discharge at the outlet. For example, BS EN 12056-3:2000 specifies that the design capacity of a ‘‘short,’’ level, semicircular gutter located on the eaves of

The water surface profile in gutters can only be assessed realistically by applying the momentum equation for channels with lateral input. In many cases, the low velocities associated with gutter flows mean that gutter

Q1

1−

This equation can be modified if frictional effects are significant (very long gutter lengths or very high flow velocities) or if the gutter outlet is not freely discharging.

Backwater Profiles

Q 2 (nonuniform)



  Q2 T   gA3    dh  Q2 So − 2 A mC2 

where Q = flow rate (m3 /s) T = surface width (m) g = gravitational constant (m/s2 ) A = flow area (m2 ) So = bed slope (-) m = hydraulic mean depth (m) C = Ch´ezy coefficient (-)

Flow Division within Gutters

(a) Q 1 (uniform)

57

Q 3 (uniform)

0.5 Q 3

0.5 Q 2

Q (uniform)

(b)

0.25 Q

0.25 Q

0.25 Q

0.25 Q

Q (uniform)

(c)

0.5 Q

0.5 Q

Figure 2. Effect of outlet positioning on flow division in gutters. (a) Flow division between multiple outlets in a gutter. (b) Flow division between evenly spaced outlets in a gutter. (c) Flow division between outlets positioned at gutter extremities.

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ROOF DRAINAGE HYDRAULICS

a building (with outlets capable of allowing free discharge) is given by (17) QL = 0.9 × QN = 0.9 × 2.78 × 105 × A1.25 E

(2)

where QN = notional gutter design capacity (L/s) AE = gutter cross-sectional area (mm2 ) Clearly not all gutters can be designed by application of Equation 2. For example, BS EN 12056-3:2000 (17) contains clauses to account for many eventualities, including • location of gutter on building that may result in varying consequences of failure; eaves gutter, valley gutter, parapet gutter • differently shaped gutter sections • ‘‘hydraulically long’’ gutters (where frictional effects may be significant) • gutters laid at a significant gradient • changes in gutter alignment (bends, offsets, etc.) • additional system elements, such as strainers or rainwater hoppers • restricted flow at outlets • gutters installed in siphonic roof drainage systems In addition to the type of clauses listed above, BS EN 12056-3:2000 also allows designers to use data obtained from experimental testing of a particular arrangement (17).

SYSTEM PIPEWORK The type and extent of pipework incorporated into a roof drainage system depends primarily on whether the system is conventional or siphonic. Conventional Rainwater Systems In conventional roof drainage systems, the aboveground pipework generally consists of vertical downpipes, connecting the gutter outlets to some form of underground drainage network, and offset pipes, used where the gutter overhang is significant. Note that an offset pipe is defined as a pipe with an angle less than 10◦ to the horizontal. The capacity of the system as a whole usually depends on the capacity of the gutter outlets rather than the capacity of the vertical downpipes. The flow within vertical downpipes is normally free surface; BS EN 12056-3:2000 (17) specifies that downpipes run no more than 33% full; this effectively installs redundant capacity within the system. If the downpipes are sufficiently long (normally greater than 5 meters), annular flow may occur. Similarly, the flow within offset pipes will also normally be free surface; BS EN 120563:2000 (17) specifies that offsets run no more than 70% full, indicating the need to install all offsets at a gradient. The design of the pipework can either be undertaken using the design tables in BS EN 12056-3:2000 or by applying the Wyly–Eaton equation for vertical downpipes (22) and the Colebrook–White equation for offset pipes (23). Siphonic Roof Drainage Systems

Numerical Models Numerical models have been developed that can accurately simulate the flow conditions in any type of gutter as a result of either steady or unsteady roof runoff. An example of this is incorporated into the ‘‘ROOFNET’’ model recently developed as part of an academic research project dealing with the effect of climate change on urban drainage (20). This model enables the user to specify data describing the relevant aspects of a particular installation, including details of the prevailing rainfall conditions, details of the roof surfaces to be drained, and details of the actual gutters. A kinematic wave model is then used to route the rainfall over the roof surfaces and into the gutters. A method of characteristics solution of the fundamental equations of one-dimensional flow in open channels is then used to route the runoff along the gutters to the outlets (21), at which point the flow enters the drainage pipework. The model automatically determines the flow conditions at the gutter outlets and, in addition to dealing with free discharge, can also simulate the effect of restricted flow and submerged outlet scenarios. Output includes depths, velocities, and flow rates along the gutter, as well as the location and severity of any gutter overtopping. At present, models such as those described before are research tools; they are normally developed and used by universities for specific research projects. However, it is envisaged that such models may soon be used as diagnostic design aids, particularly for national code development.

In contrast to conventional systems, siphonic installations depend on purging air from the system (priming) and subsequently establishing full-bore flow within the pipework connecting the outlets in the roof gutters to the downstream surface water sewer network (at ground level). Current design practice assumes that, for a specified design storm, a siphonic system fills and primes rapidly with 100% water (24). This assumption allows siphonic systems to be designed using steady-state hydraulic theory. The steady flow energy equation is normally employed (25), and the elevation difference between the gutter outlets and the point of discharge is equated to the head losses in the system. Although this approach neglects the small quantities of entrained air that always enter a siphonic roof drainage system, it reportedly yields operational characteristics similar to those observed in laboratory test rigs in the fully primed state (25,26). However, steady-state design methods are not applicable when a siphonic system is exposed to rainfall below the design criteria or with time-varying rainfall intensity. In the former case, the flow may contain substantial quantities of entrained air and exhibit pulsing or cyclical phases, a result of greatly varying gutter water levels and an indication of truly unsteady, transient flow. Such problems are exacerbated when the system incorporates more than one outlet connected to a single downpipe (multi-outlet system), as the breaking of full-bore conditions at one of the outlets (due to low gutter depths and air entry) is

ROOF DRAINAGE HYDRAULICS

transmitted throughout the system and, irrespective of the gutter depths above the remaining outlet(s), results in cessation of fully siphonic conditions. As subdesign events are the norm, it is clear that current design methods may not be suitable for assessing the day-to-day performance characteristics of siphonic roof drainage systems. This is a major disadvantage, as it is during these events that the majority of operational problems tend to occur, for example, noise and vibration. Despite any defects that current design methods may have, thousands of systems have been installed worldwide with very few reported failures. Where failures have occurred, they have invariably been the result of one or more of the following:

5. Once the conditions throughout the downpipe are full bore, any remaining air pockets are purged from the system (Phase 2). 6. Full siphonic action occurs (Phase 3) and continues until the gutter depth(s) falls below the level at which air can enter the system. The data shown in Fig. 4a illustrate the type of unsteady flow conditions that occur when a siphonic system is exposed to rainfall below the design point and the gutter flow depths are insufficient to sustain full siphonic action. The data shown in Fig. 4b illustrate the type of unsteady flow conditions that occur when an installed siphonic system is exposed to a ‘‘real’’ rainfall and the rainfall intensity varies with time. Figure 5 shows an example of the output from one of the numerical models that has recently been developed (SIPHONET). As can be seen, the model can accurately simulate the priming of a siphonic system (0–32 s) as well as steady siphonic conditions (32–62 s). These data also illustrate that the model can simulate complex operating conditions, such as the rise in system pressure when the depth in gutter 1 drops below that necessary for full-bore flow, hence allowing air to enter the system and break the siphon (at approximately 62 s).

1. a lack of understanding of operational characteristics 2. poor material specification 3. installation defects 4. a poor maintenance program In response to these perceived shortcomings, a series of research projects has recently been undertaken to augment the understanding of siphonic roof drainage systems and to develop numerical models for use as diagnostic design aids (27). The remainder of this section will present a selection of the salient points arising from this work. In contrast to the assumption made in current design methods, the priming of a typical siphonic system actually found was as follows (refer to Fig. 3):

CONCLUSION The text has illustrated how roof drainage systems are a key, but often overlooked, element of urban drainage infrastructure. It has also been shown that their design is a complex process, which relies heavily on gutter outlet performance. The following conclusions may be drawn with respect to the operation of roof drainage systems:

1. Flow conditions throughout the system are initially free surface (Phase 1). 2. Full-bore flow forms at some point within the horizontal pipework (Phase 1). 3. Full-bore flow conditions propagate downstream toward the vertical downpipe and upstream toward the gutter outlets (Phase 1). 4. Full-bore flow conditions reach the vertical downpipe, the downpipe starts to fill, and the system starts to depressurize (Phase 2).

Phase 1

Phase 2

59

1. Their operation depends on three interacting components: the roof surface, the collection gutter, and the collection pipework. 2. Green or brown roofs provide an opportunity to reduce the flow from roof surfaces, improve urban aesthetics, and increase biodiversity. 3. Outlet conditions are key to understanding how a system performs.

Phase 3

150

Gutter 1

Gutter 2

0

0.6 Pressure P1

Depth in gutter 1 Depth in gutter 2 Pressure P1 0

10

Schematic of test rig

0.0 −0.6

Pressure (mH2O)

Gutter depth, mm

75

−1.2

Air pockets leave system −1.8 20 30 40 50 60 Time since start of simulated rainfall, s

Figure 3. Priming of a laboratory siphonic drainage test rig (28).

60

ROOF DRAINAGE HYDRAULICS (a)

0.4

Pressure (mH2O)

0.0 −0.4 −0.8

Regime 1 (15−40% Qmax)

−1.2

Regime 2 (40−60% Qmax)

−1.6

Regime 3 (60−80% Qmax) 0

20

40

60

80

Time since start of simulated rainfall, s (b) 120

0.5 0.0

100

−0.5

80

−1.0

60

−1.5

40

−2.0

20

−2.5

0

1000

2000 3000 4000 5000 Time since start of rainfall, s

160

Gutter depth (mm)

0 7000

Measured gutter 1 depth Predicted gutter 1 depth

80

Figure 5. Measured and predicted system conditions within a laboratory siphonic drainage test rig: no inflow into gutter 1 between 62 and 82 s (28). Note that data refer to the system shown in Fig. 3.

6000

Rainfall intensity (mm/hour)

Figure 4. (a) Measured system pressures for subdesign rainfall within a laboratory siphonic drainage test rig (28). Note that data refer to the pressure P1, as indicated in Fig. 3. (b) Subdesign rainfall within an installed siphonic drainage system (28).

Rainfall intensity

0.4

0

−0.4 Measured pressure P1 Predicted pressure P1 0

4. Siphonic roof drainage systems present a more efficient way to drain large roof surfaces. 5. The design of siphonic roof drainage systems should consider subdesign rainfall and operational problems, such as blocked outlets. THE FUTURE Although green roofs are an attractive alternative, it is probable that conventional roof surfaces will continue to

25 50 75 Time since start of simulated rainfall, s

Pressure (mH2O)

Pressure (mH2O)

Pressure P1

−1.2 −2.0 100

dominate domestic installations. However, it is likely that green roofs will experience a step-change in acceptance by the commercial sector once more becomes known about their performance and sustainability. Similarly, the efficiencies offered by siphonic systems means that they will continue to play a significant role in draining large commercial buildings, particularly if numerical models are applied diagnostically to improve performance and reduce costly system failures. The biggest threat to roof drainage comes from climate change. Existing systems may not simply become more

SEPTIC TANK SYSTEMS

prone to flooding; changes in rainfall patterns may result in long periods of low precipitation, and self-cleansing velocities may be attained less frequently as a result. Furthermore, changes in wind patterns may also increase levels of rooftop debris and hence necessitate enhanced maintenance programs. As concern regarding climate change grows and the sustainability agenda widens, it is possible that harvesting roof runoff may become more widespread. At present, water consumption varies globally between 7 and 300 liters/household/day (L/h/d). In the United Kingdom, average consumption is 145 L/h/d, but only 1–2 liters may actually be consumed by humans, 30% may be used for WC flushing (29). Studies have shown that, when coupled with storage, roof rainwater harvesting has the potential to contribute substantially to domestic water usage in both developing and developed countries (30,31). BIBLIOGRAPHY 1. BS EN 6229:2003. (2003). Flat Roofs with Continuously Supported Coverings, Code of practice, British Standards Publishing Limited (BSPL). 2. Simmons, T. (1994). Methods for designing proper roof drainage. Prof. Roofing 22–25. 3. Singh, V.P. (1996). Kinematic Wave Modeling in Water Resources: Surface Water Hydrology. John Wiley & Sons, New York. 4. Tarr, A.R. (2003). Green roof implementation: construction and contractor issues in the UK. Green Roofs for Healthy Cities. University of Sheffield. 5. K¨ohler, M. (2004). The multi-beneficiary system of green roofs: The green roof challenge to biophilic architecture and ecology. Proc. Sheffield Conf. ‘‘Nature Enhanced’’. 6. Rosenfeld, A.H. et al. (1995). Mitigation of urban heat islands: materials, utility programs, updates. Energy Build. 22(3): 255–265. 7. Hendricks, N.A. (1994). Designing green roof systems: A growing interest. Prof. Roofing 20–24. 8. K¨ohler, M., Schmidt, M., Paiva, V.L.A., and Taveres, S. (2002). Green roofs in temperate climates and in hot-humid tropics—far beyond aesthetics. Environ. Manage. Health 13(4): 382–391. 9. Niachou A. et al. (2001) Analysis of the green roof thermal properties and investigation of its energy performance. Energy Build. 33(7): 719–729. 10. Onmura, S., Matsumoto, M., and Hokoi, S., (2001), Study on evaporative cooling effect of roof lawn gardens. Energy Build. 33: 653–666. 11. K¨ohler, M. (2004). Green roof technology—From a fireprotection system to a central instrument in sustainable urban design. Second Green Roof Conf., Portland, Oregon. 12. GVRD. (2002). Effectiveness of Stormwater Source Control. Greater Vancouver Regional Sewerage and Drainage District. 13. PSU (2001). Online Research http://www.rps.psu.edu/0105/roofs.html. 14. NJDEP (2000). Standard for Rooftop Runoff Management. 15. Pratt, C.J. and Parkar, M.A. (1987). Rainfall loss estimation on experimental surfaces. Proc. Fourth Int. Conf. Urban Storm Drainage, International Association of Hydraulic Runoff Research, Lausanne, Switzerland. 16. Beij, H. (1934). Flow in Roof Gutters. US Department of Commerce, Bureau of Standards: Research Paper RP644, Bureau of Standards Journal of Research, Vol 12.

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17. BS EN 12056-3:2000. (2000). GravitY Drainage Systems Inside Buildings. Roof drainage, layout and calculation, British Standards Publishing Limited (BSPL). 18. May, R.W.P. (1995). Design of conventional and siphonic roof drainage systems, Public Health Services in Buildings—Water Supply, Quality and Drainage, IWEM Conference, London. 19. May, R.W.P. (1984). Hydraulic Design of Roof Gutters. Proceedings of the Institution of Civil Engineers, Part 2, Vol. 77. 20. Blanksby, J., Ashley, R., Saul, A.J., Cashman, A., Packman, J., Maksimovic, C., Jack, L., Wright, G., and Kay, D. (2004). Adaptable urban drainage (AUDACIOUS), NOVATECH 2004: 5th Int. Conf. sustainable Tech. Strategies Urban Water Manage. 21. Escarameia, M. and Swaffield, J.A. (1999). Prototype monitoring and numerical simulation of roof drainage gutter systems. Proc. 8th Int. Conf. Urban Storm Drainage, Sydney, Australia. 22. Wyly, R.S. and Eaton, H.N. (1952). Capacities of Plumbing Stacks in Buildings, National Bureau of Standards Building Materials and Structures, Report BMS 132. 23. Chadwick, A. and Morfett, J. (2004). Hydraulics in Civil and Environmental Engineering, 4th Edn. Spon Press. 24. Arthur, S. and Swaffield, J.A. (2001). Siphonic roof drainage: The state of the art. Urban Water 3(1): 43–52. 25. May, R.W.P. and Escarameia, M. (1996). Performance of siphonic drainage systems for roof gutters. Report No SR 463. HR Wallingford. 26. Arthur, S. and Swaffield, J.A. (2001). Siphonic roof drainage system analysis utilising unsteady flow theory. Build. Environ. 36(8): 939–948. 27. Wright, G.B., Swaffield, J.A., and Arthur, S. (2002). Investigation into the performance characteristics of multi-outlet siphonic rainwater systems. Build. Serv. Eng. Res. Technol. 23(3): 127–141. 28. Swaffield, J.A., Wright, G.B., Jack, L.B., and Arthur, S. (2004). Pressure transient analysis to inform system design for building and roof drainage systems. Proc. The Practical Application of Surge Analysis For Design And Operation, 9th Int. Conf. 29. Butler, D. and Davies, J. (2004). Urban Drainage, 2nd Edn. Spon Press. 30. Thomas, T. (1998). Domestic water supply using rainwater harvesting. Build. Res. Inf. 26(2): 94–101. 31. Fewkes, A. (2000). Modelling the performance of rainwater collection systems: towards a generalised approach. Urban Water, 1(4): 323–333.

SEPTIC TANK SYSTEMS JOHN E. MOORE Hydrologic Consultant Denver, Colorado

According to Wilson and Moore (1), a septic tank is an ‘‘underground vessel for treating wastewater from a single dwelling or building by a combination of settling and anaerobic digestion. Effluent is usually disposed of by leaching. Settled solids are pumped out periodically and hauled to a treatment facility for disposal.’’ When properly

62

SEPTIC TANK SYSTEMS

Septic tank

Drain field

Wastewater flow Movement of gases

Unsaturated zone Groundwater flow Figure 1. Components of septic tank systems (4).

sited, constructed, and maintained, septic systems can provide a low-cost environmentally responsible method of waste disposal. Improperly sited, constructed, operated, or maintained septic systems can, however, lead to water quality degradation and threats to public health. The basic components of a septic tank system are shown in Figs. 1 and 2. The septic tank is an enclosed receptacle designed to collect wastewater, segregate floatable solids, accumulate, consolidate, and store solids; wastewater treatment is provided by septic tank systems. The tank is the most important component used in these systems (2). The waste enters the tank near the top. There is a pair of baffles in the tank to keep the solids in the tank, preventing them from flowing out of the tank with liquids. Bacteria in the tank break down the solids as much as they can into a liquid form and this with the water leaves the tank on the other side of the baffles. The liquid then flows to a leaching field where the liquid enters the soil and is absorbed. If the bacteria cannot break the solids down, they will build up over time. If these solids are not removed by periodic pumping, the tank will allow solids to be washed out to the leaching field and begin to clog the soil. When the soil is clogged, the system stops working. Septic systems fail for the following reasons: 1. Faulty design (leaching field that is too small). 2. Faulty installation (plugged lines or uneven grades).

e

us

Ho

Pumphouse or well 50 ft. 100 ft.

Septic tank

Figure 2. Setback distances (4).

Soil absorption field

3. Soil conditions (highly permeable soil or relatively impervious soil, less than 6 feet of unsaturated soil cover). 4. High water table less than 6 feet from the land surface. 5. Water overload. 6. Inadequate cleaning of the tank (should be pumped every 2–3 years). 7. Highly permeable soil. It has been estimated that 25% of the U.S. population uses septic systems for treatment and disposal of their household sewage. Septic system technology is undergoing dramatic changes in efficiency and reduced contamination (2). The American Society of Testing and Material has prepared three standards for the treatment and disposal of on-site waste (D 5879-95, D 592196, D5925-96). Bacterial and viral contamination from septic systems is the most common cause of drinking water contamination in the United States. The liquid effluent from septic systems follows the same path as precipitation moving into an unsaturated zone and aquifer. When the effluent reaches the water table, it moves downgradient to the point of discharge (lake, stream, wetland, and well). The location of the septic system in relation to the slope of the land surface is important because septic tank discharge follows the slope of the land surface. Wells downslope from septic tanks are subject to contamination (3). The septic tank effluent can contain bacteria and also toxic materials and other contaminants. Some of the contaminants adhere to the soil and aquifer material or travel with the water. Many septic systems are found in small rural homes sites and are commonly located on small narrow lots along a feeder highway. An increasing number of states are zoning suburban areas to limit the density of houses using septic tanks (4). Community sewer systems are used in some areas to substitute for septic systems. Some banks require the prospective seller of rural property to provide proof of a bacteria-free water supply. Some sellers chlorinate the water to destroy the bacteria in the well. The bacterial contamination is in the aquifer, so this treatment lasts only a short time (5). The homeowner should have the well water analyzed at least once year for bacteria. A buyer of rural property should determine the location of the well and the septic system. The buyer should also

DOMESTIC SOLAR WATER HEATERS

63

determine the age, maintenance, distance to the drinking water supply well, and depth to water at the septic system site. A wet area, lush vegetation over the leaching field, or odor of sewage is cause for further investigation. A water sample from the well at a septic system site should be obtained and analyzed for fecal coliform bacteria. BIBLIOGRAPHY 1. Wilson, W. and Moore, J.E. (1998). Glossary of Hydrology. American Geological Institute, Denver, CO. 2. Bedinger, M.S. Fleming, J.S., and Johnson, A.I. (Eds.). (1997). Site Characterization and Design of On-Site Septic Systems. ASTM, Philadelphia, PA. 3. Waller, R.W. (1959). Ground Water and the Rural Homeowner. U.S. Geological Survey, General Interest Publication, Washington, DC. 4. Wyoming Department of Environmental Regulations (DER). (1998). Septic Systems: Rural Wellhead Protection Fact Sheet. Cheyenne, WY. 5. Covin, C. (1999). Handbook of Groundwater Quality Protection for Farmers. CSIR, South Africa.

DOMESTIC SOLAR WATER HEATERS MERVYN SMYTH Centre for Sustainable Technologies Newtownabbey, United Kingdom

Figure 1. Advertisement for the climax solar-water heater, 1892 (1).

Domestic solar water heaters can be categorized as being either active or passive and can be further grouped according to the configuration of the main solar water heating components: integral or distributed. Integrated systems combine the collector and storage functions in a single unit, whereas distributed systems have a separate solar collector and hot water store connected by a piping network. Distributed systems can be either active or passive. In active systems, a pump circulates the transfer fluid between the collector and the store. Integrated systems are almost always passive as they do not require external power.

INTRODUCTION Solar water heating systems convert solar radiation into useable thermal energy in the form of hot water. Domestic solar water heaters can provide households with a large proportion of their hot water needs while reducing the amount of conventional fuel used and hence reducing home energy costs. The amount of hot water produced will depend on the type and size of the system, the climate, and location for solar access. Over the years, a variety of system designs have been developed and tested to meet specific consumer needs and environmental conditions. The following article is a brief description of the many types of system in common use today. The first solar water heaters consisted of exposed tanks of water left out to warm in the sun. Used on a few farms and ranches in the Southwestern United States in the late 1800s, they were reportedly capable of producing water hot enough for showering by the late afternoon on clear days (1). The first solar water heater, manufactured commercially under the trade name Climax Solar-Water Heater, was patented in 1891 (2). Figure 1 illustrates a reproduction of an advertisement for the Climax Solar-Water Heater. This water heater could be used from April to October in the State of Maryland in the eastern United States. It claimed to produce water hotter than 38 ◦ C on sunny days even during early spring and in late autumn when daytime temperatures sometimes approached freezing.

THE INTEGRATED COLLECTOR/STORAGE SOLAR WATER HEATER The most basic of solar water heaters is the integrated collector/storage solar water heating (ICSSWH) system or the integral passive solar water heater (IPSWH), commonly referred to as breadbox or batch water heaters. Kemp’s early Climax Solar-Water Heater was an integrated system. A simplified diagram of a typical ICS solar water heating installation is shown in Fig. 2. In its simplest form, the ICSSWH is a water tank painted black to absorb insolation (incident solar radiation). Variations consist of one or more tanks, painted black or coated with a selective absorbing surface, within a well-insulated box, possibly with reflectors and covered with single, double, or even triple layers of glass, plastic, or a combination of the two. Because of its simplicity, an integrated collector/storage system is easier to construct and install, which reduces maintenance and capital costs. In most climates, the large thermal mass of the store provides inherent resistance to freezing. However, the integrated unit has a significant problem because of its unique mode of operation. The earliest systems suffered substantially from heat losses to ambient, especially at night and at noncollection periods, which meant no matter how effective the unit was in collecting solar energy, unless the hot water was fully withdrawn at the end of the collection period, losses

64

DOMESTIC SOLAR WATER HEATERS Pressure relief value A.A.V. I.V.

Roof mounted ICSSWH

(closed) M.V. M.V.

M.V.

Hot supply to appliances Pressure relief value A.A.V.

I.V.

N.R.V.

Cold supply to appliances

I.V.

Cold mains water supply

Domestic hot water cylinder Key

M.V. I.V.

Figure 2. A simplified diagram of a typical roof-mounted ICS solar water heating installation.

A.A.V. = Automatic Air Value I.V. = Isolating value N.R.V. = Non return value M.V. = Motorised value

I.V. (closed)

I.V. To drain

to ambient led to only lukewarm water being available early the next day. This process reduced the overall solar fraction, which renders it less viable economically. Indeed this deficiency in the late nineteenth century led to the prominence of thermosyphon solar water heaters with diurnal heat storage to the detriment of the ICSSWH system. To overcome excessive heat loss and be in a position to compete with the more established distributed solar water heater systems, the ICSSWH design has had to evolve and incorporate new and novel methods of improving performance.

Roof mounted flat plate collector

I.V. Pressure relief valve A.A.V Hot supply to appliances

Cold mains water supply

I.V. N.R.V

I.V. Domestic I.V. hot water cylinder

I.V.

I.V.

I.V.

I.V.

I.V.

DISTRIBUTED SOLAR WATER HEATERS Distributed systems consist of a separate solar collector and water store, with pipes connecting the collector(s) to and from store(s). As previously mentioned, these systems can be either active or passive, with the active system using an electric pump, and the passive system relying on buoyancy forces in the form of thermosiphonic action. Active systems also require more valves and control systems, which tend to make them more expensive than passive systems but generally more efficient. Figure 3 shows a simplified diagram of a typical roof-mounted distributed (flat-plate) solar water heating installation. Active systems are often easier to retrofit than are passive systems because their storage vessels do not need to be installed above or close to the collectors. In addition, a

To drain

I.V. To drain

Key A.A.V = Automatic air valve I.V = Isolating valve N.R.V = Non return valve M.V. = Motorised valve

Figure 3. A simplified diagram of a typical roof-mounted distributed (flat-plate) solar water heating installation.

photovoltaic panel could power the pump, which results in stand-alone, proportional pump operation with reduced running costs. Distributed solar water heaters can also be characterized as being direct (open loop) or indirect (closed loop). A direct system circulates incoming mains water through the collector and into the tank, whereas an indirect system transfers collected thermal energy via a heat exchanger to

DOMESTIC SOLAR WATER HEATERS

the domestic water. Indirect systems usually contain an aqueous antifreeze solution that flows through the heat exchanger immersed in the hot water store to provide protection from freezing. This process, however, results in reduced collection efficiencies over the direct system through lower specific heat capacities and losses during the heat exchange process. Active Direct Systems Active direct systems use pumps to circulate incoming mains water through the collector and back into the tank. This design is efficient and reduces operating costs but is not appropriate where water is hard or acidic because of scale buildup and corrosion. However, direct active systems are popular in regions that do not experience freezing temperatures (Fig. 4). Active Indirect Systems Active indirect systems pump the heat-transfer fluid (usually a glycol-water antifreeze mixture) through the collector and a heat exchanger transfers the heat from the fluid to the water that is stored in the tank. Heat exchangers can be double-walled vessels or have twin coil arrangements. Indirect glycol systems are popular in areas where temperatures regularly fall below zero because they offer good protection from freezing. However, antifreeze systems are more expensive to purchase and install and require regular checking and maintenance. Indirect drainback systems do not use antifreeze mixtures, but they use pumped water as the heat-transfer fluid in the collector loop. When freezing conditions prevail or the system is not in use, the pump is switched off and the water in the collector is drained out, thus providing protection from freezing. The collector installation and plumbing arrangement must be carefully positioned to

65

allow complete drainage and the pump must have sufficient head pressure to pump the water up to the collector each time the pump starts. Thermosiphon Systems A thermosiphon system relies on warm water rising, a phenomenon known as natural convection or buoyancy forces, to circulate water to and from the collector and tank. In this type of installation, the tank must be located above the collector. As water in the collector heats, it becomes less dense and naturally rises into the tank above. Meanwhile, cooler water in the tank flows downward into the collector, thus causing circulation throughout the system. Some forms of thermosiphon solar water heaters can be described as being compact. Compact systems are close-coupled thermosyphon flat-plate or evacuated-tube collector units fabricated and installed as a single item as opposed to a separate collector, store, and pipework. Thermosiphon systems are much cheaper than are active systems as no pump or controller is required and are ideal where a low-cost solar heater is required such as holiday houses and cabins, or countries where low-cost solar heating is required. SOLAR WATER HEATING COLLECTORS Basically three types of domestic solar collector are in common use today: flat-plate, evacuated-tube, and concentrating. Flat-plate Solar Collector The flat-plate system consists of a ‘‘flat’’ absorber panel through which water or conducting fluid passes. The panel may be of formed channels in a sandwich format or may be pipes connected to expanded absorber plates. Most

(b)

(a) Collector

Collector

Hot outlet

Open collector loop

Hot outlet

Open collector loop

Auxillary heating

Auxillary heating Cold inlet

Cold inlet Direct HWC

Indirect HWC (d)

(c) Collector

Hot

Drainback collector loop

Auxillary heating Cold

Drain back cylinder

HWC

Altemative location for compact thermosyphon tank

Collector

Hot outlet

Auxillary heating Cold inlet Direct or indirect HWC

Figure 4. Schematic detail of common distributed solar water heating configurations: (a) active direct system, (b) active indirect system, (c) indirect drainback system, and (d) thermosiphon system.

66

DOMESTIC SOLAR WATER HEATERS

absorbers are covered with a selective coating to improve solar radiation absorption and reduce long-wave radiative heat loss. As the fluid flows adjacent to the heated surface, it is heated. The absorber is mounted in an insulated, weatherproof unit, and the exposed collector aperture is covered with one or more transparent or translucent covers. The make-up of a typical flat-plate solar collector is shown in Fig. 5.

Co l ma d wat nifo er ld

Outer tubes

Ho ma t wat nif er old

Evacuated-tube Collector Evacuated-tube collectors are made up of rows of parallel, glass tubes, linked to a common flow (and return) manifold depending on the collector installed. Two types of evacuated-tube collector exist: glass/glass or metal/glass. The glass/glass collector consists of two concentric glass tubes. The inner tube is covered with a selective coating to improve solar radiation absorption and reduce long-wave radiative heat loss. The transparent outer tube forms a space between the two tubes that is evacuated to eliminate conductive and convective heat loss. The metal/glass collector consists of a copper plate attached to a heat pipe or water pipe mounted within a single evacuated glass tube. Again the absorber is coated with a selective coating to improve the collection performance. Figures 6 and 7 illustrate some common evacuated-tube collectors.

Outlet manifold

Casing

Absorber plate

Glazing Collector back Insulation high temperature rigid foam

Inlet manifold

Figure 5. A typical flat-plate solar collector.

Heat exchange manifold Hot outlet Evacuated glass tubes Heat pipe and absorber plates

Vaccum

Figure 7. Metal/glass water pipe evacuated-tube solar collector.

Hot outlet manifold Tubular absorbers CPC reflectors

Cold inlet manifold

Outlet

Inlet

Inner tubes coated with selective absorbant coating

Transfer fluid

Cold inlet

Figure 8. Diagram ing collector.

of

a

compound

parabolic

concentrat-

Concentrating collector To increase the insolation on the absorber surface over that incident at the collector aperture, reflectors are employed in solar water heating systems. Concentrating reflectors can obtain higher temperatures on the absorbing surface than can those achievable by a flat absorber, and as the absorbing surface area is reduced relative to that of the aperture, a reduction in the overall heat loss from the system occurs, hence an improved thermal efficiency. Internal reflectors are contained within the unit enclosure, whereas external reflectors are located outside the sealed casing. Reflecting concentrator designs for low-to-medium concentrations can be flat or curved, line-axis or line-focus (circular, parabolic or compound parabolic) reflectors, symmetrical or asymmetrical. The concentrating collector used for domestic applications usually incorporates a concentrating reflector in the form of parabolic trough or compound parabolic concentrating (CPC) collector, using highly reflective surfaces to concentrate the insolation onto the absorber. Most absorbers are tubular, although not exclusively. Figure 8 illustrates a CPC collector. BIBLIOGRAPHY

Figure 6. Metal/glass heat pipe evacuated-tube solar collector.

1. Butti, K. and Perlin, J., (1981). A Golden Thread. Marion Boyars Publishers Ltd, London, UK. 2. Kemp, C.M. (1991). U.S. Patent 451384, April 28.

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READING LIST Duffie, J.A. and Beckman, W.A. (1991). Solar Engineering of Thermal Processes. John Wiley & Sons, New York. Norton, B. (1992). Solar Energy Thermal Technology. SpringerVerlag London Ltd, London, UK.

HOUSEHOLD DRINKING WATER TREATMENT AND SAFE STORAGE SUSAN MURCOTT Massachusetts Institute of Technology Cambridge, Massachusetts

Household water treatment is the decentralized treatment of drinking water in the home and safe storage is the protection of drinking water in specially designated household storage vessels prior to use. A safe water storage vessel is typically comprised of a container with a narrow mouth to prevent contact with potentially dipped cups or dirty hands, a lid, a spigot to access water and a flat base for easy water extraction. In many cultures and regions, household drinking water treatment and storage has been women’s work based on traditional practices stretching back for millennia. For example, an ancient Indian medical text, the Susruta Samhita, compiled over several centuries and reaching its present form in about A.D. 300, includes the prescribed water treatment and handling practices as follows: Heat contaminated water by boiling on fire, heating in the sun, by dipping hot copper into it seven times, cooling in an earthen vessel and also scenting it with flowers of nagkesara, campaka, utpala, patala, etc. (Book 1, Chapter 45, Verse 12)

Moreover, home storage in various types of containers, including skin bags, ostrich eggs, vessels of wood, ceramic, metal, glass, or stone has been a traditional practice for hundreds or even thousands of years. Thus the twin concepts of household drinking water treatment and safe storage are not new. But there are new developments arising from a global need (Figs. 1–3). Currently, about 50% of people worldwide are supplied with household connections that provide drinking water on tap in their homes. Sufficient, safe, acceptable, physically accessible, and affordable water for all is a fundamental human right essential to life and dignity. Tapped water for all is the long-term goal, but even among those with tap water today, the drinking water is not always considered safe, in terms of its water quality. In homes with a tapped water supply, household treatment devices typically provide a final ‘‘extra’’ step that begins with a well-protected source and includes a treatment process provided by a centralized water treatment system, administered by a municipal authority or private entity. In these cases, the purpose of the household treatment step is typically to improve the aesthetics of the water (e.g., chlorine odor or taste, hardness) and/or to remove certain harmful contaminants, including

Figure 1. Ceramic filter—Nepal.

Figure 2. Ceramic candle filter by Katadyne, Switzerland.

possible organic (e.g., benzene, toluene), inorganic (e.g., cadmium, lead), or microbiological (e.g., Cryptosporidium, Giardia) substances. These household drinking water

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HOUSEHOLD DRINKING WATER TREATMENT AND SAFE STORAGE

Figure 3. Ceramic filter ‘‘filtron’’—Ghana.

treatment devices take two forms—point-of-entry or pointof-use—depending on whether the device is installed at the point where the water main enters the home or whether the treatment unit is attached to or placed beside the kitchen faucet (i.e., at the point where drinking water is withdrawn). In such cases, household drinking water treatment in industrialized countries and regions is used to provide an additional barrier of safety to a water supply that has already received treatment upstream or is of known high quality. Homes lacking a tapped drinking water supply via a household connection or lacking another form of ‘‘improved’’ water supply such as a public standpipe, a borehole, a protected dug well, a protected spring, or rainwater collection are more likely to bear the burden of water-related illnesses: • 3.4 million deaths are water-related; • 1.4 million children die annually of diarrhea, making this the third highest cause of illness and the sixth highest cause of mortality globally; • 1.5–2 billion people are affected by intestinal parasites; • 1.1 billion people lack access to safe drinking water; • 2.6 billion people are without access to basic sanitation (1). These combined conditions can be addressed and an improved quality of life can be realized by applying the same principles that brought about the industrialized world public health miracle of the nineteenth and twentieth centuries—a treated drinking water supply, sanitation, and good hygiene practices—to households globally. Between 1990 and 2002, 1.1 billion more people worldwide gained access to improved water supplies. Yet that same number—1.1 billion or about one in six people—still lack access to improved water in 2004. Most of these people live in rural areas and urban and periurban slums. Their water needs are a focal point of international efforts to provide safe drinking water (2). For

these 1.1 billion people, household water treatment and safe storage is not an additional barrier, post-treatment, as it is for those who purchase and use point-of-entry or pointof-use systems, but instead it may be their main barrier in the prevention of water-related illness. And these systems work! ‘‘There is now conclusive evidence that simple, acceptable, low-cost interventions at the household and community level are capable of dramatically improving the microbial quality of household stored water and reducing attendant risks of diarrheal disease and death’’ (3). Moreover, we know that household drinking water treatment and safe storage, access to sanitation, and hygienic behavior are all interrelated activities. The combination of all these three main interventions will maximize health benefits to all. Household drinking water treatment and safe storage is one essential technology with a special role to play for households lacking a safe water supply. It was with this understanding that the World Health Organization formed the International Network to Promote Household Drinking Water Treatment and Safe Storage, a public–private partnership announced at the Kyoto World Water Forum in March 2003. While we know that household water treatment and safe storage has been practiced locally and regionally, recognition of the role that household water treatment and safe storage can play globally in securing safe drinking water is a recent development dating to the 1990s. Research and development have been a process of adapting traditional wisdom and best engineering and public health practices, applied in settings that necessitate simple, low maintenance designs, use of local materials, applications under demanding local conditions, social acceptability, and economic sustainability. Research on cost effectiveness indicates that these household water treatment and safe storage practices can avert much of the burden associated with diarrheal disease at low cost(4). Some of the treatment processes for household drinking water treatment and safe storage currently under investigation and/or in early stages of implementation include (see Figs. 4–6): Sedimentation Mechanical and/or biological filtration

Figure 4. Chlorine solution for household disinfection.

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69

Mixtures of coagulants/flocculants, weighting agents, calcium hypochlorite Adsorption Arsenic remediation household systems Fluoride remediation household systems Ion exchange processes Membrane/reverse osmosis processes Oxidation processes Disinfection Chlorine and the safe water system Solar UV or UV lamp disinfection Heat disinfection or pasteurization

Figure 5. Safe water storage container.

Distillation Combined (multiple barrier) household treatment systems Sedimentation + solar UV disinfection Pretreatment filters (strung-wound + granular activated carbon filter + chlorine disinfection) Pretreatment cloth + sand + ceramic candles with colloidal silver Coagulation/flocculation + filtration + chlorine disinfection Other combinations Beyond inactivation and/or removal of microbiological contamination—which is the major concern for those lacking access to safe drinking water—appropriately designed household drinking water treatment can effectively remove physical substances (e.g., turbidity) and/or toxic chemicals (e.g., arsenic, fluoride, pesticides) as well as microbiological contamination by one or several of the processes listed above. On every continent, there are promising household drinking water treatment and safe water storage options available. In the decade to come, we will witness new research, innovation, and scale-up of these systems from hundreds to thousands to millions to meet the enormous global need for clean, safe drinking water.

Figure 6. Household ter—Nepal).

arsenic

filter

(Kanchan

arsenic

Cloth filters Ceramic water filters Intermittent household slow sand filters Coagulation/flocculation Metal salts (e.g., alum, ferric chloride, ferric sulfate) Natural polymers

fil-

BIBLIOGRAPHY 1. WHO/UNICEF Joint Monitoring Programme on Water Supply and Sanitation (2004). Meeting the MDG Drinking Water and Sanitation Target: A Mid-Term Assessment of Progress. Available at http://www.who.int/water sanitation health/ monitoring/jmp2004/en/. 2. Millennium Development Goal #7: to reduce by half the proportion of people lacking access to safe water and sanitation by 2015. Available at www.developmentgoals.org. 3. Sobsey, M. (2003). Managing water in the home: accelerated health gains from improved water supply. Available at http://www.who.int/water sanitation health/dwq/wsh0207/ en/index.htm.

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4. Hutton, G. and Haller, L. (2004). Evaluation of the costs and benefits of water and sanitation improvements at the global level. In: Water, Sanitation and Health Protection of the Human Environment. World Health Organization, Geneva, WHO/SDE/WSH/04.04. Available at http://www.who.int/ water sanitation health.

VIRUS TRANSPORT IN THE SUBSURFACE ANN AZADPOUR-KEELEY JACK W. KEELEY United States Environmental Protection Agency Ada, Okaholma

BACKGROUND The 1986 Safe Drinking Water Act (SDWA) amendments directed the EPA to develop national requirements for drinking water disinfection. The legislation required all public water supply systems to disinfect unless they fulfill criteria ensuring equivalent protection. To provide direction for the regulations associated with ‘‘acceptable’’ health risks to the public (4), the EPA established goals for maximum contaminant levels (MCLGs) of pathogenic microorganisms in drinking water, set or setting a level of zero for viruses (5,6). On June 29, 1989, a Surface Water Treatment Rule (SWTR) was published addressing microbial contamination of drinking water from surface sources or from groundwater sources directly influenced by surface water, that had strict provisions for filtration and disinfection (5). On January 14, 2002, a SWTR was promulgated with special emphasis on the protozoan Cryptosporidium (7). The development of a corresponding rule for groundwater, Ground Water Disinfection Rule (GWDR, later designated as the Groundwater Rule), to meet SDWA requirements began in 1987 and led to a published discussion piece (8) and a deadline for the GWDR proposal upon completion of the status of public health with respect to the microbial contamination of groundwater by conducting studies to generate a more careful nationwide picture of the problem. On May 10, 2000, ‘‘US EPA proposed to require a targeted risk-based regulatory strategy for all groundwater systems addressing risks through a multiple barrier approach that relies on five major components: periodic sanitary surveys of groundwater systems requiring the evaluation of eight elements and the identification of significant deficiencies; hydrogeological assessments to identify wells sensitive to fecal contamination; source water monitoring for systems drawing from sensitive wells without treatment or with other indications of risk; a requirement for correction of significant deficiencies and fecal contamination (by eliminating the source of contamination, correcting the significant deficiency, providing an alternative source water, or providing a treatment which achieves at least 99.99 percent (4-log) inactivation or removal of viruses), and compliance monitoring to insure disinfection treatment is reliably operated where it is used (9). The Ground Water Rule will be issued in 2005.

INTRODUCTION More than 97% of all freshwater on the earth is groundwater. Of more than 100 million Americans who rely on groundwater as their principal source of potable water, over 88 million are served by community water systems and 20 million by noncommunity water systems (9). Historically, groundwater has been considered a safe source of drinking water which required no treatment. It has long been believed that this valuable resource was protected from surface contamination because the upper soil mantle removed pollutants during percolation. It was also believed that, even if contaminated, groundwater would be purified through adsorption processes and metabolism of indigenous aquifer microflora. In the United States alone, the estimated annual number of reported illnesses resulting from contact with waterborne pathogens was as low as one million and as high as seven million between 1971 and 1982, and 51% of all waterborne disease outbreaks due to the consumption of contaminated groundwater (1). It is estimated that approximately 20–25% of U.S. groundwater sources are contaminated with microbial pathogens, including more than 100 types of viruses. A literature review by Craun (2) indicated that approximately one-half of the surface water and groundwater sources tested contained enteric viruses. Even 9% of conventionally treated drinking water (coagulation, sedimentation, filtration, postfiltration disinfection using chlorine/ozone) tested positive for enteric viruses. Although water-transmitted human pathogens include various bacteria, protozoa, helminths, and viruses, agents of major threat to human health are pathogenic protozoa (Cryptosporidium and Giardia) and enteroviruses. Despite ample information regarding the fate of viruses in the subsurface, research on the persistency of pathogenic protozoa through passage in soil and groundwater is just now emerging. In the past, it was generally believed that pathogenic protozoa are confined to surface water. Contrary to that expectation, recent monitoring results from 463 groundwater samples collected at 199 sites in 23 of the 48 contiguous states suggested that up to 50% of the groundwater sites were positive for Cryptosporidium, Giardia, or both, depending on the parasite and the type of groundwater source (vertical wells, springs, infiltration galleries, and horizontal wells) (3). Viruses are small obligate intracellular parasites that infect and sometimes cause a variety of diseases in animals, plants, bacteria, fungi, and algae. Viruses are colloidal particles, negatively charged at high pH (pH >7), ranging in size from 20 to 350 nm. The smallest unit of a mature virus is composed of a core of nucleic acid (RNA or DNA) surrounded by a protein coat. Due to this unique feature of viral structure and colloidal physicochemical properties, the transport of viruses in soil and ground water can act with a combination of characteristics ranging from those of solutes, colloids, and microorganisms. Enteroviruses are a particularly endemic class of waterborne microorganisms that cause a number of ubiquitous illnesses, including diarrhea, gastroenteritis, and meningitis, to name only a few. Included in this group

VIRUS TRANSPORT IN THE SUBSURFACE

are poliovirus, hepatitis type A (HAV), Coxsackie virus A and B, and rotavirus. Although gastroenteritis is the most common disease resulting from these microorganisms, other associated illnesses include hepatitis, typhoid fever, mycobacteriosis, pneumonia, and dermatitis (10). SOURCES OF VIRUSES A number of avenues are available for the introduction of viruses to the subsurface, including land disposal of untreated and treated wastewater, land spreading of sludge, septic tanks and sewer lines, and landfill leachates. Among these, septic systems may pose a significant chemical as well as biological threat to surface and groundwaters. One trillion gallons of septic-tank waste are released into the subsurface annually. Although phosphate and bacteria are ordinarily removed by soil, nitrate and viruses may escape these processes and move through the soil into the groundwater. The presence of viral particles is even more significant in the light of studies that indicate they are not necessarily inactivated in septic tanks and may move into the groundwater where they may survive for long periods of time. It is a general consensus that the transport of pathogens in the subsurface depends on the extent of their retention on soil particles and their survival. Among the major factors that affect viral transport characteristics in the subsurface are temperature, microbial activity, moisture content, and pH. Among all the factors, temperature appears to be the only well-defined parameter that causes a predictable effect on viral survival. A direct relationship between a rise in temperature and viral inactivation rates (K = log inactivated/h) among various viruses has been suggested. Badawy et al. (11) stated that during the winter (4–10 ◦ C), viral inactivation rates for coliphage, poliovirus, and rotavirus were 0.17, 0.06, and 0.10 per hour, respectively. Whereas, during the summer (36–41 ◦ C), the inactivation rates for MS-2, poliovirus, and rotavirus were 0.45, 0.37, and 0.20 per hour, respectively. It should be pointed out that this information is based on ambient air. A more direct comparison would be correlation with temperatures in the subsurface. In this regard, the inactivation rates for enteroviruses are 0.06 (10–15 ◦ C), 0.08 (15–20 ◦ C), and 0.19 (20–25 ◦ C). This worker also indicated that viruses may remain viable for 3 to 5 weeks on crops irrigated with sewage effluent, polio and Coxsackie virus up to 4 months on vegetables during commercial and household storage; and up to 30 days on vegetables stored at 4 ◦ C. Microbial ecology may also play an important role in the inactivation of waterborne viruses. For example, microbial activity could affect viral survival by the action of proteolytic enzymes of some bacteria and protozoa in destroying the viral capsid protein. As discussed earlier, viral transport through porous media is controlled by sorption and by inactivation. However, adsorption of viruses to soil should not be confused with their inactivation because adsorption is not permanent and can be reversed by the ionic characteristics of percolating water. Viruses can remain infective after a travel distance of 67 meters vertically and 408

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meters horizontally (12). The various forces involved in attaching viruses to soil particles include hydrogen bonding, electrostatic attraction and repulsion, van der Wals forces, and covalent ionic interaction. EFFECT OF HYDROGEOLOGIC SETTINGS ON VIRAL MOVEMENT The concentration and loading of viruses and the hydrogeologic setting through which they move will control the potential for viral migration to wells to a much greater extent than biological survivability. A hydrogeologic setting often consists of a soil underlain by unconsolidated deposits of sand, silt, and clay mixtures over rock. The setting further incorporates unsaturated and saturated zones. All other factors being equal, the persistence of viruses at a well or other source of water is most likely where saturated flow transports large concentrations of the particles along short flow paths through media that contribute little to attenuation. Although the interrelated processes that control viral movement and persistence in the subsurface are not completely understood, some of the major hydrogeological factors that can be used to evaluate the potential for viral presence in groundwater wells include • transport mechanisms (unsaturated versus saturated flow conditions), • type of media through which the virus will travel (clays versus sands versus fractured media), • length of the flow path to the extraction point (well), and • time of travel. Hydrogeologic settings that have shallow water tables are more susceptible to viral transport. Viruses are attenuated or immobilized by processes such as dessication, microbial activity, and stagnation. Further, viruses commonly bind to soil particles, fine-grained materials, and organic matter. The lower transport velocities associated with unsaturated conditions (e.g., move, stop, move cycle) allow these processes more time to occur. If viruses are introduced directly into the water table (such as from leaching tile fields associated with on-site sewage disposal) or if the volume of contaminants can maintain saturated flow conditions (such as in some artificial recharge situations), the potential for contamination increases. Where the viral concentration is high, the probability of contaminant migration increases regardless of the hydrogeologic setting. Therefore, in hydrogeologic settings that have deeper water tables and where contaminants are not introduced into the aquifer through saturated flow conditions, viruses are much less likely to survive transport to a well. Hydrogeologic settings that have interconnected fractures or large interconnected void spaces that lack finegrained materials have a greater potential for viral transport and well contamination. Karst aquifers, fractured bedrock, and gravel aquifers have been identified in the proposed GWDR as sensitive hydrogeologic settings (9).

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In these settings, fractures and large void spaces allow rapid transport through the aquifer, thereby reducing the amount of time and particulate contact available for attenuation. Potential interaction with rock walls along fractures is reduced, and contact with fine-grained materials for potential sorption sites is minimal. Similar to fractured rock aquifers, gravel aquifers that have only a small fine-grained fraction have little potential for viral sorption. However, as the amount of fine-grained material increases, effective grain size decreases, the potential for sorption increases, and travel times decrease. Finer grained aquifers and aquifers where void spaces are less interconnected or smaller are therefore less likely to transport viruses significant distances. The potential for physical viral removal by filtration also appears to increase as grain size becomes smaller, although the filtration processes are not well understood due to their size. However, filtration of bacteria, which are larger than viruses, it has been shown is an effective removal mechanism. Hydrogeologic settings, where fractures are not as interconnected or where more tortuous flow paths must be followed to reach a well, also allow greater viral removal. For example, in many rock aquifers, groundwater flow follows bedding planes that may result in an elongated indirect pathway to a well. In other rock aquifers, flow must travel around and through cemented portions of the matrix thereby increasing the flow path. Similarly, sand and gravel aquifers that have fine-grained materials in the matrix will have less direct flow paths as the water flows around the finer grained materials. Generally, it can be stated that tortuosity increases the length of the flow path and decreases the hydraulic conductivity, thus decreasing viral survival. Where finer grained materials are present or fractures are less interconnected, flow paths are also longer, thereby offering some protection to wells in more permeable units. Hydrogeologic settings where time of travel is short have a greater potential for viral contamination. Where less permeable units (called aquitards) restrict or reduce vertical flow to underlying aquifers, time of travel is increased. Although inactivation rates, it has been shown, are extremely variable, time is a major factor affecting virus viability. Due to the importance of hydrogeologic settings, the Proposed Ground Water Rule thoroughly addresses this issue to identify wells that are sensitive to fecal contamination. A component of the Proposed Ground Water Rule requires states to perform hydrogeologic assessments for the systems that distribute groundwater that are not disinfected (source waters that are not treated to provide 99.99% removal or inactivation of viruses). The states are required to identify sensitive hydrogeologic settings and to monitor for indicators of fecal contamination from sensitive hydrogeologic settings (see Ref. 9 for the complete proposed strategy). VIRUS TRANSPORT MODELING One method of addressing regulations for viral exposure, such as groundwater disinfection, the application of liquid

and solid waste to the land, and wellhead protection zones, is using predictive viral transport models. Like most predictive modeling efforts, the results depend on the conceptual basis of the model as well as the quality and availability of input data. Clearly, a thorough understanding of the processes and parameters of viral transport are essential elements in their application. Some of the more important subsurface viral transport factors include, soil water content and temperature, sorption and desorption, pH, salt content, organic content of the soil and groundwater matrix, virus type and activity, and hydraulic stresses. Berger (14) indicated that the inactivation rate of viruses is probably the single most important parameter governing viral fate and transport in ground water. Some of the existing models require only a few of these parameters which limit their use to screening level activities, whereas others require input information which is rarely available at field scale and is usually applied in a research setting. One limitation of most models is that they have been developed for use in the saturated zone. It has been shown, however, that the potential for viral removal is greater in the unsaturated zone than in groundwater. Setback Distances Traditionally, state and county regulators have established fixed setback distances for all geologic settings in their jurisdictions. For example, the distance between a septic tank and a private well would, in many instances, be as little as 50 feet and would apply for tight clays as well as fractured rock. It would apply to areas where the water table is near the surface as well as at considerable depth. As discussed in this document, the travel time or transport distance of viral particles depends on a number of factors, including moisture content, geologic setting, type and depth of the soil overburden, and source loading, to name only a few. Frequently, guidelines established as minimum distances became so standard that a well was often positioned precisely 50 feet from the septic tank. In the survey conducted as part of the proposed Ground-Water Treatment Rule, setback distances were quite variable (9). Some of the distances were presumably based on scientific principles, while others were holdovers from past practices. One approach in determining setback distances for septic tanks in wellhead protection areas and bank filtration sites is to determine travel times using groundwater flow characteristics. This approach has been implemented in the Federal Republic of Germany, for example, where three concentric zones protect each drinking-water well. The zone immediately surrounding the well is faced with the most restrictive regulations which are founded on the belief that 50-day residence time is adequate for inactivation of any pathogen in contaminated water. However, a comprehensive study by Matthess et al. (15), involving the evaluation of the ‘‘50-day zone,’’ concluded that the reduction of viruses by 7 log units (current regulations) requires a much longer residence time. Matthess et al. indicated that a reduction of 7 log units occurred in about 270 days (Haltern and

WINDMILLS

Segeberger Forest) in one study and about 160–170 days (Dornach) would be required, according to another study. Another approach to this important issue is to consider the vulnerability to viral transport in the subsurface of portions of a state or county or of individual aquifers. Although there are a number of approaches to rank vulnerability, DRASTIC is one assessment methodology that uses hydrogeologic setting descriptions and a numerical ranking system to evaluate groundwater pollution potential (13). DRASTIC assumes that a potential contaminant will be introduced at the ground surface, have the mobility of water, and be flushed toward the aquifer by infiltration. Using existing information on variable scales, the methodology was designed to evaluate areas of 100 acres or larger. DRASTIC is an acronym representing seven reasonably available factors that are used to develop a numerical score. They are Depth to water, net Recharge, Aquifer media, Soil, Topography (slope), Impact of the vadose zone media, and hydraulic Conductivity of the aquifer. DRASTIC uses a weighting system to create a relative pollution potential index that varies between 65 and 223; the higher numbers express greater vulnerability. Although DRASTIC was not designed specifically to evaluate the movement of viruses in the subsurface, the major transport mechanisms and flow paths for viral transport are considered, and the flexibility of the systems’ rating scheme allows many of these factors to be taken into account. For example, depth to water addresses saturated versus unsaturated flow conditions and their importance. Aquifer media, soil, and impact of the vadose zone media all are based on descriptive soil and rock terms that allow variation due to fracturing, grain size, attenuation mechanisms, and overall characteristics that affect flow. Topography addresses the tendency of viruses to be introduced into the subsurface or to be carried away by runoff. Hydraulic conductivity addresses the relative ease of a contaminant to move with the velocity of water through the aquifer. Clearly, meaningful setback distances can be developed only by using scientific principles that allow the use of available knowledge. The establishment of setback distances from sources of viral contamination to points of extraction (wells) can be established using DRASTIC if both the hydrogeologic setting and sensitivity rankings are considered. For example, high pollution potential index signal the need for greater setback distances. However, the hydrogeologic factors that control viral movement must be evaluated within this context to establish reasonable numbers for setback distances. A matrix that incorporates the important DRASTIC factors can be used to establish setback distances that include the vulnerability concept. Setback distances must incorporate the knowledge of saturated flow, transport pathway length, transport velocities, media interaction, and potential attenuation mechanisms. These setback distances can be used on a regional scale but can be modified if site-specific information is available. The beauty of DRASTIC is that its rationale and sensitivity factors are easily displayed, so that it can be readily modified.

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BIBLIOGRAPHY 1. CDC. (1991). Waterborne-disease outbreaks, 1989–1990 MMWR 40(SS-3): 1–21. 2. Craun, G.F. (1989). Causes of waterborne outbreaks in the United States. Water Sci Technol. 24: 17–20. 3. Hancock, C.M., Rose, J.B., and Callahan, M. (1998). Crypto and Giardia in US groundwater. J. Am. Water Works Assoc. 90(3): 58–61. 4. Macler, B.A. (1996). Developing the ground water disinfection rule. J. Am. Water Works Assoc. 3: 47–55. 5. U.S. EPA. (June 29, 1989). Drinking water; national primary drinking water regulations; filtration, disinfection; turbidity; Giardia lamblia, viruses, Legionella, and heterotrophic bacteria: final rule. Fed. Reg. 54: 27486. 6. U.S. EPA. (June 29, 1989). Drinking water; national primary drinking water regulations; total coliforms (including fecal coliforms and E. coli): final rule. Fed. Reg. 54: 27544. 7. U.S. EPA. (January 14, 2002). National primary drinking water regulations: long term 1 enhanced surface water treatment rule; final rule. Fed. Reg. 67(9): 1844. 8. U.S. EPA. (1992). Draft Ground-Water Disinfection Rule. U.S. EPA Office of Ground Water and Drinking Water, Washington, DC, EPA 811/P-92-001. 9. U.S. EPA. (May 10, 2000). National primary drinking water regulations: ground water rule; proposed rules. Fed. Reg. 65(91): 30202. 10. Bull, R.J., Gerba, C.P., and Trussell, R.R. (1990). Evaluation of the health risks associated with disinfection. CRC Crit. Rev. Environ. Contr. 20: 77–113. 11. Badawy, A.S., Rose, J.B., and Gerba, C.P. (1990). Comparative survival of enteric viruses and coliphage on sewage irrigated grass. J Environ Sci Health. A25(8): 937–952. 12. Keswick, B.H. and Gerba, C.P. (1980). Viruses in groundwater. Environ. Sci. Technol. 14(11): 1290–1297. 13. Aller, L., Bennett, T., Lehr, J.H., Petty, R.J., and Hackett, G. (1987). DRASTIC: A Standardized System for Evaluating Ground-Water Pollution Potential Using Hydrogeologic Settings. EPA/600/2-87/035. 14. Berger, P. (1994). Regulation related to groundwater contamination: The draft groundwater disinfection rule. In: Groundwater Contamination and Control. U. Zoller (Ed.). Marcel Dekker Inc., New York, NY. 15. Matthess, G., Pekdeger, A., Zoller, U., and Schroeter, J. (1988). Persistence and transport of bacteria and viruses in groundwater—a conceptual evaluation. J. Contam. Hydrol. 2: 171–188.

WINDMILLS ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

Windmills are machines that convert the force of the wind into energy that is applicable to various tasks, like pumping water. Windmills that are used to pump water use the energy generated by these mills to turn the gears that propel the pump. These types of windmills, sometimes referred to as wind pumps, have been used for centuries and continue in use around the world.

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A BRIEF HISTORY The first documented use of windmills was in Persia around 500–900 A.D., however, it is widely accepted that they were invented in China more than 2000 years earlier (1). In these early windmills, vertical-axis systems were used for grinding grain and pumping water. In the vertical-axis system, the wind would have to hit the mill from a specific angle to get the desired effect; therefore, most of the area around the mill had to be shielded. Later, horizontal-axis systems were developed that proved to have greater structural efficiency. By tilting the blade to a certain degree, the system eliminated the dependence on the direction of the wind and harnessed the wind energy lost in the vertical-axis systems by the areas shielded. Horizontal systems were later used throughout Europe. Windmill technology, it is believed was introduced in Europe in the eleventh century (2). Aiming to their lowlands, the Dutch set out to develop a more efficient windmill and in the process became the driving force behind wind-machine development. Although many innovations were made, one of the fundamental improvements was designing sails that allowed optimum aerodynamic lift. This improvement made the sails rotate faster, greatly increasing their efficiency and speed in completing the task at hand. Wind energy was applied to irrigation, pumping from local wells, and drainage pumping; this in turn made areas habitable and liberated workers from these labor-intensive jobs. As Europeans sought to expand and colonize, they brought the windmill technology with them. The colonization of the Americas is a prime example. Without this technology, it would have been impossible for immigrants to settle in areas that lacked a constant water supply. Certain areas in Texas, for example, lacked the needed water

Figure 1. Multi-sailed windmill.

supply to sustain life and allow cultivating of the land. The drive to expand into these areas helped to stimulate the need to refine the windmill to solve this problem. Daniel Halladay addressed this problem in 1854 by changing the European windmill so it could operate unattended and more efficiently (3). The changes that he made appealed to many companies and small communities without water systems by providing an inexpensive way to get the water they needed (Fig. 1). By the early twentieth century, windmills were being mass-produced, and millions of them were being used around the world. However, in the 1930s, as other fuel resources such as oil were demanded, wind pumping systems were not as desirable because they were more expensive and not as reliable as these other fuels. This was true until the 1970s when a shortage of oil prompted communities to revisit the idea of wind energy. Wind pumps are currently being used for crop irrigation, drinking water supply for communities, and even individual household water supply. Between 5,000 to 10,000 wind pumps are being installed worldwide each year, and the market for these is expected to increase as wind technology advances and becomes less expensive (4). THE WIND PUMP: HOW IT WORKS The wind pump has four main features the wind turbine, the tower, the actual pumping equipment, and the storage basin. These parts can be found in almost all wind pumps. They vary in design depending on the wind conditions in which they will be used. If the wrong design is used in strong wind conditions, the pump may move too fast and malfunction, and if the wind speed is low, it may not be able to function at all. The horizontal-axis is the typical system used for wind turbines to pump water. The rotor does not have to follow any specific design but instead should be designed for the wind conditions in which it will be used. However, the pump tends to have more force when there are more blades. The next main part of the wind turbine is its transmission. The transmission of a wind turbine converts each rotation of the rotor into an up and down motion, driving the pump rod in and out of the well. The tail of the wind turbine is a piece that was added during the westward expansion in the United States. It allows the wind turbine to work without constant supervision by changing the direction of the rotor to keep it facing the wind. A tower holds the wind turbine generally between 10 m and 15 m high (5). It holds up the wind turbine and stabilize its connection with the pumping parts of the machine. The tower has either a square or triangular base, and the pump rod that moves in and out of the well is positioned in the center of the tower. The well that the pump rod enters can either be a shallow hand-dug well or a deep-drilled well. If the latter is the case, the walls of the well should be lined with a water permeable material to prevent them from caving in. A pipe called the rising main lies in the center of the pump rod. Its function is to carry the water pumped up to the surface. The actual pump is at the bottom of the pump rod submerged in the water and is attached to the rising main. The pump fills with water

WINDMILLS

during the downward motion of the pump rod and pushes the water into the rising main and up to the surface as the pump rod moves up. Finally, all wind pumps should have a storage basin that can hold the excess water that is pumped. When the water moves up into the rising main, it is then redirected into the storage tank. This stored water is also essential to water pumping systems because wind energy cannot always be relied on, and having a surplus at hand is useful. THE ADVANTAGES OF WIND PUMPS There are many advantages to using wind energy for pumping water. First, wind pumps are environmentally friendly. As with the atmosphere, wind turbines will not contaminate the land, and in the case of water pumps, there is no chance the water will be contaminated as a result of a malfunctioning wind pump. Wind turbines generally do not affect the wildlife that inhabits the area. Sheep, cattle, deer, and other wildlife are not bothered by the turbines, and in fact have been known to graze under them. The only argument that has been raised about this issue is the tendency for birds to collide with them. However, several studies suggest that the impact of the wind turbines on birds does not compare to that of other things, such as electrical lines and buildings. Wind pumps are cost-efficient. After the initial cost of installation, the owner basically has an infinite source of energy for just the cost of maintenance. Whereas using a fossil fuel to work the pump would make the owner subject to its cost. THE DISADVANTAGES OF WIND PUMPS There are, however, a few setbacks in using wind energy for pumping water. Because the wind is not reliable, if water is needed on a day where there is little to no wind, obviously the wind pump will not be able to pump water. This problem can be solved by a storage basin, as explained earlier. Many people who inhabit areas around wind turbines complain about the noise that is created by the rotation of the rotor, but this generally applies only to many wind turbines grouped together. As the rotors are

75

improved upon and become more aerodynamic, the noise level will decrease greatly. BIBLIOGRAPHY 1. Illustrated History of Wind Power Development. (7 June 2001). http://relosnet.com/wind/early.html. 2. World Meteorological Organization. (1981). Meteorological Aspects of the Utilization of Wind As An Energy Source. No. 575, p. 44. 3. The Handbook of Texas Online. (6 June 2001). http://www.tsha. utexas.edu/handbook/online/articles/view/WW/aow1.html. 4. Wind Energy & Atmospheric Physics. Review of Historical and Modern Utilization of Wind Power. http://www.risoe.dk/veawind/history.htm. 5. Jenkins, N. et al. (1997). Wind Energy Technology. John Wiley & Sons, New York, pp. 116–117.

READING LIST Andersen, P.D. (9 July 2001). Review of historical and modern utilization of wind power. Wind Energy and Atmospheric Physics. http://www.risoe.dk/vea-wind/history.htm. British Wind Energy Association. (28 June 2001). Wind Energy in Agriculture. http://www.britishwindenergy.co.uk/you/agric. html. Danish Wind Turbine Manufacturers Association. (28 June 2001). 21 Frequently Asked Questions About Wind Energy. http://www. windpower.dk/faqs.htm. Gipe, P. (1995). Wind Energy Comes of Age. John Wiley & Sons, New York. Illustrated History of Wind Power Development. (7 June 2001). http://relosnet.com/wind/early.html. Jenkins, N. and Walker, J.F. (1997). Wind Energy Technology. John Wiley & Sons, New York. Larkin, D. (2000). Mill: The History and Future of Naturally Powered Buildings. Universe, New York. Look Learn & Do. A History of Windmills. (7 June 2001). http://looklearnanddo.com/documents/history winmills.html. Windmills. The Handbook of Texas Online. (6 June 2001). http://www.tsha.utexas.edu/handbook/online/articles/view/ WW/aow1.html. World Meteorological Organization. (1981). Meteorological Aspects of the Utilization of Wind as an Energy Source, No. 575.

MUNICIPAL WATER SUPPLY MIXING AND AGITATION IN WATER TREATMENT SYSTEMS

flow. Inside the bulk of the vessel, on the other hand, the action of the impeller induces flow circulation, which follows a pattern typical of the impeller type. Thus, we distinguish mainly radial and axial impellers, depending on the direction of the flow that emerges from the impellerswept region. Radial impellers eject a liquid stream radially. In a typical stirred vessel where the impeller is mounted on a shaft, is vertical and is usually centrally located, the ejected liquid flows from the edge of the impeller blades toward the vessel walls. There, it separates into two streams; one flows in the upper part of the vessel and one in the lower part of the vessel, thus forming two flow loops. The liquid from these two streams circulates in the upper and the lower parts of the vessel and eventually is drawn back into the agitator-swept region; two primary circulation loops are established inside the stirred tank. Figure 1 presents some typical radial impellers. The Rushton turbine (RT; Fig. 1a) is one of the most widely used impellers due to its efficiency in gas–liquid and liquid–liquid mixing. Its construction is simple; usually it has six flat blades mounted on a flat disk. Figures 1b (SCABA 6SRGT turbine or Chemineer CD-6) and 1c (Chemineer BT-6) present two variants of the Rushton turbine, where blades have a parabolic shape, which is even more efficient than the RT, especially in dispersing gas inside a stirred vessel. Finally, the Narcissus (NS) impeller (2) produces an inverse radial flow; liquid is drawn in from its side and pumped out from its upper and lower parts. Figure 2, which is a 2-D plot of composite radial and axial velocities URZ —which are obtained from the vector sum of the radial (UR ) and axial (UZ ) components of the local velocity vector—illustrates the typical radial flow patterns of the Rushton turbine—radial flow directed from the impeller toward the vessel walls—and of the Narcissus (NS)—radial flow directed toward the impeller. A similar double-loop circulatory flow pattern is induced by the SCABA turbine (3). Axial impellers draw liquid mainly from one of their sides, top or bottom, and eject it from the opposite site; when liquid is ejected toward the bottom of the vessel, the impeller is said to work in the ‘‘down-pumping’’ mode, whereas when the liquid is ejected toward the surface of the liquid, this corresponds to the ‘‘up-pumping’’ mode. Often, liquid is also drawn from the side of the rotating impeller. Note that axial impellers are sometimes called ‘‘mixed-flow’’ impellers, too: in some cases, part of the ejected flow is directed sideways; this becomes more pronounced when the viscosity of the liquid increases (4). Figure 3 presents some typical axial-flow impellers. The marine propeller (Fig. 3a) has been used for the propulsion of boats, and nowadays it is its sole application; far more efficient agitators have been designed for mixing liquids. Figure 3b shows the widely used pitched-blade turbine (PBT); the number of blades and their inclination usually characterizes the PBT more specifically, for example, the PBT in the illustration is referred to as a ‘‘4-45-PBT.’’ The Mixel TT has blades, which are wider than those of

PAUL MAVROS Aristotle University Thessaloniki, Greece

Mixing is one of the primary processes involved in water and wastewater treatment. In this article, the state of the art in mixing and impeller design is presented. INTRODUCTION Stirring is provided in a wide variety of processes to blend constituents or to disperse one phase into another or several other phases. In a blend-type operation, the purpose is to obtain a homogeneous mixture, whereas in the dispersion process, the goals vary widely, depending also on the nature of the phases involved: • in gas–liquid dispersions, gas is dispersed into fine bubbles which must be distributed as evenly as possible in the vessel to take part in a subsidiary process, for example, absorption and/or reaction with a dissolved component, as in water and wastewater treatment, or flotation of hydrophobic particles, among others; • in solid–liquid distributions, it is necessary to provide the appropriate conditions for entraining all solid particles inside the bulk of the liquid, either from the bottom of the vessel or from the free surface of the liquid; • in liquid–liquid dispersions, fine droplets of one of the liquids have to be dispersed inside the other liquid to produce an emulsion or for a polymerization, among others. For each of these processes, a particular type of agitator is appropriate. These have evolved from the simple paddles used during the past centuries; modern flow visualization techniques (1) have helped in designing agitator blade shapes optimized for specific processes. In the following sections, the main types of agitators are presented, according to the processes for which they are intended; this presentation is limited to turbulent flow, which is typical in water and wastewater treatment processes, and does not describe agitators designed for viscous liquids (anchors, gates, etc.). TYPES OF IMPELLERS An impeller is a pump; by its rotation, it draws liquid from its neighborhood and then ejects it at a relatively high speed. It is typically mounted on a shaft connected to a motor, and the shaft–impeller structure is inserted in the stirred tank either axisymmetrically or sideways. Close to the impeller blades, the rotation induces a tangential 76

MIXING AND AGITATION IN WATER TREATMENT SYSTEMS (a)

(b)

(c)

(d)

77

Figure 1. Radial agitators: (a) Rushton turbine; (b) SCABA 6SRGT (or Chemineer CD-6); (c) Chemineer BT-6; (d) Narcissus.

(a)

(b)

Figure 2. Flow patterns induced by radial impellers in a stirred tank: (a) Rushton turbine (4); (b) Narcissus (NS) impeller (2).

the PBT and are profiled to be more efficient in energy consumption. The typical 2-D flow pattern induced by all axialflow impellers in their usual configuration—‘‘downpumping’’—is illustrated in (Fig. 4a). As already stated, liquid is drawn from the upper part and the side of the

impeller and is ejected downward. A single circulation loop is established in all cases; liquid flows upward close to the vessel walls and returns toward the impeller. The velocities in the upper part of the vessel are typically rather slow: the URZ vectors are much shorter than those close to the upper and lower sides of the impeller.

78

MIXING AND AGITATION IN WATER TREATMENT SYSTEMS (a)

(a)

(b)

(c)

(b)

Figure 3. Typical axial-flow impellers; (a) marine propeller; (b) pitched-blade turbine; (c) Mixel TT.

Therefore, the liquid in the stirred tank may be divided into two regions: the first corresponds to the primary circulation loop, which is established around the impeller; liquid flows fast and results in an intensive mixing process. The liquid in the second region, located mainly in the upper part of the vessel, circulates slowly; therefore the mixing process is less intense and effective there; it is often necessary to add a second impeller on the same shaft, to enhance circulation and mixing in the upper part of the vessel. The inverse configuration—‘‘up-pumping’’—again yields a single primary circulation loop, located around the impeller (Fig. 4b). A smaller, secondary circulation loop is established in the upper part of the vessel, achieving better overall circulation and mixing than the ‘‘downpumping’’ mode. Several other impellers have been tested and/or marketed, based on extensive hydrodynamic performance measurements, taking into consideration some optimization criterion; some of them are variants of the pitched-blade turbine; others have blade shapes originating from hydrofoils. Figure 5 illustrates some of these impellers. The size and location of the impeller inside the stirred vessel are dictated by the process needs and affect its performance, for example,

Figure 4. Flow patterns induced by axial impellers in a stirred tank: (a) Mixel TT in down-pumping mode [4]; (b) Mixel TT in up-pumping mode (5).

• to disperse gas effectively inside a stirred tank, it is necessary to use a radial agitator that has a large impeller diameter (D) to tank diameter (T) ratio, for example, D/T = 1/2, and to provide high rotational speed; • if it is required to provide surface aeration to the stirred tank, the impeller is located close to the free liquid surface; • if it is necessary to achieve an effective distribution of solid particles, an axial impeller having a reduced size (D/T = 1/3) should be used, located closer to the bottom of the vessel, having clearance (C), that is,

(a) (e)

67° 26°

(b) (f)

(g)

(c)

(h)

(d)

Figure 5. Examples of advanced impellers: (a) ‘‘Medek’’ PBT [6,7]; (b) Chemineer HE-3 [8]; (c) Ekato MIG; (d) Lightnin A-310; (e) DeDietrich hydrofoil; (f) Lightnin A-320; (g) Prochem Maxflo; (h) APV B2. 79

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MIXING AND AGITATION IN WATER TREATMENT SYSTEMS

the distance from the bottom of the vessel, close to T/3 or even T/4. When mixing is applied to rectangular troughs, the axis of the impeller is horizontally located at one end of the trough, and an axial-flow agitator with hydrofoil blades is used to induce longitudinal motion of the liquid in the trough. PERFORMANCE DATA The performance of the various impellers is characterized by quantitative criteria; some of these are power consumption, the amount of flow circulation caused by the pumping action of the impeller, the ability of the impeller to cause intense circulation in the stirred tank, and the time necessary to achieve homogeneity of the tank contents, among others. The power consumption depends upon the impeller type; it has been found that in turbulent conditions, where the dimensionless Reynolds number (Re), Re =

the flow number, (Fl), which also characterizes impellers may be obtained: QP (3) Fl = ND3 Table 1 presents typical values of flow numbers for the most common types of impellers. One of the purposes of an impeller is to create circulation inside a stirred vessel, so one quantitative characteristic of its efficiency is the spatial mean velocity achieved in the vessel. This mean velocity, compared to the velocity at the tip of the blades (VTIP ), yields the ‘‘agitation efficiency’’ (IG ) of each particular impeller (9). Finally, the time to obtain vessel homogeneity is termed ‘‘mixing time’’ (tMIX ); it has been found that for a wide variety of impellers it may be correlated to the power number and to the impeller-to-vessel diameter ratio (10): N tMIX = 5.3(Po)−1/3




T D

2

CONCLUSIONS

ρND2 µ

(1)

is larger than about 4000, the dimensionless power number (Po), P (2) Po = ρN 3 D5 is approximately constant and characterizes each impeller. Table 1 presents power numbers for a variety of commonly used impellers. Another feature of impellers is the amount of fluid being ‘‘pumped out’’ of the agitator-swept region; from the flow rate of this stream (Qp ), another dimensionless number,

Mixing is used in a multitude of processes, including water and wastewater treatment, to achieve several goals: to disperse another phase—gas, liquid, or solid—into the bulk of the liquid; to homogenize the stirred tank contents; and to assist and promote a reaction between some of the dissolved and/or dispersed species, among others. This is usually achieved by using rotating impellers, whose blade design has been often optimized for particular processes. Radial impellers, such as the Rushton turbine, are more suitable for homogenization and for dispersing a second phase in liquids; however, they generate high-shear flows. Axial-flow impellers are more suitable for solids dispersion and for cases where shear-sensitive material exists in the liquid, requiring benign mixing conditions.

Table 1. Characteristics of Various Impellers

Type

D/T

C/T

Power Number (Po)

A310 (Lightnin) A315 (Lightnin) A320 (Lightnin) A410 (Lightnin) 4-45-PBT 6-45-PBT (down-pumping) 6-45-PBT (up-pumping) 6SRGT (SCABA) BT-6 (Chemineer) CD-6 (Chemineer) HE-3 (Chemineer) Marine propeller Medek PBT Mixel TT (down pumping) Mixel TT (up-pumping) Narcissus MaxFloT (Prochem) Rushton turbine

1/2 NAa 0.40 0.40 1/3 1/3 1/3 1/3 1/3 1/3 1/2 1/3 1/3 1/2 1/2 1/3 0.35 1/3

1/3 NAa 0.39 0.40 1/3 1/3 1/3 1/3 1/3 1/3 1/4 NAa 1/3 1/3 1/3 1/3 0.45 1/3

0.56 0.75–0.80 0.64 0.32 1.25 1.93 2.58 2.8–3.0 2.1 2.8–3.0 0.31 0.89 0.41 0.74 0.67 1.14 1.58 4.9–5.2

a

NA: not available.

(4)

Flow Number (Fl) 0.62 0.73 0.64 0.62 0.77 0.75 0.68 NAa NAa NAa 0.41 0.79 0.60 0.67 0.61 0.31 0.82 0.78

NOTATION C:

clearance of impeller (from midplane) to vessel bottom (m) D: impeller diameter (m) Fl: dimensionless flow number (-) IG : dimensionless agitation index (-) N: impeller rotational frequency (Hz) Po: dimensionless power number (-) Re: dimensionless Reynolds number (-) T: vessel diameter (m) tMIX : mixing time (s) U: liquid velocity (m/s) VTIP : liquid velocity at the tip of the impeller blades (= π ND) (m/s) GREEK LETTERS µ: viscosity of liquid (Pa.s) ρ: density of liquid (kg/m3 )

ARSENIC IN NATURAL WATERS

INDEXES R: radial RZ: composite radial-axial Z: axial BIBLIOGRAPHY 1. Mavros, P. (2001). Flow visualisation in stirred vessels. Review of experimental techniques. Chem. Eng. Res. Des. 79A(2): 113–127. 2. Hristov, H.V., Mann, R., Lossev, V., and Vlaev, S.D. (2004). A simplified CFD for three-dimensional analysis of fluid mixing, mass transfer and bioreaction in a fermenter equipped with triple novel geometry impellers. Food and Bioproducts Processing 82(C1): 21–34. 3. Khopkar, A., et al. (2004). Flow generated by radial flow impellers: PIV measurements and CFD simulations. Int. J. Chem. Reactor Eng. 2: A18. 4. Mavros, P., Xuereb, C., and Bertrand, J. (1996). Determination of 3-D flow fields in agitated vessels by laser-Doppler velocimetry. Effect of impeller type and liquid viscosity on liquid flow patterns. Chem. Eng. Res. Design 74A: 658–668. 5. Aubin, J., Mavros, P., Fletcher, D.F., Bertrand, J., and Xuereb, C. (2001). Effect of axial agitator configuration (uppumping, down-pumping, reverse rotation) on flow patterns generated in stirred vessels. Chem. Eng. Res. Des. 79A(8): 845–856. 6. Medek, J. and Foˇrt, I. (1991). Relation between designs of impeller with inclined blades and their energetic efficiency. Proc. 7th Eur. Conf. Mixing, Brugge, Belgium, Sept. 18–20, KVIV, vol. I, pp. 95–102. 7. Medek, J., Seichter, P., Rybn´ıˇcek, Z., and Chaloupka, Z. (1987). Czechoslovak patent no. 256630. 8. Ibrahim, S. and Nienow, A.W. (1995). Power curves and flow patterns for a range of impellers in Newtonian fluids −40 < Re < 5 × 105 . Chem. Eng. Res. Design 73(A5): 485–491. 9. Mavros, P. and Baudou, C. (1997). Quantification of the performance of agitators in stirred vessels. Definition and use of an agitation index. Chem. Eng. Res. Des. 75A: 737–745. 10. Ruszkowski, S. (1994). A rational method for measuring blending performance and comparison of different impeller types. Proc. 8th Eur. Conf. Mixing. Cambridge, UK, Sept. 21–23, IChemE, Rugby, pp. 283–291.

ARSENIC IN NATURAL WATERS ROBERT Y. NING King Lee Technologies San Diego, California

arsenicals have widened use in commerce, but so have the recognition that their presence in drinking water, largely from natural sources, is a major public health problem around the world. Acute and chronic arsenic exposure via drinking water has been reported in many countries, especially Argentina, Bangladesh, India, Mexico, Mongolia, Thailand, and Taiwan, where ground (well) water is contaminated with a high concentration of arsenic of 100 to more than 2000 µg/liter (ppb) (2). Studies have linked long-term exposure to arsenic in drinking water to cancer of the bladder, lungs, skin, kidney, nasal passages, liver, and prostate. Noncancer effects of ingesting arsenic include cardiovascular, pulmonary, immunological, neurological, and endocrine (e.g., diabetes) disorders (3). Besides its tumorigenic potential, it has been shown that arsenic is genotoxic (4,5). Given the importance of arsenic as a global environmental toxicant, we will summarize the geochemistry, natural distribution, regulation, anthropogenic sources, and removal mechanisms. GEOCHEMISTRY Average concentrations of arsenic in the earth’s crust reportedly range from 1.5 to 5 mg/kg. Higher concentrations are found in some igneous and sedimentary rocks, particularly in iron and manganese ores. Common minerals containing arsenic are shown in Table 1. Arsenopyrite, realgar, and orpiment are the most important of these minerals, and they are commonly present in the sulfide ores of other metals, including copper, lead, silver, and gold. Arsenic may be released from these ores to the soil, surface water, groundwater, and the atmosphere. Natural concentrations of arsenic in soil typically range from 0.1 to 40 mg/kg; an average concentration is 5 to 6 mg/kg. Arsenic can be released to ground or surface water by erosion, dissolution, and weathering. Geothermal waters can be sources of arsenic in groundwater. In Yellowstone National Park, the arsenic concentrations in geysers and hot springs range from 900 to 3,560 ppb. Waters from these sources cause elevated arsenic levels in rivers downstream. Other natural sources include volcanism and forest fires. Volcanic activity appears to be the largest natural source of arsenic emissions to the atmosphere, estimated variously between 2,800 to 44,000 metric tons annually. The relative contributions of volcanic sources, other natural sources (Table 1), and anthropogenic sources to the atmosphere have not been definitively established. The predominant forms of arsenic in groundwater and surface water are arsenate (V) and arsenite (III). Examples

INTRODUCTION Arsenic is widely distributed in nature in air, water, and soil as a metalloid and as chemical compounds, both inorganic and organic (1). This class of compounds was known to the ancient Greeks and Romans both as therapeutic agents as well as poisons. This dual nature as useful substances as well as toxic matter to be controlled has grown over the centuries. Arsenic and

81

Table 1. Common Minerals of Arsenica Arsenopyrite, FeAsS Lollingite, FeAs2 Orpiment, As2 S3 Realgar, As4 S4 Chloanthite, NiAs2 Niciolite, NiAs a

Reference 3.

Smalite, CoAs2 Cobaltite, CoAsS Gersdorffite, NiAsS Tennantite, 4Cu2 SAs2 S3 Proustite, 3Ag2 SAs2 S3 Enargite, 3Cu2 SAs2 S5

82

ARSENIC IN NATURAL WATERS

of inorganic arsenic compounds found in the environment include oxides (As2 O3 , As2 O5 ), and sulfides (As2 S3 , AsS, HAsS2 , HAsS3 3− ). Inorganic arsenic species that are stable in oxygenated waters include arsenic acid [As(V)] species (H3 AsO4 , H2 AsO4 − , HAsO4 2− and AsO4 3− ). Arsenous acid [As(III)] is also stable as H3 AsO3 and H2 AsO3 − under slightly reducing aqueous conditions. Arsenite is generally associated with anaerobic conditions. Oxidation state, oxidation–reduction potential, pH, iron concentrations, metal sulfide and sulfide concentrations, temperature, salinity, and distribution and composition of biota appear to be the significant factors that determine the fate and transport of arsenic. In surface waters, additional factors include total suspended sediment, seasonal water flow volumes and rates, and time of day. Sorption of arsenic to suspended sediment may strongly affect the fate and transport of arsenic in surface water systems (6). Where pH and arsenic concentrations are relatively high and total suspended sediment levels are relatively low, sorption processes may be less important. However, where suspended sediment loads are higher, arsenic concentrations are lower, and pH levels are lower, arsenic is more likely to be present in the suspended particulate phase rather than in the dissolved phase. Particulate phase arsenic may settle to bottom sediment in reservoirs and areas of low flow levels. In deeper lakes, remobilization of arsenic from sediment may be minimal, whereas in shallower lakes, arsenic may be remobilized faster from wind-induced wave action and high-flow scouring. Diurnal changes of as much as 21% in arsenic concentrations have been observed in rivers, attributable to pH changes due to sunlight and photosynthesis.

NATURAL DISTRIBUTION A survey of arsenic concentration in natural waters is of importance relative to the desirable maximum limit of 10 ppb or less for human consumption. An attempt has been made to quantify the global element cycle for arsenic, based on published data (1). Arsenic concentrations in environmental media are presented in Table 2. In addition to geochemical factors, microbial agents can influence the oxidation state of arsenic in water and can mediate the methylation of inorganic arsenic to form organic arsenic compounds (8). Microorganisms can oxidize arsenite to arsonate and reduce arsenate to arsenite or even to arsine (AsH3 ). Bacteria and fungi can reduce arsenate to volatile methylarsines. Marine algae transform arsenate into nonvolatile methylated arsenic compounds such as methylarsonic acid [CH3 AsO(OH)2 ] and dimethylarsinic acid [(CH3 )2 AsO(OH)] in seawater. Freshwater and marine algae and aquatic plants synthesize complex lipid-soluble arsenic compounds (9). Organic arsenical compounds were reportedly detected in surface water more often than in groundwater. Surface water samples reportedly contain low but detectable concentrations of arsenic species, including methylarsonic acid and dimethylarsinic acid. Methylarsenicals reportedly comprise as much as 59% of total arsenic in lake water. In some lakes, dimethylarsinic acid has been reported as

Table 2. Arsenic Concentrations in Environmental Mediaa Environmental Media Air Rain from unpolluted ocean air Rain from terrestrial air Rivers Lakes Ground (well) water Seawater Soil Stream/river sediment Lake sediment Igneous rock Metamorphic rock Sedimentary rock Biota—green algae Biota—brown algae a

Arsenic Concentration Range 1.5–53 0.019 0.46 0.20–264 0.38–1,000 1, 000 0.15–6.0 0.1–1,000 5.0–4,000 2.0–300 0.3–113 0.0–143 0.1–490 0.5–5.0 30

Units ng/m3 µg/L (ppb) µg/L µg/L µg/L µg/L µg/L mg/kg (ppm) mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg

Reference 7.

the dominant species, and concentrations appear to vary seasonally from biological activity within waters. REGULATIONS In the United States, the Safe Drinking Water Act (SDWA) of 1974 called for establishing Maximum Contaminant Levels (MCL) as national drinking water standards and required the Environmental Protection Agency (EPA) to revise the standard periodically. Based on a Public Health Service standard established in 1942, the EPA established a standard of 50 µg/liter (50 ppb) as the maximum arsenic level in drinking water in 1975. In 1984, the World Health Organization (WHO) followed with the same 50 ppb recommendation. Since that time, rapidly accumulated toxicity information prompted a revision of the standard, and a provisional guideline of 10 ppb was recommended by WHO in 1993. In January 2001, EPA published a revised standard that would require public water supplies to reduce arsenic to 10 ppb by 2006. Perceived hardships in implementation and uncertainty in setting the standard at 3, 5, 10, or 20 ppb has led the EPA to announce temporary delays in the effective date for the January 2001 rule to allow for further cost–benefit analysis and public input. The rule is significant because it is the second drinking water regulation for which the EPA has used its discretionary authority under the SDWA to set the MCL higher than the technically feasible level, which is 3 ppb for arsenic, based on the determination that the costs would not justify the benefits at this level. ANTHROPOGENIC SOURCES Arsenic is released from a variety of anthropogenic sources, including metal and alloy manufacturing, petroleum refining, pharmaceutical manufacturing, pesticide manufacturing and application, chemical manufacturing, burning of fossil fuels, and waste incineration.

EVALUATION OF MICROBIAL COMPONENTS OF BIOFOULING

Most agricultural uses of arsenic are banned in the United States. However, sodium salts of methylarsonic acid are used in cotton fields as herbicides. Organic arsenic is also a constituent of feed additives for poultry and swine and appears to concentrate in the resultant animal wastes. About 90% of the arsenic used in the United States is for the production of chromated copper arsenate (CCA), the wood preservative. CCA is used to pressure treat lumber and is classified as a restricted use pesticide by the EPA. A significant industrial use of arsenic is in the production of lead-acid batteries; small amounts of very pure arsenic metal are used to produce gallium arsenide, which is a semiconductor used in computers and other electronic applications. The U.S. Toxics Release Inventory data indicated that 7,947,012 pounds of arsenic and arseniccontaining compounds were released to the environment in 1997; most of that came from metal smelting. The data did not include some potentially significant arsonic sources associated with herbicides, fertilizers, other mining facilities, and electric utilities. REMOVAL MECHANISMS At the regulated maximum arsenic level of 10 ppb, the U.S. EPA estimated that 5% of all U.S. community water systems would have to take corrective action to lower the current levels of arsenic in their drinking water. In high arsenic areas of the world, the need for removal from water supplies is even more acute. Due to their predominance in natural waters, arsenic(V) acid (H3 AsO4 ) and arsenous(III) acid (HAsO2 ) and their salts can serve as the model for these and alkylated species for consideration of removal mechanisms. The pK values of arsenic acid = 2.26, 6.76, 11.29 (10) and arsenious acid = 9.29 (10) or 8.85 (11) are of prime importance in determining the degree of ionization at the pHs of the water from which removal strategies are considered. It is readily apparent that at a natural pH of 7 to 8, arsenic acid is extensively ionized as the divalent ion; arsenious acid remains largely un-ionized. Due to the ionic charge, arsenate(V) is more easily removed from source waters than arsenite(III). In particular, activated alumina, ion exchange, and reverse osmosis may achieve relatively high arsenate removal rates, but they show lower treatment efficiencies for arsenite. Elevating the pH such as by caustic injection into reverse osmosis system feedwater would be one approach to greater removal of arsenite(III) compounds. Arsenite can also be oxidized to arsenate to improve removal efficiencies. In water that contains no ammonia or total organic carbon, chlorine rapidly (in less than 5 seconds at chlorine concentrations of 1.0 mg/L) oxidizes approximately 95% of arsenite to arsenate. Monochloramine at a concentration of 1.0 mg/L oxidized 45% of arsenite to arsenate. Potassium permanganate performs this oxidation rapidly; oxygen does so slowly unless activated by light and sensitizer. In contrast to other heavy metals, As3+ and As5+ are not precipitated as hydroxides, only as sulfides. Alkyl and arylarsonic acids are precipitated by quadrivalent metals such as tin, thorium, titanium, and zirconium.

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The tendencies of dissolved arsenic species to adsorb on inorganic particle surfaces such as iron, ferric hydroxide, iron oxide, alumina, sulfur, and sulfides allow for removal strategies involving fixed-bed reactors or adsorption/coagulation/filtration schemes. Sulfate, fluoride, and phosphate ions are known to be strong competitors of arsenic adsorption in some systems. BIBLIOGRAPHY 1. Matschullat, J. (2000). Arsenic in the geosphere—a review. Sci. Total Environ. 249(1–3): 297–312. 2. Tchounwou, P.B., Wilson, B., and Ishaque, A. (1999). Rev. Environ. Health 14(4): 211–229. 3. Various reports available from the US-Environmental Protection Agency website: www.epa.gov/safewater/arsenic.html and the World Health Organization website: www.who.int/ water sanitation health/water quality/arsenic.htm. 4. Basu, A., Mahata, J., Gupta S., and Giri, A.K. (2001). Mutat. Res. 488(2): 171–194. 5. Gebel, T.W. (2001). Int. J. Hyg. Environ. Health 203(3): 249–262. 6. Nimick, D.A., Moore, J.N., Dalby, C.E., and Savka, M.W. (1998). Wat. Res. Research 34(11): 3051–3067. 7. US-Environmental Protection Agency. (2000). Arsenic Occurrence in Public Drinking Water Supplies, EPA-815-R-00-023, December 2000, available from EPA Website (Ref. 3). 8. Tamaki, S. and Frankenberger, W.T., Jr. (1992). Rev. Environ. Contam. Toxicol. 124: 79–110. 9. Shibata, Y., Morita, M., and Fuwa, K. (1992). Adv. Biophys. 28: 31–80. 10. Lide, D.R. (Ed.). (1995). CRC Handbook of Chemistry and Physics, 76th Edn. CRC Press, New York, pp. 8–43. 11. Meites, L. (Ed.). (1963). Handbook of Analytical Chemistry. McGraw-Hill, New York, pp. 1–29.

EVALUATION OF MICROBIAL COMPONENTS OF BIOFOULING STUART A. SMITH Smith-Comeskey Ground Water Science LLC Upper Sandusky, Ohio

Detecting the occurrence of biofouling and assessing its impact involve a range of analytical techniques, including informed observation and inspection of well components, interpretation of hydraulic and water quality testing, and direct analysis of microbial components of biofouling. The latter is the subject of this article. SYMPTOMS: INDIRECT ANALYSIS Symptoms can be used as qualitative indicators of biofouling but not specifically of the microbial component. Observable symptoms include the following: • clogging (both formation/well and pump/discharge systems).

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EVALUATION OF MICROBIAL COMPONENTS OF BIOFOULING

• corrosion • alteration of water quality in pumped samples • distinct coating on surfaces visible during inspections, such as borehole television surveys. Such symptoms are indications of biofouling and should trigger analysis to determine the nature of the biofouling. DIRECT ANALYSIS OF MICROBIAL BIOFOULING AND COMPONENTS Standard Methods Some existing ‘‘standard methods’’ for analyzing aspects of biofouling are described in: • ASTM Test Method for Iron Bacteria in Water and Water-Formed Deposits (D 932) (1) • Section 9240—Iron and Sulfur Bacteria, Standard Methods for the Examination of Water and Wastewater (APHA-AWWA-WEF, currently 20th ed. Supplement) (2). • Additional Standard Methods (Part 9000) microbiological methods (and others accepted by technical and regulatory bodies) for analysis of heterotrophic bacteria and specific groups of interest such as the total coliform group. All relevant aquatic and public health microbiological methods are applicable in identifying microbiological components of biofouling. • Microscopic particulate analysis (MPA) ((2), consensus method) can identify biofilm components as part of its larger scope and represents a systematic approach. Microscopic Examination and Analysis. The presence of filamentous (e.g., Leptothrix, Thiothrix, or Crenothrix) or stalked (Gallionella) iron or sulfur bacterial forms is accepted as a positive indicator of biofouling. Examination by light microscopy has traditionally been the method of choice for confirming and identifying these ‘‘iron bacteria’’ or ‘‘sulfur bacteria’’ (1,2). However, the absence of such visible structures does not necessarily mean the absence of biofouling: 1. Samples may not include enough recognizable materials to provide the basis for a diagnosis of biofouling. 2. Samples examined may not include the filamentous or stalked bacteria normally searched for in such analyses. It is generally understood that the morphologically distinct types are only part of the biofouling present. 3. Analysts vary in their skill and opinions in interpreting what they see. In addition, the existing standard microscopic tests (1,2) are specified as qualitative. Attempts have been made to quantify degrees of biofouling by microscopy. However, spatial variations and the pulsating, three-dimensional nature of active biofilms make them unreliable (4). An

available semiquantitative method using microscopy is described by Barbic et al. (5). Culturing Methods for Detection. Culturing enriches biofouling microflora that cannot be identified microscopically and helps provide more complete information on the nature of biofilm samples. It can also be used to draw reasonable conclusions about biofouling in the absence of microscopy. Standard Methods (2) describes a range of media formulations, and more for other purposes are found in other sections of Part 9000. With regard to Section 9240 formulations, a number of limitations have been identified. These weaknesses have all been recognized by the Standard Methods Section 9240 joint technical group: • No reported effort has been made to standardize these media with reference cultures from well water. Thus the efficiency of recovery of ironprecipitating bacteria from groundwater samples remains unknown at present (4) but is probably very low. • Except for the modified Wolfe’s medium for Gallionella enrichment (6), the environmental conditions for growth in these media do not seem to match well with groundwater environmental conditions (7,8). • Useful Mn-precipitation media have been one area of weakness in common practice (7), although they are being refined (9). • The available sulfur oxidizer media are still nonisolating, enrichment media. • Employment of all of these culture methods requires preparation from raw materials (no packaged agar or broth media specific to IRB are available), sterilization, and maintenance in a microbiological laboratory by a skilled person. Thus, they are rarely used in operational practice. Cultural Media Improvements. An important innovation of the last 20 years in biofouling cultural recovery has been the development of prepackaged cultural media that permit (1) practical use in operational monitoring in addition to more academic analyses and (2) recovery of a range of microflora, resulting in a more complete understanding of the microbial ecology of wells and more refined maintenance and management of water quality (see entry GW-1311 Well Maintenance). Two groups using slightly different approaches developed these methods independently:

MAG Method (MAG Laboratorio Ambiental, La Plata, Argentina). MAG tests for heterotrophic iron-related bacteria (IRB) and sulfate-reducing bacteria (SRB) consist of a prepared liquid medium contained in small septum bottles (10,11). The MAG medium for ironrelated heterotrophic bacteria (BPNM-MAG) uses ferric ammonium citrate [like W-R and R2A + FAC (7)] and the SRB medium (BRS-MAG) uses Postgate C medium under a reducing atmosphere, supplemented with iron filings (11). Inoculation of a single bottle provides a presence-absence (P-A) result. Dilution to extinction provides a semiquantitative (MPN) result (11).

EVALUATION OF MICROBIAL COMPONENTS OF BIOFOULING

BART Method. Currently, the most commonly used cultural approach for routine biofouling monitoring is the Biological Activity Reaction Test (BARTTM ) Method (Droycon Bioconcepts Inc., Regina, Saskatchewan). • BART tubes contain dehydrated media formulations and a floating barrier device, which is a ball that floats on the hydrated medium of the sample. • These devices and their proposed use are described in detail in Cullimore (12). They can be used as an enrichment method to provide a presence-absence (PA) or semiquantitative (MPN) detection of biofouling factors (7,12–14). BART tubes are available in a variety of media mixtures. The IRB-BARTTM test, for example, is designed to recover microaerophilic heterotrophic Fe- and Mnprecipitating microorganisms and is (like the BPNM-MAG test) derived from the W-R iron bacteria medium (15). This method, which is gaining wide operational acceptance as a means of detecting and characterizing biofouling symptoms, has provided useful qualitative information in well biofouling in various field trials (7) and has proven useful in a range of applications. These methods have proved to be significant advances in making microbial ecology an important factor in routine operational monitoring as they 1. Reliably provide results to the level of detection needed to make operational decisions. 2. Are relatively easy to use and interpret. 3. Offer a means for precise scientific characterization of the microbial system if properly used and interpreted within their limitations. At present, neither of these systems is included in Standard Methods, although they have become operationally de facto standard methods within the water operational and hydrogeologic communities (16). A definitive comparability test among these and various Standard Methods media is yet to be conducted. Hybrid Methods. A subcategory of cultural methods consists of field-usable enrichment procedures to increase the potential for successful detection by microscopy. For example, Alcalde and co-workers (10,17) describe a simple enrichment and staining technique to enhance the numbers and visibility of filamentous bacteria on glass slides. Nonculturing Biofouling Analytical Methods. All cultural enrichment methods recover only a fraction (1% is an often cited value) of microflora in environmental samples. Employing a range of cultural media and using media that better approximate ideal growth conditions (and making them easier to use) helps to overcome these problems; however, many microflora types may remain unrecoverable. Nonculturing analytical methods that directly detect interpretable evidence of microflora bypass these limitations.

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Phospholipid Fatty Acid (PLFA) or Phospholipid Fatty Acid Methyl Ester (PL-FAME) Analysis. The method described by White and Ringelberg (18,19) and its variations provide a nonculturing method of characterizing the microbial components of environmental samples (solids such as sediment cores as well as fluids). In PLFA or PL-FAME, ‘‘signature’’ lipid biomarkers from the cell membranes and walls of microorganisms are extracted from the sample. Particular types of biomarkers are linked to groups of microorganisms. Amounts of biomarkers can be reliably associated with viable biomass, and compound ratios can be linked to community nutritional status. Once extracted from cells with organic solvents, the lipids are concentrated and fractionated using gas chromatography/mass spectrometry (GC/MS). A profile of the fatty acids and other lipids is then used to determine the characteristics of the microbial community. This ‘‘fingerprint’’ represents the living portion of the microbial community because phospholipids degrade rapidly following cell death (20). A variation is TC-FAME (total-cell FAME), in which the lipids of whole cells are extracted and characterized from microbial isolates. These fatty acid profiles are compared to libraries of known isolates to identify unknowns in environmental samples or pure cultures. Adenosine Triphosphate (ATP) and Adenylate Kinase (AK) Detection. ATP is ubiquitous in cellular life and can be detected rapidly using bioluminescence methods. The ATP bioluminescence test (21), detects ATP, which is present only in living cells. The amount of light emitted in the reaction can be correlated with the amount of ATP that can be extracted from a known number of bacteria. The test is rapid ( Q02 . Equations 4 and 5 are as follows: q1 =

Q02 b1 C1 (Q1 0 − Q2 0 )b1 C1 + 1 + b1 C1 1 + b1 C1 + b2 C2

(4)

q2 =

Q02 b2 C2 1 + b1 C1 + b2 C2

(5)

 kj qi = 1+

N j=1

Ci ηi aj

 

Cj ηi

bj

(8)

where ηI is the interaction parameter, which is constant. In practice, this parameter varies from different equilibrium compositions. Therefore, this model is not very successful in explaining the multicomponent systems. Fritz and Schlunder Multicomponent Model

where q1 and q2 are the amounts of solute 1 and 2, respectively, adsorbed per unit weight of activated carbon/adsorbent at equilibrium concentrations of C1 and C2 , respectively. Q01 and Q02 are the maximum value of adsorption for solute 1 and 2, respectively, determined from the respective single-solute systems, whereas b1 and b2 are the constants related to the energy of adsorption from solute 1 and 2, respectively, in their pure solution systems. This model also has some limitations. It is only appropriate to describe the competitive adsorption between the molecules having very different single-solute adsorption capacity. In spite of its limitations, a number of workers have applied this model in their multicomponent adsorption studies. In one of the studies, Huang and Steffens (11) applied this model to determine the competitive adsorption of organic materials by activated carbons. It was concluded that mutual suppression of equilibrium adsorption because of competition between acetic and butyric acids has shown that the observed data are somewhat closer to the values predicted by Jain and Snoeyink’s model than by the original Langmuir equation. However, the actual degree of suppression is greater than the prediction for acetic acid and smaller for butyric acid. Multicomponent Isotherm of Mathews and Weber Mathew (12) proposed this model in 1975. This model is a modified and extended form of the Redlich–Peterson model (13), which can be given by the expressions below, which is basically three parameter, single-solute adsorption isotherm model KCe qe = 1 + a R Cb e

The parameters Kj , aj , and Cj can be determined from single-solute isotherm data. A new constant, ηi , has been introduced in this model, which has to be determined from adsorption data in the mixtures. With the addition of this constant, the equation may be written as

(6)

where K, aR , and b are the Redlich–Peterson constants.

This model was given by Fritz and Schlunder in 1974 (14,15). For modeling multicomponent systems comprising species whose single-solute isotherms follow the Freundlich isotherm, a multicomponent Freundlich equation may be used. The first model of this type, as proposed by Fritz and Schlunder (14), can be expressed by Equations 9–12: q1 = q2 = α1,2 α2,1

K1 C1 n1 +β11 β

C111 + α12 C2 β21 K2 C2 n2 +β22 β

β

C222 + α21 C121 α1,2 = α1,1 α2,1 = α22

(9)

(10) (11) (12)

where q1 and C1 are the concentrations of solute 1 in the solid and liquid phase, respectively; q2 and C2 are the concentrations of the solute 2 in solid and liquid phase, respectively; and K1 , n1 , and K2 , n2 are the Freundlich constants in single-solute 1 and solute 2 systems. Equations 9 and 10 consist of ten adjustable variables; however, K1 and K2 , n1 and n2 can be determined from single-solute isotherms using Freundlich model for singlesolute systems. Dastgheib and Rockstraw Model Very recently Dastgheib and Rockstraw (16) proposed a multicomponent Freundlich equation for binary systems, as given by Equations 13–15:  q1 =  q2 =

K 1 C1 n1 K 2 C2 n2

 K 1 C1 n1 K 1 C1 n1 + α12 K2 C2 n2 + b12 C2 n12  K 2 C2 n2 K 2 C2 n2 + α21 K1 C1 n1 + b21 C1 n21

(13) (14)

COMPETITIVE ADSORPTION OF SEVERAL ORGANICS AND HEAVY METALS ON ACTIVATED CARBON IN WATER

where q1 and C1 are the concentrations of solute 1 in the solid and liquid phase, respectively; q2 and C2 are the concentrations of solute 2 in the solid and liquid phase, respectively; K1 , n1 , K2 , and n2 are the Freundlich constants in single-solute system; and α12 , α21 , b12 , b21 , n12 , and n21 are the interaction constants obtained from a least-squares analysis of the binary data. The term in brackets on the right-hand side represents the overall competition and interaction factor and has a value of less than or equal to unity (where C2 tends to zero, it is equal to 1). The terms α12 K2 and α21 K1 can be condensed to single terms and are considered as constants. For the ith component systems, the general equation model can be written as qi =

K i Ci

ni

(Ki Ci ni )2 N + j=1 (αij Kj Cj nj + bij Cj nij )

(15)

where qi and C1 are the concentrations of solute I in the solid and liquid phase, respectively; Cj is the concentration of other solutes in liquid phase; Ki and ni , Kj and nj are the single component Freundlich constants; αIJ , bij , and nij are the binary interaction constants obtained from a least-squares analysis of the multicomponent data having αii = bii = 0; and N is the number of solute. This model is different from the Fritz and Schlunder model in the sense that this demonstrates equal or stronger performance. Sheindorf et al. Model

109

mixtures (19) using the parameters calculated from singlesolute systems. This model was later modified and applied to calculate various multicomponent adsorption parameters by Radke and Praunitz (20). In the IAST model, the following five basic equations (18–22) are used to predict multicomponent behavior from single-solute adsorption isotherms (21). The total surface loading can be defined by Equation 18: qT =

N 

qi

(18)

i=1

where qI is the single-solute solid phase concentration for component i, which is evaluated at spreading pressure of the mixture, and N is the number of components. The mole fraction on the carbon surface for component i can be calculated by Equation 19: Zi =

qi qT

Ci = Zi C0i

and and

I = 1 to N I = 1 to N

(19) (20)

where CI is the single-solute liquid phase concentration for component i, which is evaluated at the spreading pressure of the mixture. The single-solute liquid phase concentration in equilibrium with q0i is  1 = Zi /q0i qT N

This model was given by Sheindorf et al. (17,18) for the multicomponent systems comprised of species whose single-solute isotherm obeys the Freundlich isotherm. This equation was based on the following assumptions: 1. each component in single system obeys the Freundlich model 2. each component in multicomponent system and the adsorption energies of different sites are disturbed exponentially, with the distribution function being identical to that for single-component systems.

(21)

i=1

πm = RT



q0i

d ln C0i

q0j

d ln C0j

0

 = 0

π10 A 0 dq = 1 RT d ln q01 πj0 A 0 dq = = for · j = 2 · to · N (22) j RT d ln q0j

where A is the surface area of carbon per unit mass of adsorbent, R is the gas constant, T is the absolute temperature, πi is the spreading pressure of the single solute i, and πm is the spreading pressure of the mixture.

The model equations can be given by Equations 16 and 17: LeVan and Vermeulen Model

q1 = K1 C1 (C1 + η12 C2 )n1 −1

(16)

q2 = K2 C2 (C2 + η21 C1 )n2 −1

(17)

where q1 and C1 are the concentrations of solute 1 in the solid and liquid phase, respectively; q2 and C2 are the concentrations of solute 2 in the solid and liquid phase, respectively; K1 , n1 , K2 , and n2 are the Freundlich constants in single-solute system; and η12 and η21 are the interaction constants. Ideal Adsorbed Solution Theory (IAST) The ideal adsorbed solution theory is based on the thermodynamics of adsorption, which is analogous to Roul’s law in a liquid-gas system. The only difference is that it is applied to a solid-liqid system. Initially, this model was used to calculate multicomponent adsorption of gaseous

LeVan and Vermeulen (22) have modified the competitive Langmuir-like model. IAS theory was considered in modifying the model. This model predicts the equilibrium relationships of solute mixture only from the data derived from single adsorption isotherms. It is the simplest isotherms derived from the IAS model. Statistical Design for Competitive Adsorption 24 factorial experimental design was used to study the competitive adsorption of Fe(II), Mn(II), Ca(II), and Zn(II) on selected activated carbons B3, W2, W3, and lignite by Mohan and Chander (23). These designs are important for the following reasons: 1. They require relatively few runs per factor studied; and although they are unable to explore fully a

110

2.

3.

4.

5.

COMPETITIVE ADSORPTION OF SEVERAL ORGANICS AND HEAVY METALS ON ACTIVATED CARBON IN WATER

COMPETITIVE SORPTION OF INORGANICS ON ACTIVATED CARBON

wide region in the factor space, they can indicate major trends and therefore determine a promising direction for further experimentation. When a more thorough local exploration is needed, they can be suitably augmented to form composite designs. These fractional designs are often of great value at an early stage of an investigation, when it is frequently good practice to use the preliminary experimental efforts to look at a large number of factors superficially rather than a small number. These designs and the corresponding fractional designs may be used as building blocks so that the degree of complexity of the family-constructed design can match the sophistication of the problem. The interpretation of the observations produced by the designs can proceed largely by using common sense and elementary arithmetic.

The adsorption of Pb2+ , Cu2+ , Zn2+ , and Cd2+ from aqueous solutions by three activated carbons in single and multicomponent systems were studied by Budinova et al. (24). These three activated carbons were obtained from apricot stones (A), coconut shells (C), and lignite coal (L). The results of the individual metal ions from an aqueous solution containing all four metal ions together in equal concentration are presented in Table 1. It is clear from Table 1 that the presence of foreign ions diminishes the adsorption of each of the ions. The effect is greatest for the lead ions and smallest for the copper ions. The authors did not mention any mechanism for the multicomponent sorption of these ions. The only reason given was that, apart from the properties of the cations, the chemical nature of the metal ions is of great importance for the adsorption process. They also concluded that a selective adsorption of the metals is observed; the ones preferentially adsorbed do not completely prevent the adsorption of other ions. Johns et al. (25) reported the sorption of cadmium, copper, lead, nickel, and zinc in single and multicomponent systems on various granular activated carbon developed from agricultural waste materials. A study on the competitive effect of metal ions was carried out from a solution having 2.5 mM of each metal at pH 5.0 and was also unbuffered to reduce solution species complexation. The uptake of various metals from a mixed solution is presented in Table 2. Bansode et al. (26) evaluated the adsorption effectiveness of pecan shell-based granular activated carbons (GACs) in removing metal ions Cu(II), Pb(II), and Zn(II) commonly found in municipal and industrial wastewater. Pecan shells were activated by phosphoric acid, steam, or carbon dioxide activation methods. Metal ion adsorption

The authors have used a total run of 24 = 16. The variables were the concentration levels of various metal ions, with high level (+) 100 ppm and low level (−) 0 ppm. The experiments were conducted at pH 3.5. The experiments were arranged as the design matrix. General Factorial Design To perform a general factorial design, a fixed number of ‘‘levels’’ or (versions) for each of the variables (factors) can be selected, and then experiment is run with all possible combinations. If l1 levels exist for the first variable, l2 for the second, . . ., and lk for the kth, the complete arrangement of l1 × l2 × l3 × . . . lk experimental runs is called an l1 × l2 × l3 × . . . lk factorial design, e.g., a 2 × 3 × 5 factorial design requires 2 × 3 × 5 = 30 runs and a 2 × 2 × 2 = 23 factorial design.

Table 1 Adsorption from Solution Containing All the Four Ions (mol g−1 )

Decrease of Ion Adsorption in the Presence of the ion (%)

Carbon

Cu2+

Pb2+

Zn2+

Cd2+

Cu2+

Pb2+

Zn2+

Cd2+

A C L OA

434.4 430.7 403.5 398.5

355.4 354.0 328.7 1550.4

410.0 390.0 390.0 73.6

385.0 360.0 360.3 47.6

11.6 11.6 15.4 12.1

28.3 28.1 32.1 37.3

11.0 17.2 17.0 27.1

19.9 19.4 21.0 32.2

Table 2

GAC Calgon GAC Norit RO3515 Norit vapure Soybean hulls Peanut shells Sugar cane bagasse Rice straw

BET Surface Area (m2 g−1 )

Ni(II)

Cu(II)

Zn(II)

Cd(II)

Pb(II)

Total

783 827 876 479 275 162 460

0 0 0 14 9 7 2

97 117 98 127 195 132 144

0 0 0 29 31 21 24

30 11 4 36 39 29 32

113 67 66 190 236 206 174

240 195 168 396 510 395 376

µ Moles of Metals Adsorbed per Gram of GAC

COMPETITIVE ADSORPTION OF SEVERAL ORGANICS AND HEAVY METALS ON ACTIVATED CARBON IN WATER

of shell-based GACs was compared with the metal ion adsorption of a commercial carbon, namely Calgon’s Filtrasorb 200. Adsorption experiments were conducted using solutions containing all three metal ions in order to investigate the competitive effects of the metal ions as would occur in contaminated wastewater. The results obtained from this study showed that acid-activated pecan shell carbon adsorbed more lead ion and zinc ion than any of the other carbons, especially at carbon doses of 0.2–1.0%. However, steam-activated pecan shell carbon adsorbed more copper ion than the other carbons, particularly using carbon doses above 0.2%. In general, Filtrasorb 200 and carbon dioxide-activated pecan shell carbons were poor metal ion adsorbents. The results indicate that acidand steam-activated pecan shell-based GACs are effective metal ion adsorbents and can potentially replace typical coal-based GACs in treatment of metal contaminated wastewater. The surface complex formation model was used successfully to describe the surface change density, as well as the single and multispecies metal adsorption equilibrium by Chen and Lin (27). Choi and Kim (28) studied the adsorption characteristics of zinc and cadmium ion on granular activated carbon in singular and binary systems. Features of binary adsorption were discussed for several influential parameters, and experimental observations for both ions were correlated with a predicted adsorption isotherm based on a Langmuir multicomponent model. Yu and Kaewsarn (29) used the multicomponent model on the sorption of heavy metals on low-cost adsorbents. An equilibrium isotherm was predicted by the extended Langmuir model using Langmuir parameters as determined from a single component system. Copper ions were found to have adsorption affinity, and the separation factor αCu/Cd was determined as 3.05. Trujullo et al. (30) reported the competitive adsorption of six metal ions from a single solution, which led to a model applicable to their batch and semicontinuous packed beds. Binding capacity was highest for copper, independent of the other ions, and copper also exerted the largest competing effect. Bunzle et al. (31) carried out the studies in the pH range of 3.5–4.5, and the order was found to be Pb2+ > Cu2+ > Cd2+ ≈ Zn2+ > Ca2+ , whereas Masslenilov and Kiselva (32) reported the adsorption capacity order as Cu2+ > Zn2+ > Fe3+ > Ca2 . Ho et al. (33) reported that competitive effect affected the sorption of three metals in the order Ni2+ > Pb2+ > Cu2+ . It was concluded that all the metals are not necessarily adsorbed by exactly similar mechanisms for all the biosorbents, and that each needs to be tested to determine its characteristics. Adsorption of lead is usually greater than of copper, although copper is a more aggressive competitor and the adsorption of nickel is usually weaker than that of others. The influence of a range of commercially available, water-soluble surfactants on the uptake of heavy metal ions (Cu, Zn, Cd, and Pb) by three types of clay (kaolinite, illite, and a montmorillonite) was reported by Beveridge and Fickering (34). The adsorption of Cu, Pb, Cd, and Zn was significantly reduced in the presence of small amounts of cationic surfactants, particularly with montmorillonite

111

suspensions. The addition of anionic surfactants led to increase metal loss from the solution. Studying multicomponent adsorption system equilibria must commence with an accurate description of each component in its single (or pure) component equilibrium state. Allen and Brown (35) studied the single component and multicomponent metal sorption onto lignite. A comparison was made between the single component saturation uptake and multicomponent uptakes The multicomponent systems were equimolar binary solutions solutions Cu-Cd, Cu-Zn, and Cd-Zn and a ternary mixture of equimolar Cu-Cd-Zn. In single component systems, the adsorption capacity followed the order Cu > Zn > Cd. These capacities were reported on a molar basis. Despite the competition, the total sorption capacity was found to increase even though the adsorption capacity of a single ion may be less than if it were to present alone. It has been pointed out that a substantial effect of multicomponent mixtures was observed on the capacity of lignite for cadmium and zinc. There appears to be slight increases in capacity in binary Cu-Cd, Cd-Zn, and ternary mixtures and a decrease in the capacity of Cu-Zn mixture compared with single component data. The order of sorption of metals in multicomponent systems is as follows: Cu > Cd > Zn. The preference of the sorbents for metals uptake is related to the electronegativity of the ions. Copper possessing the greatest ionic potential has the strongest attraction to the adsorbent, followed by cadmium then closely by zinc. The sorption capacities for single as well as multicomponent systems as reported by Allen and Brown are presented in Table 3. Tan et al. (36) reported the uptake of metal ions in single and multicomponent systems by chemically treated human hairs. Various suppressors and promoters were identified and given in Table 4. Beveridge and Pickering (34) reported the effect of various water-soluble surfactants on the uptake of Cu, Zn, Cd, and Pb ions by three types of clays, kaolinite, illite, and montmorillonite, over the pH range 3–10. Adsorption of Cu(II), Pb(II), Cd(II), and Zn(II) was significantly reduced in the presence of small amounts of cationic surfactants, particularly with montmorillonite suspensions. No multicomponent sorption modeling was reported. The effect of Cu(II), Hg(II), and Pb(II) on the uptake of Cd(II) by activated carbon was investigated by Krishnan

Table 3 Metal Systems Copper alone Cu in Cu-Cd Cu in Cu-Zn Cu in Cu-Cd-Zn Cadmium alone Cadmium in Cu-Cd Cadmium in Cd-Zn Cadmium in Cu-Cd-Zn Zinc alone Zinc in Cu-Zn Zinc in Cd-Zn Zinc in Cu-Cd-Zn

µmol/gram 440 350 370 360 360 90 250 85 375 50 130 70

COMPETITIVE ADSORPTION OF SEVERAL ORGANICS AND HEAVY METALS ON ACTIVATED CARBON IN WATER Table 4

Hg(II) Ag(I) Pb(II) Cd(II) Cu(II) Cr(VI) Ni(II)

Promoters

Suppressor

Ag(I)>Pb(II) Cd(II)>Cu(II) Cd(II)>Cu(II) − Ni(II)∼Cd(II) Cu(II) −

Cu(II) Hg(II) Hg(II)>Cu(II) Cu(II)>Ag(I)>Ni(II) Ag(I)>Hg(II)>Cr(VI)>Pb(II) − Ch(II)>Cd(II)

and Anirudhan (37). The removal of Cd(II) was reported to be 98.8% in absence of any co-ions. The same decreases to 83.3%, 79.1%, and 72.1%, respectively, when Cu(II), Hg(II), and Pb(II) ions are present in a 1:1 ratio. The results further showed that a 72.2%, 70.5%, and 60.6% reduction in Cd(II) removal was observed when Cu(II), Hg(II), Pb(II) ions were present at a molar ratio of 1:2. The reduction may be because of the competitive ion effect between Cd(II) and co-ions for the adsorption sites available on the carbon surface. Based on these results, it was concluded that Pb(II) ions may be stronger competitive ions than Hg(II) and Cu(II) removal by SA-S-C. The results can also be explained by the selectivity sequence of the most common cations on the adsorbent surface. It was also observed that, among the cations used, interference of Pb(II) ion is highest, followed by Hg(II) and Cu(II). The observed order of interference was the same as that of their increasing ionic radii, i.e., their decreasing hydrated ionic radii. The smaller the hydrated ionic radii, the greater its efficiency to active groups of the adsorbent, which suggests that the energy required in the dehydration of the metal ions, in order that they could occupy a site in the adsorbent, plays an important role in determining the selectivity series for the metal ions. Recently, Mohan and Singh (38) reported the batch sorption isotherm studies to obtain the data required in the design and operation of column reactors for treatment of cadmium- and zinc-bearing wastewater both in single and multicomponent systems. The metals chosen for the investigation in single component studies were Cd (II) and Zn(II). In multicomponent system investigations, four binary systems, Cd(Cd-Cu), Cd(Cd-Zn), Zn(Zn-Cu), and Zn(Zn-Cd), and two ternary systems, Cd(Cd-Cu-Zn) and Zn(Zn-Cu-Cd), were selected. The adsorption isotherms for binary, ternary, and multicomponent systems were obtained at pH 4.5. A 1:1 ratio was used to determine the effect of other metal ions on the adsorption of Cd(II) and Zn(II) on the prepared carbon. The Freundlich and Langmuir adsorption isotherms for Cd(II) and Zn(II) in binary and ternary systems are presented in (Figs. 1 and 2), respectively (38). The results clearly revealed that the presence of other metal ions compete with Cd(II) and Zn(II) ions. It was observed that Cu(II) had the least interfering capacity among Cd(II), Zn(II), and Cu(II) ions in binary systems. Both Langmuir and Freundlich isotherms adequately described the data over the entire range of concentration, and corresponding parameters are presented in Table 5. The effect of ionic interaction (36,39) on the sorption process may also be represented by the ratio of the sorption capacity for one

Amount adsorbed (mg/gram)

Metal Ions

(a) 50 Cd only Cd(Cd-Cu) Cd(Cd-Zn) Cd(Cd-Cu-Zn)

40

30

20

10

0

0

200 400 600 800 Equilibrium concentration (mg/l)

1000

(b) 50 Amount adsorbed (mg/gram)

112

Cd only Cd(Cd-Cu) Cd(Cd-Zn) Cd(Cd-Cu-Zn)

40

30

20

10

0

0

200 400 600 800 Equilibrium concentration (mg/l)

1000

Figure 1. Multicomponent adsorption of Cd(II) on activated carbon developed from bagasse. Solid lines represent the fitting of data by (a) Freundlich and (b) Langmuir isotherms.

metal ion in the presence of the other metal ions, Qmix , to the sorption capacity for the same metal when it is present alone in the solution, Q0 , such that when: Qmix > 1, the sorption is promoted by the Q0 presence of other metal ions Qmix = 1, no observable net interaction exists Q0 Qmix < 1, sorption is suppressed by the Q0 presence of other metal ions Qmix were found to be less than 1, as Q0 presented in Table 5. The prepared activated carbon followed the same trend, that is, Qmix decreased in the following order for the adsorption of Cd(II) and Zn(II) in multicomponent systems:

The values of

Cd(II) < Cd-Cu < Cd-Zn < Cd-Cu-Zn

for Cd(II)

Zn(II) < Zn-Cu < Zn-Cd < Cd-Cu-Zn

for Zn(II)

COMPETITIVE ADSORPTION OF SEVERAL ORGANICS AND HEAVY METALS ON ACTIVATED CARBON IN WATER

113

Table 5. Freundlich and Langmuir Isotherm Constants for Singe and Multicomponent Adsorption of Cd(II) and Zn(II) on Activated Carbon Developed from Bagasse Carbon Freundlich Constants

Metal Ions Cd Cd Cd Cd Zn Zn Zn Zn

System

KF

1/n

R2

KFmix /KF

Q0

b × 10−3

R2

Qmix /Q0

Cd alone Cd + Cu Cd + Zn Cd + Cu + Zn Zn alone Zn + Cu Zn + Cd Zn + Cu + Cd

5.78 4.30 1.74 0.59 5.62 3.96 1.42 0.79

0.28 0.29 0.02 0.03 0.25 0.27 0.41 0.45

0.9760 0.9706 0.9864 0.9864 0.9659 0.9674 0.9868 0.9792

– 0.74 0.30 0.10 – 0.70 0.25 0.14

38.03 33.11 30.02 29.77 31.11 26.00 23.09 19.02

13.2 11.0 8.8 2.7 14.2 13.5 8.00 5.4

0.8886 0.8829 0.9678 0.9583 0.8683 0.8723 0.9794 0.9804

– 0.87 0.79 0.78 – 0.84 0.74 0.61

Amount adsorption (mg/gram)

(a) 35 Zn only Zn(Zn-Cu) Zn(Zn-Cd) Zn(Zn-Cu-Cd)

30 25 20 15 10 5 0

0

200

400

600

800

1000

Equilibrium concentration (mg/l)

Amount adsorbed (mg/gram)

(b) 35 Zn only Zn(Zn-Cu) Zn(Zn-Cd) Zn(Zn-Cu-Cd)

30 25 20 15 10 5 0

0

Langmuir Constants

200

400

600

800

1000

Equilibrium concentration (mg/l)

Figure 2. Multicomponent adsorption of Zn(II) on activated carbon developed from bagasse. Solid lines represent the fitting of data by (a) Freundlich and (b) Langmuir isotherms.

Overall, it was concluded that the adsorption capacity of activated carbon for Cd(II) and Zn(II) decreased more in ternary systems as compared with binary systems. In acid mine wastewater, some other metal ions are always present besides iron and manganese; therefore, it is desirable to see the effect of other metal ions on the adsorption capacity of different activated carbons. A very important study in this regard was carried out by Mohan and Chander (39) where the adsorption of four metal ions, i.e., Mn(II), Fe(II), Zn(II), and Ca(II), were conducted in

binary, ternary, and multicomponent systems on different types of activated carbons. The adsorption isotherms for binary, ternary, and multicomponent systems were obtained at pH 3.5 and 25 ◦ C. The concentration range of 5.0 × 10−5 to 9.0 × 10−3 M was investigated, and a 1:1 ratio was used to determine the effect of Mn(II), Ca(II), and Zn(II) on the adsorption of Fe(II) on carbons. The Freundlich and Langmuir adsorption isotherms for Fe(II) in the absence and presence of interfering metal ions were determined. Both Langmuir and Freundlich adsorption isotherms were found to adequately describe the data over the entire range of concentration, and the Langmuir and Freundlich isotherm parameters are presented in Tables 6 and 7, respectively. The detailed analysis of the regression coefficients showed that the data was slightly better fitted by Freundlich adsorption isotherm for multicomponent systems. The adsorption isotherms for different carbons revealed that, when Ca(II), Mn(II), and Zn(II) were present in the system with Fe(II), the interference did not change the adsorption of Fe(II) in the low concentration range, whereas a competitive uptake, with Fe(II) being preferentially adsorbed by carbons, took place at the higher concentrations. Carbon B0, which showed abnormal behavior and adsorption, was found to increase in the presence of other metal ions. It was further concluded that the presence of manganese and/or zinc had limited effect on the capacity of carbons for Fe(II) in comparison with calcium. Thus, overall it was found that Ca(II) had the highest interfering capacity. The adsorption capacity of various activated carbons for Fe(II) in the presence and absence of Mn(II), Zn(II), and Ca(II) are presented in Table 6, whereas the Freundlich constants are presented in Table 7. Thus, when two or more metal ions are present in the solution, they seem to compete for the adsorption sites as the metal ions are adsorbed on the same sides. Qmix The values of are found to be less than 1, Q0 as presented in Table 6, except for carbon B0, thereby confirming the suppression in the adsorption of Fe(II) by the presence of other metal ions. These results are consistent with the adsorption isotherms obtained for Fe(II) in the absence and presence of various metal ions. It is clear from Table 6 that carbons can be divided into two different categories, i.e., wood-based activated

Table 6. Langmuir Isotherm Constants for Multicomponent Metal Ion Adsorption on Different Types of Activated Carbons Activated Carbons W1

W2

W3

B0

B4

B3

C1

Lignite

Parameters 0

Q b × 10−3 R2 mix Q /Q0 Q0 b × 10−3 R2 Qmix /Q0 Q0 b × 10−3 R2 Qmix /Q0 Q0 b × 10−3 R2 Qmix /Q0 Q0 b × 10−3 R2 Qmix /Q0 Q0 b × 10−3 R2 Qmix /Q0 Q0 b × 10−3 R2 Qmix /Q0 Q0 b × 10−3 R2 Qmix /Q0

Fe(II)

Fe(Fe-Ca)

Fe(Fe-Mn)

Fe(Fe-Zn)

Fe(Fe-Mn-Zn)

Fe(Fe-Mn-Zn-Ca)

22.27 38.63 0.9606 – 25.60 70.74 0.9409 – 21.67 53.97 0.9302 – 14.59 2.643 0.8334 – 28.78 27.560 0.8773 – 25.61 2.28 0.9287 – 46.35 36.85 0.92668 – 34.22 28.98 0.8768 –

15.35 42.14 0.8615 0.69 18.16 37.59 0.7529 0.71 14.58 82.39 0.8054 0.67 15.2 21.485 0.8525 1.0418 16.58 14.609 0.8722 0.58 12.05 1.69 0.9809 0.47 21.77 34.77 0.619730 0.47 19.59 32.87 0.8912 0.5725

17.65 38.67 0.7013 0.79 19.73 117.37 0.7474 0.77 16.86 78.06 0.6983 0.78 22.98 33.22 0.6847 1.57505 22.32 35.662 0.8257 0.7241 16.90 3.73 0.9498 0.66 27.40 77.15 0.60214 0.59 20.88 68.67 0.8714 0.6101

16.26 53.86 0.7223 0.73 18.78 90.70 0.7643 0.73 16.10 86.71 0.7206 0.74 22.02 32.558 0.6700 1.509253 20.84 34.002 0.7682 0.7755 15.80 4.46 0.9672 0.61 26.05 31.73 0.5852 0.56 19.96 33.50 0.8986 0.5832

13.79 30.36 0.9609 0.62 17.16 34.22 0.9491 0.67 13.64 34.91 0.9735 0.63 18.78 75.111 0.8929 1.287183 18.21 23.964 0.9549 0.63 16.47 1.80 0.9858 0.64 23.85 23.72 0.6432 0.51 12.39 55.21 0.9662 0.3621

12.36 44.75 0.7581 0.56 13.26 81.73 0.7516 0.52 13.06 21.71 0.6746 0.60 14.44 32.318 0.7157 0.989719 15.3184 18.943 0.7780 0.53 9.74 2.98 0.9728 0.38 18.31 43.12 0.7003 0.40 11.23 53.12 0.9238 0.3282

Table 7. Freundlich Isotherm Constants for Multicomponent Metal Ion Adsorption on Different Types of Activated Carbons Activated carbons W1

W2

W3

B0

B4

B3

C1

Lignite

Parameters

Fe(II)

Fe(Fe-Ca)

Fe(Fe-Mn)

Fe(Fe-Zn)

Fe(Fe-Mn-Zn)

Fe(Fe-Mn-Zn-Ca)

KF 1/n R2 KF 1/n R2 KF 1/n R2 KF 1/n R2 KF 1/n R2 KF 1/n R2 KF 1/n R2 KF 1/n R2

90.70 0.21 0.9299 90.34 0.18 0.9452 80.67 0.19 0.9163 57.16 0.25 0.8257 281.88 0.41 0.9447 1.87 0.87 0.3759 444.84 0.28 0.9325 214.24 0.2455 0.9215

36.90 0.13 0.8715 33.67 0.09 0.8170 29.37 0.10 0.8653 29.04 0.10 0.84426 123.33 0.42 0.9667 41.94 0.14 0.8618 37.36 0.08 0.6869 66.190 0.1732 0.9061

34.78 0.10 0.7857 38.77 0.09 0.8235 28.80 0.08 0.7800 67.88 0.15 0.9054 112.86 0.32 0.9726 70.03 0.16 0.8376 – – – 69.79 0.1618 0.9030

27.00 0.10 0.7957 39.91 0.10 0.8343 25.42 0.07 0.7953 57.86 0.14 0.9787 109.11 0.32 0.9725 53.18 0.13 0.8419 – – – 60.358 0.1580 0.8971

44.30 0.17 0.9173 57.00 0.18 0.8712 51.54 0.19 0.8671 63.04 0.16 0.9938 265.72 0.47 0.9656 101.13 0.24 0.9675 – – – 8.345 0.1367 0.9070

34.12 0.14 0.7240 39.37 0.52 0.7423 27.83 0.12 0.7989 27.60 0.097 0.7268 100.18 0.39 0.9614 44.18 0.16 0.7561 38.1925 0.11 0.7738 9.5887 0.1396 0.9067

114

Normalized concentration (Ce/Co)

carbons follow the same trend (Fe-Mn 2300; Shi = 0.0149Rei 0.88 Sc 1/3 ; Sc = Schmidt number = ν/D; and ν = kinematic viscosity (m2 /s); Li = length of pipe i. The reaction coefficients for other combinations of bulk and wall reactions are given in Ref. 25. If Nsn is the number of source nodes, then Equation 1 with appropriate reaction kinetics results in (Njn − Nsn) number of linear equations. An iterative procedure based on the Gauss–Siedel algorithm can be adopted to solve the equations. The converged solution gives steady-state chlorine concentrations at all nodes for given source chlorine concentrations. DYNAMIC-STATE MODELING The unsteady advection–reaction process for the transport of chlorine in a pipe flowing full is given by the following classic equation: ∂Ci (x, t) ∂Ci (x, t) = −vi − R[Ci (x, t)] ∂t ∂x

changes in a series of discrete parcels of water as they travel through the pipe network. Event-driven simulation updates the state of the system only at times when a change actually occurs, such as when a new parcel of water reaches the end of a pipe and mixes with water from other connecting pipes, and at the output reporting time (30). The Eulerian methods include the finite difference method (FDM) (30) and the discrete volume element method (DVEM) (21), and the Lagrangian methods are the timedriven method (TDM) (19) and the event-driven method (EDM) (22). The Lagrangian methods are more efficient for simulating the chemical transport in a water distribution system (30). The input to the dynamic simulation model essentially consists of the system demands, source chlorine concentrations, and reaction parameters (bulk and wall) of the network. The outputs of this model are the spatially and temporally varying nodal chlorine concentrations. APPLICATION OF THE CHLORINE TRANSPORT MODEL The Brushy plains zone of the South Central Connecticut Regional Water Authority is chosen to illustrate the application and is shown in Fig. 1. The pipe and node data of the network are taken from EPANET [distributed by USEPA, (29)] example networks. The first-order bulk

27

(4)

29

36 28

where Ci (x, t) = chlorine concentration in pipe i (mg/L) as a function of distance x and time t; vi = mean flow velocity in pipe i (m/s); and R[Ci (x, t)] = reaction rate expression. For the first-order bulk and wall chlorine reaction kinetics, it is given by the following equation: kwi kfi R[Ci (x, t)] = kbi Ci (x, t) + Ci (x, t) rhi (kwi + kfi )

31 26 Tank

35

25 23 24 15

17 18

30 34 33 20

32

(5)

13

16

12 19

11

Ninpj

Cncj,t =

Qi Ci (Li , t) + QE CE

i=1 Ninpj



9

;

j = 1, · · · Njn

22

14

kfi can be calculated using expressions as described earlier. Instantaneous and complete mixing of chlorine at the node j and time t is given by the following expression: 

133

7

(6)

Qi + QE

i=1

where Qi = flow in the pipe i (m3 /s); QE = external source flow in to node j (m3 /s); CE = external source chlorine concentration into node j (mg/L); and Njn = total number of nodes in the network. The numerical approaches adopted to solve the aboveformulated problem can be classified spatially as either Eulerian or Lagrangian and temporally as time driven or event driven (30). Eulerian models divide the pipe into a series of fixed, interconnected control volumes and record changes at the boundaries or within these volumes as water flows through them. Lagrangian models track

4

8 5 6

3

10

2

1

Pump station

Figure 1. Brushy plains network.

21

134

MODELING CHLORINE RESIDUALS IN URBAN WATER DISTRIBUTION SYSTEMS

(a)

Node 3

(b) Chlorine concentration (mg/L)

Chlorine concentration (mg/L)

Simulated Observed

1.4 1.2 1 0.8 0.6 0.4 0.2 0

0

5

10

15

20

(c)

25 30 Time (h)

35

40

45

50

1.2 1 0.8 0.6 0.4 0.2 0

5

10

15

20

25

30

35

40

45

50

55

Time (h) (d)

Node 19

Node 34 1.6

Simulated Observed

1.4

Chlorine concentration (mg/L)

Chlorine concentration (mg/L)

Simulated Observed

1.4

0

55

1.6

1.2 1 0.8 0.6 0.4 0.2 0

Node 11 1.6

1.6

0

5

10

15

20

25 30 Time (h)

35

40

45

50

55

Simulated Observed

1.4 1.2 1 0.8 0.6 0.4 0.2 0

0

5

10

15

20

25

30

35

40

45

50

55

Time (h)

Figure 2. Simulated and observed chlorine concentration at nodes (a) 3, (b) 11, (c) 19, and (d) 34.

and wall reaction parameters used are 0.55 d−1 and 0.15 m/d, respectively. The system has been hydraulically well calibrated, with most pipes having been assigned roughness coefficients (HWC) of 100. The chlorine input at node 1 (Fig. 1) to the network has a constant value of 1.15 mg/L. The chlorine transport model is run using the time-driven method for a hydraulic time step of 1 hr and a water quality step of 3 min. The results are obtained for a simulation period of 55 hr. The chlorine concentrations at a few network nodes are represented in Fig. 2. CHLORINE REACTION PARAMETER ESTIMATION The chlorine transport model described earlier predicts the constituent concentrations throughout the distribution system under steady or dynamic state. The reliability of these predicted concentrations when compared with the field observations depends on the assigned parameter values involved in the type of reaction kinetics used in the model. The parameters that control the chlorine reaction kinetics within the system can be broadly classified into the bulk and wall reaction parameters. Bulk reaction parameters (first- or non-first-order) are associated with individual pipes and storage tanks, assigned to groups of pipes in an area, which is contributed more by a particular source, or applied globally. Wall reaction

parameters (first- or zero-order) are associated with individual pipes, applied globally, or assigned to a group of pipes with similar material/age/roughness factors. The wall reaction parameter can also be related inversely to the Hazen–Williams roughness coefficient and represented in the wall reaction pipe-roughness parameter. The advantage of using this sort of representation is that it requires only a single parameter to allow wall reaction parameters to vary throughout the network in a physically meaningful way (5). The parameters involved in the single-parameter or multiple-parameter bulk reaction expressions can be determined with the data sets observed by conducting the bottle tests on the water samples. By performing measurements in the distribution system, the researchers have calculated the decay rate constant (overall firstorder) for site-specific tests (such as fixed pipe diameter, pipe material, or water source). This process may yield reasonable results; however, a wide range of values for this constant is obtained, thus severely limiting its use as a predictive tool (31). The determination of an overall reaction parameter (which represents the combined effect of bulk and wall reactions) and a wall reaction parameter is much more difficult than is establishing a bulk reaction parameter. Hence, these reaction parameters are more a product of calibration. Calibration is a process of

MODELING CHLORINE RESIDUALS IN URBAN WATER DISTRIBUTION SYSTEMS

adjusting a model so that the simulation reasonably predicts system behavior. The objective of water quality calibration is to capture the steady/transient dynamic behavior of the network. Wall reaction parameters are similar to pipe roughness coefficients in that they can and do vary from pipe to pipe. Unfortunately, unlike the head loss test for pipe roughness values, direct measurements of wall reaction parameters are extremely difficult to make (32). Because these parameter values are difficult to measure, they need to be estimated with the field measurements. The various techniques adopted to estimate these parameters include the trial-error (33), gradient-descent-based search technique (34), method of Lagrange multiplier (35), Gauss–Newton sensitivity analysis technique (36), and stochastic-based genetic algorithm (GA) technique (37). Gradient-based methods are generally faster, but they are more difficult to formulate because either an analytical expression must be derived or the gradient must be approximated. Stochastic search methods are more robust and simpler to formulate and use, but they are generally slower (32). The parameter estimation can be formulated as an optimization problem so that the difference between the observed and computed chlorine concentrations at the monitoring nodes are minimized in the least-squares sense. Thus, the objective function is given by Minimize E =

N(j) M  

[Cnoj,tk − Cncj,tk ]2

(7)

j=1 k=1

where M = number of monitoring nodes; N(j) = number of monitoring times at node j; Cnoj,tk = computed chlorine concentration at node j at time tk (mg/L); and Cnoj,tk = measured chlorine concentration at node j at time tk (mg/L). The simulation-optimization inverse modeling technique, which uses the field measurements and simulated chlorine concentrations at monitoring nodes, can be

Genetic Algorithm

Initial/New Set of Reaction Parameters

Chlorine Simulation Model

Apply GA Operators: Fitness Scaling Niche Operator Reproduction Crossover Mutation Creep Mutation Elitism

adopted to solve this unconstrained optimization problem. The flow diagram in Fig. 3 illustrates the simulationoptimization procedure if GA is used in its optimization module. The various unknown reaction parameters constitute a set of decision variables to be evaluated by GA, and this can work by evaluating the fitness of each potential solution that consists of values for the set of unknown reaction parameters. Fitness is determined by comparing how well the simulated chlorine concentrations that result from the candidate solution match the measured values collected in the field. The computationally intensive step involved in the GA technique is the determination of fitness that is somehow related to the objective function value. The GA continues to spawn generations of potential solutions until comparison of solutions from successive generations no longer produces a significant improvement. In addition, the GA process eliminates most routine and tedious aspects of the calibration process. GA will generally achieve better fits to the available data if the correct set of variables is included in the solution and it can establish the correct range of possible solutions. APPLICATION OF THE INVERSE MODEL FOR PARAMETER ESTIMATION The network used earlier is chosen for applying the inverse model. The time-varying chlorine concentrations were observed for this network at the nodes 3, 6, 10, 11, 19, 25, 28, and 34 for a period of 55 hr (10). These measurements become input observed data to the inverse model, and the unknown global wall reaction parameter for the first-order reaction kinetics, which is assumed to be applicable for the system, is estimated. The inverse model is run with a zero concentration tolerance, quality time step of 3 min, and hydraulic step of 1 hr. The estimated value of the global first-order wall reaction parameter is found to be 0.3654 m/d, which results in the lowest possible RMS

Select Best of Generation Reaction Parameters

Nodal Chlorine Concentrations

135

Evaluation of Fitness Function

Repeat for Each String (Set of Reaction Parameters)

Figure 3. GA implementation.

MODELING CHLORINE RESIDUALS IN URBAN WATER DISTRIBUTION SYSTEMS (a) Chlorine concentration (mg/L)

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Figure 4. Simulated and observed chlorine concentration at nodes (a) 3, (b) 11, (c) 19, and (d) 34.

residual error of 0.172 mg/L without any tedious trialand-error computations. The comparison of observed and simulated chlorine concentrations for this parameter estimated at node 3, 11, 19, and 34 are represented in Fig. 4. CLOSURE The maintenance of chlorine residual is needed at all points in the distribution system supplied with chlorine as a disinfectant. Chlorine is subjected to bulk flow and pipe wall reactions as it propagates through the pipes of the distribution system. The bulk flow reaction term depends on the organic content of the water, whereas the wall reaction term is related to the material and age of the distribution pipe. Because of these reactions, the loss of chlorine is significant between the outlet of the treatment plant and the consumer end. Thus, the study of spatial and temporal distribution of chlorine forms an important aspect of modeling. The chlorine simulation model forms the base for the reaction parameter estimation. The application of the inverse model is more relevant as any combination of bulk and wall reaction kinetics is possible for best fit with field chlorine observations. The inverse model is significant in deciding the operational strategy for real life systems. The chlorine simulation together with the inverse model provides the water supply agencies a tool for better management of their systems.

BIBLIOGRAPHY 1. Larson, T.E. (1966). Deterioration water quality in distribution systems. J. AWWA. 58(10): 1307–1316. 2. Pathak, P.N. (1969). Deterioration of potable water quality in distribution system and its remedial measures. Environmental Health 11(10): 220–228. 3. Boulos, P.F., Altman, T., Jarrige, P.A., and Collevatti, F. (1995). Discrete simulation approach for network water quality models. J. Water Resour. Plng. Mgmt., ASCE 121(1): 49–60. 4. Rossman, L.A., Clark, R.M., and Grayman, W.M. (1994). Modeling chlorine residuals in drinking water distribution systems. J. Envir. Eng., ASCE 120(4): 803–820. 5. Vasconcelos, J.J. et al. (1997). Kinetics of chlorine decay. J. AWWA. 89(7): 55–65. 6. Frateur, I. et al. (1999). Free chlorine consumption induced by cast iron corrosion in drinking water distribution systems. Water Res. 33(8): 1781–1790. 7. Powell, J.C. et al. (2000). Performance of various kinetic models for chlorine decay. J. Water Resour. Plng. Mgmt., ASCE 126(1): 13–20. 8. Clark, R.M. and Sivaganesan, M. (1998). Predicting chlorine residuals and the formation of TTHMs in drinking water. J. Envir. Eng., ASCE 124(12): 1203–1210. 9. Ki´en´e, L., Lu, W., and L´evi, Y. (1998). Relative importance of the phenomena responsible for chlorine decay in drinking water distribution systems. Water Res. 38(6): 219–227.

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10. Ozdemir, O.N. and Ucak, A. (2002). Simulation of chlorine decay in drinking-water distribution systems. J. Envir. Eng., ASCE 128(1): 31–39.

30. Rossman, L.A. and Boulos, P.F. (1996). Numerical methods for water quality in distribution systems: A comparison. J. Water Resour. Plng. Mgmt., ASCE 122(2): 137–146.

11. Vasconcelos, J.J., Boulos, P.F., Grayman, W.M., Ki´en´e, L., Wable, O., Biswas, P., Bhari, A., Rossman, L.A., Clark, R.M., and Goodrich, J.A. (1996). Characterization and modeling of chlorine decay in distribution systems. AWWA Research Foundation, Denver, CO.

31. Biswas, P., Lu, C.S., and Clark, R.M. (1993). Chlorine concentration decay in pipes. Water Res. 27(12): 1715–1724.

12. Zhang, G.R., Ki´en´e, L., Wable, O., Chan, U.S., and Duguet, J.P. (1992). Modeling of chlorine residuals in the water distribution network of Macao. Environ. Technol. 13: 937–946. 13. Chun, D.G. and Selznick, H.L. (1985). Computer modeling of distri-bution system water quality. Proc. Comp. Appl. Water Resour., ASCE, NewYork, pp. 448–456. 14. Males, R.M., Clark, R.M., Wehrman, P.J., and Gates, W.E. (1985). Algorithm for mixing problems in water systems. J. Hydr. Eng., ASCE 111(2): 206–219. 15. Clark, R.M., Grayman, W.M., and Males, R.M. (1988). Contaminant propagation in distribution system. J. Envir. Eng., ASCE 114(4): 929–943. 16. Shah, M. and Sinai, G. (1988). Steady state model for dilution in water networks. J. Hydraulic Eng., ASCE 114(2): 192–206. 17. Wood, D.J. and Ormsbee, L.E. (1989). Supply identification for water distribution systems. J. AWWA. 81(7): 74–80. 18. Boulos, P.F., Altman, T., and Sadhal, K. (1992). Computer modeling of water quality in large networks. J. Appl. Math. Modeling 16(8): 439–445. 19. Liou, C.P. and Kroon, J.R. (1987). Modeling the propagation of waterborne substances in water distribution networks. J. AWWA. 79(11): 54–58. 20. Grayman, W.M., Clark, R.M., and Males, R.M. (1988). Modeling distribution system water quality: Dynamic approach. J. Water Resour. Plng. Mgmt., ASCE 114(3): 295–311. 21. Rossman, L.A., Boulos, P.F., and Altman, T. (1993). Discrete volume element method for network water quality models. J. Water Resour. Plng. Mgmt., ASCE 119(5): 505–517. 22. Boulos, P.F., Altman, T., Jarrige, P.A., and Collevatti, F. (1994). An event-driven method for modeling contaminant propagation in water networks. J. Appl. Math. Modeling 18(2): 84–92. 23. Islam, M.R. and Chaudhary, M.H. (1998). Modeling of constituent transport in unsteady flows in pipe networks. J. Hydr. Eng., ASCE 124(11): 1115–1124. 24. Shang, F., Uber, J.G., and Polycarpou, M.M. (2002). Particle backtracking algorithm for water distribution system analysis. J. Environ. Eng., ASCE 128(5): 441–450. 25. Tzatchkov, V.G., Aldama, A.A., and Arreguin, F.I. (2002). Advection-dispersion-reaction modeling in water distribution networks. J. Water Resour. Plng. Mgmt. 128(5): 334–342. 26. Munavalli, G.R. and Mohan Kumar, M.S. (2004). Modified Lagrangian method for water quality in distribution systems. Water Res. 38(8): 2973–2988. 27. Axworthy, D.H. and Karney, B.W. (1996). Modeling low velocity/high dispersion flow in water distribution systems. J. Water Resour. Plng. Mgmt., ASCE 122(3): 218–221. 28. Ozdemir, O.N., and Ger, A.M. (1999). Unsteady 2-D chlorine transport in water supply pipes. Water Res. 33(17): 3637–3645. 29. Rossman, L.A. (2000). EPANET 2 Users manual. Risk Reduction Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio.

32. Walski, T.M., Chase, D.V., and Savic, D.A. (2001). Water Distribution Modeling. Haestad Press, Baltimore, MD. 33. Clark, R.M., Rossman, L.A., and Wymer, L.J. (1995). Modeling distribution system water quality: Regulatory implications. J. Water Resour. Plng Mgmt., ASCE 121(6): 423–428. 34. Al-Omari, A.S. and Chaudhry, M.H. (2001). Unsteady-state inverse chlorine modeling in pipe networks. J. Hydrol. Eng., ASCE 127(8): 669–677. 35. Zeirolf, M.L., Polycarpou, M.M., and Uber, J.G. (1998). Development and autocalibration of an input-output model of chlorine transport in drinking water distribution systems. IEEE Trans. Control Syst. Technol. 6(4): 543–553. 36. Munavalli, G.R. and Mohan Kumar, M.S. (2003). ‘‘Water quality parameter estimation in a steady state distribution system’’, J. Water Resour. Plng and Mgmt., ASCE, 129(2), 124–134, 2003. 37. Mohan Kumar, M.S. and Munavalli, G.R. (2003). Autocalibration of chlorine transport model for steady state distribution system by genetic algorithm. Seventh International Water Technology Conference IWTC-2003, Egypt, May 1–3.

PARTICULATE MATTER REMOVAL BY COAGULATION RASHEED AHMAD Khafra Engineering Consultants Atlanta, Georgia

Particulate matter in natural water varies in size, concentration, and surface chemistry. The particle size may range from a few tens of nanometers to a few hundred micrometers. Discrete particles less than one micron in size are called colloidal. Colloidal particles have significantly higher external surface area per unit area and move in a random diffusional motion known as Brownian motion. In colloidal suspension, surface phenomena dominate over mass phenomena. The most important surface property is the accumulation of electrical charges at the particle surface. Loss of atoms due to abrasion, molecular arrangement within the crystal, and imperfections within the molecular structure may result in surfaces being charged. The colloidal particles in most surface water are negatively charged. Because of hydration and/or electrostatic surface charges, colloidal particles repel other material and thereby remain suspended. Surface waters that are turbid due to colloidal particles cannot be clarified without special treatment. Coagulation is a process for enhancing the tendency of particulate matter in aqueous suspension to attach to one another and/or to attach to collector surfaces. Coagulation promotes destabilization of surface charges on colloidal particles. Destabilization and aggregation of particulate matter and precipitation or adsorption of NOM in subsequent solid–liquid separation processes are

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PARTICULATE MATTER REMOVAL BY COAGULATION

the primary functions of the coagulation process. The coagulation process involves two steps: (1) the addition of chemical coagulants to destabilize particulate matter and react with NOM and (2) the physical transport of collisions among particulate matter, resulting in aggregation or floc formation. In the water treatment literature, coagulation refers to all reactions and mechanisms that result in aggregation, and the physical transport step of producing interparticle aggregation is called flocculation. In a water treatment plant, coagulation is achieved by rapid or flash mixing of coagulants followed by flocculation. The two most common types of coagulants are metallic salts and polymers; the most common metallic salt coagulants are aluminum sulfate (alum) and ferric chloride. The selection of a particular coagulant depends on the required level of effectiveness. A standard jar test is a recommended method for determining the relative effectiveness of coagulants for a particular raw water supply. The factors that are considered normally in selecting a coagulant include cost, availability, overall safety, ease of storage, handling, and application. Alum is the most widely used coagulant because of its availability, low cost, ease of use, and ease of storage. Ferric chloride, other metallic salts, and polymers are less widely used. Alum’s performance, however, is greatly affected by the pH of the influent. The commonly used dosage of alum ranges from 5 to 150 mg/L, but the problem of sludge disposal increases at higher alum dosages. Due to special raw water characteristics and because of health concerns about aluminum, some water utilities use ferric chloride. Although ferric chloride is not always as effective as alum in reducing trihalomethane formation potential (THMFP) and total organic carbon (TOC), it is more effective than alum for water that has high dissolved color, low turbidity, and a moderate pH. Polymers are effective coagulants, coagulant aids, and filter aids. They consist of monomers and are classified according to their charge or lack of charge. A polymer that has a charge is an ionized polymer, or a polyelectrolyte. Polymers can be cationic, anionic, or nonionic. In applications where polymers are effective, dosages are generally lower than alum dosages for the same effect. Typical polymer dosages range from 1.5 to 10 mg/L. Consequently, polymer coagulants produce less residual sludge than alum. Coagulant aids are added to the influent after or simultaneously with the primary coagulants to improve particle capture efficiency during flocculation, sedimentation, and filtration. Nonionic and anionic polymers are commonly used as coagulant aids. The ratio of alum to coagulant aid dosages ranges from 100:1 to 50:1. Standard jar tests are required to determine precise coagulant aid dosages. There are four coagulation mechanisms that, it is thought, occur in destabilizing colloidal particles: double layer compression, surface charge neutralization, sweep coagulation, adsorption and interparticle bridging. The double layer model is used to understand the ionic environment near a charged colloid particle. The surface charge on the colloid attracts ions of opposite charge and forms a dense layer adjacent to the particle known as

the Stern layer. Excess positive ions are still attracted by the negatively charged colloids but are repelled by the Stern layer. This dynamic equilibrium results in creating a diffuse layer of counterions. The Stern and the diffuse layer in the interfacial region around colloidal particles are referred to collectively as the double layer. The electrical potential at the junction of the Stern layer and the diffuse layer, called the zeta potential, can be measured experimentally. It correlates with colloid particle stability. Highly stable colloidal systems are characterized by a high zeta potential, whereas lower zeta potentials reflect less stable systems. The DLVO theory (named after Derjaguin, Landau, Verwey, and Overbeek) governs the net interactive force between colloidal particles by combining the van der Waals attractive force and the electrostatic repulsion force. The double layer can be compressed by adding a coagulant that has a positive charge (to counteract negatively charged colloids). In water treatment practice, destabilization by double layer compression is not a dominant mechanism because it requires an extremely high salt concentration. This is an important destabilization mechanism in natural systems, for instance, delta formation in estuaries. Destabilization by surface charge neutralization involves reducing the net charge of colloidal particles in the suspension. The net surface charge can be reduced by adjusting the solution chemistry. In other cases, colloidal particles can be destabilized by neutralizing using counterions of coagulants. In water treatment practice, a similar type of surface charge destabilization occurs that is called heterocoagulation. The distribution of charges on a colloidal surface is not uniform. Large particles that have high negative surface charges may come in contact with smaller particles that bear relatively low positive charges. These particles may be destabilized by simple electrostatic interaction. Sweep coagulation or sweep-floc coagulation is also known as enmeshment in a precipitate. At higher coagulant doses, excess metal salts hydrolyze into metallic hydroxides. These hydroxides are extremely insoluble in water, amorphous, heavier than water, and gelatinous. As the hydroxide precipitate forms and accumulates, the colloidal particles are enmeshed or entrapped in the hydroxide floc. This destabilization mechanism is called sweep coagulation. Interparticle bridging destabilization occurs when highmolecular-weight polymers are used as coagulants or coagulant aids. These polymers are highly surface-active, and their surface structure may be linear or branched. The polymers destabilize particles by first adsorbing at one or more sites on the colloidal particle surface and then extending the chain length into solution and attaching to other particles. This results in forming an interparticle bridge. Sometimes an excessive dosage of polymer may cause restabilization due to surface saturation or sterical stabilization. READING LIST ASCE and AWWA. (1998). Water Treatment Plant Design, 3rd Edn. McGraw-Hill, New York.

SELECTIVE COAGULANT RECOVERY FROM WATER TREATMENT PLANT RESIDUALS USING THE DOMAIN MEMBRANE PROCESS

10000

DOC, Al(III) concentration (mg/L)

AWWARF. (1998). Treatment Process Selection for Particle Removal. American Water Works Association Research Foundation, Denver, CO. AWWA. ( 1999). Water Quality and Treatment, 5th Edn. McGrawHill, New York. Hudson, H.E. (1981). Water Clarification Processes. Van Nostrand Reinhold Company, New York. Kawamura, S. (2000). Integrated Design of Water Treatment Facilities. John Wiley & Sons, New York. Qasim, S.R., Motley, E.M., and Zhu, G. (2000). Water Works Engineering. Prentice-Hall, Upper Saddle River, NJ. USEPA. (1990). Technologies for Upgrading Existing or Designing New Drinking Water Treatment Facilities. Office of Drinking Water, Center for Research Information, Cincinnati, OH.

Al(III) 1000

100

SELECTIVE COAGULANT RECOVERY FROM WATER TREATMENT PLANT RESIDUALS USING THE DOMAIN MEMBRANE PROCESS PRAKHAR PRAKASH Pennsylvania State University University Park, Pennsylvania

ARUP K. SENGUPTA Lehigh University Bethlehem, Pennsylvania

BACKGROUND During the last two decades, pressure-driven membrane processes, namely, reverse osmosis (RO), nanofiltration (NF), and ultrafiltration (UF), have found increased applications in water utilities and chemical industries. Unlike RO, NF, and UF, Donnan membrane process (DMP) or Donnan dialysis is driven by an electrochemical potential gradient across an ion exchange membrane. Theoretically, the Donnan membrane process is not susceptible to fouling because particulate matter or large organic molecules do not concentrate on the membrane surface, as commonly observed with pressuredriven membrane processes. Although information on several applications of DMP is available in the open literature (1,2), no work is reported on the use of DMP to treat a sludge or slurry with a high concentration of suspended solids or large organic molecules. It was conceived that a single-step Donnan membrane process could selectively recover coagulant alum (Al2 (SO4 )3 · 14H2 O) (3,4) from water treatment plant sludge or water treatment plant residuals (WTR), which are the endproduct of coagulation. WTR contain insoluble aluminum hydroxide (50–75%) along with suspended inorganic particles, natural organic matter (NOM), and trace amounts of heavy metal precipitates (5). Several efforts were made to recover alum from WTR. The acid digestion process is the most commonly tried process at the laboratory, pilot-scale, and plant level (6). In this process, WTR are sufficiently acidified with sulfuric acid, dissolving insoluble aluminum hydroxide in the form of alum up to aluminum concentration levels of 360–3700 mg/L. However, the process is nonselective;

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Equilibrium pH Figure 1. Variation of dissolved organic carbon (DOC) and aluminum concentration with pH for water treatment residuals (WTR) from Allentown Water Treatment Plant (AWTP).

with the dissolution of aluminum hydroxide, NOM-like humates and fulvates get dissolved too, and the resulting dissolved organic carbon (DOC) concentration ranges from 326 to 1800 mg/L (7). This recovered alum, if reused as a coagulant, may impart a high trihalomethane formation potential (THMFP) during the chlorination stage of water treatment. The trihalomethanes are suspected carcinogens regulated by the USEPA (8). As an alternative to the acid digestion process; the amphoteric nature of aluminum oxide also permits alum recovery from the WTR under alkaline conditions. However, the alkali digestion process suffers from the same limitation as the acid digestion process; i.e., NOM concentration is very high in the recovered solution. Figure 1 shows both DOC and aluminum concentrations of the Allentown Water Treatment Plant (AWTP) in WTR at different pH levels. The Donnan membrane process is uniquely capable of recovering alum from WTR in a single-step process using sulfuric acid and a cation-exchange membrane.

THEORY Let us consider solutions of aluminum sulfate (feed) and sulfuric acid (recovery) in a Donnan membrane cell divided into two chambers by a cation-exchange membrane that allows only cations to migrate from one side to the other but rejects any passage of anions according to Donnan’s co-ion exclusion principle (9). At equilibrium, the electrochemical potential of aluminum ion Al3+ ion (µ) in the feed solution will be the same as that in the recovery solution for both aluminum and hydrogen ions, which corresponds to the following Donnan equilibrium condition: 

CR Al CLAl



 =

CR H CLH

3

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Water treatment residuals (WTR) Impermeable Natural organic matter (NOM): anions & neutral molecules

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Figure 2. A schematic of Donnan membrane process illustrating selective alum recovery from WTR.

recovery side contained 1.5 L of 10% sulfuric acid solution, separated by a cation-exchange membrane Nafion 117. At the start, pH of the WTR side was between 3.0 and 3.5. With the progress of the run, aluminum ions from the WTR side moved to the recovery side through the cation-exchange membrane, whereas an equivalent amount of hydrogen ions permeated to the WTR side, thus further reducing the pH. Under the experimental conditions of the Donnan run, free aluminum ions, Al3+ , was the predominant aluminum species. Figure 3 shows the results of the process for a period of 24 hours; the percentage aluminum recovery and the concentration of aluminum in the two chambers were plotted against time. It can be seen that over 70% recovery (72%) was attained in 24 hours. The noteworthy observation is that the

R

CH is 10, it means CR L Al is 1000 times CH L greater than CAl . Thus, by maintaining high hydrogen ion concentration in the recovery solution, aluminum ions can be driven from the feed to the recovery side even against a positive concentration gradient, i.e., from a lower concentration region to a higher concentration one. Figure 2 depicts the conceptualized selective alum recovery from WTR, highlighting the following. If the ratio

KEY FINDINGS In the Donnan membrane cell, the feed side of the membrane contained 6 L of the decanted and slightly acidified WTR collected from the AWTP, whereas the

(a) 7000

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PHYSICAL WATER CONDITIONING

recovered aluminum concentration was 6,650 mg/L as Al, and it was significantly greater than the total aluminum concentration (2400 mg/L) present in the parent sludge. It was also noted that the recovery was selective with respect to trace heavy metal ions. The recovered alum did not contain any suspended solids, whereas NOM expressed as DOC was consistently less than 5 mg/L. The ratio of individual contaminants to aluminum in the recovered alum was comparable, and in some cases lower, than in the commercial alum currently being used in AWTP. Similar results were obtained with WTR received from the Baxter Plant (Philadelphia, PA), which used FeCl3 as a coagulant, where over 75% recovery was made in 24 hours. Figure 4(a,b) show the visual comparison of recovered coagulants, both alum and ferric sulfate, between the traditional acid digestion process and the Donnan membrane process. Higher transparency of the coagulants from AWTP and the Baxter Plant, recovered by Donnan membrane process, is readily noticeable because of the absence of turbidity and NOM. CONCLUSIONS In this work, it was worthy to note that (a) aluminum (ferric) hydroxide precipitates could be dissolved and coagulant ions concentrated in the recovery solution; (b)

(a)

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negatively charged NOM, sulfate, and chloride could not permeate the membrane because of Donnan exclusion; and (c) the recovered alum was sufficiently pure and reuseable in water treatment plants. BIBLIOGRAPHY 1. Wallace, R.M. (1967). Ind. Eng. Chem. Process Des. Dev. 6(4): 423–431. 2. Kim, B.M. (1979). AIChE Symposium Series 76(197): 184–192. 3. Prakash, P. and SenGupta, A.K. (2003). Environ. Sci. Technol. 37: 4468–4474. 4. SenGupta, A.K. and Prakash, P. (2002). Process for selective coagulant recovery from water treatment plant sludge. U.S. Patent 6,495,047. 5. Cornwell, D.A and Westerhoff, G.P. (1981). Management of water treatment plant sludges. In: Sludge and Its Ultimate Disposal, J.A. Borchardt (Ed.). Ann Arbor Science, Ann Arbor, MI. 6. Bishop, M.M., Rolan, A.T., Bailey, T.L., and Cornwell, D.A. (1987). J. AWWA. 79(6): 76–83. 7. Saunders, F.M. (1989). Coagulant Recovery from Alum WTR at North Area Plant. Final Report for Bureau of Water, City of Atlanta, GA. 8. Christman, R.F., Norwood, D.L., Millington, J.D.S., Stevens, J.D., and Stevens, A.A. (1983). Environ. Sci. Technol. 17: 625–628. 9. Helfferich, F. (1995). Ion Exchange. Dover Publication, New York, pp. 134–135.

PHYSICAL WATER CONDITIONING CHARLES H. SANDERSON Magnatech Corporation Fort Wayne, Indiana

(b)

Figure 4. (a) Visual comparison of recovered alum coagulant from AWTP residuals by acid digestion process (left) and Donnan membrane process (right) (b) Visual comparison of recovered ferric coagulant from Baxter Plant residuals by acid digestion process (left) and Donnan membrane process (right).

Magnetic water treatment technology that was used two centuries ago to control hard water scale is now being used in hundreds of different kinds of beneficial applications throughout the world. Permanent magnetic water conditioning (PMWC) does not change the chemistry of the water; only a physical change takes place. It is therefore referred to as physical water conditioning (Fig. 1) (1). Physical water conditioning (PWC) has been highly controversial for many years; however, due to numerous successful installations on boilers, cooling towers, and other HVAC equipment, the technology has gained credibility throughout the industry in recent years. The ASHRAE Handbook (considered the ‘‘bible’’ of the industry) states, ‘‘Equipment based on magnetic, electromagnetic, or electrostatic technology has been used for scale control in boiler water, cooling water, and other process applications.’’ Several laboratory tests have been conducted to evaluate the results of magnetically treated water for scale and corrosion control in heat transfer equipment. The purpose of one such test was to determine the physical or chemical differences between treated and untreated residues. Emission spectrographic analyses of

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PHYSICAL WATER CONDITIONING + − + + − − + + − + − − Arrangement of nonpolar molecules in the absence of a magnetic field + −

− +

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+

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+ − Arrangement of polar molecules in the absence of a magnetic field

− +

(a) − +

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− + − + Arrangement of nonpolar molecules under the influence of a magnetic field − +

− + − +

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Arrangement of polar molecules under the influence of a magnetic field

Figure 1. The effect of a magnetic field. (b)

Magnetically treated water

Figure 2. Partially evaporated water droplets magnified to the 40th power.

Increase in energy cost Percent increase in power cost

the two water residues, it was found, have the same chemical constituents; however, a distinctive difference was observed in the crystalline structure. The residue from the magnetically conditioned water was a soft powder (when dry), whereas the untreated water deposit was typical boiler scale (2). Disassociated dissolved molecules of CaCO3 in water have a tendency to recombine by forming scale that adheres to the inner walls of the piping system, containers, steam vessels, etc. When water flows through a magnetic field of relatively low intensity, the formation of scale in the treated water is prevented in many instances. Instead, aragonite is formed within the flowing bulk water (aragonite forms a dilute slurry in the water, and the sediment can be easily removed by blowdown or bleed-off). The magnetochemical reaction is only one of the many cross effect reactions that enable the transformation of calcite to aragonite. Other reactions include thermochemical and mechanochemical reactions (Fig. 2). One of the benefits of PWC is energy savings from keeping the heat transfer area clean and free of scale, which prevents efficient transfer from the energy side to the water (see Fig. 3). Billions of dollars are lost in the United States every year due to corrosion. Laboratory tests and field installations, using corrosion coupons, as a comparison, have proven that magnetic water treatment reduces corrosion rates in HVAC equipment. This is accomplished by eliminating the aggressive chemicals used to control scale, and aragonite talc, a by-product of calcium, which puts a microscopic film on all wetted parts and provides protection against oxygen pitting. The CaCO3 with the PMWC does not build upon itself; however, a thin transparent coating deposits on the metal water side and dries as a fine white or gray powder. The National Aeronautics and Space Administration (NASA), using their Dynamic Corrosion Test System in an evaluation study, compared magnetically conditioned water to chemical treatment. Corrosion coupon #42 in the magnetically conditioned water loop had a corrosion rate of 0.0 mils per year; #41 and #43, treated with chemical corrosion inhibitors in the loop, had 5 and 6 mils loss per year respectively (Fig. 4) (3). An independent laboratory used a test rig consisting of two 48 inch glass cylinders filled with steel wool and

Untreated raw water

70 60 50 40 30 20 10 0

1/16

1/8

3/16 1/4 5/16 3/8 7/16 1/2 Scale thickness inches (Above data from the University of Illinois and the Bureau of Standards)

Figure 3. Additional fuel consumed with scale accumulation.

tapered at one end to evaluate the effect of magnetically treated water on corrosion. The flow rate of the water entering the top of each cylinder (one treated and one untreated) was adjusted to make the water head in the cylinders about 10 inches. After 72 hours, water in the untreated system overflowed the glass cylinder. The flow rate was then determined (4).

PHYSICAL WATER CONDITIONING NASA

Figure 4. NASA dynamic corrosion test stand and # C-75-1495 photos of coupons.

Untreated Tap Water Treated Tap Water Flow Rate Water height

25 mL per 58.0 s 46 inches

25 mL per 7.5 s 20 inches

A PMWC system was installed on a 4-year-old, 1500ton cooling tower and chillers that had moderate scale buildup in the tower and condenser tubes. There was also a considerable amount of corrosion and pitting on the inside of the chiller end caps. The chemical company that treated the system during the first four years of operation suggested that they monitor the results of the PMWC system. Six months after the installation, the chillers were opened for inspection. The end caps did not require wire brushing to remove the hard brittle scale as previously found, and the small accumulation of mud in the tubes was easily flushed out of the system by a garden hose. Prior to the PMWC installation, the average corrosion rate of the copper coupon ID # 10008 16-S was 0.01 MPY. After the PMWC installation, the copper coupon ID # 10007 A0882 corrosion rate was also 0.01 MPY; however, the loss was even and general and had an average penetration of 0.15 mils versus 0.50 mils penetration using the previous chemical treatment. The prior steel coupon ID #10008 48 V had a corrosion rate of 0.36 MPY and an average pit depth of 0.500 mils. The PMWC steel coupon ID # 10007 A3379 had a corrosion rate of 0.22 MPY and an even and general average penetration of 0.150 mils. A 100% return on investment took only 27 months in chemical cost alone (Fig. 5) In recent years, pollution prevention has taken on a new meaning. The transfer of polluted material from one

place to another was, at one time, considered ‘‘prevention’’ if the contaminated water was prevented from entering our freshwater streams. However, true prevention is now understood as eliminating the generation of toxic waste. The 200-year-old technology is now being used to replace chemicals previously used for water treatment. Research has been spurred recently due to public awareness and concern for the environment. Several different types of applications have been discovered since the turn of the century. The science laid dormant for many years; however, it is now considered ‘‘state of the art technology’’ and is gaining momentum in numerous industrialized countries. The arrangement of the magnetic fields is a very important factor to provide adequate and effective treatment for most applications. To a multiple reversing polarity field that has a N–SS–NN–S arrangement is the most successful in controlling hard water scale deposition (Fig. 6) (5). In addition to the multiple reversing fields, research has shown that scale can be best controlled when the water containing the minerals cuts through the magnetic lines of force at right angles. A steel pipe (magnetic material) surrounding the magnet pulls the magnetic lines of force at right angles through the water passageway. The pipe directs the water flow perpendicular to the magnetic fields (Figs. 7 and 8) (6). Since the turn of the century, there have been dozens of new beneficial applications for magnetically treated water found through new research discoveries and hundreds more in the field. There are numerous case histories of PMWC installations on various types of equipment that have been documented by third parties as successful applications (7). Most convincing are the original equipment manufacturers (OEM) who install magnetic water conditioners on their equipment to extend the period of time between maintenance service calls.

N

S

N SN SN S

(a)

(b)

N S S NN S

N SS

(c) Copper and mild steel corrosion coupons Dearborn Laboratory After 96/01284

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NN S

(d)

Figure 6. (a) One 2 pole, single field; (b) Three 2 pole, single field; (c) Three 2 pole, three separate fields; (d) One 6 pole, three dense fields.

Dearborn Laboratory After 96/01283 S

Figure 5. Nissan/Dearborn corrosion coupons.

NN

SS

N

Figure 7. One 6 pole, three dense fields, with steel shield.

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PHYSICAL WATER CONDITIONING

(Top view cutaway) Female NPT fittings

Copper shell

Galvanized steel shield

Copper stabilizer bars

NN

S

Raw water

SS

Parabolic copper housing for magnet

Cobalt alloy permanent magnet N

Air space

Magnetically treated water Figure 8. Cutaway view of a PMWC unit showing the water path perpendicular to the magnetic fields.

Water softeners that weighed over 100 pounds, plus bags of salt that were carried in carpet cleaning vans, are being replaced by small magnetic water conditioners (only 1 14 inches in diameter and 12 inches long), weighing less than 2 lb. Thousands of these small specially designed units for truck-mounted applications have been installed to prevent scale in steam boilers and related equipment by the manufacturer of a turnkey carpet steam cleaning equipment package. Hard water is one of the most common causes of truckmounted heating system failure. Untreated hard water causes loss of performance in heat exchangers and, many times, permanent damage to components (8). Several manufacturers of carpet cleaning systems throughout the United States have installed thousands of PMWC units on their truck-mounted systems during the last 5 years and reported excellent scale control and corrosion protection. Several manufacturers of the truck-mounts now encourage the use of PMWC by their customers who purchased their equipment with water softeners prior to the changeover. Magnetic water conditioning is believed to be the wave of the future by many proponents, especially environmental groups. It is nonpolluting, does not require energy for continued operation, and needs very little, if any, maintenance. The total value of the overall benefits of PMWC is still not known. There are hundreds of different types of successful applications now in use, and there are dozens more evaluations of laboratory studies and field installation presently underway. The PMWC system

has the ability to enhance our environment by reducing chemical use (true pollution prevention) and also helps save energy. The technology is cost-effective and plays a big role in water conservation. (9). The steel industry is reportedly the single largest consumer of PMWC technology in total gallons treated per day due to the high demand for water needed to cool its furnaces. Units are presently available in sizes up to 50,000 gallon per minute with 72 inch flanges. Economics is also one of the deciding factors in reducing chemical usage; the cost of preventing contaminated process wastewater from entering our freshwater rivers, lakes, and underground aquifers is very high. Transportation to another location can be as much as three times the cost of the chemical, depending on how hazardous the discharge water is. The food service industry is the largest user of the PMWC in total number of units; installations are on ice machines, coffee makers, dish washers, proofers, steamers, and drink dispensers. Sizes usually range from 1 gpm up to 15 gpm (10). Original equipment manufacturers that supply the food service industry have found that adding a small specially designed unit to their water using product eliminates the majority of service calls and replacement of parts under warranty that are prone to scaling or corrosion (11). As a result of millions of successful installations reported (many documented by qualified professionals), the question of whether or not magnetic water

CONSUMER CONFIDENCE REPORTS

conditioners work has been put to rest. However, the issue of how they work is still being debated. ASHRAE funded a comprehensive research study in 2000, titled ‘‘Efficiency of Physical Water Treatment for the Control of Scale’’ This report, published in 2003, confirms the results of several PMWC units performing under very rigid conditions and identifies the most effective velocity through each unit to control scale on heat transfer surfaces (12). BIBLIOGRAPHY 1. Quinn, C.J. (1986). Professor Emeritus, Indiana UniversityPurdue University Chairman, Manufacturing Technology. Magnetic Water Treatment For Heating, Refrigeration, and Air-Conditioning Systems. DOE/CE/40568-T1 (DE86014306). 2. Schmutzer, M.A. and Hull, G.W. (1969). Water Residues After Evaporation. Report # 28258. United States Testing Company, Inc. 3. Kuivinen, D.E. (1975). Dynamic Corrosion Test System. Report # 70112. National Aeronautics and Space Administration, Lewis Research Center. 4. Levinsky, H.L. and Schmutzer, M.A. (1968). Evaluation Of Steel Wool Corrosion With Magnetic Treated Water versus Untreated Water. Report # 27421-2. United States Testing Company, Inc. 5. Lin, I.J. (1982). Faculty of Civil Engineering, Mineral Engineering Department, Technion Institute of Technology, Haifa, Israel. 6. Young, I.C. (2002). Efficiency of Physical Water Treatments in Controlling Calcium Scale Accumulation in Recirculating Open Cooling Water System. Department of Mechanical Engineering, and Mechanics, Drexel University, College of Engineering. 7. Craine, J.E. (1984). Professor Emeritus, Ohio State University, International Foodservice Consultants Society, Control Lime/Scale Build-Up With Magnetic Water Treatment, Volume XVII, Number 2. 8. HydraMaster Corporation, Manufacturers of Truck Mounted Steam Machine Carpet Cleaning Systems, Instructions How To Retrofit Magnetic Water Conditioners (1998). 9. Evers, D. (1998). Technology for Improving Energy Efficiency Through the Removal or Prevention of Scale. DOE/EE-0162 Battelle Columbus Operations, United States Department of Energy—Federal Technology Alert. 10. Ferlin, J. (1982). Magnetic Water Conditioning Will Do Many Thing For The Foodservice Operator. President, Ferlin and Hopkins, USPWCA National Seminar Presentation. 11. Kappus, J. (1997). Taylor Freezer Food Service, Henny Penny Steamers. 12. ASHRAE Research Project 1155 TRP (2000).

CONSUMER CONFIDENCE REPORTS SUSAN L. FRANKLIN Tetra Tech MPS Ann Arbor, Michigan

On August 19, 1998, the U.S. Environmental Protection Agency (USEPA) promulgated a final rule (40 CFR Parts 141 and 142) requiring community water systems to

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prepare and provide annual water quality reports to their customers. Mandated by the 1996 amendments to the Safe Drinking Water Act, these ‘‘Consumer Confidence Reports’’ (CCRs) provide the missing link for consumers to have easy access to valuable water quality information. Just as food products have labeling requirements, CCRs are the public’s ‘‘right-to-know’’ resource. The federal rule outlines a number of required elements, including information on the source(s) of drinking water and any source water assessment that has been completed, plainly worded definitions of industry terminology, information on all regulated contaminants detected in the water, health effects language, notice for non-Englishspeaking residents, and specific language on vulnerable subpopulations and the reasonably expected presence of some contaminants. The CCR also gives the public direct access to their water system professionals by requiring the inclusion of information on public participation opportunities, local contacts, and additional water-related resources, such as USEPA’s toll-free Safe Drinking Water Hotline. In addition to the federal requirements, some states mandate supplementary guidelines. The first CCRs, due to customers on October 19, 1999, covered test results from January 1, 1998—December 31, 1998. Subsequent annual reports are due to consumers by July 1 each year and include data from the prior calendar year. Detections of contaminants that are tested for less frequently should also be included; however, no data more than 5 years old are permitted. All community water systems that have at least 15 service connections serving residents year-round must prepare a CCR. Water wholesalers must provide monitoring data and other required information to their retail customers by April 1 of each year. The retail system is responsible for ensuring that its customers receive a CCR containing all required content, whether they produce the report themselves or contract with their wholesaler to provide the report. After completing the CCR, each utility must certify to the state that they have complied with the CCR rule; each state determines the method for certification. A copy of the CCR is shared with the local health department, and the utility must keep the report on file for 5 years. CCR mailing and distribution requirements are based on the size of the water utility as outlined in the federal guidelines or as modified by the state. The governor of the state has the option to waive the mailing requirements for systems that serve fewer than 10,000 people; however, these systems must still prepare a CCR, inform customers that the report is available upon request, and publish the report in a local newspaper each year. Systems that receive the waiver and serve fewer than 500 customers are not required to publish the report in the newspaper. Systems that serve more than 10,000 people must mail one copy of the report to each customer must and make a ‘‘good faith effort’’ (mailing to all postal patrons in the service area, posting the CCR in public places, delivering multiple copies to apartment buildings or to large employers for distribution) to get reports to non-bill-paying customers. Systems that serve 100,000 or more people must also post their CCR on the Internet.

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CCRs are promoted by industry organizations, such as the American Water Works Association (AWWA), as ‘‘a timely opportunity for utilities to deliver straight talk to the consumer on quality and other issues affecting drinking water’’ (1). The AWWA offered training on how to prepare a CCR when the rule was issued, and it commissioned a series of six qualitative focus groups in 1997 to assess public perception and reaction to prototype CCRs in anticipation of the final rule. Their findings suggested that CCRs be called ‘‘Water Quality Reports’’ to let customers decide for themselves whether or not they are confident in the water system; to keep the report basic, simple, and honest; to avoid self-serving or alarming statements; and to gear the materials toward the needs of the system’s customers. According to the USEPA (2), there are more than 54,000 community water systems in the United States. Most Americans (263 million) get their drinking water from one of these community water systems and use about 370 billion gallons of water daily. Approximately $22 billion is spent annually to make tap water fit to drink. CCRs are the first nationwide commitment to educate the public about these important drinking water quality issues.

systems serving 10,000 or fewer people. EPA’s Basic guidelines suggest (1) metering, (2) water accounting and loss control, (3) pricing and costing, and (4) education or information. EPA’s Guidelines are not regulations, but recommendations that suggest 11 different conservation methods. How appropriate and desirable any given method is must, in the end, be accepted by the individual community and utility. Pricing may be the primary way to encourage conservation, however, utilities should not automatically rely on any single method.

METER ALL WATER Metering is a most important part of water demand management. In fact, unless a utility is 100 percent metered, it is difficult to enforce any conservation program. According to a U.S. Housing and Urban Development document, metered customers use an average of 13–45 percent less water than unmetered customers because they know they must pay for any misuse or negligence. A U.S. General Accounting Office report states that metering also assists in managing the overall water system, since it can help to:

BIBLIOGRAPHY 1. AWWA. (1998). Preparing the Consumer Confidence Reports— One Day Seminar, August. 2. USEPA Office of Water. (2001). Factoids: Drinking Water and Ground Water Statistics for 2000, EPA 816-K-01-004, June.

• locate leaks in a utility’s distribution system by identifying unaccounted-for blocks of water, • identify high use customers, who can be given literature on opportunities for conserving, and • identify areas where use is increasing, which is helpful in planning additions to the distribution system.

WATER CONSERVATION MEASURES National Drinking Water Clearinghouse

Water is a finite resource, and in many areas, future water supplies are uncertain. Individuals are usually aware when there is a drought; however, because water is inexpensive, there are often few incentives to reduce water loss. Water has no viable substitutes, and its depletion bodes profound economic and social impacts. Citizens and utilities need to consider water conservation programs. This fact sheet considers the role of water conservation as an integral part of long-term resource planning. It might be more appropriate to use the term ‘‘water demand management.’’ Traditional water supply management seeks to provide all the water the public wants, which, in some sections of the country, translates to a constant search for untapped sources. WHAT METHODS CONSERVE WATER? The water demand management methods described in this fact sheet incorporate the methods the August 1998 U.S. Environmental Protection Agency (EPA) Water Conservation Plan Guidelines recommend for water

Once water meters are installed, equipment begins to deteriorate. Eventually meters will fail to measure flows accurately. The question of how long to leave a meter in service has long troubled the waterworks industry. According to a Journal of the American Water Works Association (AWWA) article by Tao and a Community Consultants report, average losses of accuracy, for periods greater than 10 years, range from 0.03–0.9 percent per year. To be fair to both customers and the utility, meters must be maintained at regular intervals.

ACCOUNT FOR WATER, REPAIR LEAKS The EPA Guidelines recommend that all water systems—even smaller systems—implement a basic system of water accounting. The cost of water leakage can be measured in terms of the operating costs associated with water supply, treatment, and delivery. Water lost produces no revenues for the utility. Repairing larger leaks can be costly, but it also can produce substantial savings in water and expenditures over the long run. Water accounting is less accurate and useful when a system lacks source and connection metering. Although the system should plan to meter sources, unmetered source water can be estimated by multiplying the pumping rate by the time of operation based on electric meter readings.

WATER CONSERVATION MEASURES

A utility may want to consider charging for water previously given away for public use or stepping up efforts to reduce illegal connections and other forms of theft. Drinking water systems worldwide have begun to implement programs to address the problem of water loss. Utilities can no longer tolerate inefficiencies in water distribution systems and the resulting loss of revenue associated with underground leakage, water theft, and under registration. As pumping, treatment, and operational costs increase, these losses become more and more expensive. If a utility does what it can to conserve water, customers will tend to be more cooperative in other water conservation programs, many of which require individual efforts. In Economics of Leak Detection, Moyer states that of the many options available for conserving water, leak detection is a logical first step. A highly visible leak detection program that identifies and locates water system leakage encourages people to think about water conservation before they are asked to take action to reduce their own water use. When leaks are repaired, water savings result in reduced power costs to deliver water, reduced chemicals to treat water, and reduced costs of wholesale supplies. According to Le Moigne’s technical paper Using Water Efficiently: Technologies Options, old and poorly constructed pipelines, inadequate corrosion protection, poorly maintained valves and mechanical damage are major factors contributing to leaks. In addition to loss of water, water leaks reduce pressure in the supply system. Raising pressure to compensate for such losses increases energy consumption and can make leaking worse, as well as causing adverse environmental impacts. A World Bank technical paper by Okun and Ernst shows that, in general, it is normal to be unable to account for 10–20 percent of water. However a loss of more than 20 percent should raise a red flag. It should be noted that percentages are great for guidelines, but volume of water lost is probably more meaningful. According to AWWA’s Leak Detection and Water Loss Reduction, once a utility knows the volume of water lost, it can determine revenue losses and decide the best way to correct the problem. EPA’s Guidelines recommend that each system institute a comprehensive leak detection and repair strategy. This strategy may include regular onsite testing using computer-assisted leak detection equipment, a sonic leak-detection survey, or another acceptable method for detecting leaks along water distribution mains, valves, services, and meters. Divers can inspect and clean storage tank interiors. Increasingly, water systems are using remote sensor and telemetry technologies for ongoing monitoring and analysis of source, transmission, and distribution facilities. Remote sensors and monitoring software can alert operators to leaks, fluctuations in pressure, problems with equipment integrity, and other concerns. Each system should institute a loss-prevention program, which may include pipe inspection, cleaning, lining, and other maintenance efforts to improve the distribution system and prevent leaks and ruptures. Whenever

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possible, utilities might also consider methods for minimizing water used in routine water system maintenance procedures. COSTING AND PRICING In a Journal of the American Water Works Association article ‘‘Long-Term Options for Municipal Water Conservation,’’ Grisham and Fleming stress that water rates should reflect the real cost of water. Most water rates are based only on a portion of what it costs to obtain, develop, transport, treat, and deliver water to the consumer. Experts recommend that rates include not only current costs but those necessary for future water supply development. Only when rates include all costs can water users understand the real cost of water service and consequently, the need to conserve. When utilities raise water rates, among other factors, they need to consider what members of the community can afford. According to Schiffler, the ability to pay for water depends on a number of variables, including its intended use. In households, the assumption is that if the share of water costs does not exceed 5 percent of total household revenue it can be considered as socially acceptable. This rule of thumb has no specific foundation, but is widely used. Many utility managers argue, correctly, that an effective water conservation program will necessitate rate increases. In Water Conservation, Maddaus states that a reduction in water use by customers in response to a water conservation program can decrease a water utility’s revenues, and the utility may need to re-examine the water rate structure needs and possibly raise rates to compensate for this effect. Water charges have typically been looked at as a way of financing the operation and maintenance (O&M) costs of a water agency, rather than as a demand management measure to encourage water-use efficiency. As a World Bank document states, political objections and constraints to increasing water charges are often seen as insurmountable. However, low water charges encourage consumption and waste and can put pressure on O&M budgets, leading to poor water treatment and deterioration in water quality. In Water Strategies for the Next Century, Rogers et al. advocate a positive price for water that is less than the cost of desalination, but not zero. Desalination presently costs about $2 a cubic meter. The ideal is to charge a reasonable amount that sends the message to the users. EPA suggests that systems consider whether their current rate structures promote water usage over conservation. Nonpromotional rates should be implemented whenever possible. Systems that want to encourage conservation through their rates should consider various issues, such as the allocation between fixed and variable charges, usage blocks and breakpoints, minimum bills and whether water is provided in the minimum bill, seasonal pricing options, and pricing by customer class. Numerous sources recommend tying sewer prices to water prices. Billing for wastewater is not included in

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this analysis; however, it is expected to become a more significant motivation for reducing water use over the next 15 years.

• Representatives of major local economic interest groups—major industries, chambers of commerce, builders’ associations, farm bureaus, boards of realtors, and landscape contractors;

INFORMATION AND EDUCATION

• Representatives of major community forces, such as federated civic associations, neighborhood associations, school boards, local unions, churches, and local press and media owners;

According to Maddaus, water conservation initiatives are more likely to succeed if they are socially acceptable. Measuring social acceptability, an exercise in anticipating public response to a potential water conservation measure, may be measured with a two-part survey technique. First, conduct interviews with community leaders to assess the political and social atmosphere. Second, assess the response to selected specific measures via a questionnaire mailed to a random sample of water customers. The public tends to accept lawn watering restrictions, education, home water-saver kits, low-flush toilet rebates, and a low-flow fixtures ordinance for new construction. Overall acceptance of conservation is strongly related to attitudes about the importance of water conservation, as well as to age, income, and type of residence. Howe and Dixon note that, ‘‘Public participation is now widely understood to be a necessary input for both efficiency and equity.’’ Public participation should be part of any long-term public education program, as well as an element of plan development. A plan responsive to public needs usually receives continuing support. The EPA Guidelines state that water systems should be prepared to provide information pamphlets to customers on request. Consumers are often willing to participate in sound water management practices if provided with accurate information. An information and education program should explain to water users all of the costs involved in supplying drinking water and demonstrate how water conservation practices will provide water users with long term savings. An informative water bill goes beyond the basic information used to calculate the bill based on usage and rates. Comparisons to previous bills and tips on water conservation can help consumers make informed choices about water use. Systems can include inserts in their customers’ water bills that provide information on water use and costs or tips for home water conservation. School programs can be a great way to get information out. Systems can provide information on water conservation and encourage the use of water conservation practices through a variety of school programs. Contacts through schools can help socialize young people about the value of water and conservation techniques, as well as help systems communicate with parents. Workshops and seminars can be used to solicit input, and water equipment manufacturers can be invited to these sessions to exhibit their equipment. Maddaus suggests that a number of groups may have a role in water conservation planning: • Elected officials from all jurisdictions immediately affected by the process; • Staff persons from private water companies, key personnel from local government agencies, and state agency people;

• Representatives of local government interest groups; • Local professionals, such as economists and engineers; and • Representatives of major water users, for example, food processing plants and homeowners’ associations.

WHERE CAN I FIND INFORMATION? Information in this fact sheet was obtained from the following sources: American Water Works Association. 1986. Leak Detection and Water Loss Reduction. Distribution System Symposium Proceedings, Minneapolis, MN. American Water Works Association Leak Detection and Water Accountability Committee, 1996. ‘‘Committee Report: Water Accountability.’’ Journal of American Water Works Association. American Water Works Association. 1992. ‘‘Alternative Rates.’’ Manual of Water Supply Practices, Manual No. 34. Denver, CO. Baumann D., J. Boland, and M. Hanemann. 1998. Urban Water Demand Management and Planning: McGraw Hill. Community Consultants. 1986. ‘‘Testing of Residential Meters.’’ Consultants Report for the City of Tempe, Arizona. Grisham, A., and M. Fleming. 1989. ‘‘Long-Term Options for Municipal Water Conservation.’’ Journal of American Water Works Association. Howe, C. and J. Dixon. 1993. ‘‘Inefficiencies in Water Project Design and Operation in the Third World: An Economic Perspective.’’ Water Resources Research 29: 1889–1894. Le Moigne, G., U. Kuffner, M. Xie, et. al. 1993. ‘‘Using Water Efficiently: Technological Options.’’ Technical Paper 205, World Bank. Maddaus, W. 1987. Water Conservation. Denver: American Water Works Association. Moyer, E. E. 1985. Economics of Leak Detection: A Case Study Approach. Denver: American Water Works Association. Okun, D. A., W. Ernst. 1987. Community Piped Water Supply Systems in Developing Countries: A Planning Manual. World Bank Technical Paper 60. Rogers, P., K. Frederick, G. Le Moigne, D. Seckler, and J. Keller. 1994. Water Strategies for the Next Century: Supply Augmentation vs. Demand Management. A debate sponsored by the U.S.

PREVENTING WELL CONTAMINATION

Agency for International Development and ISPAN. Washington, DC.: U.S. Department of State. Schiffler, M. 1995. ‘‘Sustainable Development of Water Resources in Jordan: Ecological and Economic Aspects in a Long-Term Perspective.’’ in J. A. Allan and C. Mallat, eds. Water in the Middle East: Legal, Political and Commercial Implications. New York: I.B. Tauris Publishers. Tao, P., 1982. ‘‘Statistical Sampling Technique for Controlling the Accuracy of Small Water Meters.’’ Journal of American Water Works Association. U.S. Department of Housing and Urban Development. 1984. Residential Water Conservation ProjectsSummary Report. Report No. HUD-PDR-903, Prepared by Brown and Caldwell Consulting Engineers for the Office of Policy Development and Research, Washington, DC. U.S. Environmental Protection Agency. 1998. Water Conservation Plan Guidelines. EPA-832-D-001. Office of Water. Washington, DC. U.S. General Accounting Office. 1978. Municipal and Industrial Water Conservation—The Federal Government Could Do More. Report CED-78-66; B114885. Report to the Congress of the U.S. by the Comptroller General, Washington, DC. Vickers, A. 1990. ‘‘Water Use Efficiency Standards for Plumbing Fixtures: Benefits of National Legislation.’’ Journal of American Water Works Association 82: 53. World Bank. 1994. A Strategy for Managing Water in the Middle East and North Africa. Washington, DC. For further information or comments about this fact sheet, call the National Drinking Water Clearinghouse (NDWC) at (800) 624-8301 or (304) 293-4191. Additional copies of the Water Conservation Measures fact sheets are free; however, postal charges are added to orders. To order, call one of the above numbers. You may also order online at ndwc [email protected], or download it from our Web site at http://www.ndwc.wvu.edu where the fact sheet is available in the Products section.

PREVENTING WELL CONTAMINATION VIPIN BHARDWAJ NDWC Engineering Scientist

Nearly 80% of communities rely on groundwater as their primary drinking water source. Wells extract groundwater for use in homes and businesses. In addition, about 42 million Americans use private wells for drinking water. In light of this information, preventing groundwater contamination is of utmost importance, especially since a number of factors can contribute to groundwater contamination. To prevent well contamination, one of the first steps is to construct it properly. This Tech Brief presents tips about how to site a well and includes information about design issues; material selection and

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location, such as screens and filter pack; appropriate well sealing methods; and the use of pitless adaptors to prevent contamination INTRODUCTION To prevent well contamination, one of the first steps is to construct it properly. This tech brief presents tips on siting a well, its design, choosing proper materials, proper location of screens, filter pack and appropriate method of sealing a well and use of pitless adaptors to prevent contamination. SITE SELECTION To prevent groundwater contamination, the first step is to locate the well so that surface water and contaminants cannot flow into it. Site engineers try to install the well uphill from any potential contamination source. This means avoiding potential pollution sources, such as industrial plants, home septic systems, landfills, and underground storage tanks. Hiring a qualified hydrogeologist to investigate potential contaminant sources and likely subsurface conditions makes locating a well easier. For most private wells, the primary contaminant source is the owner’s septic system. The best protection practice is to locate the well above the area where contaminants can enter it, usually about 50 to 100 feet away. In addition, install a surface seal into a fine-grained layer or nonfractured zone above the aquifer. To prevent water from collecting near the casing, the ground surrounding the well should slope away from the wellhead on all sides. In addition, most states regulate how far a well must be located from potential contamination sources. For instance, most states require that wells be a minimum of 50 feet away from a septic system. WELL DESIGN Proper selection of well casings, seals, screens, filter packs, and pump chamber casings are important factors that determine the efficiency of the well and prevent contamination. Figure 1 shows the components of a well that prevent pollutants from entering the well. Most states have well construction standards and permitting processes that must be followed. The American Water Works Association has a standard A100-90 that deals with construction design. CASING A casing is a pipe that is usually made of steel or plastic. It lines the borehole dug in the earth and keeps the well from caving in and prevents runoff and other material from getting into the well. When contactors select casing, they must take into account the forces that are exerted while installing. In addition, the surrounding materials, such as soil and rocks, tend to collapse into the hole. If possible, the driller should

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Well cap Surface protective measures (SEE WAC 173-160-510) Gas vent tubes

Lock Well identifier Well protector cap with lock Drain holes (Outer casing) Concrete surface seal

Ground surface

Frost zone

Well seal

Vadose zone

Borehole diameter 4" larger than well casing (normal dimension)

Bentonite plug (2 feet thick minimum) Filter pack (3 feet above well screen)

Saturated zone

Well screen

Figure 1. General resource protection well—cross section Definitions Well Seal: A seal is a cylindrical layer of material, usually cement, bentonite, or clay, that surrounds the casing up to a certain depth in the well. It prevents runoff or other contaminants from entering the well, and serves to further protect the casing. Well Screen: Well screen is a cylindrical sieve-like structure that serves as the intake portion of the well. It is a metallic pipe that has holes or perforated sections or slotted sections that is placed on the water- carrying zones of the aquifer. Filter Pack: A filter pack is made up of sand or gravel that is smooth, uniform, clean, well-rounded, and siliceous. It is placed in the annulus of the well between the bore-hole wall and the well screen to prevent formation material from entering the screen. Vadose Zone: This is the zone that contains water under pressure less than that of the atmospheric pressure. It is the layer of soil between the water table and the ground surface. Potentiometric Surface: This is an imaginary surface representing the total head of groundwater in a confined aquifer that is defined by a level to which water will rise in the well.

use a temporary casing for the borehole. The temporary casing diameter must be at least four inches larger than the permanent casing to provide sufficient space for a good well seal. The American Society for Testing and Materials, the American Petroleum Institute, and the American Iron and Steel Institute have specifications for casings. Most state standards require steel casing of a specified wall thickness for wells, whether for a community or private individual. The diameter of the casing must leave enough room to install the submersible pump and still have space for

maintenance. The size of the pump depends upon the desired well yield. Casing depth also helps prevent well contamination. Logs of any other nearby wells and the local geology can help determine how deep the casing should go. The casing should extend at least 12 inches above the ground for sanitary protection. Reducing the casing’s diameter requires a minimum of eight feet of casing overlap. A watertight well cap should be placed on top of the casing. The Water Systems Council (WSC) has standards for well caps and other well components.

PREVENTING WELL CONTAMINATION

WELL SCREEN A well screen is a cylindrical sieve-like structure that serves as the intake portion of the well. It is a metallic pipe that has holes or perforated sections or slotted sections that is placed on the water carrying zones of the aquifer. Proper selection, design, placement, and development of the screened section are very important and determine the well’s efficiency and yield. Since certain sections of the ground are more porous than others and, hence, carry more water, placing the screens in these sections will yield higher flow rates. By looking at the data collected during drilling, a good well driller can locate and place the screen in the proper zones. To better understand conditions at the site, use borehole geophysical logs to grasp the subsurface conditions. In addition, visual inspection of the cuttings or samples can show if the layers of earth are sandy, coarse, or clayey. And to help determine well yield, use sieve analysis and hydraulic conductivity tests. FILTER PACK A filter pack is typically made up of sand or gravel that is smooth, uniform, clean, well rounded. It is placed in the area between the borehole wall and the well screen to prevent formation material from entering the screen. To enhance the permeability of the zone surrounding the screen, place a filter pack around it. A good filter pack keeps sediment out and decreases friction losses around the screen and is especially important if the aquifer consists of uniform fine sand. A filter pack allows for larger openings in the screen and improves well yield. To install a filter pack, start from the bottom of the screen, filling in to at least three feet above the top of the screen. Domestic wells do not require a filter pack. WELL SEALS The most important components that prevent contaminants from entering the well are well seals. A seal is a cylindrical layer of material, usually cement, bentonite, or clay, that surrounds the casing up to a certain well depth. It prevents runoff or other contaminants from entering the well and serves to further protect the casing. The drilled hole must be four inches larger in diameter than the outer diameter of the casing so that the seal can be placed in the space between casing and the hole. Well construction standards specify the material that well installers must use to seal the well, as well as the depth to which the well is grouted. Typically, public water supply wells are grouted to a depth of 50 feet. A cement slurry is pumped in the ring-shaped space between casing and hole and the well is sealed from the bottom up. Grout is placed using a small diameter pipe called a tremie. A layer of bentonite two feet thick should be placed on top of the filter pack. PITLESS ADAPTORS Pitless adapters and pitless units are devices that attach to the well casing below the frost line and provide sanitary

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connections. They prevent entry of contaminants into the well near the surface. These devices provide access to the well for servicing. The adapter connects the casing with a horizontal line that supplies water through a removable seal joint. This connection allows the drop pipe and pumping equipment in the well to be easily removed for repair or maintenance work without digging the ground around the well. WSC has performance standards for pitless adapters, pitless units, and watertight well caps. A list of manufacturers that meet those standards can be obtained from the WSC. DISINFECTION PROCEDURES Well installers must disinfect all equipment and tools using a chlorine solution before any drilling operation to prevent bacterial contamination. The well must be disinfected after it’s completed. Some types of bacteria, such as E. coli, are found in soils and can contaminate the well. By dissolving calcium hypochlorite or sodium hypochlorite, installers can make a chlorinated water solution. The strength of the solution can range from 50–200 milligrams per liter of available chlorine. WHERE CAN I FIND MORE INFORMATION? American Water Works Association. 1999. Design and Construction of Water Systems, An AWWA Small System Resource Book, Second Edition. Denver, Colorado: AWWA. Driscoll, FG. 1995. Groundwater and Wells. St. Paul, MN: U.S. Filter/Johnson Screens. U.S. Environmental Protection Agency. 1991. Manual of Individual and Non-Public Water Supply Systems. Washington, DC: EPA. (Available from the National Drinking Water Clearinghouse, order product #DWBKDM06). U.S. Environmental Protection Agency. 1975. Manual of Water Well Construction Practices. Washington, DC: EPA. (Available from the National Drinking Water Clearinghouse, order product #DWBKDM01). National Ground Water Association, Westerville, OH: NGWA. (www.ngwa.org) Water Systems Council, National Programs Office, Washington, DC: WSC. (www.wellcarehotline.org) Acknowledgments The author wishes to thank Dr. Dale Ralston, Ralston Hydrogeologic Services, for his comments and corrections to this Tech Brief. NDWC Engineering Scientist Vipin Bhardwaj has a BS in Chemical Engineering, and Masters degrees in Environmental engineering and agriculture from West Virginia University.

HAVE YOU READ ALL OF OUR TECH BRIEFS? Tech Briefs, drinking water treatment and supply fact sheets, have been a regular feature in the National Drinking Water Clearinghouse (NDWC) publication On Tap for more than seven years.

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A package of Tech Briefs is now available as a product. A three-ring binder holds all the current Tech Briefs in print. New selections can easily be added to the package as they become available. To order this free product, call the NDWC at (800) 624-8301 or (304) 2934191 and ask for item #DWPKPE71. (Additional copies of fact sheets are free; however, postal charges may be added.) You can also order copies of one or all of the free Tech Briefs listed below. Tech Brief: Organics Removal, item #DWBLPE59; Tech Brief: Package Plants, item #DWBLPE63; Tech Brief: Water Treatment Plant Residuals Management, item #DWBLPE65; Tech Brief: Lime Softening, item #DWBLPE67; Tech Brief: Iron and Manganese Removal, item #DWBLPE70; Water Conservation Measures Fact Sheet, item #DWBLPE74; Tech Brief: Membrane Filtration, item #DWBLPE81; Tech Brief: Treatment Technologies for Small Drinking Water Systems, item #DWPSPE82; Tech Brief: Ozone, item #DWBLPE84; Tech Brief: Radionuclides, item #DWBLPE84; Tech Brief: Slow Sand Filtration, item #DWBLPE99; Tech Brief: Ultraviolet Disinfection, item #DWBLPE101; Tech Brief: Leak Detection and Water Loss Control, item #DWBLPE102; Tech Brief: Diatomaceous Earth Filtration for Drinking Water, item #DWBLPE108. Tech Brief: Reservoirs, Towers, and Tanks, item #DWFSOM15 Tech Brief: System Control and Data Acquisition (SCADA), item #DWFSOM20 Tech Brief: Valves, item #DWFSOM21 Tech Brief: Water Quality in Distribution Systems, item #DWFSOM25 Tech Brief: Water Treatment Plant Residuals Management, item #DWBLPE65

CORROSION CAN CAUSE SYSTEM PROBLEMS What Problems Does Corrosion Cause? Corrosion can cause higher costs for a water system due to problems with: • decreased pumping capacity, caused by narrowed pipe diameters resulting from corrosion deposits; • decreased water production, caused by corrosion holes in the system, which reduce water pressure and increase the amount of finished water required to deliver a gallon of water to the point of consumption; • water damage to the system, caused by corrosionrelated leaks; • high replacement frequency of water heaters, radiators, valves, pipes, and meters because of corrosion damage; and • customer complaints of water color, staining, and taste problems. How is Corrosion Diagnosed and Evaluated? The following events and measurements can indicate potential corrosion problems in a water system: Consumer Complaints: Many times a consumer complaint about the taste or odor of water is the first indication of a corrosion problem. Investigators need to examine the construction materials used in the water distribution system and in the plumbing of the complainants’ areas (See Table 1). Corrosion Indices: Corrosion caused by a lack of calcium carbonate deposition in the system can be estimated using

Table 1. Typical Water Quality Complaints That Might Be Due to Corrosion Customer Complaint Red water or reddish-brown staining of fixtures and laundry Bluish stains on fixtures Black water

CORROSION CONTROL Foul taste and/or odors National Drinking Water Clearinghouse

Corrosion occurs because metals tend to oxidize when they come in contact with water, resulting in the formation of stable solids. Corrosion in water distribution systems can impact consumers’ health, water treatment costs, and the aesthetics of finished water. Various methods can be used to diagnose, evaluate, and control corrosion problems. Techniques for controlling it include distribution and plumbing system design considerations, water quality modifications, corrosion inhibitors, cathodic protection, and coatings and linings.

Loss of pressure

Lack of hot water

Short service life of household plumbing

Possible Cause Corrosion of iron pipes or presence of natural iron in raw water Corrosion of copper lines Sulfide corrosion of copper or iron lines or precipitations of natural manganese Byproducts from microbial activity Excessive scaling, tubercle (buildup from pitting corrosion), leak in system from pitting or other type of corrosion Buildup of mineral deposits in hot water system (can be reduced by setting thermostats to under 60 degrees C [140 degrees F]) Rapid deterioration of pipes from pitting or other types of corrosion

Source: U.S. Environmental Protection Agency

CORROSION CONTROL

indices derived from common water quality measures. The Langelier Saturation Index (LSI) is the most commonly used measure and is equal to the water pH minus the saturation pH (LSI = pH water − pH saturation). The saturation pH refers to the pH at the water’s calcium carbonate saturation point (i.e., the point where calcium carbonate is neither deposited nor dissolved). The saturation pH is dependent upon several factors, such as the water’s calcium ion concentration, alkalinity, temperature, pH, and presence of other dissolved solids, such as chlorides and sulfates. A negative LSI value indicates potential corrosion problems. Sampling and Chemical Analysis: The potential for corrosion can also be assessed by conducting a chemical sampling program. Water with a low pH (less than 6.0) tends to be more corrosive. Higher water temperature and total dissolved solids also can indicate corrosivity. Pipe Examination: The presence of protective pipe scale (coating) and the condition of pipes’ inner surfaces can be assessed by simple observation. Chemical examinations can determine the composition of pipe scale, such as the proportion of calcium carbonate, which shields pipes from dissolved oxygen and thus reduces corrosion. Can System Design Affect the Potential for Corrosion? In many cases, corrosion can be reduced by properly selecting distribution and plumbing system materials and by having a good engineering design. For example, water distribution systems designed to operate with lower flow rates will have reduced turbulence and, therefore, decreased erosion of protective layers. In addition, some piping materials are more resistant to corrosion in a specific environment than others. Finally, compatible piping materials should be used throughout the system to avoid electrolytic corrosion. Other measures that help minimize system corrosion include: • using only lead-free pipes, fittings, and components; • selecting an appropriate system shape and geometry to avoid dead ends and stagnant areas; • avoiding sharp turns and elbows in the distribution and plumbing systems; • providing adequate drainage (flushing) of the system; • selecting the appropriate metal thickness of piping, based on system flow and design parameters; • avoiding the use of site welding without replacing the pipe lining; • reducing mechanical stresses, such as flexing of pipes and ‘‘water hammer’’ (hydraulic pressure surges); • avoiding uneven heat distribution in the system by providing adequate coating and insulation of pipes; • providing easy access for inspection, maintenance, and replacement of system parts; and • eliminating the grounding of electrical circuits to the system, which increases the potential for corrosion.

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alkalinity; softening the water with lime; and changing the level of dissolved oxygen (although this is not a common method of control). Any corrosion adjustment program should include monitoring. This allows for dosage modification, as water characteristics change over time. pH Adjustment: Operators can promote the formation of a protective calcium carbonate coating (scale) on the metal surface of plumbing by adjusting pH, alkalinity, and calcium levels. Calcium carbonate scaling occurs when water is oversaturated with calcium carbonate. (Below the saturation point, calcium carbonate will redissolve: at the saturation point, calcium carbonate is neither precipitated nor dissolved. See the section on ‘‘corrosion indices,’’.) The saturation point of any particular water source depends on the concentration of calcium ions, alkalinity, temperature, and pH, and the presence of other dissolved materials, such as phosphates, sulfates, and some trace metals. It is important to note that pH levels well suited for corrosion control may not be optimal for other water treatment processes, such as coagulation and disinfection. To avoid this conflict, the pH level should be adjusted for corrosion control immediately prior to water distribution, and after the other water treatment requirements have been satisfied. Lime Softening: Lime softening (which, when soda ash is required in addition to lime, is sometimes known as limesoda softening) affects lead’s solubility by changing the water’s pH and carbonate levels. Hydroxide ions are then present, and they decrease metal solubility by promoting the formation of solid basic carbonates that ‘‘passivate,’’ or protect, the surface of the pipe. Using lime softening to adjust pH and alkalinity is an effective method for controlling lead corrosion. However, optimum water quality for corrosion control may not coincide with optimum reduction of water hardness. Therefore, to achieve sound, comprehensive water treatment, an operator must balance water hardness, carbonate levels, pH and alkalinity, as well as the potential for corrosion. Dissolved Oxygen Levels: The presence of excessive dissolved oxygen increases water’s corrosive activity. The optimal level of dissolved oxygen for corrosion control is 0.5 to 2.0 parts per million. However, removing oxygen from water is not practical because of the expense. Therefore, the most reasonable strategy to minimize the presence of oxygen is to: • exclude the aeration process in the treatment of groundwater, • increase lime softening, • extend the detention periods for treated water in reservoirs, and • use the correct size water pumps in the treatment plant to minimize the introduction of air during pumping.

How Can System Corrosion be Reduced?

What About the Use of Corrosion Inhibitors?

Corrosion in a system can be reduced by changing the water’s characteristics, such as adjusting pH and

Corrosion inhibitors cause protective coatings to form on pipes. Although they reduce corrosion, they may not totally

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arrest it. Therefore, the success of any corrosion inhibitor hinges upon the water operator’s ability to: • apply double and triple the design doses of inhibitor during initial applications to build a protective base coat that will prevent pitting; (Note that initial coatings typically take several weeks to form.) • maintain continuous and sufficiently high inhibitor doses to prevent redissolving of the protective layer; and • attain a steady water flow over the system’s metal surfaces to allow a continuous application of the inhibitor. There are several commercially available corrosion inhibitors that can be applied with normal chemical feed systems. Among the most commonly used for potable water supplies are inorganic phosphates, sodium silicates, and mixtures of phosphates and silicates. Inorganic Phosphates: Inorganic phosphate corrosion inhibitors include polyphosphates, orthophosphates, glassy phosphates, and bimetallic phosphates. Zinc, added in conjunction with polyphosphates, orthophosphates, or glassy phosphates, may help to inhibit corrosion in some cases. Silicates: The effectiveness of sodium silicates depends on both pH and carbonate concentrations. Sodium silicates are particularly effective for systems with high water velocities, low hardness, low alkalinity, and pH of less than 8.4. Typical coating maintenance doses of sodium silicate range from 2 to 12 milligrams per liter. They offer advantages in hot-water systems because of their chemical stability, unlike many phosphates. Before installing any technology for delivering corrosion inhibitors, several methods or agents first should be tested in a laboratory environment to determine the best inhibitor and concentration for each water system. Is Cathodic Protection an Option? Cathodic protection is an electrical method for preventing corrosion of metallic structures. However, this expensive corrosion control method is not practical or effective for protecting entire water systems. It is used primarily to protect water storage tanks. A limitation of cathodic protection is that it is almost impossible for cathodic protection to reach down into holes, crevices, or internal corners. Metallic corrosion occurs when contact between a metal and an electrically conductive solution produces a flow of electrons (or current) from the metal to the solution. The electrons given up by the metal cause the metal to corrode rather than remain in its pure metallic form. Cathodic protection stops this current by overpowering it with a stronger, external power source. The electrons provided by the external power source prevent the metal from losing electrons, forcing it to be a ‘‘cathode,’’ which will then resist corrosion, as opposed to an ‘‘anode,’’ which will not. There are two basic methods of applying cathodic protection. One method uses inert electrodes, such as

high-silicon cast iron or graphite, which are powered by an external source of direct current. The current impressed on the inert electrodes forces them to act as anodes, thus minimizing the possibility that the metal surface being protected will likewise become an anode and corrode. The second method uses a sacrificial anode. Magnesium or zinc anodes produce a galvanic action with iron, so that the anodes are sacrificed (or suffer corrosion), while the iron structure they are connected to is protected. Are Commercial Pipe Coatings and Linings Effective? The nearly universal method of reducing pipe corrosion involves lining the pipe walls with a protective coating. These linings are usually mechanically applied, either when the pipe is manufactured or in the field before it is installed. Some linings can be applied even after the pipe is in service, but this method is much more expensive. Mechanically applied coatings and linings differ for pipes and water storage tanks. The most common types of pipe linings include coal-tar enamels, epoxy paints, cement mortar, and polyethylene. Water storage tanks are most commonly lined to protect the inner tank walls from corrosion. The most common types of water storage tank coatings and linings include coal-tar paints and enamels, vinyls, and epoxy. Where Can I Find More Information? Information for this fact sheet was obtained from three primary sources: Technologies for Upgrading Existing or Designing New Drinking Water Treatment Facilities, EPA/625/4-89/023; Corrosion Manual for Internal Corrosion of Water Distribution Systems, EPA/570/9-84/001; and Corrosion in Potable Water Supplies, EPA/570/983/013. All of these documents are free and may be ordered from the U.S. Environmental Protection Agency (EPA) Office of Research and Development by calling (513) 5697562. If these publications are no longer available from the EPA, call the National Drinking Water Clearinghouse (NDWC) at (800) 624-8301. A photocopied version of the 209-page document Technologies for Upgrading Existing or Designing New Drinking Water Treatment Facilities, item #DWBKDM04, costs $30.05. There is no charge for the other two documents listed above; however, postage charges apply to all orders. Also, the NDWC’s Registry of Equipment Suppliers of Treatment Technologies for Small Systems (RESULTS), version 2.0, is a public reference database that contains information about technologies—including those related to corrosion—in use at small water systems around the country. For further information about accessing or ordering RESULTS, call the NDWC at (800) 6248301 or (304) 293-4191. You may also obtain more information from the NDWC’s World Wide Web site at www.ndwc.wvu.edu. For additional copies of ‘‘Tech Brief: Corrosion Control,’’ item #DWBRPE52, or for a copy of the previously published ‘‘Tech Brief: Filtration,’’ item #DWBRPE50, or ‘‘Tech Brief: Disinfection,’’ item #DWBRPE47, call the NDWC at the number printed above.

CROSS CONNECTION AND BACKFLOW PREVENTION

CROSS CONNECTION AND BACKFLOW PREVENTION VIPIN BHARDWAJ NDWC Engineering Scientist

BETHANY REED NESC Graphic Designer

When drinking water is transported to a consumer, it is possible for contaminants to be introduced in the distribution system. This situation may occur due to connections between potable water lines and non-potable water sources or by a water flow reversal, resulting in contaminated water. This Tech Brief, discusses cross connections and backflow, and explores ways to prevent these situations.

• • • • •

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air gaps, reduced pressure principle devices, double check valves, vacuum breakers, and barometric loops

Air Gap Air gaps are one of the most effective ways to prevent backflow and backsiphonage. An air gap is a vertical separation between a water outlet and the highest level of a potential fluid contamination source. However, because of air gaps, flow of water is interrupted and loss of pressure occurs. Because of this, air gaps are used at the end of a pipe. Air gaps should be twice the size of the supply pipe diameter or at least one inch in length, whichever is greater (see Fig. 1). Reduced Pressure Principle Backflow Preventer

WHAT IS A CROSS CONNECTION? A cross connection is a link or structural arrangement where potable water in a distribution system can be exposed to unwanted contaminants. It is the point at which it is possible for a non-potable substance to come in contact with the drinking water system. Cross connections are generally unintentional and can happen anywhere pipes supply water. WHAT IS A BACKFLOW? Backflow is the reverse flow of undesirable materials and contaminants into the water mains. Backflow can happen because of two conditions: backpressure and backsiphonage. Backpressure occurs when pressure in a pipe connected to a main pipe in the distribution system becomes greater than the pressure in the main pipe itself. When this happens, a net force acts on the volume of liquid in the connecting pipe, allowing unwanted material to enter the main pipe. Backsiphonage refers to a situation where the pressure in a service pipe is less than the atmospheric pressure. If water in a supply line is turned off, such as when a pump fails, backsiphonage can cause contamination to be sucked into the system due to a vacuum in the service line. If a cross connection exists in a system, it does not mean that there will be a backflow every time. But, where cross connections exist, there is always the possibility.

The reduced pressure zone backflow-preventing (RPBP) device has two spring check valves with a pressurerelief valve located between them that can be vented to the atmosphere. During normal flow of water through this arrangement, the water flows through the two valves (see Fig. 2). The spring action of the first valve opposes the pressure of water as the water flows from left to right and enters the central chamber. Pressure in the central chamber is maintained lower than that in the incoming line by the operation of the relief valve. The second check valve to the right is designed to open with a pressure drop of one pound per square inch (psi) in the direction of flow and is independent of the pressure required to open the relief valve. If the pressure downstream from the device increases for some reason (backpressure), the second check valve will close because of the spring action. Reverse flow of water or backflow is thus prevented. In case the pressure in the supply line on the left decreases abruptly or if there is a vacuum in the supply line (backsiphonage), the check valves close because of spring action, and backflow is prevented.

BACKFLOW CONTROL METHODS AND DEVICES If possible, cross connections must first be eliminated before installing any backflow prevention devices. The device chosen depends on the degree of hazard involved, accessibility to the location of the device, and whether the backflow is due to backpressure or backsiphonage. Basic types of backflow prevention devices are:

Figure 1. Air gap.

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Figure 2. Reduced pressure principle backflow preventer. Source: U.S. Environmental Protection Agency. 2003. Cross-Connection Control Manual, Washington DC: EPA.

In case the valve leaks and the second check valve on the right does not close fully, water will leak back into the central chamber and increase the pressure in the chamber. The relief valve then opens and discharges water to the atmosphere. Keep in mind, if you see water coming from the relief valve on an RPBP, don’t panic. It’s actually working as designed. The valve assembly should be checked by a certified tester. This type of device is usually installed on high hazard locations, such as hospitals, plating plants, and car washes. Double Check Valve The double check valve has two single check valves coupled within one body, and has test cocks (to determine if there’s any leakage) and two closing gate valves (to isolate each

section). It is essentially the same reduced pressure zone backflow-preventing (RPBP) device but without the relief valve (see Fig. 3). The absence of the relief valve reduces the effectiveness of the device. Double check valves are used in low- to medium-level hazard installations. The check valves are spring-loaded and require one pound of pressure to open.

Vacuum Breakers Vacuum breakers provide protection against backsiphonage. When the pressure in a service pipe is less than the atmospheric pressure, a vacuum can form in the pipe and cause contamination to be sucked into the system. Vacuum breakers have an element, such as a check valve, that glides on a supporting shaft and seals in the

Figure 3. Double check valve. Source: AWWA, Water Transmission and Distribution, Second Edition.

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Figure 4. Vacuum breakers. Source: AWWA, Water Transmission and Distribution, Second Edition.

uppermost position from the push of water pressure (see Fig. 4). If the flow in the pipe is stopped, the valve drops down, closes the water supply entry, and opens an air vent. This opens up the downstream piping to atmospheric pressure and prevents backsiphonage. Vacuum breakers do not protect against backpressure. Barometric Loop A barometric loop is formed by having a section of the pipe in the shape of an inverted ‘‘U’’ upstream of a cross connection (see Fig. 5). Based on a physics principle, the height of a water column open to the atmosphere at the bottom will not be greater than 33.9 feet at sea level pressure.

If the loop is greater than that height, backsiphonage cannot occur through it. However, the barometric loop is not effective against backpressure. ‘‘YELLOW GUSHY STUFF’’ FROM FAUCETS A small town in Maryland provides a dramatic example of what can happen because of a cross connection. One fateful day, yellow gushy stuff poured from some faucets there and the state warned residents from using the water for cooking, drinking, or bathing. The incident drew widespread attention and was a big story in the local media. An investigation revealed that water pressure in the town mains was reduced temporarily due to a water pump failure in the distribution system. A gate valve between a herbicide chemical holding tank and the town water supply mains had been left open. A lethal cross connection was created that allowed the herbicide to flow into the water supply. When the normal water supply pressure returned, water containing the herbicide was pumped into many faucets. Door to door public notification, extensive flushing, sampling, and other measures were taken to get the situation under control. Fortunately, no one was seriously injured in this incident. Source: U.S. Environmental Protection Agency. 2003. Cross-Connection Control Manual. Washington DC: EPA. CROSS CONNECTION CONTROL PROGRAMS

Figure 5. Barometric loop. Source: AWWA, Water Transmission and Distribution, Second Edition.

Numerous, well-documented cases about illnesses and other hazards posed by cross connections have been documented. (See the sidebar above.) More information about the health risks cross connections may present and methods to prevent them is needed. Water utility personnel (managers, operators, local officials), plumbers, public health officials, and consumers need to be aware of the risks and understand prevention methods.

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As mandated by the Safe Drinking Water Act, water suppliers are responsible for ensuring that the water they supply meets federal primary drinking water regulations and is delivered to consumers without compromising water quality due to its distribution system. Water utilities may want to implement a cross connection program to stave off any problems that could occur. Such a program would include informing consumers, conducting inspections of the distribution system, analyzing and recognizing potential cross connections, and installing backflow prevention devices where needed. WHAT ABOUT TERRORISM AND WATER SECURITY? Concern that U.S. water supplies could be vulnerable to terrorist attacks has increased in the post-9/11 era. Because they are a place where lethal substances could be introduced into the water distribution system, cross connections should be identified and corrected. The Public Health Security and Bioterrorism Preparedness and Response Act of 2002 (Section 1433) requires systems that serve more than 3,300 people to assess their vulnerability to a terrorist attack. (See the article ‘‘Security and Emergency Planning: Community-Wide Efforts Require Preparation’’ in the Winter 2003 On Tap.) The act states that the vulnerability assessment shall include but not be limited to ‘‘a review of pipes and constructed conveyances, water collection, pretreatment, storage and distribution facilities, electronic, computer or other automated systems.’’ WHERE CAN I FIND MORE INFORMATION? American Water Works Association. 1996. Water Transmission and Distribution, Principles and Practices of Water Supply Operations, 2nd Edn. Denver: AWWA. U.S. Environmental Protection Agency. 2003. CrossConnection Control Manual. Washington, DC: EPA. (Available as a product from NDWC, item #DWBLDM03) Montana Water Center. 2002. Sanitary Survey Fundamentals Preparation Course CD. Bozeman, MT: Montana Water Center. (Available as a product from NDWC, item #DWCDTR19)

the package as they become available. To order, call the NDWC at (800) 624-8301 or (304) 293-4191 and ask for item #DWPKPE71. (Additional copies of fact sheets are free; however, postal charges may be added.) You can also order copies of one or all of the free Tech Briefs listed below. Tech Brief: Organics Removal, item #DWBLPE59 Tech Brief: Package Plants, item #DWBLPE63 Tech Brief: Water Treatment Plant Residuals Management, item #DWBLPE65 Tech Brief: Lime Softening, item #DWBLPE67 Tech Brief: Iron and Manganese Removal, item #DWBLPE70 Water Conservation Measures Fact Sheet, item #DWBLPE74 Tech Brief: Membrane Filtration, item #DWBLPE81 Tech Brief: Treatment Technologies for Small Drinking Water Systems, item #DWPSPE82 Tech Brief: Ozone, item #DWBLPE84 Tech Brief: Radionuclides, item #DWBLPE84 Tech Brief: Slow Sand Filtration, item #DWBLPE99 Tech Brief: Ultraviolet Disinfection, item #DWBLPE101 Tech Brief: Leak Detection and Water Loss Control, item #DWBLPE102 Tech Brief: Diatomaceous Earth Filtration for Drinking Water, item #DWBLPE108 Tech Brief: Reservoirs, Towers, and Tanks, item #DWFSOM15 Tech Brief: System Control and Data Acquisition (SCADA), item #DWFSOM20 Tech Brief: Valves, item #DWFSOM21 Tech Brief: Water Quality in Distribution Systems, item #DWFSOM25 Tech Brief: Water Treatment Plant Residuals Management, item #DWBLPE65

MOLECULAR-BASED DETECTION OF CRYPTOSPORIDIUM PARVUM IN WATER NICHOLAS J. POKORNY CHRISTINE M. CAREY JEANINE L. BOULTER-BITZER HUNG LEE JACK T. TREVORS

NDWC Engineering Scientist Vipin Bhardwaj has a bachelor’s degree in chemical engineering, and master’s degrees in environmental engineering and agriculture from West Virginia University.

University of Guelph Guelph, Ontario, Canada

HAVE YOU READ ALL OF OUR TECH BRIEFS? Tech Briefs, drinking water treatment and supply fact sheets, have been a regular feature in the National Drinking Water Clearinghouse (NDWC) publication On Tap for more than seven years. A free package of Tech Briefs is available as a product. A three-ring binder holds all the current Tech Briefs in print. New selections can easily be added to

INTRODUCTION Cryptosporidium is an apicomplexan protozoan parasite that is becoming increasingly recognized as an important infectious pathogen in water. The parasite is transmitted by the fecal–oral route, usually by ingestion of oocystcontaminated water or food (1–3). A C. parvum oocyst is

MOLECULAR-BASED DETECTION OF CRYPTOSPORIDIUM PARVUM IN WATER

small, averaging 5 µm × 4.5 µm, and its coat is bilayered and highly proteinaceous, allowing it to resist chlorinebased disinfection in water treatment facilities (4). Within the oocyst are four motile sporozoites (Fig. 1), which are the infectious forms of Cryptosporidium. After passing through the stomach, the oocyst releases the infectious sporozoites, which attach to epithelial cells of the small intestine. Asexual and sexual reproduction of Cryptosporidium within the infected host results in autoinfection and self-perpetuation of the disease. Sexual reproduction also results in the production of infectious oocysts that are excreted along with fecal matter into the environment. Cryptosporidiosis, the globally occurring disease caused by Cryptosporidium infection, can result in gastrointestinal distress with copious amounts of diarrhea, abdominal cramping, and fever. In healthy individuals, infection is an acute but self-limiting illness that generally lasts 1 to 2 weeks. However, in immunocompromised individuals, symptoms are more severe, and the illness may become chronic and can result in death. It has been recognized that two distinct C. parvum genotypes, referred to as the human genotype (genotype I or genotype H) and the cattle genotype (genotype II or genotype C), are responsible for human cryptosporidiosis. Morgan-Ryan et al. (5) proposed a new species, Cryptosporidium hominis, to denote the human genotype. Cryptosporidium hominis infects humans primarily, whereas C. parvum infects both humans and numerous other mammalian hosts such as mice, dogs, cattle, sheep, goats, and pigs. There is no cure therapy currently available for Cryptosporidium infection, so prevention is the key to containing and managing this disease. Prevention requires accurate monitoring and detection of the parasites in water and food samples. As intestinal parasites, C. parvum cannot be cultured in vitro. Therefore, traditional culture methods used in routine microbiology labs are not suitable to detect and enumerate C. parvum oocysts. Classical methods to detect Cryptosporidium are based on microscopy of

Figure 1. A scanning electron micrograph of a C. parvum oocyst. The four crescent-shaped sporozoites are visible within the oocyst. The deflated appearance of the oocyst is the result of dehydration required for sample preparation.

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clinical samples that can be combined with different stains for improved visualization. However, microscopy requires training and is subject to human error in both sample preparation and viewing. The use of fluorogenic dyes can make detection easier, but microscopy cannot indicate oocyst viability or infectivity. Rather than relying on the results of clinical samples to identify a Cryptosporidium outbreak, recent research has focused on analyzing water samples for oocysts. The U.S. EPA has developed a method that is recommended for detecting Cryptosporidium in raw and treated waters. This method requires the concentration of a water sample, immunomagnetic separation (IMS) of the oocysts from the concentrated debris, and determination of oocyst concentrations by immunofluorescence assay (IFA). Potential oocysts are further stained with vital dyes such as DAPI and propidium iodide, followed by differential interference contrast microscopy. The number of oocysts within water sources is generally low and variable, so large volumes of water need to be concentrated for quantitative analysis. This can be accomplished by filtration, flocculation, or centrifugation. Oocysts are then purified from background detrital material by flotation, immunomagnetic separation, or flow cytometry. The concentrated water samples can be screened by molecular techniques to determine the presence of oocysts. A desired detection method should differentiate between dead or nonviable oocysts, which pose no threat to public health, and those oocysts that are viable and infective. Because only a few oocysts are needed to cause disease, efficient detection methods should be sensitive and amenable to quantification. Additionally, the detection method should differentiate between the Cryptosporidium species because not all species are harmful to humans. Molecular detection techniques have the advantage of being more specific, sensitive, reproducible, and often more rapid. DETECTION OF CRYPTOSPORIDIUM INFECTIVITY Although human infectivity assays are true representations of the disease, they are not practical for routine evaluations of oocyst infectivity. The most direct method for assessing oocyst viability and infectivity is to administer the oocysts experimentally to an animal (typically a neonatal mouse). Following a period of incubation, the intestine is sectioned longitudinally and the cells are examined for histological evidence of infection. Commonly used animal surrogates are neonatal or immunosuppressed rodents, typically CD-1 or BALB/c strains. In addition to ethical concerns, animal infectivity experiments are time-consuming, expensive, and not amenable to routine environmental testing. Alternatively, in vitro infectivity involves exposing oocysts to excystation stimuli followed by inoculating them into a cultured adherent mammalian intestinal cell line, which supports infection and asexual development. Following a suitable infection period, samples are examined by fluorogenically labeled antibodies or nucleic acid sequences specific to Cryptosporidium. Based on the presence or absence of infection, the number of infective

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oocysts can be statistically determined according to a standardized most-probable number (MPN) table or MPN calculator (6). The success of an in vitro cell culture assay depends on the type of cell line, the use of oocysts or sporozoites, preincubation excystation treatment, and centrifugation of the sample. Although cell culture requires intensive labor to grow and maintain the cell monolayer and a prolonged incubation period, the technique remains a good substitute for animal infectivity models. A number of cell lines can support the asexual development of C. parvum, but human ileocecal adenocarcinoma (HCT-8) cells that have been found superior in their ability to support parasite growth and represent a practical and accurate alternative to animal infectivity models, yield results similar to those with CD-1 mice in assessing C. parvum infectivity (7). ANTIBODY TECHNOLOGY Antibodies are produced by immune systems that bind to foreign antigens within the host organism. Strong target affinity, specificity, and sensitivity of antibodies make them ideal for targeting and identifying Cryptosporidium. Various types of antibodies can be used for molecular diagnostic methods: monoclonal antibodies (mAbs), polyclonal antibodies (pAbs), and recombinant antibodies (rAbs). Monoclonal antibodies are identical because they were produced by one type of immune cell, all clones of a single parent cell. In contrast, pAbs are produced by different immune cells and differ from one another. Recombinant antibodies are constructed within a laboratory and are based on the artificial random recombination of antibody genes. This may yield antibodies that may not occur naturally. Conjugation of antibodies to fluorophores or enzymes allow for the production of a visual signal. Many immunologic methods are available for detecting C. parvum, including flow cytometry, immunofluorescence assays (IFA), and enzyme-linked immunosorbent assays (ELISA). Immunofluorescence Assay (IFA) Routine detection of Cryptosporidium in environmental samples is most commonly performed using the IFA. In detecting the parasite, the characteristics used for identification include size, shape, and fluorescence. The cells become fluorescent via antibody staining, which is carried out either in suspension, on a microscope slide, or on a membrane prior to microscopic examination. Both fluorophore-conjugated pAbs and mAbs are used to identify purified oocysts. Nonspecific fluorescence is a problem with IFAs, along with background fluorescence of detritus, algae, and some freshwater diatoms and cyanobacteria. Foci Detection Method (FDM) An in vitro cell culture infectivity assay involves exposing oocysts to excystation stimuli followed by inoculating them into a cultured mammalian cell line, which supports infection and asexual development. An infectious focus

is the site of infection of at least one sporozoite in the cell culture. After 24–48 h, samples are examined for the presence of infection. Foci are labeled with fluorophore-conjugated antibodies specific to the various life stages of Cryptosporidium and are visualized by epifluoresence microscopy. Enzyme-Linked Immunosorbent Assay (ELISA) ELISA has been developed to detect and evaluate the growth of C. parvum in culture systems using antibodies developed against its various life-cycle stages. ELISA can be used to test multiple samples simultaneously and does not require a high level of technical skill compared with that for identifying parasites based on morphological and staining characteristics by microscopic examination. Tests commonly use 96-well microtitration plates coated by antibodies specific for Cryptosporidium. Tests samples are then added to the plate. After washing away nonbinding materials, a second enzyme-conjugated antibody specific to Cryptosporidium is added to the sample wells. Addition of the enzyme substrate results in a color reaction to indicate the presence of the parasite. There may be problems of cross-reactivity with algae and turbidity interfering with results, but this can be overcome by the development and use of more specific antibodies. Flow Cytometry A flow cytometer analyzes particles in a suspension as they pass by a laser beam. The light scattering pattern is analyzed and correlated to size and internal complexity along with the fluorescent light emitted by each particle. Flow cytometer cell sorters can also sort particles of interest from unwanted particles by using the binding of a fluorescein-conjugated antibody to antigens present on Cryptosporidium oocysts. Particles within the sample that do not match the criteria for Cryptosporidium are filtered out. POLYMERASE CHAIN REACTION (PCR) DETECTION METHODS PCR involves using oligonucleotide primers to amplify a DNA fragment specific to the target organism. The desired fragment is identified by gel electrophoresis and can be subsequently sequenced. PCR is reproducible, cost-effective, sensitive and amenable to quantitation, and capable of differentiating among Cryptosporidium species that infect humans. Common gene targets for PCR detection include 18S ribosomal DNA, heat-shock proteins (i.e., hsp70), and oocyst wall proteins. Real-Time PCR Real-time PCR involves virtual real-time visualization of fluorescence emitted by a fluorogenic probe accumulated during PCR. Quantification of amplified DNA during the exponential phase of the reaction, when reagents are not limiting, allows precise determination of the initial quantity of the target sequences. Real-time PCR offers higher throughput, reduced turnaround time, and minimal

MOLECULAR-BASED DETECTION OF CRYPTOSPORIDIUM PARVUM IN WATER

amplicon contamination due to a closed-vessel system. The quantification range of real-time PCR methods is greater (5–6 log units) than conventional PCR (2–3 log units). Cryptosporidium species and genotypes can be differentiated through melt curve analysis, which is based on melting temperature differences of PCR-probe complexes, and reflects the extent of complementation of the probes to the amplified PCR fragments. Real-time PCR investigations for environmental detection have demonstrated that the method is rapid, sensitive, and specific. Reverse Transcription-PCR (RT-PCR) Only viable organisms can produce messenger RNA (mRNA), so RT-PCR selectively detects viable organisms. In this method, a specific fragment of the complementary DNA (cDNA), which is produced from an mRNA template by reverse transcriptase, is amplified by PCR. Reverse transcriptase is extremely sensitive to chemical contamination, so the RT step is the source of greatest variability. Common gene targets for RT-PCR analysis of C. parvum viability include hsp70, amyloglucosidase, and B-tubulin. Cell Culture-PCR (CC-PCR) Comparable with FDM in detecting oocyst infectivity, a CC-PCR assay involves the extraction of genomic DNA from an infected cell culture and amplification of specific DNA fragments by PCR. A cell culture infectivity assay, coupled with either real-time PCR or an MPN assay, can be used to quantify infective oocysts. Alternatively, mRNA can be extracted from the infected cell culture and subjected to RT-PCR to detect infectious oocysts. Most Probable Number-PCR (MPN-PCR) In a standard MPN assay, replicate samples are serially diluted and evaluated by PCR for the presence of Cryptosporidium-specific nucleic acid sequences. Used in

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conjunction with a cell culture infectivity assay, MPNPCR could provide a good alternative to using FDM alone for enumerating infective oocysts. Because MPN assays are costly, labor-intensive, and require a large volume of samples with multiple preparation steps, alternative detection strategies may be more suitable for quantitative analyses. NUCLEIC ACID PROBES Small nucleotide chains (oligonucleotides) can be constructed to a desired DNA sequence. The binding nature of DNA is such that complementary single-stranded DNA (ssDNA) fragments strongly bind to one another, whereas a single nucleotide mismatch drastically reduces the binding strength of the two fragments. Fluorescent In Situ Hybridization (FISH) FISH uses fluorescently labeled oligonucleotide probes targeted to C. parvum-specific sequences, such as ribosomal RNA (rRNA). Ribosomal RNA is used as a target because it should be present in high quantities in viable and potentially infective oocysts, but due to its short halflife, it would be degraded in nonviable oocysts. Oocysts are collected and made permeable to the labeled probes by the addition of organic solvents. Species-specific probes hybridize with target rRNA within the permeabilized oocysts and are detected by epifluorescence microscopy or flow cytometry. Microarray (DNA Chip) Microarray technology combines PCR with hybridization of oligonucleotide probes. An array of oligonucleotide probes is synthesized and fixed to a chip. Oocyst genomic DNA is extracted and highly specific, and conserved target genes, such as rRNA and hsp70, are amplified by PCR and then fluorescently labeled. For

Table 1. Summary of the Molecular Detection Methods That Can Be Used for Detecting Cryptosporidium parvum in Environmental Samples

Detection Method IFA FDM—animal FDM—cell culture FDM-MPN PCR Real-time PCR Real-time CC-PCR RT-PCR CC-RT-PCR CC-PCR MPN-PCR Flow cytometry FISH Microarray a

Differentiate Among Cryptosporidium Species

Determine Infectivity

Quantifiable

Detection Limita

− − − − + + + + + + + − + +

− + + + − − + + + + − − − −

+ + + + − + + − − − + + + +

10 oocysts 60 oocysts 5–10 oocysts 2–10 oocysts 1–10 oocysts 1–5 oocysts 1 oocyst 1 oocyst 10 oocysts 10 oocysts 10 oocysts 10 oocysts 100 oocysts Not available

Detection limits are not directly comparable as the conditions used by different researchers may not be similar.

Reference (8) (10) (10) (6,10) (18) (11) (12) (13) (14) (15,16) (17) (18) (8) (19,20)

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identification at the species level, highly variable or species-specific genes are used. PCR products hybridize with their complementary probe sequences, and the presence or absence of Cryptosporidium can be determined by computerized detection of fluorescent emissions. The high throughout capability of microarrays can allow simultaneous detection of numerous different genera of pathogens or species within a genus. A complete list of the methods described is presented in Table 1. The table indicates whether each method can distinguish among Cryptosporidium species, detect oocyst infectivity, quantify detected oocysts, and has a limitation on detection. A reference for each of the methods is also included in the table. A recent review article by Carey et al. (22) provides a more detailed description of Cryptosporidium detection methods. Acknowledgment This research was supported by the Canadian Water Network (CWN) Centre of Excellence.

BIBLIOGRAPHY 1. Carreno, R.A. et al. (2001). Phenotypic and genotypic characterization of Cryptosporidium species and isolate. J. Ind. Microbiol. Biotechnol. 226: 95–106. 2. Millard, P.S. et al. (1994). An outbreak of cryptosporidiosis from fresh-pressed apple cider. JAMA 272(20): 1592–1596. 3. Smith, H.V. (1998). Detection of parasites in the environment. Parasitology 117(7): S113–S141. 4. Korich, D.G. et al. (1990). Effects of ozone, chlorine dioxide, chlorine and monochloramine on Cryptosporidium parvum oocyst viability. Appl. Environ. Microbiol. 56(5): 1423–1428. 5. Morgan-Ryan et al. 6. Weir, S.C. et al. (2001). Improving the rate of infectivity of Cryptosporidium parvum oocysts in cell culture using centrifugation. J. Parasitol. 87(6): 1502–1504. 7. Higgins, J.A. et al. (2003). Recovery and detection of Cryptosporidium parvum oocysts from water samples using continuous flow centrifugation. Water Res. 37(15): 3551–3560. 8. Rochelle, P.A. et al. (2002). Comparison of in vitro cell culture and a mouse assay for measuring infectivity of Cryptosporidium parvum. Appl. Environ. Microbiol. 68(8): 3809–3817. 9. U.S. Environmental Protection Agency. (1999). Method 1622: Cryptosporidium in water by filtration/IMS/FA. Office of Water EPA-821-R-99-061. 10. Slifko, T.R. et al. (2002). Comparison of tissue culture and animal models for assessment of Cryptosporidium parvum infection. Exp. Parasitol. 101(2–3): 97–106. 11. Johnson, D.W. et al. (1995). Development of a PCR protocol for sensitive detection of Cryptosporidium oocysts in water samples. Appl. Environ. Microbiol. 61(11): 3849–3855. 12. Limor, J.R., Lal, A.A., and Xiao, L. (2002). Detection and differentiation of Cryptosporidium parasites that are pathogenic for humans by real-time PCR. J. Clin. Microbiol. 40(7): 2335–2338. 13. MacDonald, L.M. et al. (2002). The development of a time quantitative-PCR method for characterisation Cryptosporidium parvum in vitro culturing system assessment of drug efficacy. Mol. Biochem. Parasitol. 279–282.

realof a and 121:

14. Stinear, T., Matusan, A., Hines, K., and Sandery, M. (1996). Detection of a single viable Cryptosporidium parvum oocyst in environmental water concentrates by reverse transcriptionPCR. Appl. Environ. Microbiol. 62(9): 3385–3390. 15. Rochelle, P.A., De Leon, R., Stewart, M.H., and Wolfe, R.L. (1997). Comparison of primers and optimization of PCR conditions for detection of Cryptosporidium parvum and Giardia lamblia in water. Appl. Environ. Microbiol. 63(1): 106–114. 16. Di Giovanni, G.D. et al. (1999). Detection of infectious Cryptosporidium parvum oocysts in surface and filter backwash water samples by immunomagnetic separation and integrated cell culture-PCR. Appl. Environ. Microbiol. 65(8): 3427–3432. 17. LeChevallier, M.W., Abbaszadegan, M., and Di Giovanni, G.D. (2000). Detection of infectious Cryptosporidium parvum oocysts in environmental water samples using an integrated cell culture-PCR (CC-PCR) system. Water Air Soil Pollut. 123(1/4): 53–65. 18. Tsuchihashi, R., Loge, F.J., and Darby, J.L. (2003). Detection of Cryptosporidium parvum in secondary effluents using a most probable number-polymerase chain reaction assay. Water Environ. Res. 75(4): 292–299. 19. Chung, J., Vesey, G., Gauci, M., and Ashbolt, N.J. (2004). Fluorescence resonance energy transfer (FRET)-based specific labeling of Cryptosporidium oocysts for detection in environmental samples. Cytometry 60(A): 97–106. 20. Straub, T.M. et al. (2002). Genotyping Cryptosporidium parvum with an hsp70 single-nucleotide polymorphism microarray. Appl. Environ. Microbiol. 68(4): 1817–1826. 21. Wang, Z., Vora, G.J., and Stenger, D.A. (2004). Detection and genotyping of Entamoeba histolytica, Entamoeba dispar, Giardia lamblia, and Cryptosporidium parvum by oligonucleotide microarray. J. Clin. Microbiol. 42(7): 3262–3271. 22. Carey, C.M., Lee, H., and Trevors, J.T. (2004). Biology, persistence and detection of Cryptosporidium parvum and Cryptosporidium hominis oocyst. Water Res. 38: 818– 862.

CRYPTOSPORIDIUM JACQUELINE BRABANTS University of New Hampshire Durham, New Hampshire

Cryptosporidia are small, spherical, obligate, intracellular sporozoan parasites that infect the intestinal tract of a wide range of mammals, including humans. Cryptosporidium oocysts range in size from 2 to 8 µm, depending on the species and the stage of the life cycle. Oocysts are commonly found in many of the lakes and rivers that supply public drinking water as a result of runoff from sewage and animal wastes applied to nearby fields and pastures or from areas of wildlife or livestock activity (4,7,10). Infection, resulting from ingesting oocysts, manifests as cryptosporidiosis. Although Cryptosporidium was first described in 1907, human infection was not reported until 1976. Today, Cryptosporidium parvum is well recognized as the cause of the disease cryptosporidiosis in humans and cattle and is one of the more opportunistic agents seen in patients with AIDS that occurs primarily in individuals

CRYPTOSPORIDIUM

with compromised immune systems. Currently, there is no totally effective therapy for cryptosporidiosis (3,8,11,12). LIFE CYCLE AND MORPHOLOGY Cryptosporidia undergo alternating life cycles of sexual and asexual reproduction that are completed within the gastrointestinal tract of a single host. The developmental stages of the life cycle occur intracellularly and extracytoplasmically and include schizogony, gametogony, fertilization, and sporogony (1). The developmental stages of the organism are contained within a host cell parasitophorous vacuole, located at the microvillous surface of the host cell. The cycle begins when infectious oocysts containing four sporozoites are discharged in the feces of a parasitized animal. These thick-walled oocysts remain viable for months unless exposed to extremes of temperature, desiccation, or concentrated disinfectants (1). Following ingestion by another animal, most likely from food or water that has become fecally contaminated, the oocyst excysts and releases sporozoites that attach to the microvilli of the small bowel epithelial cells, where they develop into trophozoites (2). Trophozoites divide asexually (schizogony) to form schizonts that contain eight daughter cells known as type I merozoites. Upon release from the schizont, these cells attach to another epithelial cell, and the schizogony cycle is repeated to produce schizonts that contain four type II merozoites. Type II merozoites develop into male (microgametocyte) and female (macrogametocyte) sexual forms. Fertilization results in a zygote (oocyte) that develops into an oocyst, which is ultimately shed into the lumen of the bowel. The oocysts undergo sporulation to the infective stage within the brush border of the enterocytes and are excreted as infectious oocysts in the stool (3–5). The majority of oocysts generated possess a thick, protective cell wall that ensures their intact passage in the feces and survival in the environment; however, approximately 20% of the oocysts generated fail to develop the thick wall, and following release from a host cell, the thin cell membrane on these oocysts ruptures and releases four infectious sporozoites. These sporozoites penetrate the intestinal lumen and initiate a new autoinfective cycle within the original host. The presence of this thin-walled, autoinfective oocyst can lead to an overwhelming infection that creates a persistent, life-threatening infection in an immunocompromised individual (4). CLINICAL DISEASE The pathogenesis of Cryptosporidia is not completely understood; age and immune status at the time of primary exposure do not appear to influence susceptibility to infection. However, once the primary infection has been established, the immune status of the host plays an extremely important role in determining the length and severity of the illness (4). Cryptosporidia undergo their life cycle in the enteric epithelial cells and also in the gallbladder, respiratory,

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and renal epithelium, especially in immunocompromised hosts (1). The symptom found in all reported cases is acute diarrhea (2). Clinical symptoms found in immunocompetent individuals include nausea; low-grade fever; abdominal cramps; anorexia; and profuse watery, frothy bowel movements that may be followed by constipation. Other individuals may be asymptomatic, particularly later in the course of the infection. In patients who have the typical watery diarrhea, the stool specimen will contain very little fecal material, consisting mainly of water and mucus flecks, and the organisms are trapped in the mucus. In most cases, a patient who has a normal immune system will have a self-limited infection lasting 1 to 2 weeks, whereas a patient who is immunocompromised may have a chronic infection with symptoms ranging from asymptomatic to severe (2,4). Studies to examine susceptibility and serologic responses to reinfection have demonstrated that previous exposure of immunocompetent adults to Cryptosporidium is not entirely protective but may decrease the severity of disease and the number of oocysts shed (6). Cryptosporidium can generate life-threatening infections in immunocompromised individuals, particularly in human AIDS patients (1). In these individuals, the duration and severity of diarrheal illness will depend on the immune status of the patient. It is believed that cryptosporidiosis in AIDS patients causes malabsorption and intestinal injury in proportion to the number of organisms that infect the intestine. Most severely immunocompromised patients cannot overcome the infection; the illness becomes progressively worse with time and leads to death. The disease is prolonged; profuse, watery diarrhea persists from several weeks to months or years, as a result of the autoinfective nature of the organism, reportedly resulting in fluid losses as high as 25 L/day (7). In such patients, infections in areas other than the gastrointestinal tract may cause additional symptoms such as respiratory problems, cholecystitis, hepatitis, and pancreatitis (4). Diagnosis Oocysts recovered in clinical specimens usually represent the 80% that are thick-walled. The oocysts are difficult to visualize because of their small, colorless, transparent appearance, and may be confused with yeast cells. In the past, cryptosporidiosis has been diagnosed following examination of small or large bowel biopsy material, under both light and electron microscopy; however, in Cryptosporidium infections, the entire mucosa may not be infected uniformly; therefore, biopsy specimens may miss the infected area. As a result, cases have recently been diagnosed by recovering the oocysts from fecal material using flotation or fecal concentration techniques. Diagnosis is achieved by demonstrating the oocysts in feces, sputum, or possibly respiratory secretions (3). Special staining techniques such as the modified acidfast, Kinyoun’s, and Giemsa methods may be employed to enhance visualization, along with the direct fluorescentantibody (FA) or enzyme-linked immunosorbent assay (ELISA) techniques that incorporate monoclonal antibody reagents (4).

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Recently, the issues of water quality and water testing have become more important and controversial. Today, most water utilities have developed their own water quality testing laboratories or contracted with commercial water laboratories for the recovery and identification of Cryptosporidium. Currently, the most common methods for capturing and recovering oocysts from water employ polypropylene cartridge filters or membrane filtration. The oocysts are subsequently eluted and may be concentrated on a percoll-sucrose gradient by flotation or by immunomagnetic separation (IMS) and visualized with commercially available immunofluorescence assay kits. Oocysts are often present in small numbers in environmental samples; therefore, molecular techniques involving the polymerase chain reaction (PCR) may also be used to detect oocysts in water samples (2,8). Integrated cell culture-PCR infectivity assays address the drawbacks of alternative methods such as vital dye staining by allowing for both detection of organisms and the determination of viability and infectivity (9). Considering that, under favorable conditions of moisture and moderate temperature, oocysts can remain viable and infectious for a relatively long time and have been reported viable after storage for 12 months, a great need still exists for simple, efficient, and reliable procedures for capturing and recovering Cryptosporidium oocysts from water (10) (Fig. 1). Treatment There is no effective specific treatment for Cryptosporidium infection, despite testing of hundreds of compounds (1,6). Cryptosporidiosis tends to be self-limiting in patients who have an intact immune system; the clinical course of infection varies, depending on the immune status of the host. Treatment with antidiarrheal drugs along with rehydration therapy may reduce the severity of acute cryptosporidiosis but is less effective for chronic cryptosporidiosis that involves the colon and extraintestinal tissues (4). The antibiotic paromomycin, it has been

Figure 1. Cryptosporidium parvum oocysts viewed at 1000× magnification by Nomarski differential interference contrast (DIC) magnification.

shown, slightly reduces parasite numbers and stool frequency and may be combined with azithromycin as a course of treatment (6,11). The establishment of a parasitophorous vacuole within the host cell may somehow protect the parasite from antimicrobial drugs (6). Epidemiology and Prevention Cryptosporidium is transmitted by oocysts that are usually fully sporulated and infective when they are passed in stool. The principal transmission route is direct fecal–oral spread and transmission by contaminated water. Calves and other animals such as livestock, dogs, cats, and wild mammals are potential sources of human infections, and contact with these animals or their feces may be an unrecognized cause of gastroenteritis in humans (1). Generally, young children tend to have higher infection rates, and there is a high prevalence of cryptosporidiosis in children in areas where sanitation and nutrition are poor (10). Direct person-to-person transmission is likely and may occur through direct or indirect contact with stool material. Outbreaks of human disease in daycare centers, hospitals, and urban family groups indicate that most human infections result from person-to-person contact (5). Indirect transmission may occur from exposure to positive specimens in a laboratory, from contaminated surfaces, or from consuming contaminated food or water. In healthy adults who have no serologic evidence of past infection by Cryptosporidium parvum, as few as thirty Cryptosporidium parvum oocysts is sufficient to cause infection, and in some cases, infection has occurred from just one oocyst (11). In the United States, the parasite has been identified in 15% of patients who have AIDS and diarrhea (5). The potential contamination of water supplies by Cryptosporidium oocysts is a considerable issue for the drinking water industry. Oocysts can penetrate physical barriers and withstand the conventional disinfection processes used for drinking water treatment. Waterborne outbreaks of Cryptosporidium are an increasing public health problem and have resulted from untreated surface water, filtered public water supplies, and contaminated well water (3,12). Large-scale outbreaks of cryptosporidiosis in industrialized countries have been associated with contamination of community drinking water (9). Disease transmission through the waterborne route is especially important because of the capacity for affecting large communities of susceptible individuals. A massive waterborne outbreak was reported in Milwaukee, Wisconsin, where contamination of the public water supply during March and April of 1993 resulted in more than 400,000 infections and about 50 deaths (12,13). Additional outbreaks involving public swimming pools and wading pools demonstrate the ability of Cryptosporidium to cause infection even when ingested in small amounts of fully chlorinated water (11). The increase in the number of reported waterborne disease outbreaks associated with the Cryptosporidium species can be attributed to improved techniques for oocyst recovery and identification resulting in the demonstration of oocysts in surface and drinking water and in sewage effluents. It is very likely that cryptosporidiosis is

MEASURING CRYPTOSPORIDIUM PARVUM OOCYST INACTIVATION FOLLOWING DISINFECTION

underdiagnosed, especially in immunocompetent adults and children, as analysis for the oocysts is not normally included in a routine stool analysis (6). The importance of agricultural wastewater and runoff, particularly from lambs and calves, is also now recognized as a potential source of infective Cryptosporidium oocysts (4). Prevention involves taking proper steps to reduce the likelihood of waterborne contamination. Properly drilled and maintained wells that tap into groundwater are unlikely to contain pathogens because of the natural filtration that takes place as water passes through the soil; however, contamination may still occur if surface water can move through coarse soils or fractured bedrock into groundwater aquifers. Shallow or poorly constructed wells and springs are at risk of contamination from surface water runoff; therefore, wells should be protected from surface contamination by an intact well casing, proper seals, and a cap above ground (14). Human and animal waste contamination are minimized by protecting the watershed, controlling land use, creating and enforcing septic system regulations, and best management practices in an effort to control runoff (14). Cryptosporidium oocysts are susceptible to ammonia, 10% formalin in saline, freeze-drying, exposure to temperatures below freezing or above 65 ◦ C for 30 minutes, and 50% commercial bleach (4). For individuals who wish to take extra measures to avoid waterborne cryptosporidiosis in their drinking water, according to the EPA and the CDC, boiling the water is the most effective way of killing the organism (15). In addition to boiling water, oocysts can be removed by certain types of filters to ensure that drinking water is safe (14). BIBLIOGRAPHY 1. Bowman, D. (1999). Cryptosporidium, Georgis’ Parasitology for Veterinarians, 7th Edn. W.B. Saunders Company, New York, pp. 99–100. 2. Leventhal, R. and Cheadle, R. (1989). Protozoa, Medical Parasitology: A Self Instructional Text, 3rd Edn. F.A. Davis Company, Philadelphia, PA, pp. 96–97. 3. Ash, L. and Orihel, T. (1997). Cryptosporidium parvum. In: Atlas of Human Parasitology, 4th Edn. American Society of Clinical Pathologists, Chicago, IL, pp. 124–125. 4. Garcia, L. and Bruckner, D. (1997). Intestinal Protozoa (Coccidia and Microsporidia) and Algae. In: Diagnostic Medical Parasitology, 3rd Edn. ASM Press, Washington, DC, pp. 54–66. 5. Ryan, K.J. (1994). Cryptosporidia. In: Sherris Medical Microbiology: An Introduction to Infectious Diseases, 3rd Edn. Appleton & Lange, Norwalk, CT, pp. 655–657. 6. Clark, D. (1999). New insights into human cryptosporidiosis. Clin. Microbiol. Rev. 12(4): 554–563. 7. Balows, A., Hausler, W., Herrmann, Jr., K., Isenberg, H., and Shadomy, H. (1991). Cryptosporidium species. Manual of Clinical Microbiology, 5th Edn. ASM, Washington, DC, pp. 764–766. 8. Widmer, G., Orbacz, E., and Tzipori, S. (1999). β-Tubulin mRNA as a marker of Cryptosporidium parvum oocyst viability. Appl. Environ. Microbiol. 65(4): 1584–1588. 9. DiGiovanni, G. et al. (1999). Detection of infectious Cryptosporidium parvum oocysts in surface and filter backwash

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water samples by immunomagnetic separation and integrated cell culture-PCR. Appl. Environ. Microbiol. 65(8): 3427–3432. 10. Dubey, J.P., Speer, C.A., and Fayer, R. (1990). Cryptosporidiosis of Man and Animals. CRC Press, Boca Raton, FL, pp. 1–82. 11. Guerrant, R. (1997). Cryptosporidiosis: an emerging, highly infectious threat. Emerging Infect. Dis. 3(1): 51–57. 12. Sartory, D.P. et al. (1998). Recovery of Cryptosporidium oocysts from small and large volume water samples using a compressed foam filter system. Lett. Appl. Microbiol. 27: 318–322. 13. Sreter, T. et al. (2000). Morphologic, host specificity, and molecular characterization of a Hungarian Cryptosporidium meleagridis isolate. Appl. Environ. Microbiol. 66(2): 735–738. 14. Avery, B.K. and Lemley, A. (1996). Cryptosporidium: A Waterborne Pathogen. U.S. Department of Agriculture Water Quality Program. 329WQFS6. 15. Guidance for people with severely weakened immune systems, United States Environmental Protection Agency, Office of Water, 1999, EPA 816-F-99-005.

MEASURING CRYPTOSPORIDIUM PARVUM OOCYST INACTIVATION FOLLOWING DISINFECTION WITH ULTRAVIOLET LIGHT ZIA BUKHARI American Water Voorhees, New Jersey

Historically, ultraviolet light inactivation of Cryptosporidium parvum oocysts was considered ineffective; however, recently it was demonstrated that the methods for measuring oocyst inactivation can yield erroneous results and that neonatal mouse infectivity assays, which indicated that very low UV doses are highly effective for oocyst inactivation, are needed for determining inactivation of UV-treated oocysts. The moral, ethical, and financial constraints of using mouse infectivity has generated the need for more user-friendly alternative methods for measuring oocyst inactivation. Cell culture infectivity assays are considered promising alternatives, and this article discusses a cell culture-immunofluorescence (IFA) procedure, which following optimization, yielded results similar to those expected from mouse infectivity assays. This cell cultureIFA procedure will be an invaluable analytical tool for control or bench scale studies using C. parvum oocysts and the same water matrices as those intended for use in UV reactor validation studies. HISTORICAL PERSPECTIVE A number of investigators have examined the effect of low-pressure UV light on viruses and bacteria (1,2) and found that MS2 bacteriophage requires approximately 70 mJ/cm2 of UV light to render 4-log inactivation and that hepatitis A virus requires fourfold lower UV doses to yield similar levels of inactivation. From the studies of Wilson et al. (1), MS2 bacteriophage was approximately two times more resistant than viruses

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and three to ten times more resistant than bacteria. Protozoan parasites, which display greater resistance than bacteria or viruses to chemical disinfectants, were examined for their susceptibility to UV light by Karanis et al. (3). Reportedly delivery of 400 mJ/cm2 yielded 3log inactivation of Trichomonas vaginalis, and Giardia lamblia required 180 mJ/cm2 to yield 2-log inactivation. In 1993, Lorenzo-Lorenzo et al. (4) examined the impact of UV light on Cryptosporidium oocysts; however, due to an inadequate description of their disinfection experiments, the effectiveness of UV light for Cryptosporidium oocyst inactivation was not recognized. Two years later, a unit known as the Cryptosporidium inactivation device (CID), which delivered a total UV dose of 8748 mJ/cm2 from lowpressure lamps, was examined and 2- to 3-log inactivation of oocysts was reported using the fluorogenic vital dye assay (4 -6 -diamidino-2-phenylindole and propidium iodide) and in vitro excystation (5). Following this, Clancy et al. (6) confirmed the findings of Campbell et al. (5); however, the apparent requirement for high UV doses for significant oocyst inactivation continued to present an obstacle to the regulatory and water industry in supporting implementation of this technology. From 1998 to 1999, comparative bench scale and demonstration scale (215 gpm) Cryptosporidium oocyst inactivation studies were performed that used mediumpressure UV lamps and examined oocyst inactivation using in vitro viability assays (DAPI/PI and in vitro excystation) as well as mouse infectivity assays (7). These studies conclusively demonstrated the effectiveness of low UV doses for inactivating Cryptosporidium in finished water. These data rapidly gained attention from both regulatory agencies and the water industry and instigated the revolution in the U.S. water industry for UV disinfection to inactivate Cryptosporidium oocysts effectively. In addition to demonstrating the effectiveness of low levels of UV light for oocyst inactivation, Bukhari et al. (7) also found that in vitro viability assays underestimated oocyst inactivation following their exposure to either UV light or ozonation (8) and that mouse infectivity assays, albeit cumbersome, needed to be the methods of choice for future studies of this nature. Following these initial studies, a number of investigators confirmed the effectiveness of UV light for inactivating Cryptosporidium oocysts as well as Giardia cysts (9) and Microsporidia spores (10).

mice. Unfortunately the ‘‘gold’’ standard mouse infectivity assays have various limitations, which in addition to the moral, ethical, and financial constraints of animal experimentation, include the high degree of variability of using outbred strains of neonatal mice. As a result, the water industry has been seeking alternative, more user-friendly procedures for measuring oocyst viability/infectivity. In vitro infectivity assays for determining C. parvum oocyst inactivation have the potential to fill this void. One such in vitro infectivity assay uses human ileocecal adenocarcinoma (HCT-8) cells in conjunction with quantitative polymerase chain reaction (q-PCR) and has been used previously to determine the infectivity of environmentally derived oocysts (11). This assay has the advantage of providing a direct indication of the amount of amplifiable DNA with a specific set of primers. Furthermore, this assay also can help avoid the high variability of using the most probable number format of mouse infectivity assays. Using this assay to measure inactivation of bench-scale, UV-treated C. parvum oocysts indicated log inactivation values of 1.16, 1.24, and 1.84 logs for 10, 20, and 40 mJ/cm2 , respectively. These inactivation values for C. parvum oocysts were considerably lower than those reported in previous studies using mouse infectivity assays (7,12). Previous research has indicated that oocysts exposed to UV doses ranging between 10 and 40 mJ/cm2 continue to respond to excystation stimuli and release their sporozoites. This suggests that UV-treated oocysts inoculated onto cell monolayers could potentially excyst, and then released sporozoites could invade cell monolayers. Examination of HCT-8 monolayers confirmed the presence of pinpoints of invasion, which probably originated from invasive sporozoites (Fig. 1). A number of different oocyst pretreatment steps were used to promote differentiation between UV affected and unaffected sporozoites; however, these pretreatments did not nullify the background signal detected by the quantitative PCR procedures.

MEASURING C. PARVUM OOCYST INACTIVATION FOLLOWING EXPOSURE TO UV LIGHT In vitro assays for determining the viability of C. parvum oocysts offer several advantages over the traditional animal infectivity assays in that results can be generated in a short time. Although these assays are relatively simple to use and relatively inexpensive, recently it was shown that they do not accurately demonstrate whether oocysts are capable of infectivity in neonatal mice. For example, using fluorogenic vital dyes (i.e., DAPI/PI, SYTO-9, and SYTO-59) or in vitro excystation, the viability information for oocysts subjected to ultraviolet light (7) or ozone (8) disinfection was grossly overestimated compared with infectivity data using neonatal

Pin point invasion

Figure 1. UV-treated oocysts yielding pinpoints of invasion.

MEASURING CRYPTOSPORIDIUM PARVUM OOCYST INACTIVATION FOLLOWING DISINFECTION

CELL CULTURE-IFA FOR MEASURING INACTIVATION OF UV-TREATED C. PARVUM OOCYSTS Bukhari and LeChevallier (13) used an infectivityenhancing oocyst pretreatment step consisting of preacidification and 0.05% bile treatment followed by inoculation onto HCT-8 monolayers, incubation at 37 ◦ C for 72 h, and quantitative detection of infection by immunofluorescence microscopy. The rationale behind this assay was that UV-treated oocysts would undergo excystation and subsequently invade HCT-8 cells, leading to discrete pinpoints of invasion. In contrast, untreated organisms would invade the HCT-8 cells but would continue to differentiate further to generate clusters of secondary infection (Fig. 2). Enumerating secondary clusters of infection for various inocula of oocysts (i.e., 10,100, and 1000 oocysts) has enabled development of a dose response curve (Fig. 3). This curve, which was generated from multiple trials (n = 75 to 115) using predetermined oocyst inocula, was analyzed by linear regression to derive an equation for calculating the number of infectious organisms present in an inoculum of UV-treated oocysts after the number of clusters had been determined by immunofluorescence microscopy. UV disinfection experiments using the Iowa isolate of C. parvum oocysts were conducted, and following cell culture infectivity, the infectious clusters were extrapolated from the dose response curve generated for the untreated oocysts to calculate levels of inactivation (Fig. 4). A

Infection cluster

Figure 2. Infection clusters from infectious oocysts.

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UV dose of 1 mJ/cm2 rendered 0.44 log inactivation of oocysts, and UV doses between 1 and 4 mJ/cm2 yielded a linear decline in oocyst inactivation; 3 mJ/cm2 rendered 2.79–2.84 log inactivation, and 4-log inactivation occurred at 4 mJ/cm2 . Between 4 and 20 mJ/cm2 , measurements by the cell culture procedure continued to indicate oocyst inactivation levels of 4 logs. It is highly probable that the actual levels of oocyst inactivation after UV doses between 4 and 20 mJ/cm2 were in excess of 4 logs; however, as the determinable oocyst inactivation by the cell culture procedure is a factor of the highest original inoculum of UV-treated oocysts applied onto the monolayers (i.e., 1 × 105 oocysts), the maximum measurable levels of inactivation by this cell culture procedure were limited to 4 logs or lower in these experiments. Although theoretically it would be possible to use higher oocyst inocula (i.e., 1 × 106 –1 × 108 oocysts per monolayer) to accurately determine the levels of oocyst inactivation at individual UV doses between 4 and 20 mJ/cm2 , this would be of little empirical value from the perspective of the disinfection needs of the water industry or the logistics of conducting experiments with adequate quantities of infectious oocysts. As a result of the investigations by Bukhari and LeChevallier (13), it has been demonstrated that UVtreated C. parvum oocysts can undergo excystation and that the sporozoites from these inactivated oocysts can invade monolayers of HCT-8 cells to generate pinpoints of invasion. Optimization of excystation triggers and cell culture incubation periods led to development of a cell culture-IFA procedure that enabled detecting as few as 10 infectious oocysts. Using this cell cultureIFA allowed discriminating pinpoints (generated from noninfectious but invasive sporozoites) from secondary structures (generated from infectious sporozoites). This phenomenon is the primary reason that the cell culture quantitative PCR procedure described by DiGiovanni et al. (11) cannot be used to measure inactivation of UVtreated oocystes because the assay cannot discriminate between DNA originating from invasive sporozoites and that from infectious sporozoites. Using the optimized cell culture-IFA procedure, it was confirmed that low UV doses (i.e., 2–5 mJ/cm2 ) can be very effective for inactivating C. parvum oocysts. Comparison of data from this current study with previously published reports, using either tissue culture infectivity or mouse −5

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infectivity assays, also demonstrates excellent agreement. Numerous studies (9,12,14,15) have assessed levels of oocyst inactivation from UV doses ranging between 1 and 4-log inactivation of oocysts from UV doses ranging between 5 and 20 mJ/cm2 , >4-log oocyst inactivation was noted only when UV doses exceeded 10 mJ/cm2 in the previous studies cited. At 5 mJ/cm2 , although our data indicated >4-log inactivation, the previous studies indicated >1log and approximately 3-log inactivation of oocysts. These differences need to be interpreted with caution and do not necessarily imply differences in susceptibility between the isolates of C. parvum used. As an example, it is known that the oocyst inactivation levels can be censored as a result of the original oocyst inoculum administered in the infectivity assay, which in turn influences the outcome of the calculated levels of oocyst inactivation. In (Fig. 5), examining the inactivation data at 5 mJ/cm2 suggests that a single oocyst inactivation value (>1 log) may be an outlier. Perhaps this inactivation level was derived from administering a lower inoculum of oocysts (i.e., 1 × 103 oocysts) than the inoculum size (i.e., 1 × 105 oocysts) used in our study. Should this be the case, then it would argue that our data provide a more robust indication of the levels of oocyst inactivation than those generated from

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censored data. Furthermore, examining the general oocyst inactivation patterns for increasing UV doses in (Fig. 5) adds further credibility to the previous discussion, as data accrued during the course of our study both overlap and extend the general trend in oocyst inactivation provided by previous UV disinfection studies. This would further support the finding that oocyst inactivation levels at 5 mJ/cm2 are more likely to be greater than 3 or 4 logs. In conclusion, UV disinfection is a promising technology for rapid and effective inactivation of waterborne C. parvum oocysts. Numerous manufacturers can supply equipment using medium-pressure, low-pressure, or lowpressure, high-output lamps within their UV reactors. For water utilities to select the appropriate UV reactors for their disinfection requirements, it is imperative that reactors are validated on-site or off-site using the matrices intended for disinfection. Although the validation studies are likely to be conducted using surrogate organisms such as the MS2 bacteriophage, it is imperative to do control- or bench-scale studies also using the same matrix and C. parvum oocysts. For the latter, the cell cultureIFA procedure described in this article is an invaluable analytical tool. Acknowledgments American Water and its subsidiary utilities funded this research. The author gratefully acknowledges the technical assistance of Felicia Abrams, Janice Weihe, and Rita Scanlon from the American Water Quality Control and Research Laboratory. Dr. Mark LeChevallier, American Water, Voorhees, NJ, is thanked for tirelessly sharing insights into water industry needs.

BIBLIOGRAPHY 1. Wilson, B.R., Roessler, P.F., Van Dellen, E., Abbaszadegan, M., and Gerba, C.P. (1992). Coliphage MS 2 as a UV water disinfection efficacy test surrogate for bacterial and viral pathogens. Proc. Water Qual. Technol. Conf., Nov. 15–19, Toronto. 2. Wiedenmann, A. et al. (1993). Disinfection of hepatitis A virus and MS 2 coliphage in water by ultraviolet irradiation: Comparison of UV susceptibility. Water Sci. Technol. 27: 335–338. 3. Karanis, P. (1992). UV sensitivity of protozoan parasites. J. Water Supply Res. Technol. Aqua 41: 95. 4. Lorenzo-Lorenzo, M.J., Ares-Mazas, M.E., Villacorta-Martinez de Maturana, I., and Duran-Oreiro, D. (1993). Effect of ultraviolet disinfection of drinking water on the viability of Cryptosporidium parvum oocysts. J. Parasitol. 79: 67–70. 5. Campbell, A.T., Robertson, L.J., Snowball, M.R., and Smith, H.V. (1995). Inactivation of oocysts of Cryptosporidium parvum by ultraviolet irradiation. Water Res. 29: 2583–2586. 6. Clancy, J.L., Hargy, T.M., Marshall, M.M., and Dyksen, J.E. (1998). Inactivation of Cryptosporidium parvum oocysts in water using ultraviolet light. J. Am. Water Works Assoc. 90: 92–102. 7. Bukhari, Z., Hargy, T.M., Bolton, J.R., Dussert, B., and Clancy, J.L. (1999). Inactivation of Cryptosporidium parvum oocysts using medium-pressure ultraviolet Light. J. Am. Water Works Assoc. 91: 86–94. 8. Bukhari, Z. et al. (2000). Comparison of Cryptosporidium parvum viability and infectivity assays following ozone treatment of oocysts. Appl. Environ. Microbiol. 66: 2972–2980.

DECHLORINATION 9. Craik, S.A., Finch, G.R., Bolton, J.R., and Belosevic, M. (2000). Inactivation of Giardia muris cysts using medium pressure ultraviolet radiation in filtered drinking water. Water Res. 34: 4325–4332. 10. John, D.E., Nwachuku, N., Pepper, I.L., and Gerba, C.P. (2003). Development and optimization of a quantitative cell culture infectivity assay for the microsporidium Encephalitozoon intestinalis and application to ultraviolet light inactivation. J. Microbiological Methods 52: 183–196. 11. Di Giovanni, D.D., Hashemi, F.H., Shaw, N.J., Abrams, F.A., LeChevallier, M.W., and Abbaszadegan, M. (1999). Detection of infectious Cryptosporidium parvum oocysts in surface and filter backwash water samples by immunomagnetic separation and integrated cell culture-PCR. Appl. Environ. Microbiol. 65: 3427–3432. 12. Clancy, J.L., Bukhari, Z., Hargy, T.M., Bolton, J.R., Dussert, B., and Marshall, M.M. (2000). Using UV to inactivate Cryptosporidium. J. Am. Water Works Assoc. 92: 97–104. 13. Bukhari, Z. and LeChevallier, M.W. (2004). Finished water disinfection with UV light: Overview of validation studies at American Water. IUVA News 6: 15–20. 14. Landis, H.E., Thompson, J.E., Robinson, J.P., Blatchley, E.R. (2000). Inactivation responses of Cryptosporidium parvum to UV radiation and gamma radiation. Proceedings of the Water Quality Technology Conference, Salt lake City, November 5–9. 15. Shin, G.A., Linden, K.G., Arrowood, M., and Sobsey, M. (2001). Low pressure UV inactivation and subsequent repair potential of Cryptosporidium parvum oocysts. Appl. Environ. Microbiol. 67: 3029–3032.

DECHLORINATION RASHEED AHMAD Khafra Engineering Consultants Lilburn, Georgia

Dechlorination is the practice of removing all or a specified fraction of total residual chlorine. In potable water practice, dechlorination is used to reduce the residual chlorine to a specified level at a point where the water enters the distribution system. Dechlorination has been beneficial for waters that are burdened with high concentrations of ammonia nitrogen and organic nitrogen. In some cases where taste and odor control is a severe problem, control is achieved by complete dechlorination, followed by rechlorination. This removes the taste-producing nuisance residuals and prevents the formation of nitrogen trichloride (NCl3 ) in the distribution systems. Dechlorination of wastewater and power plant cooling water is required to eliminate residual chlorine toxicity, which is harmful to the aquatic life in the receiving waters. Other special applications requiring dechlorination are ahead of demineralizers, boiler makeup water, certain food plant operations, bottled water, and the beverage industry. In these cases, the dechlorination process is arranged to remove all residual chlorine. The most practical method of dechlorination is by sulfur dioxide and/or aqueous solutions of sulfite compounds. Among sulfur/sulfite compounds, sulfur dioxide (SO2 ) is the most widely used chemical for dechlorination. Other

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compounds such as sodium bisulfite (NaHSO3 ) and sodium metabisulfite (Na2 S2 O5 ) can be used as practical alternatives to sulfur dioxide. Sodium thiosulfate (Na2 S2 O3 ) is another chemical that is used entirely as a laboratory dechlorinating chemical. Other methods used for dechlorination are granular-activated carbon and aeration. Free and combined residual chlorine can be effectively reduced by sulfur dioxide and sulfite salts. Free chlorine is the amount of chlorine available as dissolved chlorine gas (Cl2 ), hypochlorous acid (HOCl), and hypochlorite ion (OCl− ) that is not combined with ammonia (NH3 ) or any other compounds in water. Combined chlorine is the sum of species composed of free chlorine and ammonia, including monochloramine (NH2 Cl), dichloramine (NHCl2 ), and trichloramine or nitrogen trichloride (NCl3 ). The sulfite ion is the active agent when sulfur dioxide or sulfite salts are dissolved in water. Their dechlorination reactions are identical. Sulfite reacts instantaneously with free and combined chlorine. Reactions yield small amounts of acidity, which is neutralized by the alkalinity of water (2.8 milligrams of alkalinity as calcium carbonate is consumed per milligram of chlorine reduced). Most potable waters and wastewaters have sufficient alkalinity buffering power that there is no cause for concern about lowering the pH by sulfur dioxide addition. The amount of sulfur dioxide required per part of chlorine is 0.9, but in actual practice, this ratio can be as high as 1.05. Owing to the low vapor pressure of sulfur dioxide, special precaution must be taken when using ton containers to prevent reliquefication. Unlike chlorination, dechlorination with sulfur dioxide does not require any contact chamber as the reaction occurs in a matter of seconds, probably 15–20 s at the most. There has been some apprehension about the possibility that excess sulfur dioxide might consume a significant amount of dissolved oxygen in the receiving waters downstream from a dechlorinated water discharge. In properly controlled systems, this reaction does not have sufficient time for completion. Hence, little effect on dissolved oxygen concentration has been reported. Granular and powdered carbon may be used to dechlorinate free, and some combined, chlorine residuals. The carbon requirements for dechlorination are typically determined by on-site pilot testing. The parameters of significance include mean particle diameter of carbon (pressure drop within the carbon contactor) and influent quality (pH, organics, and colloids). Carbon doses in the range of 30 to 40 mg/L have been reported. Granular-activated carbon (GAC) has proved effective and reliable as a dechlorination agent in potable water treatment. In addition, carbon provides filtration that removes other undesirable materials. In wastewater treatment, however, GAC has not been successful as a dechlorinating agent, possibly, because GAC is poor at removing organochloramines that form when significant concentrations of organic nitrogen are present. Because of the higher cost of carbon systems, their use is typically limited to specific sites or effluent with special discharge limitations. An aeration process can be used for dechlorination. Chlorine, hypochlorous acid, chlorine dioxide, and nitrogen trichloride are sufficiently volatile to be removed by aeration.

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READING LIST AWWA. (1999). Water Quality and Treatment, 5th Edn. McGrawHill, New York. Fair, G.M., Geyer, J.C., and Okun, D.A. (1968). Waste and Wastewater Engineering. John Wiley & Sons, New York. WEF. (2002). Chlorination/Dechlorination Handbook. Water Environment Federation, Alexandria, VA. WEF/ASCE. (1998). Design of Municipal Wastewater Treatment Plants, 4th Edn. Water Environment Federation, Alexandria, VA. White, G.C. (1999). Handbook of Chlorination and Alternative Disinfectants, 4th Edn. John Wiley & Sons, New York.

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the International Desalination Association (1), by the end of 2001, worldwide there were 15,233 desalination plants whose total installed treatment capacity is 32.4 million cubic meters per day (8560 MGD). Seawater or brackish waters are typically desalinated using two general types of water treatment technologies—thermal evaporation (distillation) and membrane separation. Currently, approximately 43.5% of the world’s desalination systems use thermal evaporation technologies. This percentage has been decreasing steadily during the past 10 years due to the increasing popularity of membrane desalination, which is driven by remarkable advances in membrane separation and energy recovery technologies and associated reduction of overall water production costs.

NIKOLAY VOUTCHKOV

THERMAL DESALINATION

Poseidon Resources Corporation Stamford, Connecticut

All thermal desalination technologies use distillation (heating of source water) to produce water vapor that is then condensed into low-salinity potable water. Thermal desalination is most popular in the Middle East, where seawater desalination is typically combined with power generation, which provides low-cost steam for distillation. Thermal desalination requires a large quantity of steam. The ratio of the mass of potable water produced to the mass of heating steam used to produce this water is commonly referred to as a gain output ratio (GOR) or performance ratio. Depending on the thermal desalination technology used, site-specific conditions, and the source water quality, the GOR varies between 2 and 24. The thermal desalination technologies most widely used today are multistage flash distillation (MSF), multiple effect distillation (MED), and vapor compression (VC).

INTRODUCTION Desalination is the production of fresh, low-salinity potable water from a saline water source (seawater or brackish water) via membrane separation or evaporation. The mineral/salt content of the water is usually measured by the water quality parameter, total dissolved solids (TDS), in milligrams per liter (mg/L) or parts per thousand (ppt). The World Health Organization and the United Sates Environmental Protection Agency (EPA), under the Safe Drinking Water Act, have established a maximum TDS concentration of 500 mg/L as a potable water standard. This TDS level can be used as a classification limit to define potable (fresh) water. Typically, water of TDS concentration higher than 500 mg/L and lower or equal to 15,000 mg/L is classified as brackish. Natural water sources such as sea, bay, and ocean waters that usually have TDS concentrations higher than 15,000 mg/L are generally classified as seawater. For example, Pacific Ocean seawater along the U.S. West Coast has a TDS concentration of 35,000 mg/L, of which approximately 75% is sodium chloride. Approximately 97.5% of the water on our planet is located in the oceans and therefore is classified as seawater. Of the 2.5% of the planet’s fresh water, approximately 70% is polar ice and snow, and 30% is groundwater, river and lake water, and air moisture. So even though the volume of the earth’s water is vast, less than 10 million of the 1400 million cubic meters of water on the planet are of low salinity and are suitable for use after using only conventional water treatment. Desalination provides a means for tapping the world’s main water resource—the ocean. During the past 30 years, desalination technology has made great strides in many arid regions of the world such as the Middle East and the Mediterranean. Today, desalination plants operate in more than 120 countries worldwide, and some desert states, such as Saudi Arabia and the United Arab Emirates, rely on desalinated water for more than 70% of their water supply. According to the 2003 desalination plant inventory report prepared by

Multistage Flash Distillation In the MSF evaporator vessels (flash stages or effects), high-salinity source water is heated, whereas the vessel pressure is reduced to a level at which the water vapor ‘‘flashes’’ into steam. Each flash stage (effect) has a condenser to turn the steam into distillate. The condenser is equipped with heat exchanger tubes that are cooled by the source water fed to the condensers. An entrainment separator removes high-salinity mist from the low-salinity rising steam. This steam condenses into pure water (distillate) on the heat exchanger tubes and is collected in distillate trays from where it is conveyed to a product water tank. Historically, MSF is the first commercially available thermal desalination technology used to produce potable water on a large scale, which explains its popularity. More than 80% of the thermally desalinated water today is produced in MSF plants. The GOR for the MSF systems is typically between 2 and 8. The pumping power required to operate MSF systems is 2.0 to 3.5 kWh per cubic meter of product water. Multiple Effect Distillation In the MED process, the source water passes through a number of evaporators (effects or chambers) connected in series and operating at progressively lower pressures. In MED systems, the steam vapor from one evaporator

DESALINATION

(effect) is used to evaporate water from the next effect. MED desalination systems typically operate at lower temperatures than MSF plants (maximum brine concentrate temperature of 75 ◦ C vs. 115 ◦ C) and yield higher GORs. The newest MED technologies, which include vertically positioned effects (vertical tube evaporators or VTEs), may yield GORs up to 24 kilograms of potable water per kilogram of steam. The pumping power required to operate MED systems is also lower than that typically needed for MSF plants (0.8 to 1.4 kWh per cubic meter of product water). Therefore, MED is now increasingly gaining ground over MSF desalination, especially in the Middle East where thermal desalination is still the predominant method for potable water production from seawater. Vacuum Compression The heat source for VC systems is compressed vapor produced by a mechanical compressor or a steam jet ejector rather than a direct exchange of heat from steam. In these systems, the source water is evaporated and the vapor is conveyed to a compressor. The vapor is then compressed to increase its temperature to a point adequate to evaporate source water sprayed over a tube bundle through which the vapor is conveyed. As the compressed vapor exchanges its heat with the new source water that is being evaporated, it condenses into pure water. VC desalination has been used mostly for small municipal and resort water supply systems and industrial applications. The power required to operate mechanical VC systems is typically between 10 and 20 kWh per cubic meter of product water. Further discussion of the applicability of the thermal desalination technologies described is presented elsewhere (2,3). MEMBRANE DESALINATION Membrane desalination is a process of separating minerals from source water using semipermeable membranes. Two general types of technologies are currently used for membrane desalination—reverse osmosis (RO) and electrodialysis (ED). In reverse osmosis, the product water (permeate) is separated from the salts in the source water by pressure-driven transport through a membrane. By the RO process, desalinated water is transported under pressure through the membrane while the minerals of the source water are concentrated and retained by the membrane. Applying high pressure for desalination is mainly needed to overcome the naturally occurring process of osmosis, which drives the desalinated water back through the membrane into the water of more concentrated mineral content. Nanofiltration (NF) is a process similar to RO, where membranes whose orderof-magnitude larger pore size is used to remove highmolecular-weight compounds that make water hard (i.e., calcium and magnesium). Desalination by Electrodialysis In ED-based treatment systems, the mineral–product water separation is achieved by applying electrical direct current (DC) to the source water, which drives the mineral ions in the source water through membranes to a pair of electrodes of opposite charge. A commonly

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used desalination technology that applies the ED principle is electrodialysis reversal (EDR). In EDR systems, the polarity of the electrodes is reversed periodically during the treatment process. The energy for ED desalination is proportional to the amount of salt removed from the source water. TDS concentration and source water quality determine to a great extent which of the two membrane separation technologies (RO or ED) is more suitable and cost-effective for a given application. Typically, ED membrane separation is cost-competitive for source waters whose TDS concentration is lower than 2000 mg/L. The TDS removal efficiency of ED desalination systems is not affected by nonionized compounds or objects of weak ion charge (i.e., solids particles, organics, and microorganisms). Therefore, the ED membrane desalination process can treat source waters of higher turbidity, biofouling, and scaling potential than RO systems. However, the TDS removal efficiency of ED systems is typically lower than that of RO systems (15 to 90% vs. 99.0% to 99.8%), which is one key reason why they are used mainly for brackish water desalination. Reverse Osmosis Desalination Reverse osmosis desalination is the most widely used membrane separation process today. Currently, there are more than 2000 RO membrane seawater desalination plants worldwide whose total production capacity is in excess of 3 million cubic meters per day (800 MGD). For comparison, the number of ED plants in operation is less than 300, and their total production capacity is approximately 0.15 million cubic meters per day (40 MGD). RO membrane desalination plants include the following key components: a source water intake system, pretreatment facilities, high-pressure feed pumps, RO membrane trains, and a desalinated water conditioning system. The source water intake system could be an open surface water intake or a series of seawater beach wells or brackish groundwater wells. Depending on the source water quality, the pretreatment system may include one or more of the following processes: screening, chemical conditioning, sedimentation, and filtration. Figure 1 shows a typical configuration of a seawater RO membrane system. The filtered water produced by the plant’s pretreatment system is conveyed by transfer pumps from a filtrate water storage tank through cartridge filters and into the suction pipe of the high-pressure RO feed pumps. The cartridge filters are designed to retain particles of 1 to 20 microns that have remained in the source water after pretreatment. The main purpose of the cartridge filters is to protect the RO membranes from damage. The high-pressure feed pumps are designed to deliver the source water to the RO membranes at a pressure required for membrane separation of the freshwater from the salts, typically 15 to 35 bars (200 to 500 psi) for brackish source water and 55 to 70 bars (800 to 1,000 psi) for seawater. The feed pressure required is site-specific and is mainly determined by the source water salinity and the configuration of the RO system. The ‘‘engine’’ of every desalination plant that turns seawater into fresh potable water is the RO membrane element (Fig. 2). The most widely used RO membrane element consists of two membrane sheets glued together

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Filtered seawater Product water (central collection pipe) Filtered seawater Figure 2. RO membrane vessel. 172

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and spirally wound around a perforated central tube through which the desalinated water exits the membrane element. The first membrane sheet, which actually retains the source water minerals on one side of the membrane surface, is typically made of thin-film composite polyamide and has microscopic pores that can retain compounds smaller than 200 daltons. This sheet, however, is usually less than 0.2 microns thick, and to withstand the high pressure required for salt separation, it is supported by a second thicker membrane sheet, which is typically made of higher porosity polysulfone that has several orders-of-magnitude larger pore openings. The commercially available membrane RO elements today are of standardized diameter, length, and salt rejection efficiency. For example, the RO membrane elements most commonly used for potable water production in large-scale plants are 8 in. in diameter, 40 in. long, and can reject 99.5% or more of the TDS in the source water. Standard membrane elements have limitations with respect to a number of performance parameters such as feed water temperature (45 ◦ C), pH (minimum of 2 and maximum of 10), silt density index (less than 4), chlorine content (not tolerant to chlorine in measurable amounts), and feed water pressure [maximum of 80 to 100 bars (1100 to 1400 psi)]. During reverse osmosis, the water molecules move through the RO membranes at a rate commonly referred to as flux. Membrane flux is expressed in cubic meters per second per square meter (m3 /sm2 ) or gallons per day per square foot (gfd) of active membrane surface area. For example, a typical seawater membrane RO element is operated at 8 to 10 gfd. Membrane performance tends to deteriorate over time due to a combination of wear and tear and irreversible fouling of the membrane elements. Typically, membrane elements have to be replaced every 3 to 5 years to maintain their performance in water quality and power demand for salt separation. Improvements in membrane element polymer chemistry and production processes have made the membranes more durable and have extended their useful life. Elaborate conventional media pretreatment technologies and ultra- and microfiltration membrane pretreatment systems prior to RO desalination are expected to extend membrane useful life to 7 years and beyond, thereby reducing the costs for their replacement and the overall cost of water. RO membrane elements are installed in pressure vessels that usually house six to eight elements per vessel (see Fig. 2). Multiple pressure vessels are arranged on support structures (racks) that form RO trains. Each RO train is typically designed to produce between 10% and 20% of the total amount of the membrane desalination product water flow. Figure 1 depicts one RO train. After the RO salt/water separation is complete, a great portion of the feed water energy applied through the highpressure RO pumps stays with the more concentrated seawater and can be recovered and reused to minimize the overall energy cost of seawater desalination. Dramatic improvements in membrane elements and energy recovery equipment during the last 20 years coupled with enhancements in the efficiency of RO feed pumps and

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reduction of the pressure losses through the membrane elements have allowed reducing the power to desalinate seawater to less than 3.5 kWh/m3 (13.5 kWh/1000 gallons) of freshwater produced today. Considering that the cost of power is typically 20% to 30% of the total cost of desalinated water, these technological innovations have contributed greatly to reducing the overall cost of seawater desalination. Novel energy recovery systems working on the pressure exchange principle (pressure exchangers) are currently available in the market, and use of these systems is expected to reduce further desalination power costs by approximately 10% to 15%. The pressure exchangers transfer the high pressure of the concentrated seawater directly into the RO feed water at an efficiency exceeding 95%. Future lower energy RO membrane elements are expected to operate at even lower pressures and to continue to yield further reduction in the cost of desalinated water. The ratio between the volume of the product water produced by the membrane desalination system and the volume of the source water used for its production is commonly called recovery and is presented as a percentage of the plant RO system feed water volume. The maximum recovery that can be achieved by a given pressure-driven membrane desalination system depends mainly on the source water salinity and is limited by the magnitude of the osmotic pressure to be overcome by the RO system highpressure feed pumps and by the scaling potential of the source water. Scaling occurs when the minerals left behind on the rejection side of the RO membrane are concentrated to a level at which they begin to form precipitates (crystalline compounds), which in turn plug the membrane pores and interfere with fresh water transport through the membrane. Typically, desalination plants using brackish source water can achieve 65% to 85% recovery. Seawater desalination plants can turn only 40% to 60% of source water into potable water because seawater typically has an order of magnitude higher salinity than brackish water. Detailed guidelines for designing membrane desalination plants are provided elsewhere (4,5). DESALINATED WATER QUALITY When membrane desalination is used to produce potable water, the product water quality depends mainly on the concentration of salts in the source water, the saltrejection capability of the RO membranes, and the applied membrane feed pressure. In thermal desalination, the product water quality depends mostly on the heating system and temperature, the pressure applied, and the type of distillation technology. The product water quality of thermal desalination systems is significantly less dependent on source water salinity than that produced by a membrane system. This makes thermal distillation processes an attractive alternative for parts of the world, such as the Middle East, where source water TDS is very high (40,000 to 46,000 mg/L) and low-cost steam is readily available. Typically, thermal desalination processes produce water of very low TDS concentration (1 to 60 mg/L). A single-stage, reverse osmosis system for seawater

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desalination produces permeate of 250 to 350 mg/L. More elaborate RO systems can produce water of quality matching that of a thermal desalination plant. Usually, the temperature of thermally desalinated water is 2 ◦ C to 5 ◦ C higher than the temperature of water produced by membrane separation. Desalinated water (permeate) produced by both thermal and membrane desalination is highly corrosive and has to be chemically conditioned (typically by lime addition) to increase product water alkalinity and to adjust pH (usually by addition of chlorine dioxide and/or acid to the lime conditioned permeate) to meet potable water quality regulations. CONCENTRATE MANAGEMENT The two main streams produced by every desalination plant are low-salinity freshwater and high-salinity concentrate. The results of a recent study completed by the U.S. Bureau of Reclamation (6) on the concentrate disposal methods most widely used in the United States (in order of decreasing frequency) are shown in Table 1. Sanitary sewer and surface water discharge are the two most popular and cost-effective methods for concentrate disposal. Depending on the site-specific conditions, deep well injection, evaporation ponds, and spray irrigation can also be competitive concentrate disposal alternatives. The zero liquid discharge system typically has the highest construction and operating costs. However, under specific circumstances, such as cold climate, low evaporation and soil uptake rates, high land costs, and low power costs, the zero liquid discharge system can be cost-competitive with evaporation pond and spray irrigation disposal. COST OF DESALINATED WATER Desalination water costs depend on many factors, including the type of treatment technology; the source water quality; the target product water quality; the size of the desalination plant; and the costs of energy, chemicals, labor, and membranes. The all-inclusive cost of potable water produced using a brackish water source treated by membrane separation usually varies between US$0.25 and $0.65/m3 (US$1.0 to $2.5/1000 gallons) of product water. Seawater desalination cost is estimated at US$0.65 to $1.20/m3 (US$2.5 to $4.5/1000 gallons). For very small plants located in isolated areas, this cost could be significantly higher. For comparison, the Table 1. Concentrate Disposal Methods and Their Frequency of Use in the United States Concentrate Disposal Method Surface water discharge Sanitary sewer discharge Deep well injection Evaporation ponds Spray irrigation Zero liquid discharge

Frequency of Use, % of Plants Surveyed 45 42 9 2 2 0

cost of water produced from fresh natural water sources (i.e., low-salinity groundwater, lake, and river water) is typically between US$0.15 and $0.40/m3 ($0.6 and $1.5/1000 gallons). This cost is driven mainly by the source water quality, the type of treatment technologies used, and the size of the water treatment plant. Developments in seawater desalination technology during the past two decades combined with the transition to construction of large capacity plants, colocation with power plant generation facilities, and enhanced competition by using the build-own-operate-transfer (BOOT) method of project delivery have resulted in a dramatic decrease in the cost of desalinated water (7). These factors are expected to continue to play a key role in further reducing the production cost of desalinated water. BIBLIOGRAPHY 1. Wagnick Consulting. (2002). IDA Worldwide Desalination Plant Inventory. International Desalination Association, pp. 5–1. 2. Water Desalting Committee of AWWA. (2004). Water Desalting—Planning Guide for Water Utilities. John Wiley & Sons, New York, p. 14. 3. Pankratz, T. and Tonner, J. (2003). Desalination.com—An Environmental Primer, Lone Oak, Houston, TX, p. 29. 4. AWWA. (1999). Manual of Water Supply Practices—M46. Reverse Osmosis and Nanofiltration, 1st Edn. American Water Works Association, Denver, CO, p. 21. 5. Watson, I.C. et al. (2003). Desalting Handbook for Planners. 3rd ed. Desalination and Water Purification Research Program Report No. 72, U.S. Bureau of Reclamation, p. 85. 6. Mickley, M.C. (2001). Membrane concentrate disposal: Practices and regulation. Desalination and Water Purification Research Program Report, U.S. Bureau of Reclamation, p. 229. 7. Voutchkov, N. (2004). The ocean—a new resource for drinking water. Public Works, June: 30.

DIATOMACEOUS EARTH FILTRATION FOR DRINKING WATER VIPIN BHARDWAJ NDWC Engineering Scientist

MEL J. MIRLISS International Diatomite Producers Association

Diatomaceous Earth (DE) filtration is a process that uses diatoms or diatomaceous earth—the skeletal remains of small, single-celled organisms—as the filter media. DE filtration relies upon a layer of diatomaceous earth placed on a filter element or septum and is frequently referred to as pre-coat filtration. DE filters are simple to operate and are effective in removing cysts, algae, and asbestos from water. DE has been employed in many food and beverage applications for more than 70 years and was used specifically to filter potable water during WWII. Since then, it has been used to produce high-quality, lowcost drinking water. DE filtration is currently one of the

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desalination produces permeate of 250 to 350 mg/L. More elaborate RO systems can produce water of quality matching that of a thermal desalination plant. Usually, the temperature of thermally desalinated water is 2 ◦ C to 5 ◦ C higher than the temperature of water produced by membrane separation. Desalinated water (permeate) produced by both thermal and membrane desalination is highly corrosive and has to be chemically conditioned (typically by lime addition) to increase product water alkalinity and to adjust pH (usually by addition of chlorine dioxide and/or acid to the lime conditioned permeate) to meet potable water quality regulations. CONCENTRATE MANAGEMENT The two main streams produced by every desalination plant are low-salinity freshwater and high-salinity concentrate. The results of a recent study completed by the U.S. Bureau of Reclamation (6) on the concentrate disposal methods most widely used in the United States (in order of decreasing frequency) are shown in Table 1. Sanitary sewer and surface water discharge are the two most popular and cost-effective methods for concentrate disposal. Depending on the site-specific conditions, deep well injection, evaporation ponds, and spray irrigation can also be competitive concentrate disposal alternatives. The zero liquid discharge system typically has the highest construction and operating costs. However, under specific circumstances, such as cold climate, low evaporation and soil uptake rates, high land costs, and low power costs, the zero liquid discharge system can be cost-competitive with evaporation pond and spray irrigation disposal. COST OF DESALINATED WATER Desalination water costs depend on many factors, including the type of treatment technology; the source water quality; the target product water quality; the size of the desalination plant; and the costs of energy, chemicals, labor, and membranes. The all-inclusive cost of potable water produced using a brackish water source treated by membrane separation usually varies between US$0.25 and $0.65/m3 (US$1.0 to $2.5/1000 gallons) of product water. Seawater desalination cost is estimated at US$0.65 to $1.20/m3 (US$2.5 to $4.5/1000 gallons). For very small plants located in isolated areas, this cost could be significantly higher. For comparison, the Table 1. Concentrate Disposal Methods and Their Frequency of Use in the United States Concentrate Disposal Method Surface water discharge Sanitary sewer discharge Deep well injection Evaporation ponds Spray irrigation Zero liquid discharge

Frequency of Use, % of Plants Surveyed 45 42 9 2 2 0

cost of water produced from fresh natural water sources (i.e., low-salinity groundwater, lake, and river water) is typically between US$0.15 and $0.40/m3 ($0.6 and $1.5/1000 gallons). This cost is driven mainly by the source water quality, the type of treatment technologies used, and the size of the water treatment plant. Developments in seawater desalination technology during the past two decades combined with the transition to construction of large capacity plants, colocation with power plant generation facilities, and enhanced competition by using the build-own-operate-transfer (BOOT) method of project delivery have resulted in a dramatic decrease in the cost of desalinated water (7). These factors are expected to continue to play a key role in further reducing the production cost of desalinated water. BIBLIOGRAPHY 1. Wagnick Consulting. (2002). IDA Worldwide Desalination Plant Inventory. International Desalination Association, pp. 5–1. 2. Water Desalting Committee of AWWA. (2004). Water Desalting—Planning Guide for Water Utilities. John Wiley & Sons, New York, p. 14. 3. Pankratz, T. and Tonner, J. (2003). Desalination.com—An Environmental Primer, Lone Oak, Houston, TX, p. 29. 4. AWWA. (1999). Manual of Water Supply Practices—M46. Reverse Osmosis and Nanofiltration, 1st Edn. American Water Works Association, Denver, CO, p. 21. 5. Watson, I.C. et al. (2003). Desalting Handbook for Planners. 3rd ed. Desalination and Water Purification Research Program Report No. 72, U.S. Bureau of Reclamation, p. 85. 6. Mickley, M.C. (2001). Membrane concentrate disposal: Practices and regulation. Desalination and Water Purification Research Program Report, U.S. Bureau of Reclamation, p. 229. 7. Voutchkov, N. (2004). The ocean—a new resource for drinking water. Public Works, June: 30.

DIATOMACEOUS EARTH FILTRATION FOR DRINKING WATER VIPIN BHARDWAJ NDWC Engineering Scientist

MEL J. MIRLISS International Diatomite Producers Association

Diatomaceous Earth (DE) filtration is a process that uses diatoms or diatomaceous earth—the skeletal remains of small, single-celled organisms—as the filter media. DE filtration relies upon a layer of diatomaceous earth placed on a filter element or septum and is frequently referred to as pre-coat filtration. DE filters are simple to operate and are effective in removing cysts, algae, and asbestos from water. DE has been employed in many food and beverage applications for more than 70 years and was used specifically to filter potable water during WWII. Since then, it has been used to produce high-quality, lowcost drinking water. DE filtration is currently one of the

DIATOMACEOUS EARTH FILTRATION FOR DRINKING WATER

U.S. Environmental Protection Agency’s (EPA) approved technologies for meeting the requirements of the Surface Water Treatment Rule (SWTR) and is most suitable for small communities that need to comply with the rule. WHAT IS DE FILTRATION? DE contains fossil-like skeletons of microscopic water plants called diatoms, which are a type of algae. These diatoms range in size from less than 5 micrometers to more than 100 micrometers, and have a unique capability of extracting silica from water to produce their skeletal structure. When diatoms die, their skeletons form a diatomite deposit. In its natural state, diatomite is 85 percent inert silica. The soluble portion of diatomite is extremely low (less than 1 percent). The odorless, tasteless, and chemically inert characteristics make DE safe for filtering water or other liquids intended for human consumption. APPLICATION AND HISTORICAL BACKGROUND During WWII, the U.S. Army needed a new type of water filter suitable for rapid, mobile military operations. The U.S. Army Engineer Research and Development Laboratories (ERDL) developed a DE filter unit that was lightweight, easily transported, and able to produce pure drinking water. Later, DE filtration technology was applied to filtering swimming pool water and more gradually to producing drinking water. The earliest municipal DE filter installation was a 75,000 gallons per day (gpd) system in Campbell Hills, Illinois, that began operating in 1948. By 1977, municipalities had constructed more than 145 plants. Today, nearly 200 DE plants are successfully operating. HOW DOES DE FILTRATION WORK? DE filtration strains particulate matter from water, and the process rarely uses coagulant chemicals. First, a cake of DE is placed on filter leaves. A thin protective layer of diatomaceous earth builds up, or accumulates, on a porous

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filter septum (a permeable cover over interior collection channels) or membrane. Recirculating DE slurry through the filter septum establishes this layer. The septum is most often plastic or metallic cloth mounted on a wire meshcovered steel frame. The DE process is also called pre-coat filtration because the solids separation at the start of a run takes place on the built-up pre-coat layer of DE. After the pre-coat forms on the filter leaves (usually 1/8 inch thick) raw water containing a low dose of DE, which is called body feed, is fed through the filter. Particulate solids in the product flow are separated on the pre-coat surface. With such separation, the unwanted particulate matter actually becomes part of the filter media. During a filter run, removing particulate matter from raw water causes head loss to gradually build up in the filter. The accumulation of DE body feed on the filter reduces the rate of head loss. When maximum head loss is reached, the flow of water into the filter is stopped and the filter cake is cleaned. High-pressure sprays, directed at the accumulated cake, detach the cake and provide dilution for draining the slurry suspension from the filter vessel. When cleaned, the filtration operation is repeated, beginning with the pre-coat cycle (see Fig. 1). Operators typically discard the DE removed from the filter leaves. APPROPRIATE FEED WATER QUALITY AND PERFORMANCE CAPABILITIES The use of DE filters is limited to treating source waters with an upper limit of turbidity at 10 NTU. Also, filtration rates range from 0.5 to 2 gallons per minute per square foot (gpm/ft2 ). The particle size that DE filtration removes relies upon the size distribution of the DE particles used for the pre-coat and body feed. DE filters are very effective for removing Giardia and Cryptosporidium cysts. In some cases, studies have reported up to a 6-log reduction of these cysts under routine operating conditions. Because DE filtration usually does not involve coagulation, its potential for removing dissolved constituents, such as color, is low. Therefore, the utility or its engineer must determine raw water quality before considering DE filteration.

Liquid to be filtered Pre-coat media layer

Pre-coat liquid

Cake of removed impurities and DE particles

Septum/filter media

Direction of flow

Diatomaceous earth pre-coat

Backwash water

Filtered liquid Cake of removed impurities and DE particles

Septum/filter media

Septum/filter media

Direction of flow

DE filter in operation

Direction of flow

DE filter in backwashing mode

Figure 1. Diatomaceous earth filter. Source: Fulton, George P. 2000. Diatomaceous Earth Filtration for Safe Drinking Water.

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WHAT ARE THE MONITORING AND OPERATING REQUIREMENTS? Monitoring requirements for DE filtration are simpler than requirements for coagulation and filtration because operators rarely ever use coagulant chemicals for DE filtration. However, operators must continuously monitor raw and filtered water turbidity. Operators also must monitor filter head loss so that they can determine when to backwash the filter. In general, DE filter plant operators need mechanical skills to operate the body feed pumps, pre-coat pumps, mixers, pipes, and valves. They also must be skilled in preparing the body feed and precoat slurries. Also, keeping DE filter leaves clean is of primary importance. A leaf filter that is not properly cleaned at the end of a filter run can accumulate dirt and slime on the filter cloth, which prevents a uniform pre-coat from forming when the filter is restored to service. ELEMENTS OF A DE FILTER Figure 2 shows the common elements in the manufacture of any flat leaf filter used in treating drinking water. The principal elements of a DE filter include the following: • containment vessel, • baffled inlet, • filter leaves mounted on an effluent manifold, • a method of cleaning the filter leaves at the end of a run, • a drain to receive the backwash water,

TYPES OF DE FILTRATION Two types of DE filters exist: (1) pressure filters, which have a pump or high-pressure water source on the influent side and (2) vacuum filters, which have a pump on the effluent side. Vacuum filters are open to the atmosphere. Pressure filters are enclosed within pressure vessels. The two basic groupings of DE filter designs are essentially defined by the hydraulic mode of operation, and are shown in (Fig. 3a and 3b). The principal advantages of pressure filters over the vacuum filters are related to the significantly higher differential head available.

Table 1. Pressure and Vacuum Filters Pressure Filters • Operates at higher flowrates, resulting in smaller, more compact filter units. • Longer filter runs, reducing the use of pre-coat material and backwash water because of less frequent cleaning cycles. • Less likelihood that gas bubbles will disrupt the media.

Vacuum Filters • Lower capital fabrication cost.

• Lower maintenance costs.

• Tanks are open at the top, making access and observation easy.

Source: Mel J. Mirliss, Vipin Bhardwaj, and the National Drinking Water Clearinghouse.

• open top or access mode, and • DE slurry preparation tank and pump feed. (a) Pressure difference

Pump discharge pressure

Discharge pressure to system

Backwash header

Filtrate

Baffled inlet Filter leaf

(b)

Inlet manifold Outlet manifold

Pressure difference

Discharge pressure to system Filtrate

Containment vessel Backwash drain

Figure 2. Elements of a flat leaf filter. Source: Fulton, George P. 2000. Diatomaceous Earth Filtration for Safe Drinking Water.

Pump suction

Figure 3. (a) Pressure filter. (b) Vacuum filter. Source: Fulton, George P. 2000. Diatomaceous Earth Filtration for Safe Drinking Water.

EMERGING WATERBORNE INFECTIOUS DISEASES

IS DE SUITABLE FOR SMALL SYSTEMS? DE filtration is well-suited to small systems, because it does not require chemical coagulation, so operators do not need to learn about this complex aspect of water treatment. In addition, installation costs for DE systems are less than those for other technologies, such as membranes. DE filtration is currently one of the EPA’s approved technologies for meeting SWTR requirements. An ideal, cost-effective DE filtration application is for well water supplies under the influence of surface waters, but that are otherwise acceptable in quality. Superior cyst removal capability makes the DE filter more advantageous than other alternatives. However, there is a potential difficulty in maintaining complete and uniform thickness of DE on the filter septum. WHERE CAN I FIND MORE INFORMATION? • Fulton, George P., P.E. (2000). Diatomaceous Earth Filtration for Safe Drinking Water. American Society of Civil Engineers, Reston, VA. • ‘‘Precoat Filtration.’’ 1988. AWWA M30, Manual of Water Supply Practices. American Water Works Association, Denver, CO. • Technologies for Upgrading or Designing New Drinking Water Treatment Facilities, EPA/625/489/023. (Available from the EPA) or contact the National Drinking Water Clearinghouse. HAVE YOU READ ALL OUR TECH BRIEFS? Tech Briefs drinking water treatment fact sheets have been a regular feature in the National Drinking Water Clearinghouse (NDWC) newsletter On Tap for more than five years. A package of Tech Briefs is now available as a product. A three-ring binder holds all the current Tech Briefs in print. New selections can be easily added to the package as they become available. To order this free product, call the NDWC at the numbers listed below and ask for item #DWPKPE71. Additional copies of fact sheets are also free; however, postal charges may be added. To order, call the NDWC at (800) 624-8301 or (304) 2934191. You also may order online at ndwc [email protected]. wvu.edu or download fact sheets from our Web site at www.ndwc.wvu.edu. Also, the NDWC’s Registry of Equipment Suppliers of Treatment Technologies for Small Systems (RESULTS) is a public reference database that contains information about technologies used by small waters systems around the country. For further information about accessing or ordering RESULTS, call NDWC. READING LIST Fulton, George P. 2000. Diatomaceous Earth Filtration for Safe Drinking Water. American Society of Civil Engineers, Reston, VA. Safe Water from Every Tap, 1997. Improving water service to small communities. National Research Council, National Academy Press, Washington, DC.

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Schuler, P.F., Gosh M.M., and Boutros S.N. 1988. Comparing the removal of Giardia and Cryptosporidium using slow sand and diatomaceous earth filtration. pp. 789–805, Proceedings of 1988 AWWA Conference, Denver, CO. Ongerth, J.E., Hutton, PE December 1997. DE filtration to remove Cryptosporidium. Journal of American Water Works Association. American Water Works Association, Denver, CO. Small System Compliance Technology List for the Surface Water Treatment Rule and Total Coliform Rule, 1998. U.S. Environmental Protection Agency, 815-R-98-001.

Vipin Bhardwaj, technical assistance specialist, has a B.S. in chemical engineering and has recently finished his master’s degree in agriculture at West Virginia University. He expects to receive a master’s degree in civil engineering this fall. Mel Mirliss has been associated with the diatomaceous earth industry for the past 28 years. For the past nine years, he has served as the executive director of the International Diatomite Producers Association (IDPA) where he has been a director since the association’s founding in 1987. He holds a B.S. in chemistry from the University of California, Berkeley.

EMERGING WATERBORNE INFECTIOUS DISEASES LOUIS H. NEL University of Pretoria Pretoria, Gauteng, South Africa

Humanity is plagued at present by at least 1709 different infectious diseases. These diseases are caused by pathogens, which are microscopic and parasitic organisms of diverse natures, including infectious proteins (prions), viruses, bacteria, fungi, and protozoa. Compared to diseases like cancer and metabolic diseases that have genotypic roots, including heart disease, diabetes, and the like, progress in the fight against infectious diseases has been rapid during the last half century. Whereas infectious diseases such as smallpox, measles, polio, rabies, plague and numerous other bacterial diseases have once been a global scourge, many of these diseases have now become controllable due to spectacular advances in public health practices, including improvements in the quality of water and sanitation and by immunization, education, early diagnosis, and the use of antibiotics/antimicrobials and other drug therapies. However, globally, 45% of all deaths and 63% of early childhood deaths are still caused by infectious diseases. INFECTIOUS DISEASES, EMERGING DISEASES, AND ZOONOSES New, emerging, reemerging, and resurgent infections contribute significantly to the infectious disease problems that the world is experiencing now. At least 156 such diseases have now been identified. Taken from a report by the Institute of Medicine (1), emerging infections may be defined as follows: ‘‘emerging infections are those

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whose incidence in humans has increased within the past two decades or threatens to increase in the near future. Emergence may be due to the spread of a new agent, to the recognition of an infection that has been present in the population but has gone undetected, or to the realization that an established disease has an infectious origin. Emergence may also be used to describe the reappearance (or ‘‘reemergence’’) of a known infection after a decline in incidence.’’ Generally, new diseases are synonymous with emerging diseases and can be regarded as diseases first described in the last 10 to 30 years. Most of these diseases are caused by specific modifications (mutation/evolution, species jump, etc.) of agents that are already in the environment. Reemerging diseases are those that have persisted at a subdued level in the population and recur as a result of antimicrobial drug resistance or other changes that might favor marked increases in disease incidence. Reemerging diseases can also be described as resurgent, pertinently referring to an abrupt increase in incidence or geographic distribution of the particular disease. The emergence and reemergence of diseases are clearly related to changes in the infectious pathogen, the vector or transmission system, and the host population. Such epidemiologically important changes may include drug resistance and mutations resulting in increased virulence, changes in the distribution or activity of vectors, globalization and increased travel, war, population explosions, climatic and ecological changes, geographical displacement of species, movement into previously uninhabited areas, poverty and breakdown of health care systems, and changes in agriculture and industrialization. Six infectious diseases can be identified as current, leading, high mortality rate diseases, and all of them conform to the criteria for emergence described before: AIDS, acute respiratory infection, diarrheal diseases, malaria, measles, and tuberculosis. A great many infectious diseases are zoonoses: These are diseases that are naturally transmitted between vertebrate animals and man—49% or 832 of all known infectious diseases can be considered zoonoses. A fair number of infectious diseases of man started out as zoonotic events but have since established a much more important human-to-human epidemiological cycle (e.g., HIV-AIDS, dengue hemorrhagic fever). Strikingly, however, 73% of all the emerging diseases considered (114 of 156) are zoonoses. Infectious diseases are typically transmitted through direct contact, insect vectors, sexual contact, respiratory tract (aerosolized microbes), and by contaminated food and water. Here we are concerned with waterborne infection, with pathogen entry through the mouth and alimentary track or, in some instances, through epithelial cells of the respiratory system.

among children) every year. Clearly, improved water sanitation that leads to safe drinking water for all of humanity would be the ultimate solution. Some emerging waterborne infections do not enter through the oral route but through the respiratory route. These infections are attributed to specific pathogens that have been made airborne through water spraying in nurseries, flower and vegetable markets, personal showers, and the like.

WATERBORNE INFECTIOUS DISEASES: THE SPECTRUM OF PATHOGENS

SPECIFIC EMERGING WATERBORNE INFECTIOUS DISEASES

Human and animal fecal pollution of water sources is the leading cause of waterborne infections. This fecal–oral route of infection contributes to hundreds of millions of cases of diarrhea and millions of deaths (particularly

Aseptic Meningitis and Various Other Syndromes

Viral Agents Many waterborne diseases are caused by viruses, which include adenoviruses (types 40 and 41), astroviruses, caliciviruses (including Norwalk and hepatitis E viruses), Enteroviruses (picornaviruses), and reoviruses (including rotaviruses). By far the most common medical condition associated with waterborne viral infection is diarrhea. The medical and economic importance of viral diarrhea should not be underestimated, millions of deaths (particularly, of children) are caused by these infections. At present, only a handful of these viral diseases is considered emerging and the most important of them, astrovirus enteritis, calicivirus enteritis, and hepatitis (E), are specifically discussed here. Bacterial Agents Increased resistance to antibiotics is a major factor in the resurgence of common bacterial infectious diseases and epidemic bacterial diarrhea such as those caused by waterborne Escherichia coli, Shigella, and Vibrio. The use of antimicrobials in agriculture has significantly contributed to this phenomenon, although evolution and adaptation of organisms have also led to converting nonpathogens into pathogens by adding of toxin producing capability. In addition, new bacterial zoonoses emerge, and some human pathogens are newly recognized, thanks to rapidly improving diagnostic techniques. This, as well as the reemergence of pathogens due to immune deficiencies related to HIV/AIDS, applies to the whole spectrum of infectious disease agents. Protozoal Agents These parasites typically are contaminants of potable water supplies. Whereas inadequate sanitation and chlorination of water supplies should prevent most of the waterborne bacterial diseases, large outbreaks of protozoan enteritis are usually associated with surface water supplies that are inadequately flocculated and filtered or not treated at all (many of these pathogens are chlorine resistant). In the last decade, enteric protozoa have become the leading cause of waterborne disease outbreaks for which an etiologic agent could be identified. The most common of these infections are caused by Cryptosporidium and Giardia, Cyclospora has also recently emerged as an enteric protozoan.

These diseases are caused by enteroviruses, viruses that—as their name suggests, replicate in the intestinal tract. Enteroviral infection is one of the most

EMERGING WATERBORNE INFECTIOUS DISEASES

common types of viral infections in humans, but they do not often result in serious disease, although some serotypes of enteroviruses may cause serious clinical syndromes. Such syndromes may include acute paralysis, encephalitis, meningitis, myocarditis, hepatitis, and chronic infection (particularly in immunocompromised individuals). Enteroviruses belong to the Enterovirus genus of the Picornaviridae family and include the polioviruses, group A Coxsackieviruses, group B Coxsackieviruses, echoviruses, and (newer) enteroviruses. There are 3 different polioviruses, 61 nonpolio enteroviruses, 23 Coxsackie A viruses, 6 Coxsackie B viruses, 28 echoviruses, and 4 other enteroviruses. Enteroviruses evolve rapidly and emerge as new strains, but others are of emerging importance due to drug resistance. Some enterovirus serotypes have been responsible for serious, large epidemics throughout the world in recent years. Echovirus infections have been responsible for several large epidemics of aseptic meningitis in Japan, Europe, and the Middle East in the past years. These epidemics involved different highly infectious serotypes of the virus, types 4, 9, and 30, but also of new genetic variants of these viral serotypes. Enterovirus 71 (EV71) has caused major disease outbreaks in North America, Europe, Malaysia, Japan, and Australia since 1995. In 1998, this virus infected at least a million individuals in Taiwan and caused a 20% case fatality among children under 5 years of age. Astrovirus Enteritis Today, astroviruses are considered much more prevalent than previously thought (e.g., 75% of 5 to 10-yearold children in Britain demonstrate antibodies against astroviruses). These are small, round viruses that have a starlike appearance under the electron microscope. Following a 1 to 4-day incubation period, the clinical symptoms present as watery diarrhea that lasts 2–3 days. These gastrointestinal disorders are usually not serious, but astroviruses are a leading cause of childhood diarrhea, and dehydration may be especially severe in immunocompromised individuals. Calicivirus Enteritis This disease, caused by a group of small cuplike viruses, is very common—serological evidence indicates that most people are infected by age 12. The symptoms are diarrhea with nausea and vomiting. Fever and respiratory infection occur in a small number of cases. It is thought that these viruses emerged from ocean reservoirs, with subsequent zoonotic and interspecies movement. Among the known calicivirus pathogens of humans are the Norwalk and Norwalk-like viruses and the Sapporo and Sapporolike viruses. Although it is thought that pathogenic caliciviruses are likely to continue emerging from the world’s oceans in various forms, only the Norwalk-like viruses are now and from a public health point of view, of emerging importance. Norwalk-like viruses cause sporadic and epidemic gastroenteritis in all age groups of humans, but losses of fluids and electrolytes are particularly serious in the very young and elderly. Transmission is by the fecal–oral route, and shellfish-related outbreaks

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frequently occur where appropriate standards for proper disposal of human sewage are not applied. Contaminated water, ice, eggs, salad ingredients, and ready-to-eat foods are other sources of infection. Campylobacteriosis This diarrheal disease, caused by Campylobacter jejuni, occurs worldwide. Infection by Campylobacter species has been known in livestock for many decades, but it has also been recognized as an important pathogen in humans since the middle 1970s. The development of selective media contributed to this recognition, and drug resistance and the emergence of toxin producing strains have recently become of considerable importance. Although this disease is usually self-limiting, treatment with antibiotics may reduce bacterial numbers and the duration of the infection and may be of particular value in cases involving immunocompromised individuals or where complications are evident. From this perspective, the emergence of antibiotic resistant strains of Campylobacter all over the world is disconcerting. Campylobacter jejuni infections, it is thought are more common than infections by other enteric bacteria such as Shigella, Salmonella and Escherichia coli O157:H7. Whereas other Campylobacter species (e.g., C. laridis, C. hyointestinalis) may also be involved in disease, C. jejuni is responsible for 99% of the cases. The bacterium is gram-negative, has a curved rod-shaped morphology and is motile. It is highly infectious and symptoms are diarrhea, fever, abdominal pain, nausea, headache, and myalgia. Complications are rare, but the infection may progress toward reactive arthritis, hemolytic uremic syndrome, and septicemia. Guillain-Barre syndrome is also recognized as a rare (1 per 1000 infected individuals) complication leading to disease of the peripheral nervous system. Cholera Cholera is caused by the gram-negative, motile, rodshaped bacterium, Vibrio cholerae. The disease, essentially an acute and serious infection of the small intestine with diarrhea that can be severe and is accompanied by dehydration and shock, occurs worldwide. The symptoms associated with cholera are caused by the enterotoxins produced by Vibrio cholerae during the infection. Accurate records of cholera date back almost a century, but cholera may have been known in India and elsewhere for thousands of years. This disease may be considered emerging due to new pandemics (serotype O1) and the appearance of pathogenic strains that have evolved by genetic recombination (serotype 0139). The emerging serotype O139 appeared in 1992 and has as yet only been reported from Southeast Asia where it causes epidemics in populations that were previously exposed to other serotypes of the same pathogen. In similar evolutionary fashion, Vibrio cholerae can become resistant to antibiotics by acquiring a transposon element and/or a plasmid that confers resistance to multiple antibiotics. Vibrio cholerae O1, however, caused a pandemic spread over six continents during the past 40 years. The bacterium initially (1960s) spread from Indonesia to Eastern Asia, India, the former USSR, Iran, and Iraq.

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Thereafter, in 1970, serotype O1 took hold in West Africa and then became endemic in most of the African continent. In 1991, it spread throughout South and Central America, and involved more than a million known cases at a mortality rate of about 1%. Much higher mortality rates (33%) were observed in 1994 in a large epidemic among Rwandan refugees in Zaire (33%). In recent times, Africa continued to host large cholera epidemics. These included a 1997 outbreak in Kenya, Tanzania, and Mozambique, a 1998 outbreak in Congo and Zaire, and in 2000/2001, a large epidemic in South Africa. Cryptosporidiosis The etiological agent of this persistent diarrheal disease is the protozoal parasite Cryptosporidium parvum. Together with the AIDS epidemic, cryptosporidiosis emerged in the early 1980s as a zoonosis, after having been known as a veterinareal disease for more than a century. Apart from persistent diarrhea, symptoms may include abdominal pain, nausea, and low-grade fever, and significant weight loss. Infectious oocysts may still be excreted long after diarrheal disease has ended (up to 5 weeks). In the growing number of immunocompromised individuals worldwide, this has become a life-threatening and highly infectious enteric disease. Compounding the threat of this disease is the fact that the pathogen is difficult to filter from water resources, is resistant to chlorine, is ubiquitous in many animals, can survive in infected water for long periods of time, and is highly infectious. It is therefore not surprising that C. parvum has now become the major protozoan pathogen of humans. A wide spectrum of water supplies has been implicated and these include contaminated and inefficiently treated/flocculated/filtered drinking water such as from water plants and boreholes or from swimming pools, wading pools, hot tubs, jacuzzis, fountains, lakes, rivers, springs, ponds, and streams. A high rate of secondary person-to-person transmissions in households and institutions has also been documented. Cyclosporiasis This protracted, relapsing gastroenteritic disease is similar to cryptosporidiosis and is also caused by a coccidian parasite, in this case, Cyclospora cayetanensis. This organism is food- and waterborne, and it emerged in the 1990s as a serious and widespread gastrointestinal parasite. In contrast to the zoonotic C. parvum, C. cayetanensis is specific to human hosts and is not a known pathogen of vertebrate animals. Cyclosporiasis is most common in tropical and subtropical regions and endemic in Central and South America, Southeast Asia, the Caribbean islands, and parts of Eastern Europe, but large outbreaks have recently been reported in the United States and Canada, reports also came from the United Kingdom and various African countries. Apart from fecally contaminated water, various types of fresh produce such as raspberries, basil, field greens, and salad mixes have been indicated as sources of the pathogen. Although Cyclospora oocysts, like Cryptosporidium oocysts, are resistant to chlorine, they are double the size of Cryptosporidium oocysts and may be more easily removed by flocculation and filtration methods used in water plants.

Further, due to the noninfectious nature of newly formed oocysts, secondary person-to-person spread of Cyclospora cayetanensis is a much less likely route of transmission than that for cryptosporidiosis. Dermatitis This is an inflammation of the skin marked by redness, pain, and itching (skin rashes) that is caused by the gram-negative and rod-shaped bacterium, Pseudomonas aeruginosa. Until the recent recognition of waterborne dermatitis caused by Pseudomonas aeruginosa, this organism was rarely implicated in disease. The bacterium typically survives in biofilms and antibiotic resistant strains are also appearing, but effective control should be possible by proper water treatment. Gastritis Helicobacter pylori, a spiral gram-negative bacterium was first recognized in 1982 as a cause of gastritis. It appears that the human stomach is the only recognized reservoir for this bacterium and infection is common in the general population. Although most infected individuals are asymptomatic and are unlikely to develop a serious medical problem from the infection, H. pylori causes 90% of duodenal ulcers and up to 80% of gastric ulcers, and infected persons have a twofold to sixfold increased risk of developing gastric cancer. In fact, H. pylori is classified as a group I (or definite) carcinogen by the World Health Organization’s International Agency for Research on Cancer. It is not known why some patients become symptomatic and others do not. The transmission of H. pylori also remains unclear, although the bacteria are most likely to spread from person to person through fecal–oral or oral–oral routes. Giardiosis This disease, one of the most common diarheal diseases spread by drinking and recreational water in the United States and probably worldwide, is caused by the intestinal protozoan parasite Giardia intestinalis, also known as G. lamblia or G. duodenalis. This agent also infects domestic and wild animals (e.g., cats, dogs, cattle, deer, and beavers). Emerging as a widespread and common disease during the past two decades, giardiasis has been traced back to contaminated swimming pools, hot tubs, fountains, lakes, rivers, springs, streams and ponds. As for the other parasitic diseases discussed here, the life cycle begins when cysts are ingested through person-toperson transmission or ingestion of fecally contaminated food or water. The ingested cysts release trophozoites in the duodenum where they attach to the surface of the intestinal epithelium. Giardia cysts can be excreted in fecal stools intermittently for weeks or months, and infection of a new host can result from ingestion of as few as 10 cysts. Hemorrhagic Colitis and Complications Enterohemorrhagic Escherichia coli (EHEC) is another bacterial pathogen that has recently emerged as a major waterborne enteric pathogen. After recognition of

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waterborne infection by this agent in 1985, the disease has been reported worldwide, and case numbers are ever increasing. Whereas E. coli bacteria are usually symbiotic as part of the normal intestinal flora of animals and humans, some strains of the organism can cause disease by producing large quantities of Vero toxins. It is a bacterial virus (or bacteriophage) that carries the toxin gene. When infecting the bacteria, the bacteriophage, integrates into the bacterial genome and from there provides the toxin producing capability. Today, a large number of E. coli serotypes that can produce Vero toxins are recognized, but the serotype O157:H7 is the predominant pathogen that is most frequently associated with hemorrhagic colitis in humans. The disease is serious, usually characterized by bloody diarrhea, abdominal cramps and nausea, but more severe complications, known as hemolytic uremic syndrome (HUS)—leading to complete renal failure, may follow. Case fatality rates for E. coli O157:H7 range from 3–36%, the fate of patients depends on age and immunocompetence. Waterborne transmissibility through both drinking and recreational waters together with antibiotic and chlorine resistance and an ability to survive in water for long periods of time, are major factors in the widespread emergence of EHEC. In addition, these organisms are highly infectious and are often spread to humans as zoonoses, EHEC has a very wide host range among domestic animals and wildlife. Outbreaks continue to be recognized worldwide in the United States, Japan (1996—6000 cases), United Kingdom, Australia, Europe, Argentina, Chile, and throughout Africa (e.g., South Africa, Swaziland, Kenya, Nigeria, Cote d’Ivoire, Central Africa Republic, and Egypt). In recent years, a number of efficient diagnostic techniques have been developed for accurate detection of isolate E. coli O157:H7, which also allows for distinction from other bacterial enteropathogens (e.g., Shigella) with which it may easily be confused due to clinical similarities of disease manifestation. Hepatitis (Hepatitis E-Like Viruses) Hepatis virus E (HEV) was isolated for the first time in 1988. This virus was then characterized in subsequent years, and it has recently been placed in its own taxonomic group, ‘‘hepatitis E-like viruses’’, within the class IV (+) sense RNA viruses. The disease is found most frequently in the developing world (e.g., South East Asia, India, Middle East, Africa, and Central America) and is transmitted primarily through the fecal–oral route in contaminated drinking water. Person-to-person transmission of HEV may occur during large epidemics, but this route of transmission appears to be uncommon. The incubation period following exposure to HEV is, on average a lengthy 40 days but may range from 15 to 60 days (mean, 40 days). Diarrhea is a less common symptom in this case, malaise, anorexia, abdominal pain, arthralgia, fever, anorexia, hepatomegaly, jaundice, nausea, and vomiting are more common. The disease is most often seen in young to middle-aged adults (15–40 years old) where the fatality rate is relatively low at 0.1–1%. However, pregnant women appear to be exceptionally susceptible to severe disease, and excessive mortality has been reported in this

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group (up to 30%). Recent genetic and serological evidence suggests frequent transmission of HEV between animals and humans, and zoonoses may explain the mechanism of HEV maintenance in populations during the periods between epidemic outbreaks. Legionnaires’ Disease Legionnaires’ disease, or legionnellosis, is an acute pneumonia caused by the bacterium Legionnella pneumophila. This is a flu-like illness, typically occurring 3 days after exposure and followed within a week by high fever, chills, dry cough, muscle aches and headache. The disease is usually self-limited but can lead to higher mortality in the presence of various risk factors such as immunodeficiency, cigarette smoking, chronic lung diseases, lung malignancies, heart disease, and old age. When this disease was first recognized in the late 1970s, the associated etiological agent was hard to isolate, given that Legionnella is a particularly fastidious organism that does not grow on typical culture media. The importance of these waterborne gram-negative, rod-shaped bacilli has since become clear, following the development of suitable diagnostic techniques. Of importance is the fact that these bacteria are parasites of protozoa (e.g., Amoeba sp.) that are common in rivers, lakes, and streams. Within these hosts, Legionnella will proliferate and be protected from environmental hazards, including chlorination. In addition, following intracellular replication in protozoal cells, L. pneumophila converts to a virulent form and expresses a number of traits (including enhanced motility) that will assist in extracellular survival and transmission to new cells. This phenomenonant is thought, explains the virulence of L. pneumophila for the macrophages in the human lung, given the similarities in cell biology of these cells and protozoal cells. If the virulent form of the bacteria is inhaled, replication may occur within the alveolar macrophages, and disease will ensue. This transmission occurs most easily in the presence of airborne water droplets, so Legionnellosis has become particularly important and common following the various practices leading to aerosolization of water (e.g., showers, spraying of produce in large markets, and air conditioning). Shigellosis Colitis Shigella is a highly infectious, gram-negative, rod-shaped bacterium, most species are implicated in waterborne infections. Although not new, it is the virulence and ability to develop drug resistance through plasmids that confers resistance to a large spectrum of antibiotics, which characterize a resurgence of shigellosis colitis. Disease symptoms include abdominal pains, fever, and rectal pain, and complications may include sepsis, seizure, renal failure, and hemolytic uremia syndrome. Shigella dysenteriae type 1 (Sd1) is the cause of epidemic dysentery that usually originates from water polluted with human feces. This type is a major cause of dysentery in Asia and Central and South America and has only recently arrived in Africa. After major pandemics in southern Africa in the early and middle 1990s, Shigella dysenteriae type 1 has now become endemic in many countries in Africa. Other species of Shigella that are frequently

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responsible for waterborne infections are Shigella sonnei and Shigella flexneri. Toxoplasmosis This infection, which is symptomless in most cases, is caused by the coccidian protozoan Toxoplasma gondii. However, the parasite is considered an important emerging pathogen in the AIDS patient group worldwide, where infection can cause lymphadenopathy, central nervous system disorders, myocarditis, pneumonitis, and cranial lesions. Domestic cats are the definitive host for Toxoplasma but, apart from humans, there are numerous additional intermediate hosts in the animal kingdom. Infection takes place by ingesting contaminated food and water or by inhaling airborne Toxoplasma oocysts. Tuberculosis (TB) Although most individuals infected with Mycobacterium tuberculosis are asymptomatic carriers of TB, this should be considered a serious long-term pulmonary disease. In symptomatic cases, listlessness, chest pain, loss of appetite, fever, and weight loss are early symptoms of pulmonary tuberculosis. This may progress into night sweats, bleeding in the lungs, coughing up of sputum with pus, and shortness of breath. Exposure usually occurs through inhalation of infected droplets and thus, similar to legionnellosis, the aerosolization of water contributes to the transmissibility of Mycobacterium where water supplies are contaminated with the organism (often in hospitals). Tuberculosis is not a new infection, but multidrug-resistant tuberculosis (MDRTB) has become a major public health problem during the last 10 years, particularly as an opportunistic infection in concert with the AIDS epidemic. CONCLUDING REMARKS In considering the reasons for the emergence of the specific waterborne infectious diseases described, it is easy to identify a number of common and very specific causal elements. These include (1) the rising incidence of diseases or other factors that affect immune competence in groups or populations, including HIV/AIDS, malnutrition, and the like; (2) genetic adaptation of pathogens, including the acquisition of resistance plasmids, toxin genes, and other virulent factors; (3) overcrowding combined with poor sanitation and deteriorating or inadequate public health infrastructure; (4) increased exposure to animal reservoirs (as a result of geographical expansion, overpopulation, and competition for dwindling water resources) and continual zoonoses. Based on these factors, the potential for the continued emergence and reemergence of infectious diseases seems inexhaustible. Advances in vaccination, therapeutic intervention, and surveillance are crucial elements in the fight against these diseases, but even so, much may be achieved through education, improved hygiene, and proper water sanitation. BIBLIOGRAPHY 1. Ackelsberg, J., Sivapalasingam, S., Olsen, S.J. et al. (1999). A Large Waterborne Outbreak of Escherichia coli O157:H7

and Campylobacter Infections in County Fair Attendees. NY. Presented at the 37th Annual meeting of the Infectious Disease Society of America, Philadelphia, PA, November 18–21, 1999. Abstract 728. 2. Allos, B.M. (2001). Campylobacter jejuni infections: update on emerging issues and trends. CID 32: 1201–1206. 3. Barrett, T.J. et al. (1994). Laboratory investigation of a multistate food-borne outbreak of Escherichia coli O157:H7 by using pulsed Field gel electrophoresis and phage typing. J. Clin. Microbiol. 32: 3013–3017. 4. Basu, A. et al. (2000). Vibrio cholerae O139 in Calcutta, 1992—1998. Emerging Infect. Dis. 6(2): 139–147. 5. Berke, T. et al. (1997). Phylogenetic analysis of the Caliciviruses. J. Med. Virol. 52: 419–424. 6. Chalmers, R.M., Aird, H., and Bolton, F.J. (2000). Waterborne Escherichia coli 0157. J. Appl. Microbiol. Symp. Supple. 86: 124S–132S. 7. Colwell, R.R. (1996). Global Climate and infectious disease: The Cholera paradigm. Science 274: 2025–2031. 8. DaSilva, E. and Iaccarino, M. (1999). Emerging diseases: A global threat. Biotechnol. Adv. 17: 363–384. 9. Emmerson, A.M. (2001). Emerging waterborne infections in health-care settings. Emerging Infect. Dis. 7(2): 272–276. 10. Enserink, M. (2000). Malaysian Researchers trace Nipah virus outbreaks to bats. Science 289: 518–519. 11. Ewald, P.W. (1991). Waterborne transmission and the evolution of virulence among gastrointestinal bacteria. Epidemiol. Infect. 106: 83–119. 12. Ewald, P.W. (1998). The evolution of virulence and emerging diseases. J. Urban Health 75(3): 480–492. 13. Farmer, J.J. and Davis, B.R. (1985). H7 antiserum-sorbitol fermentation medium: A single-tube screening medium for detecting Escherichia coli O157:H7 associated with hemorrhagic colitis. J. Clin. Microbiol. 22: 620–625. 14. Feng, P. (1995). Escherichia coli serotype O157:H7: Novel vehicles of infection and emergence of phenotypic variants. Emerging Infect. Dis. 1(2): 47–52. 15. Fields, P.I. et al. (1997). Molecular characterization of the gene encoding H antigen in Escherichia coli and development of a PCR restriction fragment length polymorphism test for identification of E. coli O157:H7 and O157: NM. J. Clin. Microbiol. 35: 1066–1070. 16. Fratamico, P.M., Buchanan, R.L., and Cooke, P.H. (1993). Virulence of an Escherichia coli O157:H7 sorbitol positive mutant. Appl. Environ. Microbiol. 59: 4245–52. 17. Goma Epidemiology Group. (1995). Public health impact of Rwanda refugee crisis: what happened in Goma, Zaire, in July 1994, Lancet 345: 339–344. 18. Gomez-Lus, R., Clavel, A., Castillo, J., Seral, C., and Rubio, C. (2000). Emerging and re-emerging pathogens. Int. J. Antimicrobial Agents 16: 335–339. 19. Beckett, G. et al. (2000). Pseudomonas Dermatitis/Folliculitis Associated with Pools and Hot Tubs—Colorado and Maine, 1999—2000. Morbidity and Mortality Weekly Report. Centres for Disease Control and Prevention MMWR 49(48): 1087–1091. 20. Besser, R.E., Griffin, P.M., and Slutsker, L. (1999). Escherichia coli 0157:H7 gastroenteritis and the haemolytic uremic syndrome: An emerging infectious disease. Annu. Rev. Med. 50: 355–367. 21. Herwaldt, B.L. (2000). Cyclospora cayatanensis: A review, focusing on the outbreaks of Cyclosporiasis in the 1990s. Clin. Infect. Dis. 31: 1040–57.

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24. Iverson, E.R. et al. (1998). Epidemic Shigella dysenteriae in Mumiao, western Kenya. Trans. Roy. Soc. Trop. Med. Hyg. 92: 30–31. 25. Jaykus, L. (1997). Epidemiology and detection as options for control of viral and parasitic foodborne disease. Emerging Infect. Dis. 3(4): 529–539. 26. Karas, J.A., Pillay, D.G., Naicker, T., and Storm, A.W. (1999). Laboratory surveillance of Shigella dysenteriae type 1 in Kwa Zulu-Natal. S. Afr. Med. J. 89: 59–63.

45. Yamai, S., Okitsu, T., Shimada, T., and Katsube, Y. (1997). Distribution of serogroups of Vibrio cholera non O1 non O139 with specific reference to their ability to produce cholera toxin and addition of novel serogroups. J. Japanese Assoc. Infect. Dis. 71: 1037–1045. 46. Yamasaki, S. et al. (1997). Cryptic appearance of a new clone of Vibrio cholerae O1 biotype El Tor in Calcutta, India. Microbiol. Immunol. 41: 1–6. 47. http://www.cdc.gov/. 48. http://www.who.int/.

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REBECCA L. CALDERON, PH.D. National Health and Environmental Effects Laboratory Research Triangle Park, North Carolina

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INTRODUCTION Public health surveillance has played a key role in controlling the spread of communicable disease and identifying the need for specific public health practices, such as the filtration and chlorination of drinking water supplies. However, the characteristics of waterborne outbreaks since the early 1990s have raised questions about whether current water treatment practices can prevent transmission of some enteric pathogens (1–5). In addition, one analysis suggested that a significant fraction of all enteric disease in the United States may be due to drinking water (3). Another study found evidence that consuming surface-derived drinking water which meets current U.S. Environmental Protection Agency (USEPA) drinking water standards may significantly increase the risk of enteric illness (6). These concerns have motivated the U.S. Congress to require USEPA to prepare a report

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on the magnitude of epidemic and endemic waterborne disease in the United States. Even as the needs increase for better information about waterborne disease occurrence and causes, some have suggested that our disease surveillance system is in a state of crisis and may possibly collapse (7). Another study revealed that state health departments often cannot dedicate any staff to enteric disease surveillance (8). Current concerns over the preparedness for detecting and controlling bioterrorism attacks have also motivated interest in the adequacy of waterborne disease surveillance. In this chapter, issues relating to disease surveillance and outbreak investigations are presented to assist readers in understanding the strengths and weaknesses of current waterborne disease surveillance and outbreak detection programs and to suggest additional steps to strengthen the system. With limited public health resources available, it is important to carefully consider the goals and approaches to waterborne disease surveillance. In addition to addressing the information needs of governmental disease control programs, it is essential to ensure that the information needs of the drinking water industry, the regulatory agencies, and the public are best served. It may also be essential for drinking water utilities to participate in and, perhaps, help fund these surveillance systems.

BACKGROUND It is increasingly accepted that additional information is needed about the occurrence and causes of waterborne disease, both epidemic and endemic. The Centers for Disease Control (CDC) funded ‘‘emerging pathogen’’ surveillance projects in selected state health departments, in part to improve surveillance for several important waterborne agents. In New York City (NYC), the Department of the Environment (DEP), responsible for drinking water treatment and delivery, convened a panel of public health experts in 1994 to evaluate current health department disease surveillance programs. The panel recommended specific waterborne disease surveillance activities and epidemiologic studies to determine endemic waterborne disease risks associated with use of unfiltered surface water sources (Table 1) (9). Efforts to improve NYC waterborne disease surveillance are funded by the NYC DEP, the first time this has occurred for a drinking water utility in the United States.

Table 1. New York City Panel Recommendations on Waterborne Disease Surveillance Designate an individual who is specifically responsible for coordinating waterborne disease surveillance Conduct special surveillance studies of nursing and retirement home populations Conduct surveillance in managed care populations Monitor visits to emergency rooms Conduct surveillance of high-risk populations Monitor sales of prescription and nonprescription medications

An option for improving waterborne disease surveillance is to build on the current surveillance programs in place in most state and local health departments. This system is based on voluntary disease reporting by healthcare providers and clinical laboratories. However, a number of limitations of the system have been identified, and other factors may have already significantly reduced the effectiveness of traditional disease surveillance programs. Some pathogens, such as Cryptosporidium, are often difficult to diagnose, and other pathogens may exist for which there are no known diagnostic tests or no tests available for routine use. Changes in healthcare access and delivery practices may reduce the number of patients seeking healthcare and, also, the chances that medically attended diseases are confirmed by laboratory tests. An outbreak resulting in many medically attended illnesses in a large city could be unrecognized, as almost happened in the Milwaukee outbreak. In that outbreak, a large increase in the occurrence of diarrheal illness occurred around March 30–31, 1993. On Thursday, April 1, 1993 a pharmacist noted a dramatic increase in sales of over-the-counter antidiarrheal and anticramping medications. Normally his pharmacy sold $30 a day of such medications. Starting that Thursday, drug sales increased to approximately $500–$600 a day, or 17–20 times the normal sales. The increased sales continued on Friday, as a result of which the pharmacy sold most of its supply of antidiarrheal medications. The pharmacist called the health department to inquire about excessive reports of diarrhea or intestinal illness. The health department was unaware of any outbreak. On Saturday the increased sales continued so the pharmacist contacted the three local television stations to report what he believed to be a major occurrence of diarrheal disease in the city. On Sunday night his report was carried on the evening news for one station and by Wednesday, April, 7, the outbreak was confirmed by the Milwaukee Health Department. In the case of the Milwaukee outbreak, few of the people sought medical care for their diarrhea. However, even in situations where care was sought, it is possible that no one physician would notice an outbreak. For example, if many different healthcare providers treated the patients, it is possible that no one provider would recognize excess occurrences of illness. In addition, the existence of health effects in a small but extremely susceptible subpopulation might be difficult to detect because of the small number of people at risk. As some changes have made it more difficult to detect outbreaks, other changes present new disease surveillance opportunities. Computerization of patient records, healthcare and laboratory workloads, prescription and nonprescription pharmaceutical sales, and calls to nurse hotlines are potential new tools for more effective and less costly disease surveillance. Technological advancements, such as detection of antigen or antibodies specific to a pathogen in sera, stools, and other secretions, may improve detection of etiological agents. These may also allow detection of infections in the absence of disease.

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To better evaluate the current and alternative surveillance opportunities, five questions have been selected for discussion in this chapter: 1. What are the limitations of our current disease surveillance systems? 2. Should the early detection of outbreaks be the primary goal of a surveillance system and, if so, how can it be best achieved? 3. What is meant by endemic or background rates of disease, can some of this endemic disease be attributable to drinking water, and what should water utilities do to better understand these risks? 4. Can findings from outbreak investigations be used to estimate the unreported burden of enteric disease attributable to drinking water? 5. Since only a fraction of infected persons become ill from most enteric infections, should expanded surveillance programs monitor infection rather than illness? LIMITATIONS OF THE CURRENT DISEASE SURVEILLANCE SYSTEMS What are the limitations of our current disease surveillance systems? Detection of waterborne disease outbreaks depends, in part, on a state-federal system of notifiable or reportable diseases. Disease reporting is primarily the responsibility of healthcare providers and diagnostic laboratories. State or local laws require the reporting of certain diseases. Primary responsibility for disease surveillance rests with the state or local public health authorities. Most state surveillance systems are ‘‘passive,’’ in that reports are sent to the state or local health department by cooperative health care providers or laboratories. Providers and laboratories usually receive little encouragement from the health department to report illnesses. Government enforcement of reporting requirements is minimal. An ‘‘active’’ system will routinely contact some or all healthcare providers and laboratories, asking for illness reports (Table 2) (10). It has long been recognized that both passive and active disease reporting incompletely ascertain the level of disease in the community. The level of completeness varies by disease, by state, and by areas or populations within a state (11). For example, reporting is likely to be more complete for severe diseases such as hemorrhagic E. coli than for milder infections, such as Norwalk virus gastroenteritis. Laboratories tend to be much better at reporting their findings than are physicians (10). Even within an area, there can be great variations in reporting, depending on the interest of clinical laboratories and the dedication of diagnosing physicians (11). For example, for pathogens that are new or where there are questions about the mode of transmission, reporting may be more complete than for agents that are common, where the mode of transmission is well known and where public health intervention is less necessary.

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Table 2. Surveillance System Definitions Mandatory reporting A diagnosed case of disease is required, by law, to be reported; for example, in the case of cryptosporidiosis, all diagnosed cases are to be reported Passive Disease reports are submitted by providers and/or laboratories without specific follow-up by the health department Active Providers and/or laboratories are contacted to encourage diseases reporting; because of resource requirements, this is usually done as a special project for a limited duration of time Enhanced Special additional efforts are made to encourage disease reporting; this might include news releases, posters at strategic locations, presentation to special populations, or health surveys in communities with water quality problems

In addition to incomplete reporting of diagnosed illnesses, only a portion of all infections will ever be medically attended. As illustrated in (Fig. 1), only a fraction of infections will lead to illness. These infected persons may be unaware of their infection. In other cases, such as sometimes occurs as a result of childhood Giardia infection, the child fails to thrive but experiences none of the classic symptoms of giardiasis. When symptoms occur, they may be mild and/or may resolve in a short period of time. In this case, the person may not seek medical care or may simply visit a pharmacy to obtain medication to alleviate their symptoms. In the case of Milwaukee, despite the large number of reported cases of cryptosporidiosis, very few people visited their physician and few stool specimens were positive for Cryptosporidium oocysts.

Deaths Hospitalizations Doctor′s Office

Symptomatic

Asymptomatic or Infected

Figure 1. Disease pyramid.

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If the person seeks healthcare, the physician may fail to correctly diagnose the infection, since in many cases symptoms are not sufficiently specific to accurately identify the pathogen. If misdiagnosed and the infection resolves itself, the patient may not seek additional healthcare and no report of an infection will be generated. Even when the physician correctly diagnoses the illness and prescribes the appropriate medication, a confirmatory laboratory test may not be ordered. If ordered, the patient may not submit the sample to the laboratory, since many patients are unwilling to submit stool specimens for laboratory analysis. Since laboratories are the primary source of disease reports for surveillance systems, without a laboratory-confirmed diagnosis, a report may never be filed. When a stool or blood sample is submitted for laboratory analysis, it can also test negative because of analytical or specimen collection error, untimely collection or because the material submitted was, by chance, free of the pathogen (12). Laboratory proficiency can vary considerably. This may be more of a problem for laboratories that run only a small number of the ordered test. For persons infected with enteric parasites, single stools may often be free of the parasite or have insufficient numbers of parasites to assure laboratory detection. In some cases, even multiple stools may be pathogennegative. If a sufficient number of cases of illness from the same pathogen are reported to the health department at about the same time and if the epidemiologist is alert to an increase in case reports, an outbreak may be identified. Because of the time required to perform the diagnostic tests and to report the results, outbreak recognitions may occur weeks after the onset of the actual outbreak. Many outbreaks are first detected by an alert clinician. For example, in 1976, a Camas, Washington physician’s son had returned from Russia with giardiasis. The physician later recognized that several of his patients had similar symptoms. This lead to the identification of a waterborne giardiasis outbreak (13). As mentioned earlier, in Milwaukee, Wisconsin a pharmacist noted a dramatic increase in sales of antidiarrheal medication. In California and Arizona, diarrheal illnesses reported to health agencies by 65 campers who had visited an Arizona park initiated an investigation that implicated contaminated water as the source of an outbreak that affected 1850 people (14). The fortuitous circumstances surrounding the detection of many outbreaks raises concerns about how many medium to large outbreaks are never detected. Small outbreaks may seldom be detected, especially among travelers who consume water from noncommunity systems or who swim in multiple locations. Limitations of the current disease surveillance systems prompted a series of studies in the early 1980s to evaluate potential improvements in disease reporting and to evaluate the efficacy of active surveillance programs. A three state study of various approaches to active disease surveillance, funded by USEPA, detected no additional waterborne disease outbreaks in two states (Washington and Vermont) (15). However, in one state (Colorado) a greater than threefold increase in the number of detected waterborne outbreaks occurred (16). The reasons why

Colorado was able to identify so many more outbreaks than either Washington State or Vermont are unclear. An intense effort was made to increase disease reporting in all states and dramatic increases in reports of enteric diseases were observed in all three states. It is possible that a combination of poor quality water supplies plus an exposed tourist population, without protective immunity, may have allowed Colorado to identify more outbreaks than the other two states. In summary, active disease reporting can increase reporting of diagnosed illnesses only from providers and laboratories. All the other barriers to disease identification and reporting will still remain (Fig. 1). If healthcare access declines over time or, to reduce healthcare costs, physicians use fewer laboratory diagnostic services, then the number of diagnosed reportable illnesses will decline. This will occur despite the efforts of health departments to insure that most diagnosed illnesses are reported.

EARLY DETECTION OF OUTBREAKS Should the early detection of outbreaks be the primary goal of a surveillance system, and, if so, how can it be best achieved? The occurrence of a waterborne disease outbreak is an exciting, newsworthy, and politically important event. Affected populations may experience severe illness and a large number of people may become ill. As a result of the investigation, much is often learned about the cause of major failures in water treatment or distribution. However, when the excitement has subsided, water system deficiencies have been corrected and the outbreak is officially said to be over, has the problem been solved or is disease continuing to occur but at a reduced level, below what is detectable by traditional surveillance activities? For example, a waterborne disease outbreak investigation detected major problems with the filtration system of an anonymous small community water supply. The system was, at the time of its installation, considered adequate. However, high turbidity levels were observed in treated water at the time of the outbreak, suggesting poor operation of the filtration facility. Optimization of treatment by consulting engineers allowed the plant to dramatically improve pathogen removal. This improvement reduced the number of new cases of disease, and the outbreak officially ended. However, 2 years later a serological survey of the town’s residents revealed the continued occurrence of infection by the same etiologic agent responsible for the earlier outbreak. These new data presented both philosophical and technical problems. Should all outbreaks be followed by such a survey? Is evidence of continuing infection sufficient reason for further intervention? If the serological survey were not conducted, there would be no evidence of increase risk of infection. If the plant was already optimized, what are the remaining intervention options without new filtration or disinfection technology? This scenario assumes that the continued high serological levels resulted from waterborne transmission. In fact, without a follow-up epidemiologic investigation, it is not possible to distinguish waterborne from other

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ENDEMIC DISEASE What is meant by endemic or background rates of disease and can some of this endemic disease be attributed to drinking water? Endemic level of disease is defined by the CDC as a persistent low to moderate level of disease occurrence. A persistently high level of occurrence is called hyperendemic while an irregular pattern of occurrence is called sporadic (Fig. 2). For most enteric infections, endemic disease results from a statistical averaging of small to moderate-sized undetected outbreaks or clusters of infection. There is little information to suggest that endemic levels of disease remain constant over time or across geographic areas, nor is there reason to believe the endemic level of disease is unimportant. Over the past century, the importance of endemic disease has become increasingly recognized. Following

Epidemic Number of cases

routes of transmission. In addition, without improved surveillance activities, we know little about the absence of symptomatic disease. Low levels of disease from exposure to waterborne microbes over a period of many years can result in a much larger health burden for a community than the number of disease cases that might occur during a detected outbreak. However, exposure to some waterborne pathogens at levels that boost the immune response may prevent symptomatic illness. These concerns must all be considered when developing a surveillance system. Without clear goals and a commitment to conduct epidemiologic investigations and take appropriate actions, a better surveillance system will not improve public health. Failure to detect low levels of disease transmission may provide a false sense of security. For example, why should an outbreak such as occurred in Milwaukee not have been preceded by many smaller outbreaks? Is it possible that in each of the cities experiencing a large waterborne cryptosporidiosis outbreak, prior undetected smaller outbreaks occurred? In fact, is it possible that lower levels of waterborne Cryptosporidium infection had occurred years prior to the outbreak? At the time of the detected outbreak, a higher number of oocysts may have passed through the treatment system or a more virulent strain of the pathogen emerged. If so, relying on disease surveillance systems that can only detect large outbreaks will seldom provide public health officials and the industry early warnings of emerging new diseases. This may be equivalent to basing the science of meteorology only on the study of hurricanes. The detection of an outbreak can also affect future disease reports in an area. For example, it is possible that overreporting of symptoms consistent with the disease of interest could occur. If so, similar outbreaks may be detected in neighboring areas. Given the increased popularity of bottled water use, it is possible that the at-risk population could change following an outbreak if a significant fraction of the population discontinued drinking tapwater. Therefore, decreases in the occurrence of reported waterborne disease may not reflect better control of the contamination but a reduction in the number of exposed individuals.

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Outbreak detection level Undetected epidemic

Endemic

Hyperendemic

Sporadic

Time

Figure 2. Epidemic versus endemic disease.

World War I, an attempt was made to estimate the prevalence of parasite infections in both the returning British soldiers and the British population who remained at home (17). To the surprise of the researchers, a high prevalence of asymptomatic infection was found among persons who had never left Britain. Later, a survey of Wise County, Virginia in 1930 revealed that half of the population carried Entamoeba histolytica and that 38% carried Giardia lamblia (18). A study to determine the incidence of Cryptosporidium infection among Peace Corp workers to be sent overseas revealed that almost 30% had possibly experienced infection prior to leaving the United States (19). More recent work we conducted suggests that endemic rates of Cryptosporidium infection may be very high, but that rates of cryptosporidiosis may be low (20,21). Data derived from disease surveillance systems cannot be used to compare endemic disease levels between areas or populations with different water systems. Whether observed differences in disease reports are due to the differences in the completeness of reporting or to differences in the occurrence of the disease or the infection cannot be answered, even with improved surveillance systems. In addition, it has become increasingly recognized that populations can develop protective immunity to infectious agents. If so, rates of infection may remain high while rates of illness remain low (21). The absence of disease in a population may, therefore, not mean that there is an absence of infections. Epidemiologic studies must be specifically designed and conducted to address the association of endemic disease with water system type or quality. Several epidemiologic studies have reported waterborne disease associated with public water systems in the absence of a reported waterborne outbreak. In New Zealand, the incidence of laboratory-confirmed giardiasis was found to be higher in a part of the city receiving chlorinated, unfiltered surface water compared to the part where surface water was treated by coagulation, flocculation, granular filtration, and chlorination (22). In Vermont, a higher incidence of endemic giardiasis was found in municipalities using unfiltered surface water or wells than in municipalities with filtered surface water (23). A Canadian study attempted to estimate how much endemic enteric illness was due to drinking water (6). The fraction of illness attributable to drinking water was estimated by

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comparing rates of reports of ‘‘highly credible gastrointestinal illnesses’’ among persons drinking tapwater with rates among people drinking water from reverse osmosis filtration units. Although different rates of illness could have resulted from reporting biases, if the findings are confirmed by future studies, then drinking water may significantly contribute endemic disease in at least one community. Unfortunately, a study using a similar design conducted in Melbourne, Australia, did not provide evidence of endemic waterborne disease (24). A variety of approaches have been proposed for estimating the burden of endemic diarrheal disease from drinking water sources. In addition to the Australian replication of the Payment design, a small pilot household intervention study in California has recently been completed (25). That study concluded that it was possible to blind families as to the type of treatment device they had, and although the study was not powered to examine illness rates, the families with true home treatment devices reported a lower rate of illness. A larger randomized household intervention study is under way in the United States. The advantage of the randomized household interventions is that the design precludes reporting biases and assignment biases, assuming that people do not know whether they are in the intervention or the control group. A major disadvantage of this approach is that only household drinking water quality is altered. Drinking water from other sources, such as work or at restaurants, is not altered. Another limitation is that long-term healthy residents are usually recruited and these people may have the lowest risk of suffering illness from waterborne infections. Therefore, negative results are difficult to interpret. Household intervention studies are limited in generalizability because they are conducted in single communities, although the study design would be amenable to national randomized trial. Another proposed approach is to relate variations in the occurrence of health events, such as emergency room visits and hospitalization, with variation in drinking water turbidity levels (26,27). This approach has some merit; however, the results are difficult to interpret since no causal agents are identified. There are also concerns that the optimized statistical modeling cannot be statistically evaluated. Therefore, many of the claimed associations may be spurious. Another approach uses planned changes in drinking water treatment and then evaluates the occurrence of potentially waterborne disease before and after intervention. The advantage of this approach is that most or all drinking water from an area is changed. This avoids one of the problems with household interventions. One disadvantage of this approach is that the sites receiving new water treatment technologies are not randomly assigned. For example, most unfiltered drinking water systems in the United States use high-quality source water. Adding filtration may not dramatically change the health risks from the drinking water. Another is that the community intervention looks at only one city or one pair of cities, so the sample size is restricted.

APPLICABILITY OF OUTBREAK INVESTIGATIONS Can findings from outbreak investigations be used to estimate the burden of enteric disease attributable to drinking water? Epidemic disease is defined as an unusual occurrence or clustering of a specific illness. Between 1971 and 1994 there were 737 documented waterborne disease outbreaks (28–30). Almost half of these were due to unknown etiological agents that caused acute gastrointestinal illness. Among these outbreaks, the relative importance of different etiologic agents (viruses, bacteria, protozoa, and chemicals) can be estimated. For example, the etiologic agents most commonly associated with waterborne disease in the United States include, in descending order, undefined gastroenteritis, giardiasis, shigellosis, viral gastroenteritis, and hepatitis A. This ranking is based on outbreaks and may or may not reflect the relative importance of these etiologic agents for all waterborne disease. For diseases where outbreaks account for the majority of illnesses, the outbreak is of primary interest. However, for many waterborne pathogens, outbreaks account for only a small fraction of all illnesses. For example, in a 1.5-year period during the late 1970s in Washington State, 1347 laboratory confirmed cases of giardiasis were reported to the state health department (31). Extensive follow-up of these cases (Table 3) revealed that clusters or possible small outbreaks accounted for only 16% of all cases of giardiasis reported during this time period. These data suggest that ‘‘endemic giardiasis’’ was overwhelmingly more abundant than ‘‘epidemic giardiasis’’ in Washington State during this time period. There are a number of problems with extrapolating the characteristics of cases involved in outbreaks to revise all cases of illness, including the following: 1. If there is variation in the virulence of a pathogen, then detected outbreaks may predominantly be caused by the more virulent strains of the pathogen. This may overestimate the severe morbidity or mortality associated with the pathogen. 2. By examining only detected outbreaks, one may overestimate the importance of drinking water as a

Table 3. Case Clusters of Giardiasis in Washington State 1977–1978 Number of Cases

Etiology

10 14 11 12 17 8 24 73 51 220

Untreated streamwater consumption Untreated water consumption at a work camp One small community water system Tourists returning from a resort in Mexico One-daycare center outbreak One-daycare center outbreak Among 10 different daycare centers Multiple cases among 21 families Nonfamily association with another case Total in all clusters

IMPROVING WATERBORNE DISEASE SURVEILLANCE

route of transmission. Because of the large number of cases often involved, waterborne outbreaks may be more detectable than outbreaks from other routes of transmission. Even a severe day care outbreak would involve only a few cases. Within family clusters usually involve too few cases to be a detectable outbreak. 3. Outbreak detection is often more difficult for common or endemic diseases than for uncommon diseases. For example, two cases of cholera anywhere in the United States might be considered an outbreak whereas 50 cases of cryptosporidiosis widely dispersed in a large U.S. city during a week might easily be absorbed as expected background cases of diarrhea and not recognized as an outbreak (9). Outbreaks of short duration of illnesses (e.g., some viruses) are more difficult to detect and study than are outbreaks of long duration illnesses (e.g., giardiasis, shigellosis, hepatitis A). Therefore the importance of acute, selflimited gastrointestinal illness of undetermined etiology and short duration may be underestimated relative to outbreaks of parasitic infections and some bacterial or viral pathogens with a longer duration of symptoms. Pathogens with long incubation periods are difficult to investigate since the conditions that allowed transmission of the pathogen may have changed between the time of infection and the time when the outbreak was detected. Underascertaining waterborne sources for disease outbreaks caused by these agents is likely. MONITORING INFECTION VERSUS DISEASE Only a fraction of infected persons become ill from the most commonly occurring enteric infections. Of the people that become ill, only a fraction of cases will be reposted (Fig. 3). Should expanded surveillance programs attempt to monitor infection rather than disease? The existence of asymptomatic carriers of infections has been known for some time (e.g., Typhoid Mary). However, the number of asymptomatic carriers for many infections has only relatively recently been appreciated. The parasite prevalence surveys in Britain (17) and in Virginia (18) found more asymptomatic infected persons than expected. Even as late as 1952, in New Hope, Tennessee, 10.6% of the general population was infected with Giardia lamblia (32). Following a 1966 giardiasis outbreak in Aspen, Colorado, a stool survey found that 5% of the population was infected with Giardia (33). A survey of Boulder, Colorado, also conducted following an outbreak, found a prevalence of 5% (34). Most of the individuals participating in these surveys were asymptomatic. A stool survey of one to 3-year-old Washington State children was conducted in 1980 (35). This survey found that 7% of the children were infected with Giardia lamblia. All participating children were reported as healthy at the time of the survey. The Seattle Virus Watch program, conducted during the 1960s and early 1970s monitored virus infections among a sample of people in selected U.S. cities. This study found

189

Individual is infected.

Illness occured.

Ill person seeks medical care.

Appropriate clinical test (stool, blood) ordered.

Patient complies with testing.

Laboratory is proficient and provides test(s).

Clinical test positive.

Test result reported to health agency.

Report issued in timely manner.

Health agency takes appropriate action. Figure 3. Events in reporting an individual infection.

that illness was reported in less than half of all enterovirus infections (36). New serological tools have been developed since the early 1980s to better monitor the prevalence of prior infections among the population. Even though infection may not result in moderate or severe illness, there are several reasons for considering infection rather than disease, including the following: 1. Information on infections can provide a much expanded understanding of the relative importance of various routes of transmission and provide an early warning for risks of outbreaks. 2. Serological epidemiologic studies of infection can better estimate the extent of endemic waterborne disease. These studies are statistically more powerful to detect low risks in moderate-size populations. 3. Just as the occurrence of a coliform test indicates the potential of disease risk for a drinking water source, the waterborne transmission of pathogens, even when infection is predominantly asymptomatic, can provide critical information for evaluating water treatment systems and may help identify correctable problems in water source protection and/or treatment. 4. Widespread, unrecognized transmission of infection in the general population may indicate a devastating outbreak for a susceptible subpopulation.

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Another advantage of serological surveillance occurs during an outbreak. An outbreak of cryptosporidiosis was detected in Las Vegas, Nevada in Spring 1994. Although this was clearly a cryptosporidiosis outbreak, the inability to detect problems with the water treatment system and publicity prior to the investigation that suggested the outbreak was waterborne raised questions over whether the outbreak could be classified as waterborne (37–39). Since the majority of the diagnosed cases also suffered from HIV or AIDS, the extent of the outbreak was unclear. Had asymptomatic infected persons been identified serologically, the effects of reporting bias would be reduced since asymptomatic cases would have no motivation to explain an asymptomatic infection. IMPROVING DISEASE SURVEILLANCE Several options are available for enhanced waterborne disease surveillance. The option or combination of options selected will depend on the specific goals for disease surveillance. The currently used national system of surveillance, based on diagnosed illness, has a longestablished record of both performance and nonperformance for detecting outbreaks (Table 4). Because the current system is both inexpensive to maintain and currently operational, it has considerable appeal among public health practitioners. However, monitoring pharmaceutical sales, nurse hotline calls, or physician visits is a potential enhancement to the traditional disease surveillance programs (Table 5) (39,40). The goal of our current disease surveillance system is outbreak detection. Unfortunately, there is little rigorous evaluation of its capability to detect outbreaks. Furthermore, the common occurrence of fortuitous situations that lead to the outbreak detection raise questions about the sensitivity of the system. To improve the sensitivity to detect small to medium-size outbreaks or to provide early information on the occurrence of an

Table 4. Advantages and Disadvantages of the Current Waterborne Disease Surveillance System Advantages In-place and operational across the nation Extensive health department experience using the system Inexpensive to maintain An operational nationwide network, operated by the Centers for Disease Control (CDC), for summarizing and reporting findings Methodological development of algorithms for detecting excess occurrences of disease Disadvantages Inability to detect outbreaks when diagnosed cases are not reported to the health department Delays in detecting outbreaks due to the time required for laboratory testing and for reporting of findings Undetected outbreaks where health problems are not medically treated or where infection results in only mild or no illness Limited opportunities for system improvement Possible long-term trend in healthcare delivery that may reduce its efficacy

Table 5. Advantages and Disadvantages of New Waterborne Surveillance Systems Advantages They may detect outbreaks where few patients seek healthcare or where the illness is of sufficiently short duration that healthcare is unimportant They are relatively fast in reporting outbreaks since the time delay between the onset of symptoms and the purchase of drugs or calls to nurses is likely to be short They are relatively inexpensive to maintain, especially if nationwide retail pharmacies are involved or common nurse hotline software is programmed for reporting Disadvantages Since only symptoms are ascertained, they will not usually identify an etiologic agent Although inexpensive to maintain, initial computer programming and establishing data sharing agreements would require some investment The specificity of the system for outbreak detection (e.g., number of false leads) is untested

outbreak, these alternative approaches mentioned have promise. Over-the-counter pharmaceutical sales may be useful, but it has some significant limitations (40). The use of nurse hotline calls to continuously monitor the occurrence of infectious disease has tremendous promise, but no efforts have been made to use this surveillance tool (39). Better linkages with infectious disease specialists in healthcare organizations may also improve disease surveillance. None of the traditional or enhanced surveillance tools will provide much useful information on lowlevel or endemic risk of enteric pathogen infection. However, new serological tests have increased the feasibility of studies to estimate the incidence of new infections or the prevalence of antibody response to pathogens and to relate this information with modes of transmission. In the early 1970s, the Seattle Virus Watch program examined occurrences of viral infections among volunteers in selected communities (33). Similar approaches to monitoring the occurrence of Giardia (41) and Cryptosporidium (42) infections have been developed since then. More work is needed to evaluate these new tools as well as to develop other tests. We also need to design cost-effective approaches to their widespread implementation. These tools may give us an opportunity to greatly improve our understanding of the importance of various modes of transmission and identify reasons why one population group has a higher endemic level of disease than another. It is likely that as more is known about the modes of transmission, a better understanding will emerge of both drinking water and nondrinking water routes of pathogen transmission. Healthcare reforms may reduce the use of diagnostic laboratory services, reducing the value of laboratorybased disease surveillance. However, new opportunities for improved disease surveillance, including both individual and community disease reporting and surveillance of endemic infections, may also result. To fully

IMPROVING WATERBORNE DISEASE SURVEILLANCE

exploit these opportunities, a new public health partnership with distributed responsibilities may be needed between healthcare providers, health maintenance organizations (HMOs), pharmacies, and the traditional public health agencies. The increasing age of our population has resulted in increases in the number of immunosuppressed persons. Some of this immunosuppression may result from chronic diseases, while some may result from medically induced immunosuppression following treatment for other conditions. For example, many cancer patients have temporary periods of immunosuppression following treatment. These populations may be at especially high risk of adverse consequences of infection. Since diarrheal disease in this population is also relatively common, many infections may not be detected. Infectious disease surveillance systems are operated by state and local public health agencies with little or no direct contact with healthcare providers. To improve disease surveillance system, it will likely be necessary to better integrate healthcare delivery systems with those disease surveillance programs. This integration can only occur if both the state public health agencies and the healthcare providers recognize benefits from this cooperation and barriers to data sharing are reduced.

DISCLAIMER

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7. Berkelman, R.L., Bryan, R.T., Osterholm, M.T., LeDuc, J.W., and Hughes, J.M.. (1994). Infectious disease surveillance: A crumbling foundation. Science 264: 368–370. 8. Frost, F.J., Calderon, R.L., R. L., and Craun, G.L. (1995). Waterborne disease surveillance: Findings of a survey of state and territorial epidemiology programs. J. Environ. Health. 58(5): 6–11. 9. Craun, C.L., Birkhead, G., Erlandsen, S., et al. (1994). Report of New York City’s Advisory Panel on Waterborne Disease Assessment. The New York City Department of Environmental Protection, New York. 10. Foster, L.R. (1990). Surveillance for waterborne illness and disease reporting: State and local responsibilities. In: Methods for Investigation and Prevention of Waterborne Disease Outbreaks. G.F. Craun (Ed.). EPA/600/1-90/005a. USEPA Office of Research and Development, Cincinnati. 11. Chorba, T.L., Berkelman, R.L., Safford, S.K., Gibbs, N.P., and Hull, P.E. (1989). Mandatory reporting of infectious diseases by clinicians. J. Am. Med. Assoc. 262: 3018–3026. 12. Chappell, C.L., Okhuysen, P.C., Sterling, C.R., and DuPont, H.L. (1996). Cryptosporidium parvum: Intensity of infection and oocyst excretion patterns in healthy volunteers. J. Infect. Diseases 173: 232–236. 13. Kirner, J.C., Littler, J.D., and Angelo, L.A. (1978). A waterborne outbreak of giardiasis in Camas. J. Am. Water Works Assoc. 70: 35–40. 14. Starko, K.M., Lippy, E.C., Dominguez, L.B., Haley, C.E. and Fisher, H.J. (1986). Campers’ diarrhea traced to watersewage link. Public Health Reports 101: 527–531. 15. Harter, L., Frost, F., Vogt, R., Little, A., Hopkins, R., Gaspard, B., and Lippy, E. (1985). A three-state study of waterborne disease surveillance techniques. Am. J. Public Health 75: 1327–1328.

The views expressed in this chapter are those of the individual authors and do not necessarily reflect the views and policies of the USEPA. The chapter has been subject to the Agency’s peer and administrative review and approved for publication.

16. Hopkins, R.S., Shillam, P., Gaspard, B., Eisnach, L., and Karlin, R.S. (1985). Waterborne disease in Colorado: three years surveillance and 18 waterborne outbreaks. Am. J. Public Health 75: 254–257.

BIBLIOGRAPHY

17. Smith, A.M. and Matthews, J.R. (1917). The intestinal protozoa of non-dysenteric cases. Annals Trop. Med. Parasitol. 10: 361–390.

1. D’Antonio, R.G., Winn, R.E., Taylor, J.P., Gustafson, T.L., G.W. Gray, Current, W.L., Zajac, R.A., and Rhodes, M.M. (1985). A waterborne outbreak of cryptosporidiosis in normal hosts. Annals Int. Med. 103: 886–888. 2. Bennett J.V., Holmberg, S.D., and Rogers, M.F. (1987). Infectious and parasitic diseases. Closing the Gap: The Burden of Unnecessary Illness. R.W. Amler and H.B. Dull (Eds.). Oxford University Press, New York. 3. Hayes, E.P. et al. (1989). Large community outbreak of cryptosporidiosis due to contamination of a filtered public water supply. New Engl. J. Med. 320: 372–376. 4. Leland, D., McAnulty, J., Keene, W., and Terens, G. (1993). A cryptosporidiosis outbreak in a filtered water supply. J. Am. Water Works Assoc. 85: 34–42. 5. MacKenzie, W.R. et al. (1994). A massive outbreak in Milwaukee of Cryptosporidium infection transmitted through the public drinking water supply. New Engl. J. Med. 331(3): 161–167. 6. Payment, P., Richardson, L., Siemiatycki, J., et al. (1991). A randomized trial to evaluate the risk of gastrointestinal disease due to consumption of drinking water meeting current microbiologic standards. Am. J. Public Health 81(6): 703–708.

18. Faust, E.C. (1930). A study of the intestinal protozoa of a representative sampling of the population of Wise County, southwestern Virginia. Am. J. Hygiene 11: 371–384. 19. Ungar, B.L., Milligan, M., and Nutman, T.B. (1989). Serologic evidence of Cryptosporidium infection in U.S. volunteers before and during Peace Corps service in Africa. Arch. Int. Med. 149: 894–897. 20. Frost, F. (1998). Two-city Cryptosporidium study. Am. Water Works Assoc. Research Found.—Drink. Water Research 8(6): 2–5. 21. Frost, F.J., Muller, T., Craun, G.F., Calderon, R.L., and P.A. Roeffer. (2001). Paired city Cryptosporidium serosurvey in the southwest USA. Epidemiol. Infect. 126: 301–307. 22. Frasher, G.G. and Cooke, K.R. (1989). Endemic giardiasis and municipal water supply. Am. J. Public Health 79: 39–41. 23. Birkhead, G. and Vogt, R.L. (1989). Epidemiologic surveillance for endemic Giardia lamblia infection in Vermont. Am. J. Epidemiol. 129: 762–768. 24. Hellard, M.E., Sinclair, M.I., Forbes, A.B., and Fairley, C.K. (2001). A randomized blinded controlled trial investigating the gastrointestinal health effects of drinking water quality. Environ. Health Perspect. 109: 773–778.

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25. Colford, J.M., Rees, J.R., Wade, T.J., Khalakdina, A., Hilton, J.F., Ergas, I.J., Burns, S., Benker, A., Ma, C., Bowen, C., Mills, D.C., Vugia, D.J., Juranek, D.D., and Levy, D.A. (2001). Participant blinding and gastrointestinal illness in a randomized, controlled trial of an in-home drinking water intervention. Emerging Infect. Diseases 8(1): 29–36.

42. Moss, D.M. and Lammie, P.J. (1993). J. Am. Soc. Trop. Med. Hygiene 49: 393.

DISINFECTANTS

26. Schwarz, J., Levin, R., and Hodge, K. (1997). Drinking water turbidity and pediatric hospital use for gastrointestinal illness in Philadelphia. Epidemiology 8: 615–620.

RAGHURAMAN VENKATAPATHY Oak Ridge Institute for Science and Education Cincinnati, Ohio

27. Morris, R.D.F., Naumova, E.N., and Griffiths, J.K. (1998). Did Milwaukee experience waterborne cryptosporidiosis before the large documented outbreak in 1993? Epidemiology 9: 264–270.

JUNE WEINTRAUB

28. Craun, G.F. (1992). Waterborne disease outbreaks in the United States of America: Causes and prevention. World Health Statistics Quart. 45: 192–196.

City and County of San Francisco Department of Public Health San Francisco, California

29. Moore, A.C., Herwaldt, B.L., Craun, G.F., Calderon, R.L., Highsmith, A.K., and Juranek, D.D. (1994). Waterborne disease in the United States, 1991 and 1992. J. Am. Water Works Assoc. 84(2): 87–99.

J. MICHAEL WRIGHT

30. Kramer, M.H., Herwaldt, B.L., Craun, G.F., Calderon, R.L., and Juranek, D.D. (1996). Surveillance for waterborne disease outbreaks—United States, 1993–1994. J. Am. Water Works Assoc. 88: 66–80. 31. Frost, F., Harter, L., Plan, B., Fukutaki, K., and Holman, B. (1983). Giardiasis in Washington State. USEPA Report 83134-882. National Technical Information Service, Springfield, VA. 32. Eyles, D.E., Jones, F.E., and Smith, S.C. (1953). A study of Entamoeba histolytica and other intestinal parasites in a rural west Tennessee community. Am. J. Trop. Med. 2: 173–190. 33. Gleason, N.N., Horwitz, M.S., Newton, L.H., and Moore, G.T. (1970). A stool survey for enteric organisms in Aspen, Colorado. Am. J. Trop. Med. Hygiene 19: 480–484. 34. Wright, R.A., Spender, H.C., Brodsky, R.E., and Vernon, T.M. (1977). Giardiasis in Colorado: An epidemiologic study. Am. J. Epidemiol. 105: 330–336. 35. Harter, L., Frost, F., and Jakubowski, W. (1982). Giardia Prevalence among 1-to-3 year-old children in two Washington State counties. Am. J. Public Health 72: 386–388. 36. Elveback, L.R., Fox, J.P., Ketler, A., Brandt, C.D., F.E. Wassermann, and Hall, C.E. (1966). The Virus Watch program; a continuing surveillance of viral infections in metropolitan New York families. 3. Preliminary report on association of infections with disease. Am. J. Epidemiol. 83: 436–454. 37. Craun, G.F. and Frost, F.J. (2002). Possible information bias in a waterborne outbreak investigation. Int. J. Environ. Health Research 12: 5–15. 38. Craun, G.F., Frost, F.J., Calderon, R.L., Hilborne, H., Fox, K.R., Reasoner, D.J., Poole, C., Rexing, D.J., Hubbs, S.A., and A.P. Dufour. (2001). Improving waterborne disease outbreak investigations. Int. J. Environ. Health Research 11: 229–243. 39. Rodman, J.S., Frost, F.J., and Jakubowski, W. (1998). Using nurse hot line calls for disease surveillance. Emerg. Infect. Diseases 4: 1–4. 40. Rodman, J.S., Frost, F., Burchat, I.D., Fraser, D., Langer, J., and Jakubowski, W. (1997). Pharmacy sales—a method of disease surveillance. J. Environ. Health 60(4): 8–14. 41. Nulsen, M.F., Tilley, P.G., Lewis, L., Zhang, H.Z., and IsaacRenton, J.L. (1994). The humeral and cellular host immune responses in an outbreak of giardiasis. Immunol. Infect. Diseases 4: 100–105.

Harvard School of Public Health Boston, Massachusetts

Disinfection of drinking water was instituted at a time when waterborne infectious diseases were the primary focus in public health. The use of disinfectants such as chlorine, chloramine, chlorine dioxide, ozone, and ultraviolet radiation has resulted in a reduction in the outbreak of diseases such as cholera and polio in many parts of the world. The choice of disinfectant is dependent on their availability, ease of use, cost, and their efficacy in disinfecting water. Some disinfectants commonly used in water treatment plants are not stable in water for long periods of time, which reduces their disinfection efficacy. Current practices to control pathogens in drinking water include using a combination of disinfectants, coagulation, and filtration. A main disadvantage of using disinfectants in drinking water is their ability to react with organic and inorganic matter in water to form disinfection byproducts that may be of health concern. The disinfectants can also be hazardous at high concentrations, but toxicological studies suggest that their toxicity is likely not relevant at the low concentrations encountered in drinking water. PURPOSE Contaminated, untreated, or inadequately treated water is known to transmit disease-causing pathogens such as bacteria, viruses, fungi, and protozoa (1). The size of these pathogens varies from a few hundred microns for bacterial clusters to around 0.01 µm for viruses (2). Although physical water treatment processes such as sedimentation and filtration may remove pathogens that are greater than 10 µm in size, ultra- or nanofiltration may be required to remove disease-causing pathogens (3). Conventional filtration is usually not practical for smaller systems, and, ultrafiltration may not be economically viable to treat large quantities of water; therefore, chemical treatment may have to be employed to treat water. The process of reducing the number of pathogenic microorganisms through the addition of

DISINFECTANTS

chemicals (i.e., disinfectants) into water is called disinfection. A variety of disinfectants may be used in water treatment plants, with the choice dependent on the location and size of the plant, the cost effectiveness, the pathogen(s) of concern, the water source, the characteristics of the pretreated water, and the disinfectant characteristics (e.g., solubility, stability, disinfecting and deodorizing ability, corrosiveness, and the time required to disinfect). Other factors that are considered include the effectiveness of the disinfectant in killing a range of microorganisms, the potential to form disinfection byproducts (DBPs) that may cause health effects in humans and animals, and the ability of the disinfecting agents to remain effective in water throughout the water distribution system. HISTORY Prior to A.D. 1600, civilizations consumed water based on visual clarity, or after treatments such as exposing the water to sunlight, dipping heated copper or other metals in water, boiling, and filtering water through a cloth (4). Other types of water treatment such as exposure to germicidal metals (e.g., silver and copper), sand filtration, distillation, coagulation, and adsorption with alum, lime, plant extracts, charcoal, clay, or plant materials have been employed since the 1600s. Modern methods of water disinfection were first used in Europe in the mid-to-late 1800s. One of the first known uses of chlorine as a germicide was Ignac Semmelweis’ introduction of chlorine water for hand cleansing in the Vienna General Hospital maternity wards in 1846 (5). Following research advances on the effect of microorganisms on human health in the 1870s, and improvements to physical water treatment technologies such as slow sand filtration in the 1880s, chlorine (as chloride of lime) and ozone were first used as drinking water disinfectants on a plant-scale basis at water treatment facilities in Hamburg, Germany and Oudshoorn, Holland, respectively, in 1893 (5,6). In 1897, Sims Woodhead used bleach to sterilize potable water at Maidstone, England after a typhoid outbreak. In 1903, a water treatment facility in Middlekerke, Belgium was the first to use chlorine gas as a disinfectant (4,7). In North America, the first use of chemical disinfection at water treatment plants began in 1908 when chlorine was used as a disinfectant in Jersey City, New Jersey (as sodium hypochlorite) and Chicago, Illinois (as chloride of lime). Chloramination was first used in Ottawa, Canada and Denver, Colorado in 1917, whereas ozonation was first used in the United States in the 1940s (4). Today, chemical disinfection of drinking water is widely recognized as a necessity and is widely practiced at city, community, and point-of-use levels throughout the United States. TYPES OF DISINFECTANTS Chemical disinfection is considered the most effective treatment to inactivate pathogens in drinking water. A

193

majority of the disinfectants that are in use in water treatment plants today are oxidants such as chloramine, chlorine dioxide, chlorine gas, electrochemically generated oxidant from sodium chloride (NaCl), hypochlorite, ozone, and ultraviolet (UV) light. Additional chemical disinfectants used to treat drinking water, especially in households, include acids and bases such as citric juices, lime, mineral acids and hydroxide salts, metals such as copper or silver, surfactants, and permanganate (8). Since each disinfectant has its own advantages and disadvantages (summarized in CHLORINATION BYPRODUCTS and the ALTERNATIVE DISINFECTANTS), a combination of multiple treatment processes including sedimentation, coagulation, flocculation, filtration, and disinfection is used in the majority of the water treatment plants in the world today. Table 1 provides a summary of the water treatment methods used, their availability, ease of use, cost, and efficacy in neutralizing the diseasecausing pathogens. DISINFECTION EFFICACY Disinfection kinetics is usually expressed by Chick’s law (also known as the CT law), which relates the activity [e.g., 3-log (99.9%) or 5-log (99.999%) reduction] to the product of disinfectant concentration (C) and contact time (T) (12). For example, to provide a given degree of disinfection, a low concentration and high contact time may be maintained or vice versa (13). The CT law may not provide adequate disinfection due to various factors that may affect the efficacy of the disinfectant, such as the type of pathogen; type of disinfectant; chemical factors such as pH, dissolved organic matter, salts, and ions; and particulate matter (14). Among the pathogens, vegetative bacteria are generally the easiest to disinfect followed by viruses, bacterial spores, fungal spores, and protozoan parasites (8). The formation of aggregates with other microorganisms or particulates in water may reduce the disinfectant’s efficacy by preventing access to the pathogens. Among the most commonly used disinfectants, ozone’s ability to disinfect is generally the highest followed by chlorine dioxide, electrochemically generated chlorine, and chloramine. However, ozone is the least stable disinfectant in water and is hence unable to provide a stable disinfectant residual in treated water that is necessary to prevent regrowth of the disease-causing pathogens (15). The chlorine-based disinfectants on the other hand are relatively stable in water in the absence of pathogens. At neutral pH, the half-life of chlorine dioxide is 30 min and 14 h for a 0.01 M and 0.0001 M solution, respectively (15). At ambient temperatures, the half-life of chloramines is approximately 100 h (16), whereas the half-life of sodium hypochlorite varies between 60 and 1700 d for water containing 18% and 3% available chlorine, respectively (17). The characteristics of the source water also influence the efficacy of the disinfectants. Certain disinfectants such as hypochlorite are more effective at low pH, whereas chloramines are effective at a higher pH (5,15). The disinfectants may be consumed through reactions with other dissolved constituents in the water such as dissolved

194

DISINFECTANTS Table 1. Advantages of Different Water Treatment Methods and Their Efficacy Difficulty

Cost

Pathogen Efficacy

×

$$



Adsorption Aeration Boiling Chloramines

××× × × ××

$$$$ $ $$ $$$

Variable







Chlorine gas Chlorine generated electrochemically from NaCl Chlorine dioxide

××× ××

$$$ $$$





×××

$$$$



Coagulation Combination of disinfection, coagulation, and filtration Filtration Hypochlorite

××× ××××

$$$ $$$$







××× ××

$$$$ $$$





Iodine Ion exchange Ozone

×× ×××× ××××

$$$$ $$$$ $$$$







×× ×

$$ $$



Sunlight

×

$



UV lamps

×××

$$$



Method

Availability

Acids and bases

Sedimentation Silver or copper

Comments Mostly used for pH control in water treatment plants; some microbial deactivation reported. Some bioreactors remove pathogens and organics. Has to be used with other methods to be practical. Used mostly at the point-of-use level. Less effective than free chlorine when used as a primary disinfectant; more practical as a secondary disinfectant. Widely used disinfectant. Can be generated on site. Highly effective as a primary disinfectant but is a poor secondary disinfectant. Useful in settling large flocculants of bacteria. Best method, but is accompanied by high costs. Ultrafiltration needed to remove viruses. Most widely used disinfectant; highly unstable chemical. Mostly used in tablet form at the point-of-use. Can remove salts but not pathogens. Most effective as a primary disinfectant; cannot be used as a secondary disinfectant. Useful for removing adhering bacteria. Heated metal dipped into water at the point-of-use level in lieu of boiling. Large open spaces and constant sunlight required; impractical for everyday use. Effective as a primary disinfectant against pathogens but not as secondary disinfectant; space requirements per volume of water disinfected are impractical for larger systems.

Source: References 4, 5, 8–11.

organic matter and other salts to form DBPs, thereby limiting the amount of disinfectant available for pathogen inactivation.

it can penetrate cell walls (5). Hence, the pH of the water has to be lowered to improve the disinfection efficacy. 2+ − Ca(OCl)2  Ca + 2OCl +

NaOCl  Na + OCl

CHEMISTRY OF DISINFECTANTS +

The physical and chemical properties of disinfectants can affect their behavior in drinking water as well as their toxicity. Apart from reacting with natural organic matter or other solutes to form DBPs, disinfectants undergo a number of reactions to form products that may be toxic to humans and animals.

Chlorine gas (Cl2 ) when exposed to water forms hypochlor+ ous acid (HOCl) and hydrochloric acid (HCl  H and Cl− ). + − (1) Cl2 + H2 O  HOCl + H + Cl At pH > 7.5, HOCl will dissociate to form hypochlorite ion (OCl− ) (18). In addition, calcium hypochlorite [Ca(OCl)2 ] and sodium hypochlorite (NaOCl) also immediately dissociate in water to form OCl− . HOCl is a better disinfectant than OCl− due to the relative ease with which

(2) (3)

−

H + OCl  HOCl

(4)

At pH > 12, hypochlorite ions can react to form chlorite (ClO2 − ) and chlorate (ClO3 − ), which are of major health concern (19,20). OCl− + OCl− → ClO2 − + Cl− −

Free Chlorine

−

−

−

−

OCl + ClO2 → ClO3 + Cl

(5) (6)

In waters containing bromides, hypochlorite can react with bromide (Br− ) to form hypobromous acid (HOBr), thereby consuming the available free chlorine meant for disinfection (21). HOCl + Br− → HOBr + Cl−

(7)

Chlorine Dioxide Chlorine dioxide (ClO2 ) dissolves in water under alkaline conditions to form chlorite (ClO2 − ) and chlorate

DISINFECTANTS

195

Table 2. Disinfectant Use in the United States

Disinfectant Chloramines Chlorine dioxide Chlorine gas Hypochlorite Ozone

Medium and Large Systems (>10,000 People)

Small Systems ( κT

(8)

This energy is not supplied in the form of heat but through the impact either of an electron, ion, atom, molecule, or photon or by application of electric, magnetic or other fields (15). PAHs are expected to be less refractory (i.e., more reactive toward free radicals) than benzene. As previously discussed, the main problem for these processes is low solubility. However, some of them (e.g., Fenton’s reagent) do not depend on light transmittance and can be readily implemented (11). NOMENCLATURE a b E f foc G H K KD Koc Kow L 1/n R S x

Constant Constant Energy ML2 /t2 Fugacity Pa Fractional organic carbon in solid Gas rate in stripping M/t/L2 Henry’s law constant Freudlich isotherm coefficient (capacity constant) mg/g*(L/mg)1/n Distribution coefficient, L/M Organic-carbon distribution coefficient, L/M Octanol–water distribution coefficient, L/M Liquid phase or liquid rate in stripping, M/t/L2 Freundlich bonding constant Stripping factor Solid phase Concentration, M/L3

V κ

575

Vapor phase Boltzmann’s constant J/K

BIBLIOGRAPHY 1. U.S. Environmental Protection Agency. (1980). POM Source and Ambient Concentration Data: Review and Analysis. EPA Report No. 600/7-80-044, Washington, DC. 2. U.S. Environmental Protection Agency. (1975). Scientific and Technical Assessment Report on Particulate Polycyclic Organic Matter (PPOM). EPA Report No. 600/6-75-001, Washington, DC. 3. Watts, R.J. (1998). Hazardous Wastes: Sources, Pathways, Receptors. John Wiley & Sons, New York. 4. U.S. Environmental Protection Agency. (1978). Health Assessment Document for Polycyclic Organic Matter. EPA Report No. 2/102, External Review Draft. Research Triangle Park, NC, pp. 3–1 to 3–47. 5. Neff, J.M. (1979). Polycyclic Aromatic Hydrocarbons in the Aquatic Environment: Sources, Fates and Biological Effects. Applied Science, Essex, UK. 6. Morrison, R.T. and Boyd, R.N. (1978). Organic Chemistry, 3rd Edn. Compounds. Allyn and Bacon, Chap. 30. 7. WHO. (1987). Polynuclear Aromatic Hydrocarbons (PAH). Air Quality Guidelines for Europe. Copenhagen, World Health Organization Regional Office for Europe, pp. 105–117. 8. Agency for Toxic Substances and Disease Registry. (1994). Toxicological Profile for Polycyclic Promatic Hydrocarbons (PAHs): Update. US Department of Health and Human Services, Public Health Services, Atlanta, GA. 9. IARC. (1983). Polynuclear Aromatic Compounds. Part 1. Chemical, Environmental and Experimental Data. Lyon, International Agency for Research on Cancer. (IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Humans, Vol. 32). 10. Ashworth, S.C. (1998). K Basin Sludge Polychlorinated Biphenyl Removal Technology Assessment. HNF—3095 DOE Information Bridge. Available: http://www.osti.gov/bridge/. 11. Ashworth, S.C. (2004). Hydrocarbon Treatment Techniques. IW71. 12. Thibodeaux, L.J. (1979). Chemodynamics. Wiley-Interscience. 13. Hemond, H.F. and Fechner, E.J. (1994). Chemical Fate and Transport in the Environment. Academic Press. 14. Pontious, F.W. (1990). Water Quality and Treatment, 4th Edn. McGraw-Hill, New York. 15. Bugaenko, L.T., Kuzmin, M.G., and Polak, L.S. (1993). HighEnergy Chemistry. Ellis Horwood PTR Prentice-Hall.

HYDROCARBON TREATMENT TECHNIQUES SAMUEL C. ASHWORTH Idaho National Engineering & Environmental Laboratory Idaho Falls, Idaho

There are various criteria concerning hydrocarbons for water needs. Some water users have little concern for dissolved hydrocarbons (e.g., agriculture), and some have strict standards (e.g., drinking water). In general the presence of organic compounds, including hydrocarbons,

576

HYDROCARBON TREATMENT TECHNIQUES

in water can lead to a buildup of biofilms. Biofilms can be problematic in many situations including drinking water, cooling water, recirculating water, and water used for various other industrial processes. During biofilm formation, the first substances on the surface are not bacteria but trace organics (1). Almost immediately after the clean pipe surface comes into contact with water, an organic layer deposits on the water/solid interface (2). These organics form a conditioning layer that neutralizes excessive surface charge and surface free energy that may prevent a bacteria cell from approaching near enough to initiate attachment. In addition, the adsorbed organic molecules often serve as a nutrient source for bacteria. Though there are many criteria, the thrust of this article is on the treatment, including the removal and destruction of hydrocarbons for whatever purpose. It is assumed in this article that the hydrocarbons are dissolved in water, so that technologies such as incineration or decantation will not apply. The focus will be on waterphase separation and destruction. No economic evaluation is considered. In reality, the trade-offs would need to be considered, for example, waste disposal costs, electrical cost, and materials.

BACKGROUND In general, water from natural sources will lead to poor cooling performance unless adequate steps are taken to remove hydrocarbon contaminants (3). The extent of the problem, however, will depend on the quality of the raw makeup water. The deposits that occur on heat exchanger surfaces are complex and are likely to include particulate matter, crystalline salts, corrosion products, and biofilms. It is necessary to counteract this problem of deposition onto surfaces to maintain heat exchanger effectiveness. The shape and structural arrangement of a biofilm growing in a flowing fluid will influence the mass transfer characteristics of the biofilm system as well as the drag force exerted on individual biofilm structures (4). If the biofilm is a highly compliant material, the shape will vary through the growth cycle of the biofilm and also due to variations in fluid shear stress (5). Fluctuations in biofilm shape will also affect the hydrodynamic drag that in turn will influence the detachment rate and pressure losses in a flowing system. In addition, it is thought that biofilm viscoelasticity may explain the large pressure drops observed in biofilm fouled pipes (6). There are also regulatory standards for certain waters (e.g., drinking water and water discharged to aquifers and bodies of water). Some of these that might apply include the Safe Drinking Water Act (42 U.S.C. 300f et seq.), Clean Water Act (33 U.S.C. 1251 et seq.), Resource Conservation and Recovery Act (42 U.S.C. 6901 et seq.), and a host of others. Many of them provide standards for organic compounds including hydrocarbons in effluents and source waters. The above are a few reasons for removing organic compounds but there are others, including the pharmaceutical and semiconductor industries that have specific needs and criteria in this regard.

TREATMENT Separation Separation is the process of removing a compound from one phase to another (e.g., activated carbon adsorption). One of the simplest methods for separating volatile hydrocarbons and other organic compounds from water is air stripping. An example is benzene, a ring molecule shown in Fig. 1. A typical stripper is a packed system using Raschig rings or other packing that provides high surface area, as shown in Fig. 2. Descending water containing dissolved benzene enters the top of the column and is distributed to flow through the packing. The air enters at the bottom contacts the descending water, and the benzene transfers from the water to the air, an interphase mass transfer operation. The clean water is used for the purposes required, and the air is ejected to the atmosphere, regulations permitting, or treated accordingly (e.g., vapor-phase activated carbon). The driving force for mass transfer is the gradient from the water to the air. At the interface, it is normally assumed that the interfacial water concentration and the air partial pressure are in equilibrium. If the amount of dissolved organic in the water is small enough (water mole fraction not appreciably different from 1.0), equilibrium is expressed using Henry’s law (7). Henry’s law is a special case of the following relation for gas–liquid

Figure 1. Benzene molecule.

Vent system or treatment

Water and dissolved benzene

Air

Clean water Figure 2. Packed-bed air stripper.

HYDROCARBON TREATMENT TECHNIQUES

Liquid bulk concentration

yi φi PT = xi Pvap γi pi =

Interface

thermodynamic equilibriuma :

Pvap γi xi = Hxi φi Ci

Or the partial pressure of i equals Henry’s constant times the liquid mole fraction of i where H (Henry’s constant) is constant for small mole fractions in the liquid phase. The rate of mass transfer is based on using Henry’s constant. For a dissolved gas, there are two resistances, gas-phase and liquid-phase. The mass transfer (F) flux is

Pi

Air bulk partial pressure

F = KL (CB − C∗ ) = KG (p∗ − pB ) The C∗ and p∗ above are virtual properties; C∗ is the concentration in equilibrium with the bulk pressure if there were a liquid present. Similarly, p∗ is the partial pressure that would be in equilibrium with the bulk liquid. Figure 3 illustrates these relations. Because C∗ is not known, pB /H is substituted for it. Similarly, CB H is substituted for p∗ . This is the easiest form as all that needs to be known are the bulk liquid concentration and bulk partial pressure. The mass transfer based on individual coefficients is (see Fig. 4) F = kG (pi − pB ) = kL (CB − Ci ) By combining these and using Henry’s Law, the following is determined for the overall mass transfer coefficients either one can be used: KL =

1 1 1 + HkG kL

KG =

1 1 H + kG kL

p = f (C ) p*

Direction of mass transfer

Figure 4. Interphase mass transfer.

There are many references that cover this thoroughly (8,9,11) but if Henry’s Law applies, the height of a stripping column is h = HTU ∗ NTU L HTU = KL a NTU = R=

R Ci /Co ∗ (R − 1) + 1 ∗ ln R−1 R HG L

Using benzene as an example with an MCL of 5 µg/L, the height will be determined to achieve this starting with 50 µg/L and the data in Table 1. This would take a stripper approximately 5 m high (15–20 ft) to meet the MCL for benzene. The R parameter is called the stripping factor and must be greater than one (R > 1) for stripping to be feasible. As observed by its definition, a small H can be somewhat overcome by a large G/L. The trend is such that volatile materials are easy to strip, whereas the semivolatiles and large organic compounds that have low vapor pressures are difficult to strip. Adsorption

pi pB C*

Ci

CB

Figure 3. Stripper concentration and partial pressure relations. a

577

a Actually, Hi ≡ limxi →0

fiL xi

The use of activated carbon for organic removal from a water source or effluent is another very common and useful separation process. There are two types in use, granular and powdered. Only the granular type is discussed here; the powdered type is covered in other sources (10). Normally, the system is designed so that water flows into the top of the column and exits the bottom, as shown in Fig. 5. Some methods for multicomponent adsorption are available. One is the multicomponent Langmuir that doesn’t always give good predictions (11). There has also been some work on multicomponent systems (2). Some

578

HYDROCARBON TREATMENT TECHNIQUES

Water containing organic compound(s)

Receiver for usage or further treatment

Figure 5. GAC system.

Table 1. Data for Air Stripping Benzene Given Data

Values

Ci , µg/L Co , µg/L G, m3 /m2 /h H, atm L/mg H, Dimensionless KLa , hr−1 L, m3 /m2 /h Flow, gpm

50 5 1000 5.52E−05 1.80E−01 10a 20 10

Calculated R HTU, m NTU h, m A, m2 D, m

9.00 2 2.47 4.94 0.11 0.38

a This was an assumed value. Note that KL a is a combined, overall liquid mass transfer coefficient where a is the surface area to volume ratio of the packing and is usually measured or estimated as a single quantity. It is found similar to the KG and KL and there are correlations for it in many references, e.g., 8–10.

computer software is available for sizing but is limited to 8–10 components. Vendors also have programs and rule of thumb methods available. Freundlich isotherms for many of the compounds are readily available, and hence, this method is used for single organic compounds. The Freundlich isotherms are sometimes used for scoping multicomponent systems. However, the Freundlich method is not amenable to multicomponent systems. Therefore, it is assumed that they are additive, each compound adsorbs independently of the others without competition. Column dynamic testing of the actual liquid is preferred for multicomponent systems. The Freundlich isotherm is explained by qe = KC1/n e The additive type of method assumes that the organic compounds adsorb in layers; the strongest bonding

molecules profile toward the top of the equilibrium zone, and the less strongly bonded toward the bottom (K is related to capacity, and 1/n is related to bonding strength). This is shown ideally in Fig. 6. It shows that the organic compounds that have the highest affinity for GAC profile toward the top of the column and those of low affinity profile toward the bottom. Hence, breakthrough of benzene would occur first if in a multicomponent system. The procedure is to determine the amount of GAC for a component individually via the method of Snoeyink (Pontius 1990) (11) and determine the amount of GAC required. Using benzene again at 50 µg/L (it does not have a very favorable isotherm), qe,Benzene = 1 ∗ 0.051.6 = 0.0083 mg/g The amount of water per GAC, according to Snoeyink is, Y=

0.0083 mg/g qe ∗ ρGAC = ∗ 500 g/L Ci − Co 0.05–0.005 mg/L = 92 gal/galGAC

The amount of GAC is then the volume to be treated/Y. For 1000 gallons this is GAC =

1000 = 11 gal 92

or 11 gallons of GAC per 1000 gallons treated. To avoid immediate breakthrough, the bed length must be greater than the mass transfer zone (MTZ). The MTZ is calculated from column dynamic tests as MTZ = L

Arochlors PAHs

ts − tb ts

Eqilibrium zone Mass transfer zone

Other SVOCs VOCs Benzene

Unaffected zone

Figure 6. Zones in a GAC column.

HYDROCARBON TREATMENT TECHNIQUES

The most important GAC adsorber design parameter is contact time, most commonly described by its empty bed contact time (EBCT) (11), V/Q. The empty bed contact time (EBCT) is an important design parameter in GAC column design. The range for 47 plants was 3–34 minutes, and the median was 10 minutes (Pontius 1990)(11). Longer contact times are normally used for highly concentrated solutions, so this design uses the value of 5 minutes. Based on this, the minimum column volume is Vcolumn = EBCT ∗ Q = 5 min ∗ 10 gpm = 50 gal A typical hydraulic loading is 3–4 gpm/ft2 . Using 10 gpm, the area is 2.5 ft2 . The height is h=

50 gal ∗ ft3 /7.48 gal 2.5 ft2

= 2.67 ft

Destruction The final technologies discussed for removing organic compounds from water are destruction/reaction techniques based on high-energy chemistry. High-energy chemistry is defined as E > κT So-called advanced oxidation technology is a subset of high-energy chemistry that uses free radicals to oxidize the target compound. All of these processes work by generating free radicals followed by attack on the target’s bonds by the free radicals. All of these processes are commercially available. However, some of them have certain restrictions on their use (e.g., nitrates, liquid only,). The following are some of the processes comprising advanced oxidation. The UV/H2 O2 process consists of a chemical reactor (batch or flow-through) that uses UV light to produce free-radical hydroxyls (Fig. 7) H2 O2 + hν → OH· The free radicals extract electrons from the target and create a free radical target product. The ensuing mechanism is a chain reaction eventually degrading the target to CO2 and H2 O and other simple, less regulated compounds. The overall reaction can usually be modeled by first-order kinetics. The design relation used for this type of reactor is the EE/o, the electrical energy per volume per order of magnitude (in kwh/1000 gal/order). This relationship provides all of the information required to determine efficiency (i.e., once known, any efficiency can be obtained by adjusting the power). There are several similar systems that produce freeradical hydroxyls, including • • • •

UV/ozone ozone/peroxide Fentons’ reagent UV-vis/Peroxide

It has been speculated that UV is required for some compounds like polychlorinated biphenyls (PCBs), to

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activate the bonds, and OH− is required to react with the activated bonds (i.e., synergism) (13). For some compounds (e.g., NDMA), direct irradiation of organic pollutants by high-intensity UV light provides a significant destructive pathway. The target compounds degrade after absorbing UV light. The target must strongly absorb UV. The hydrated electron is a very strong reducing agent that can react with halogenated alkanes and alkenes. The hydrated electron interacts with the chlorine–carbon bonding electrons and provide energy to break the bond and demineralize the target. There are several commercially available processes involving reduction via aqueous electrons. The aqueous electron is produced by several mechanisms including nuclear and high-energy processes, photochemistry, and chemical. Some processes require transferring high-energy electrons through thin films of water because the free path or linear energy transfer (LET) is small for electrons (high voltage process). The baseline process is a UV-catalyzed process. The chemical added that is shown is a proprietary catalyst. The catalyst interacts with UV light shown by the energy arrow that produces the aqueous electron (e− aq ). An example of this production reaction is −3 − Fe(CN)−4 6 + hν → Fe(CN)6 + eaq

The aqueous electron then interacts with the chlorine–carbon bond producing chloride ion and a free radical chlorinated aromatic: − − e− aq + C12 H5 Cl5 → C12 H5 Cl4 + Cl

The free radical goes on to extract electrons from another molecule, and a chain reaction usually occurs. The rate constants for reductions are normally quite large. This process is particularly well suited to compounds not amenable to advanced oxidations, for example, CCl4 . The UV-induced process is commercially available (e.g., Calgon Advanced Oxidation Technologies). Supercritical water oxidation, sometimes known as hydrothermal waste processing, uses the solvating traits of water in its supercritical state to destroy liquid organic wastes. As water is heated beyond its critical temperature (374.1 ◦ C) and critical pressure (250 Mpa-s, about 3219 psi), the density of the water drops dramatically (typical operating densities are 0.15–0.2 g/cm3 ). These changes in density and hydrogen bonding make organics highly soluble, and inorganic substances become nearly insoluble. The organic material is dissolved in an oxygenrich environment where conversion occurs rapidly due to the high temperature of the process. Under such high pressure and temperature, organic materials are rapidly decomposed by oxidation at removal efficiencies of 99.9999% or greater. Ultrasound can induce unusual high-energy chemistry through the process of acoustic cavitation, the formation, growth, and implosive collapse of bubbles in a liquid. Cavitation can occur in both clouds of collapsing bubbles (multibubble cavitation) or with high symmetry for isolated bubble (single-bubble cavitation, SBC). Multi-bubble cavitational collapse produces localized, transient hot spots

580

HYDROCARBON TREATMENT TECHNIQUES

with intense local heating (approx > 5000 K), high pressures (>2000 atm), and short lifetimes (sub-microsecond) in an otherwise cold liquid. From hydrodynamic modeling of this cavitational collapse, it has been estimated that both the heating and cooling rates are in excess of 1010 K/s. Acoustic cavitation is a unique means of creating high-energy chemistry, easily and inexpensively. Aqueous sonochemistry produces, supercritical water conditions on a microscopic scale. In this regime of temperature and pressure, the sonochemistry of water is an extreme limiting case of supercritical phenomena and is closely related to hydrothermal oxidation. For example, the ultrasonic irradiation of water produces a variety of extraordinarily reactive species (including OH− , H+ , and HO− 2 ) that can decompose many organic compounds. The last design example is treating benzenecontaminated water to the same degree as the previous two examples using UV/H2 O2 . Figure 7 illustrates a schematic batch test reactor. Benzene has an EE/o = 2–5 kWh/1000 US gal/order (14). Therefore, the dosage of UV required in a flow-through system at 10 gal/min is (using 5 for the worse case) and the industry design relation (14):

The first-order reaction coefficient is inversely proportional to the EE/o: ln 2 ∗ P(kW) k= V ∗ EE/o ∗ log 2 The power for a batch reactor is then the same used for the industry design equation (14): P(kW) =

ln(Ci /Co ) ∗ V ∗ EE/o ∗ log 2 ln 2 ∗ t

In a plug flow reactor (see Fig. 8), a material balance on a differential volume element yields: QC − Q(C + dC) − kCdV = V

dC dt

At steady-state, ln

Ci V =k Co Q

Power(kWatt) 5 kWh ∗ 10 gal/min/order ∗ log10 (50/5) ∗ 60 min/hr 1000 = 3 kW

=

So a UV lamp of 3 kW is sufficient to meet the requirement. Of course, if the flow rate were 100 gpm, it would require 30 kW, and so on. There will be economic trade-offs, electrical power costs, and chemical costs. It is highly useful at this point to consider reaction kinetics and reactor design as there are some unusual features for this type of reactor. For a batch system, as shown in Fig. 8, the material balance and first-order reaction are dC = −kC dt   Ci ln = kt Co

Figure 7. Batch reactor with lamp and agitator.

Water containing organic compound(s)

Receiver for usage or further treatment

Figure 8. Plug flow UV system.

USE OF ANAEROBIC-AEROBIC TREATMENT SYSTEMS FOR MAIZE PROCESSING PLANTS

The power, based on this k, is then the industry design equation (14). P(kW) =

EE/o ∗ log 2 ∗ Q ∗ ln(Ci /Co ) ln 2

For a continuously stirred reactor (CSTR), the steady-state material balance yields Ci kV = +1 Co Q This does not yield the industry design equation. The answer to this seemingly paradoxical result is that the plug flow design applies, regardless of the reactor type for flow-through reactors. This is so because the reaction must be in the vicinity of the lamp surface; a solute in a CSTR must, at some time in its residence history, approach the lamp sufficiently close. This is one reason that flow-through reactors require appropriate design to obtain turbulence. Nomenclature A CB Ce Ci Co C∗ D EE/o F G H hν HTU k K KG KL kG kL KLa L NTU 1/n P PT pB pi Pvap p∗ qe Q R tb ts V xi

Cross sectional area Bulk liquid concentration Equilibrium concentration Inlet concentration Outlet concentration Fictitious concentration in vapor phase Diameter Electrical energy/1000 gallons/order Mass transfer flux Gas mass velocity Henry’s Law constant Photon energy Height of a transfer unit 1st order reaction coefficient Freundlich constant (related to capacity) Overall mass transfer coefficient based on gas Overall mass transfer coefficient based on liquid Gas phase mass transfer coefficient Liquid phase mass transfer coefficient Overall liquid phase mass transfer coefficient Liquid mass velocity, length of column Number of transfer units Freundlich exponent (related to bonding strength) Power Total pressure Bulk gas partial pressure Partial pressure at interface Pure component vapor pressure Fictitious pressure in liquid phase Quantity adsorbate/adsorbent Liquid flow rate Stripping factor Breakthrough time Saturation time (influent = effluent) Volume Liquid phase mole fraction

yi γi κ φi

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Gas phase mole fraction Activity coefficient Boltzmann’s constant Fugacity coefficient

BIBLIOGRAPHY 1. Edstrom. (2002). Available: Resources.cfm?doc id=23.

http://www.edstrom.com/

2. Mittelman, M.W. (1985). Biological fouling of purifiedwater systems: Part 1. Bacterial Growth and Replication, Microcontamination 3(10): 51–55, 70. 3. Bott, T.R. (1998). Techniques for reducing the amount of biocide necessary to counteract the effects of biofilm growth in cooling water systems. Applied Thermal Engineering 18(11). 4. Stoodley, P., Lewandowski, Z., Boyle, J.D., and LappinScott, H.M. (1999). Structural deformation of bacterial biofilms caused by short-term fluctuations in fluid shear: An in situ investigation of biofilm rheology. Biotechnology and Bioengineering 65(1). 5. Boyle, J.D., Dodds, I., Stoodley, P., and Lappin-Scott, H.M. (1997). Stress management in biofilms. In: Biofilms: Community Interactions and Control. J.W.T. Wimpenny, P.S. Handley, P. Gilbert, H.M. Lappin-Scott, and M. Jones (Eds.). BioLine, Cardiff, UK, pp. 15–22. 6. Picologlou, B.F., Zelver, N., and Characklis, W.G. (1980). Biofilm growth and hydraulic performance. J. Hydraul. Div. Am. Soc. Civ. Eng. 106(HY5): 733–746. 7. Prausnitz, J.M., Lichtenthaler, R.N., and de Azevedo, E.G. (1986). Molecular Thermodynamics of Fluid-Phase Equilibria. Prentice-Hall, Englewood Cliffs, NJ 8. Perry, R.H. and Green, D.W. (1984). Perry’s Chemical Engineers’ Handbook, 6th Edn. McGraw-Hill, New York. 9. Treybal, R.E. (1987). Mass-Transfer Operations. McGrawHill, New York. 10. Perrich, J.R. (1981). Activated Carbon Adsorption for Wastewater Treatment. CRC Press, Boca Raton, FL. 11. Pontius, F.W. (1990). Adsorption of organic compounds. Water Quality and Treatment. 4th Edn. McGraw-Hill, New York. 12. Khan, A.R., Al-Bahri, T.A., and Al-Haddad, A. (1997). Adsorption of phenol based organic pollutants on activated carbon from multi-component dilute aqueous solutions. Wat. Res. 31(8): 2102–2112. 13. Ashworth, S.C. (1998). K Basin Sludge Polychlorinated Biphenyl Removal Technology Assessment. HNF–3095 DOE Information Bridge. Available: http://www.osti.gov/bridge/. 14. Calgon Carbon Corporation. (1996). Advanced Oxidation Technologies Handbook 1(1): MI-AOT-10/96.

USE OF ANAEROBIC-AEROBIC TREATMENT SYSTEMS FOR MAIZE PROCESSING PLANTS ´ ´ ´ MARIA-DEL-CARMEN DURAN-DE-BAZ UA

UNAM, National Autonomous University of Mexico University City, M´exico

After the discovery of the American continent by the Spaniards in the fifteenth century, corn (Zea mays), an indigenous plant from Mexico, spread all over the

582

USE OF ANAEROBIC-AEROBIC TREATMENT SYSTEMS FOR MAIZE PROCESSING PLANTS

world and became a staple food for many human groups. However, the traditional ways of eating the many products derived from this cereal in Mexico, inherited from the Mayan, Aztec, and other Mexican indigenous groups, were not disseminated until the last 20 years. Due to globalization, one of these maize products, ‘‘tortillas,’’ is now an extremely popular food that can be found in most countries of the world. ‘‘Tortillas’’ are a sort of unleavened bread in circular form that, due to a mild flavor, can be combined with vegetables, meats, pulses, etc. They are the bread equivalent to ‘‘chapattis’’ for the people from India and ‘‘falafel’’ for the people of the Middle East (Fig. 1). The traditional production of ‘‘tortillas’’ involves a Precolumbian technique, known as ‘‘nixtamalization,’’ derived from the Aztec words nextli = lime ashes and tamalli = cooked corn dough. This ancient process, almost as old as corn domestication and cultivation, is a time- water- and energy-consuming technique (Fig. 2). Modernization of the traditional process to produce cornmeal instead of dough to lengthen its shelf life as well as some other changes for mass production have been introduced in the last 50 years. However, these changes affect the sensory characteristics of ‘‘tortillas,’’ mainly ‘‘rollability,’’ ‘‘sturdibility,’’ and softness, so it has been a common practice among large-scale producers to introduce some chemical additives to make the new ‘‘supermarket tortillas’’ desirable from the sensory point of view. Unfortunately, from the nutritional point of view, it has not yet been proved that these additives do not affect the health of consumers. Some innovative processes that maintain these desirable nutritional and sensory characteristics, and at the same time, reduce energy and water consumption, and most importantly, processing time, have also been developed and started to be used (2–5). However, the traditional plants (more than 100,000 small-scale and 25 large-scale in Mexico) are still in operation, so methods to treat and stabilize the wastewater generated in the traditional process (socalled ‘‘nixtamal’’ mills that process around 0.5 to 4 metric tons maize per day) or in its industrial modifications (so-called precooked cornmeal or masa-harina factories that process from 100–1,000 metric tons maize per day) and contain appreciable amounts of soluble and insoluble organic matter (Table 1) have also been developed. These methods are based on anaerobic and aerobic systems, and the interesting issue is that by-products from the treatment can be further used, giving added value to these treatment processes. The biological processes developed were screened out after testing at laboratory level. These processes are commonly used in biological wastewater treatment, such as activated sludge, aerated lagoons, facultative lagoons, low and high rate anaerobic reactors, etc. (7). Figure 3 shows the method developed for treating the wastewater from maize alkaline processing to produce ‘‘tortillas,’’ particularly for those factories that process considerable amounts of corn. After primary settling to recover broken maize pieces and peelings, the anaerobic process is carried out to transform most of the biodegradable dissolved matter in the wastewater into methane-rich biogas. This is followed by the aerobic process that polishes

(a)

(b)

Figure 1. (a) Modern ‘‘tortilla’’ making machine. (b) Traditional ‘‘tortilla’’ hand-making (1).

the treatment, removes the traces of sulfide and other undesirable compounds in treated water, and converts the remaining biodegradable matter into protein-rich biomass. A tertiary treatment using activated carbon or any organic adsorbent will render recyclable water for the process. The water absorbed by corn is added as makeup freshwater to the overall process. The recycling of the by-products generated is shown in Fig. 3. The amount of methane from biogas obtained from the anaerobic treatment of the wastewaters is 9.6–16.8 m3 per metric ton of maize processed (considering 5 m3 of wastewater produced per metric ton maize at a conversion of 80% of the dissolved biodegradable matter into biogas). If this gas contains roughly 80% methane, the yield of this energyrich gas is 7.7–13.48 m3 methane per metric ton of maize processed. This methane-rich biogas may be washed and compressed to be used as a secondary source of energy during the lime-cooking of corn grains (nixtamalization). The amount of biomass and suspended solids obtained may be processed by extrusion to form pellets or flakes for fish feedstuff, considering the mass balances from the aerobic and anaerobic processes is of 23 kg SS and 10.6 kg biomass per metric ton of maize cooked. These biomass and suspended solids pellets or flakes proved comparable

USE OF ANAEROBIC-AEROBIC TREATMENT SYSTEMS FOR MAIZE PROCESSING PLANTS

Operating conditions

583

Nixtamalization Modern Traditional 8 20 3:1 6:1 2:1 (DBO5 extremely high) 5:1 (high BOD5) 1 (basis) 0.75

Processing time, h Water consumption, water:grain ratio Wastewaters Energy consumption

Raw corn grain

Energy

Energy

Nixtamalization (first cooking)

Wastewaters (2:1) Water (3:1) Lime (1%)

Rinsing

Wastewaters (2:1) Water (3:1)

Grinding (corn dough)

Energy

Energy

Tortillas (second cooking)

Energy

Rehydration and corn dough formation

Drying and regrinding

Energy

Water (1:1)

Precooked cornmeal

Figure 2. Diagram of the traditional nixtamalizati´on process (2).

Table 1. Maize Lime-Cooking (Nixtamalization) Wastewater Average Compositiona Characteristics 3

Suspended solids, kg/m Dissolved organic carbon, kg/m3 Biochemical oxygen demand (5 days, 20 ◦ C), kg/m3 Chemical oxygen demand (dissolved), kg/m3 Nitrogen (Kjeldahl), mg N/L Phosphates, mg PO4 −3 /L pH Value Color a b

Average Values

UE Normativityb for Surface Water

2.4–4.6 3.0–5.0 1.5–3.0 7.5–11.0 80–270 7–18 10–14 Dark yellow

0.025 (25 mg/L) — chlorine (Cl2 ) > iodine (I2 ) > hydrogen peroxide (H2 O2 ). Chlorine is probably the most effective and least expensive of all oxidizing and nonoxidizing biocides (26). The level of activity of chlorine against attached biofilms

Table 2. Typical Biocide Dosage Levelsa Biocide Chlorine Ozone Chlorine dioxide Hydrogen peroxide Iodine Quaternary ammonium compounds Formaldehyde Anionic and nonionic surfactants

Dosage Level, mg/L

Contact Time, h

50–100 10–50b 50–100 10% (v/v) 100–200 300–1,000

1–2 Cu (52), and that for FeOOH sites is Pb > Zn ≥ Co ≥ Ni > Cu (34). Metals are also complexed with a range of organic material such as algae, bacterial cells, detritus, and organic coatings on mineral surfaces (40,47,53,54). The organic adsorption affinity of metals decreases in the order Cu = Fe Zn Mn (55,56). Adsorption and surface complexation are rapid processes producing pollutant attenuation from the water column but are transitory due to marked pH dependency (33,41). Ions immobilized as surface complexes are rapidly released during changes in effluent pH conditions.

the microbiology of acid mine waters. The processes that use or depend on, microbial action have been described above. Hence, listed here are the most common examples of microbial involvement in processes of mine effluent remediation: • • • • •

oxidation of Mn (71,72) oxidation and hydrolysis of Fe (73,74) oxidation of sulfides (11) sulfate reduction products of organic decay (CO2 , organic acids) • surfaces for adsorption

NH4 + ,

and

BIBLIOGRAPHY The Effect of Aquatic Vegetation The uptake and accumulation of metals in plant biomass has been extensively studied under laboratory conditions (57–59). Dunbabin and Bowmer (40) investigated the partitioning of metals within emergent hydrophytes and found that the majority of metals were stored in the roots and rhizomes. Floating wetland plants have been shown to hyperaccumulate Cu and Fe up to 78 times their concentration in wastewater (57). However, direct uptake and accumulation of metals within plant biomass usually constitutes a minor component of the overall removal processes (14,60–62). In addition to metal uptake, wetland plants attenuate metal contaminants through a number of further mechanisms: • Plants release oxygen through their roots, creating a zone of aerobic conditions in the substrate (14,40). The oxidizing conditions within the substrate increase metal oxidation with insoluble precipitate formation [commonly Fe (oxy)hydroxide plaques]. These precipitates provide a further surface for trace metal adsorption (63). • Bacterial biofilms forming on root surfaces are strong adsorbents of trace metals (63). • The turnover of aquatic vegetation is the dominant source of simple carbohydrates, which are the main substrate for sulfate-reducing bacteria and other fermentative organisms important in the production of sulfides, ammonium, carbon dioxide, and alkalinity (14,40,64,65). • The humic substances of organic decay play an important role in aquatic contaminant chemistry through the adsorption of metals and subsequent transport as colloidal complexes or incorporation into the sediment as settled solids. The Role of Microbial Processes Microbial activity is integral to many of the processes described above. The importance of the microbiology of mine effluent remediation has recently been recognized, as evidenced by numerous studies in this discipline (66–69). Nordstrom (70) provides a good review of the state of knowledge and the gaps in the present understanding of

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27. Wieder, R.K. (1993). Ion input/output budgets for 5 wetlands constructed for acid coal-mine drainage treatment. Water Air Soil Pollut. 71(3–4): 231–270. 28. Cravotta, C.A. and Trahan, M.K. (1999). Limestone drains to increase pH and remove dissolved metals from acidic mine drainage. Appl. Geochem. 14(5): 581–606. 29. Cravotta, C.A. (2004). Size and performance of anoxic limestone drains to neutralize acidic mine drainage. J. Environ. Qual. 33(3): 1164–1164. 30. Xie, J.Z., Chang, H.L., and Kilbane, J.J. (1996). Removal and recovery of metal ions from wastewater using biosorbents and chemically modified biosorbents. Bioresource Technol. 57(2): 127–136. 31. Kratochvil, D. and Volesky, B. (1998). Biosorption of Cu from ferruginous wastewater by algal biomass. Water Res. 32(9): 2760–2768. 32. Utgikar, V., Chen, B.Y., Tabak, H.H., Bishop, D.F., and Govind, R. (2000). Treatment of acid mine drainage: I. Equilibrium biosorption of zinc and copper on nonviable activated sludge. Int. Biodeterioration Biodegradation 46(1): 19–28. 33. Langmuir, D. (1997). Aqueous Environmental Geochemistry. Prentice-Hall, Englewood Cliffs, NJ. 34. Dzombak, D.A. and Morel, F.M. (1990). Surface Complexation Modelling: Hydrous Ferric Oxide. John Wiley & Sons, New York. 35. Rose, S. and Elliott, W.C. (2000). The effects of pH upon the release of sulphate from ferric precipitates formed in acid mine drainage. Appl. Geochem. 15(1): 27–34.

47. Drever, J.I. (1997). The Geochemistry of Natural Waters. Prentice-Hall, Englewood Cliffs, NJ. 48. Kay, J.T., Conklin, M.H., Fuller, C.C., and O’Day, P.A. (2001). Processes of nickel and cobalt uptake by a manganese oxide forming sediment in Pinal Creek, Globe Mining District, Arizona. Environ. Sci. Technol. 35(24): 4719–4725. 49. Kim, J.-Y. and Chon, H.-T. (2001). Pollution of a watercourse impacted by acid mine drainage in Imgok Creek of the Gangreung coal field, Korea. Appl. Geochem. 16: 1387–1396. 50. Schemel, L.E., Kimball, B.A., and Bencala, K.E. (2000). Colloid formation and metal transport through two mixing zones affected by acid mine drainage near Silverton, Colorado. Appl. Geochem. 15: 1003–1018. 51. Yu, J.-Y. and Heo, B. (2001). Dilution and removal of dissolved metals from acid mine drainage along Imgok Creek, Korea. Appl. Geochem. 16: 1041–1053. . 52. Suarez, D.L. and Langmuir, D. (1976). Heavy metal relationships in Pennsylvania soil. Geochim. Cosmochim. Acta 40: 589–598. 53. Noller, B.N., Woods, P.H., and Ross, B.J. (1994). Case studies of wetland filtration of mine waste water in constructed and naturally occurring systems in Northern Australia. Water Sci. Technol. 29(4): 257–265. 54. Stevens, S.E., Dionis, K., and Stark, L.R. (1990). Manganese and iron encrustation on green algae living in acid mine drainage. In: Constructed Wetlands for Wastewater Treatment. D.A. Hammer (Ed.). Lewis, Ann Arbor, MI, pp. 765–773.

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55. Cohen, R.H. (2000). Conjunction of geochemical modelling and pilot scale experiments for testing bioreactor removal of arsenic and chromium. WISA-Minewater Drainage Technol. Conf. Rhodes University, Grahamstown, January 23–28, 2000. 56. Tipping, E. and Hurley, M.A. (1992). A unifying model of cation binding by humic substances. Geochim. Cosmochim. Acta 56: 3627–3641. 57. Jain, S.K., Vasudevan, P., and Jha, N.K. (1989). Removal of some heavy metals from polluted water by aquatic plants: Studies on duckweed and water velvet. Biological Wastes 28: 115–126. 58. Zayed, A., Gowthaman, S., and Terry, N. (1998). Phytoaccumulation of trace metals by wetland plants: Duckweed. J. Environ. Qual. 27: 715–721. 59. Zhu, Y.L., Zayed, A.M., Qian, J.-H., de Sousa, M., and Terry, N. (1999). Phytoaccumulation of trace metals by wetland plants: Water hyacinth. J. Environ. Qual. 28: 339–344. 60. Cohen, R.H. and Staub, M.W. (1992). Technical Manual for the Design and Operation of a Passive Mine Drainage Treatment System. Report to the U.S. Bureau of Land Reclamation. Colorado School of Mines, Golden, CO. 61. Hedin, R.S., Nairn, R.W., and Kleinmann, R.L.P. (1994). Passive Treatment of Polluted Coal Mine Drainage, Bureau of Mines Information Circular 9389. United States Department of the Interior, Washington, DC. 62. Shutes, R.B., Ellis, J.B., Revitt, D.M., and Zhang, T.T. (1993). The use of Typha latifolia for heavy metal pollution control in urban wetlands. In: Constructed Wetlands for Water Quality Improvement. G.A. Moshiri (Ed.). Lewis, Ann Arbor, MI. 63. Hansel, C.M. and Fendorf, S. (2001). Characterisation of Fe plaque and associated metals on the roots of minewaste impacted aquatic plants. Environ. Sci. Technol. 35(19): 3863–3868. 64. Batchelor, A. (2000). Some perspectives on the use of wetlands for the amelioration of mine drainage. WISA-Minewater Drainage Technol. Conf. Rhodes University, Grahamstown, January 23–28, 2000. 65. Pulles, W. (2000). Development of passive mine water treatment technology. WISA-Minewater Drainage Technol. Conf. Rhodes University, Grahamstown, January 23–28, 2000. 66. Costley, S.C. and Wallis, F.M. (2001). Bioremediation of heavy metals in a synthetic wastewater using a rotating biological contactor. Water Res. 35(15): 3715–3723. 67. Glombitza, F. (2001). Treatment of acid lignite mine flooding water by means of microbial sulphate reduction. Waste Manage. 21(2): 197–203. 68. Pagnanelli, F., Toro, L., and Veglio, F. (2002). Olive mill solid residues as heavy metal sorbent material: A preliminary study. Waste Manage. 22(8): 901–907. 69. Podda, F., Zuddas, P., Minacci, A., Pepi, M., and Baldi, F. (2000). Heavy metal coprecipitation with hydrozincite [Zn-5 (CO3 )2 (OH)6 ] from mine waters caused by photosynthetic micro-organisms. Appl. Environ. Microbiol. 66(11): 5092–5098. 70. Nordstrom, D.K. (2000). Advances in the hydrogeochemistry and microbiology of acid mine waters. Int. Geol. Rev. 42(6): 499–515. 71. Chapnick, S.D., Moore, W.S., and Nealson, K.H. (1982). Microbially mediated manganese oxidation in a fresh-water lake. Limnol. Oceanogr. 27(6): 1004–1014. 72. Thornton, F.C. (1995). Manganese removal from water using limestone filled tanks. Ecological Eng. 4: 11–18.

73. Burke, S.P. and Banwart, S.A. (2002). A geochemical model for removal of iron (II) (aq) from mine water discharges. Appl. Geochem. 17(4): 431–443. 74. Zhang, C.L. et al. (1997). Physiochemical, mineralogical, and isotopic characterization of magnetite-rich iron oxides formed by thermophilic iron-reducing bacteria. Geochim. Cosmochim. Acta 61(21): 4621–4632.

SUGARCANE INDUSTRY WASTEWATERS TREATMENT ´ -DEL-CARMEN MARIA ´ ´ DURAN-DE-BAZUA

UNAM, National Autonomous University of Mexico M´exico D.F., M´exico

Sugar (sucrose) is a sweet, crystalline, white or colorless substance, derived from the juice of several plants. World sugar production amounts approximately to 120 million tonnes, of which two-thirds come from sugarcane and one third from sugarbeet. Sugarcane is a perennial grass that grows between the tropics (30◦ latitude north and 30◦ latitude south) in more than 100 countries. It is the most efficient earth grass plant for storing solar energy as biomass and reaches field yields up to 150 ton/ha. A variety of products are obtained from this versatile plant (Fig. 1). Many studies on sugarcane are published in journals and books. Sugarcane processing to produce sugar represents one of the oldest ‘‘industries’’ in modern times. In America, the first sugarcane processing plants were located in the New Spain territories as well as in the Caribbean islands since the sixteenth century. In Mexico, for example, the first sugarcane mill was established in Veracruz on the Gulf of Mexico in 1525 (1,2). The power for sugarcane mills originally came from water sources (rivers) and from cane bagasse burning. Considering that up to the first half of the twentieth century, industry was a synonym for smoke, wastewater, and wastes in general, the sugarcane industry was no exception. Therefore, considering modern sustainability concepts, the more than 2000 sugar processing plants still operate around the world with relatively poor technologies from an environmentally friendly point of view. Figure 2 shows a schematic diagram of a sugarcane factory, including the wastes generated. Roughly, one metric tonne of sugarcane renders the following products: 350 kg wet bagasse 100 kg sugar 60 kg straw and leaves 40 kg final syrup, molasses 40 kg cachasses 100 kg cane heads 310 kg evaporated water 1000 kg sugarcane

35% 10% 6% 4% 4% 10% 31% 100%

After sugar, the by-product with more added value is the final syrup, known as molasses. Molasses is a

SUGARCANE INDUSTRY WASTEWATERS TREATMENT

615

Soil amendment, animal feed, etc.

Miscellaneous

Xylitol, ethanol, and other chemicals Furfural and derivatives Cellulose and products Electricity

Bagasse

Recycling

Charcoal and activated carbon Methane-rich biogas Pulp and paper

Boiler cinders Fibrous products

Particle and moled boards

Cane tops and leaves Sugarcane

Sugar

Flue gases Direct utilization

Fertilizer (soil improver) Animal feed

Paper and fiber boards

Filter muds (cachasse)

Distilling industries

Raw wax

Animal feed, fertilizer, etc. Rum, ethanol, derivatives Spirits, anhydrous ethanol Vinegar, other food organic acids

Molasses Other biotechnological products

Acetone, butanol, glycerol Single cell protein Aconitic acid, itaconic acid

Miscellaneous

Monosodium glutamate Dextran, xanthan gum, L-Lysine

Figure 1. Versatile sugarcane use.

very inexpensive carbon source for many biotechnological products, including vaccines and antibiotics (Table 1). Of course, fermented and distilled beverages, such as rum, are more widely known biotechnological products. In Figs. 3 and 4, diagrams for sugar and ethyl alcohol production are presented. In Fig. 3, the example is the production of raw sugar. The two other most popular sugar commercial products are the so-called plantation white sugar or mill white sugar (in Mexico, it is known as ‘‘standard’’ sugar) and refined sugar. Literature presents block diagrams for the production of these different types of sugar (1). Wastewaters are the most conspicuous

wastes, and many processes have been devised to reduce their polluting impact and to gain some added value from treating them (5). Tables 2 and 3 show the average composition of some of these wastewaters, and Table 4 presents the regulatory limits that Mexico established in the 1990s for this agro-based industry. It is clear from Table 3 that liquid effluents from a distillery in the sugarcane mill have a very important impact on the overall composition of its wastewaters. Also, to comply with the regulations for both sugar production and distilled products, namely ethanol (ethyl alcohol), a removal efficiency of more than 95 and

616

SUGARCANE INDUSTRY WASTEWATERS TREATMENT

Flue gas Boiler house Cinders

Wastewaters

Cane

Bagasse

Water

Cane sugar factory

Power

Chemicals

Grease and fats

“Bagacillo”

Molasses

Ethyl alcohol

Filter cake (cachasse)

To cane fields

Sugar

Gaseous emissions

Raw materials

By-Products / Residues

Final products

Figure 2. Raw materials, products, and by-products of the sugarcane industry.

99.8% must be reached for general wastewaters and vinasses, respectively. Most of these vinasses, to date, are discharged to receiving bodies (soil, surface water sources) without any treatment or with a partial depuration because of the notion that these wastewaters improve soil. However, this is true only when soils are very poor. For good agricultural soils, they have a negative effect (9–11). To reduce costs, biological systems are preferred to physicochemical processes. For that reason, in general,

Table 1. Molasses Compositiona Parameter Water Sucrose Glucose Fructose Nitrogen compounds and other Organic products a

Reference 3.

% weight Parameter 20 32 14 14 10

% weight

K2 O CaO P2 O5 Carbonates, CO3 2− Sulfate ions

3.5 1.5 0.2 1.6 0.4

Other inorganics Density, g m/L

0.8 1.42

anaerobic systems are chosen for treating those process streams that contain biodegradable compounds because methane-rich biogas can be a plus from the bioconversion. To improve the biotreatment and reuse the water in the process, aerobic polishing and sometimes a physicochemical treatment are added. From the aerobic treatment, the biomass produced can be used in feedlots, particularly for fish (12). Leftover treated water can be used for irrigating cane fields. In tropical countries, it is very common to ‘‘cultivate’’ river shrimp (Cammarus montezumae), known in Mexico as ‘‘acociles’’ (from the Nahuatl or Aztec language, atl = water, cuitzilli = bent, that bends in the water). Then, an interesting biological cycle is rendered. Aerobic biomass together with carotenoproteins from cephalothorax and exoskeleton from river shrimp (the unedible portion), are pelleted for fish feedlots, and the unedible residues from fish are ground and used for shrimp cultivation. Figure 3 shows a complete nature-like cycle process instead of the typical man-made linear polluting processes (Fig. 2). When wastewaters contain mainly high concentrations of inorganic compound and they are highly soluble, as

SUGARCANE INDUSTRY WASTEWATERS TREATMENT

Water

Sugarcane

Washing wastewater

Ion-exchange resins and regeneration solutions

Oil and grease

Lime Ca(OH)2

Milling Cane cutting Cane shredding Cane crushing Cane pressing Cane inbibition

Aerobic biomass for feedlots

Aerobic treatment

Water recycling

Water softening

Regeneration wastewaters and spent resins

Resins to other uses To solar drying and solids to other uses

Boiler house

Fly ash, boiler blowdown,purges

Return to bagasse burners

Floor wash, bearing cooling water

Grease trap and solids separation

Cachasse filtration

Cachasse cake to cane fields

Bagasse

617

Steam

Water to cane fields Grease and solids to bagasse burners

Lime addition and heating

Bagacillo Mixed juice

Biogas to further use

Water to cane fields

Acid-caustic wastewaters

To solar drying and solids to other uses

Aerobic treatment

Condensates and excess condensates and condenser water

To cooling lagoons

Water recycling

Seed magma tank

Centrifugation

Molasses or final syrup

Clarification Juice recycling HCl + NaOH washing solutions

Multiple stage evaporation

Vacuum boiling

Anaerobic treatment

Vinasses

Ethanol production

Yeast and nutrients Massecuit receiver

Massecuit receiver

Cool water for crystals washing

Centrifugation

Dry hot air

Sugar crystals rotary drying

Syrup recrystallization

Vacuum boiling

Sugar dust

Raw sugar crystals

Biomass pelleting

Protein complexes

Fish cultivation

Carotenoid pigments

Fish byproducts

Chitin (chitosan)

River shrimps

Figure 3. Example of water use, wastewater generation, and treatment proposals for raw sugar.

happens with boilers purges and acid-caustic solutions used to clean heat transfer surfaces in evaporators and vacuum boiler pans, the most suitable process is to eliminate water by evaporation (preferably solar evaporation, taking advantage of climatic conditions in tropical areas). The resulting impure salts, collected as dry solids, may be recycled by the companies that sell

the caustic and acid products used as raw materials in the sugarcane mill, reducing the environmental impact of both enterprises (2). Figures 5 and 6 present schemes of an aerobic plant for treating wastewaters generated in a plantation sugar production plant (14). This plant wastewater treatment system had a different approach. The wastewater

(a)

Tank 1 Mixing – Dilution 18,000 L

15,000 L cane molasses, diluted to 8°Brix at pH 4.5 10 L H2SO4

5 kg urea 15 kg (bioactivator) Penicillin (for bacilli, short and long, cocci, fungi)

Tank 2 Pasteurization

Yeast /proliferation, consumes molasses sugars, addition of more molasses (duplicating the volume of preparation). It is first to 10°Brix till reaching 16ºBrix. To avoid ethanol formation, the mixture is aerated. When mixture reaches 6ºBrix, the next step comes

Tank 3 Yeast preparation

10,000 L yeasts developed at 6°Brix 40,000 L diluted molasses and 30 L H2SO4 Aeration When 15 –16°Brix is reached, molasses are added and fermentation starts (anaerobic ethanol production)

Tank 4 Mixing

(b)

Water Sulfuric acid Molasses

(NH4)2SO4 y (NH4)3PO4 Yeast

1

2

3

Heads Yeast

4

Filter

5

Air

6

7

8

Ethanol 95 – 96°

Tails

Steam

Fussel oil Vinasses Figure 4. (a) Sequential steps for ethanol production in a typical Mexican sugarcane mill using the final syrup or molasses as a carbon source (4). (b) Ethanol production flow in a typical Mexican sugarcane mill using the final syrup or molasses as a carbon source (4).

1/2″

Window

1.5 m

“Bell” biogas collector

Biogas to H2S and CO2 stripping and burning

5

1 2

6

Angle 3″ × 1/4 3

Effluent to primary settler 2″

4

Siphon

0.80 m Biogas stripper

Profile PTR 3″

3.20 m

0.80 m Manhole 1.64 m

Feed 2.1/2″ Climber

1 m Sampler

1 m 25% Concrete

1 2 3 4 5

Biogas Reagent addition Water level pipe Purge Biogas volume measuring pipe 6 Secondary settler feeding pipe

Sludge purge

Figure 5. Upflow anaerobic sludge bed reactor (UASB) and biogas absorber tower (13). 618

SUGARCANE INDUSTRY WASTEWATERS TREATMENT

619

Table 2. Liquid Effluent Characteristics of Some Cane Sugar Processing Plants in the World Parameter pH Biochemical oxygen demand (BOD5 ), mg/L Chemical oxygen demand (COD), mg/L Total suspended solids, mg/L Total nitrogen, mg/L Total phosphorus, mg/L Grease and oil, mg/L

Puerto Ricoa

Hawaiia

Philippinesa

Louisianaa

Indiaa

Mexicob

5.3–8.8 112–225 385–978 500–1400 n.r. n.r. n.r.

n.r.d 115–699 942–2340 3040–4500 n.r. n.r. n.r.

5.3–7.9 130–1220 50–1880 n.r. n.r. n.r. n.r.

n.r. 81–562 720–1430 409 n.r. n.r. n.r.

6.8–8.4 67–660 890–2236 792–2043 n.r. n.r. n.r.

6–10 20–36,700c 47–176,635c 20–46,190 0.2–1260 0.2–2000 0–570

a

Reference 6. Reference 7. c When vinasses are considered. d n.r.: not reported. b

composition is similar to those of other countries that do not contain vinasses (Table 2), so once the suspended matter is eliminated in a primary settler, an aerobic reactor is adequate for removing most of the dissolved biodegradable pollutants. Here, the feasibility of using an anaerobic reactor to ‘‘digest’’ the aerobic biomass when it is not used for feedlots is shown (13). This reactor gives the added value of methane-rich biogas production, once the gas is stripped in a column using the treated water to dissolve H2 S leaving the methane-rich gas free of this corrosive compound and ‘‘enriching’’ the treated wastewater in sulfur compounds before sending it to cane fields as irrigation water. BIBLIOGRAPHY ´ ´ C., and Cordov´es, M. (1991). 1. Chaux, D., Duran-de-Baz ua, Towards a Cleaner and More Profitable Sugar Industry. UNIDO, Vienna, Austria. ´ ´ C., Cordov´es, M., and Zedillo, L.E. (1994). 2. Duran-de-Baz ua, Demonstration of Cleaner Production Techniques for the Sugar Cane Agroindustry. Final Draft, Consultancy Report 1994. United Nations Industrial Development Office. Project US/INT/91/217/15-01-2. UNIDO/UNDP, Mexico City, Mexico. 3. Jim´enez, R., Mart´ınez, M., Espinoza, A., Noyola, A., and ´ ´ C. (1995). La Cana Duran-de-Baz ua, ˜ de Azucar, ´ su Entorno Ambiental. Parte II. Tratamiento de vinazas en una planta piloto en M´exico en un reactor anaerobio de lecho de lodos.

Table 3. Average Vinasse Compositionsa Parameter pH Biochemical oxygen demand, mg/L Chemical oxygen demand, mg/L Organic nitrogen, mg/L Total nitrogen, mg/L Ammonium ion, mg/L Grease and oil, mg/L Settleable solids, mL/L Total solids, g/L Total suspended solids, g/L Total dissolved solids, g/L a

Parameter 4.8–4.9 90,000 106,000 446 730 310 2 27 97–106 8–10 87–97

Total volatile solids, g/L Total fixed solids, g/L Dissolved volatile solids, g/L Dissolved fixed solids, g/L

76–83 21–23 69–76 18–22

Phosphorus, mg/L Calcium, mg/L Magnesium, mg/L

150 2,960 1,370

Sodium, mg/L Potassium, mg/L

310 2,550

Sulfate ions, mg/L

10,500

Reference 3.

Informe t´ecnico de proyecto VIN-01-95. GEPLACEA, CNIIAA, PIQAYQA-FQ-UNAM. Facultad de Qu´ımica, UNAM, M´exico D.F., M´exico. ´ ´ ´ 4. Castro-Gonzalez, A., Bernal-Gonzalez, M., and Duran-de´ C. (2004). Tratamiento de vinazas de plantas Bazua,

Table 4. Regulatory Limits Established in Mexico for Cane Sugar and Ethanol Production Plant Liquid Effluentsa Cane Sugar Parameter pH BOD5 , mg/L COD, mg/L Total suspended solids, mg/L Settleable solids, mg/L Total nitrogen, mg/L Total phosphorus, mg/L Grease and oils, mg/L Phenols, mg/L a b

Reference 8. n.r. not reported

Distilling Industry

Daily Average Value

Instantaneous Value

Daily Average Value

Instantaneous Value

6–9 60 n.r.b n.r. 1.0 n.r. n.r. 15 0.5

6–9 72 n.r. n.r. 1.2 n.r. n.r. 20 0.75

6–9 200 260 200 1.0 10 5 10 n.r.

6–9 240 360 240 2.0 12 6 20 n.r.

620

ESTIMATED USE OF WATER IN THE UNITED STATES IN 1990 INDUSTRIAL WATER USE

Influent

Sludge purge

Purge

Sludge aerobic treatment

Primary settler

Purge

Secondary settler

Aeration tank

Sludge

cat-walk

UASB

Feed tank

Sludge aerobic treatment

Aeration tank

Chlorine addition Purge

Purge

Sludge purge

Secondary settler effluent

General effluent

Figure 6. Complete sugarcane mill wastewater treatment plant, including a primary settling tank, an anaerobic feed tank, a UASB reactor, and a biogas absorption tank (biogas stripping) (14).

5.

6.

7.

8.

9.

10.

11.

12.

13.

destiladoras de alcohol usando consorcios microbianos anaerobios. Bebidas Mexicanas 13(3): 12–14, 16–20, 22–25. ´ ´ C. (2004). Tratamiento Biol´ogico de Aguas Duran-de-Baz ua, Residuales Industriales. PIQAyQA-Facultad de Qu´ımica, UNAM, 5th Edn. (1994). M´exico, D.F. Mexico. 6th Edn., in press. Bao-Guo-Yu and Chen-Shi-Zhi. (1990). Treatment and Utilization of Pollution Effluents of Cane Sugar Factories in China. UNIDO. Workshop on pollution control and low waste technologies in agro-based industries. Shanghai, China. Anonymous. (1976). Uso del Agua y Manejo Del Agua Residual ´ en la Industria. Vol. 8. Azucar. Secretar´ıa de Recursos ´ Hidraulicos. M´exico, D.F. Mexico. DOF. (1993, 1995). Normas Oficiales Mexicanas en Materia de Protecci´on Ambiental. NOM-002-ECOL-1993, Proy. NOM064- ECOL-1995. Diario Oficial de la Federaci´on, Secretar´ıa de Desarrollo Social (Sedesol), Mexico City, Mexico. ˜ Bautista-Zu´ niga, F., Reina-Trujillo, T. de J., Villers-Ruiz, L., ´ ´ C. (2000). Mejoramiento de Suelos and Duran-de-Baz ua, Agrı´colas Usando Aguas Residuales Agroindustriales. Caso: Vinazas crudas y tratadas. Serie: Qu´ımica Ambiental del Suelo. Vol. 1. PIQAyQA-FQ, UNAM. M´exico D.F., M´exico. ˜ ´ ´ C., Reyna-Trujillo, T., Bautista-Zu´ niga, F., Duran-de-Baz ua, and Villers-Ruiz, L. (2000). Agroindustrial organic residues: Handling options in cane sugar processing plants. Part I and Part II. Sugar y Azucar ´ 95(9): 32–45; 95(10): 23–37. Villatoro-Res´endiz, J. (1998). Estudio de la Transformaci´on de la Materia Organica ´ Biodegradable de la Vinaza Cruda y Tratada en Los Suelos Acrisol y Vertisol, Del Municipio Miguel Aleman ´ en el Estado de Veracruz, M´exico. Tesis profesional (Biolog´ıa). UNAM, Facultad de Ciencias, M´exico D.F., M´exico. ´ınguez, M.C., ´ Duran-Dom Pedroza-Islas, R., Rosas´ ´ Vazquez, C., Luna-Pabello, V.M., Sanchez-Zamora, A., Capilla-Rivera, A., Paredes-G´omez, L., Valderrama´ ˜ I. (1991). Producci´on Herrera, S.B., and Vazquez-Cede no, de alimentos para peces: Utilizaci´on de subproductos del tratamiento de aguas residuales. In: Premio Nacional Serfı´n El Medio Ambiente. J.J. De Olloqui (Ed.). Futura Eds., M´ex., M´exico, pp. 79–106. ´ ´ Castro-Gonzalez, A., Enr´ıquez-Poy, M., and Duran-de´ C. (2001). Design, construction, and starting-up of Bazua, an anaerobic reactor for the stabilisation, handling, and

disposal of excess biological sludge generated in a wastewater treatment plant. Anaerobe (Biotechnology) 7: 143–149. ´ ´ ´ M.C., 14. Castro-Gonzalez, A., Duran-de-Baz ua, Enr´ıquezPoy, M., and Pliego-Bravo, Y. (1998). Consideraciones en un ingenio azucarero para minimizar las descargas l´ıquidas y mejorar las condiciones de operaci´on de su planta de tratamiento de aguas residuales. Ingenio (M´exico) 3(34): 2–7.

ESTIMATED USE OF WATER IN THE UNITED STATES IN 1990 INDUSTRIAL WATER USE U.S. Geological Survey

Industrial water use includes water for such purposes as processing, washing, and cooling in facilities that manufacture products. Major water-using industries include, but are not limited to, steel, chemical and allied products, paper and allied products, and petroleum refining. Many States have developed permit programs that require reporting of industrial withdrawals and return flows. Estimates for 1990 are improved over those of previous years because of the availability of more comprehensive inventories of industrial facilities and more complete water-use records. Information on deliveries from public suppliers to industrial users were estimated from a variety of methods if not available directly from public suppliers. Consumptive-use estimates generally were based on coefficients ranging from 3 to 80 percent (depending on the type of industry) of withdrawals and deliveries. Industrial water use (freshwater withdrawals, publicsupply deliveries, saline water withdrawals) during 1990 was an estimated 19,300 Mgal/d of self-supplied

This article is a US Government work and, as such, is in the public domain in the United States of America.

Table 1. Industrial Water Use By Water-Resources Regions [Figures May Not Add to Totals Because of Independent Rounding. All values in Million Gallons Per Day] Self-Supplied Withdrawals By Source and Type Ground water Region

Surface water

Total

Fresh

Saline

Fresh

Saline

Fresh

Saline

Total

New England Mid Atlantic South Atlantic-Gulf Great Lakes Ohio

96 361 896 235 532

0.0 .2 0 3.7 0

382 1,370 1,920 3,950 1,840

68 1,470 94 0 0

479 1,730 2,810 4,190 2,370

68 1,470 94 3.7 0

547 3,200 2,910 4,190 2,370

Tennessee Upper Mississippi Lower Mississippi Souris-Red-Rainy Missouri Basin

23 349 501 1.3 114

0 0 .6 0 0

1,170 618 2,120 47 57

0 0 67 0 0

1,190 967 2,620 49 171

0 0 67 0 0

1,190 967 2,690 49 171

Arkansas-White-Red Texas-Gulf Rio Grande Upper Colorado Lower Colorado

67 141 11 2.9 49

0 1.1 0 0 0

301 600 1.0 2.5 124

0 1,460 0 0 0

Great Basin Pacific Northwest California Alaska Hawaii Caribbean

77 336 126 5.2 20 11

2.3 0 0 0 .6 1.2

29 691 4.8 106 23 0

0 36 25 0 0 50

106 1,030 130 111 43 11

3,260

19,300

Total

3,950

9.7

15,400

368 741 12 5.4 174

0 1,460 0 0 0 2.3 36 25 0 .6 51 3,270

368 2,200 12 5.4 174

Reclaimed Waste Water 6

2

108 1,060 156 111 44 62 22,600

9

Table 2. Industrial Water Use By States [Figures May Not Add to Totals Because of Independent Rounding. All values in Million Gallons Per Day] Self-Supplied Withdrawals By Source and Type Ground water State

Saline

Total

Reclaimed Waste Water

784 111 163 177 129

0.0 0 0 0 25

784 111 163 177 154

0.0 0 2.3 0 .8

118 80 65 .5 403

0 68 6.0 0 56

118 148 71 .5 459

0 0 0 0 0

Surface water Fresh

Total

Fresh

Saline

Saline

Alabama Alaska Arizona Arkansas California

31 5.2 39 99 125

0.0 0 0 0 0

753 106 124 78 3.4

0.0 0 0 0 25

Colorado Connecticut Delaware D.C. Florida

33 19 18 .5 282

0 0 0 0 0

85 61 47 0 121

0 68 6.0 0 56

Georgia Hawaii Idaho Illinois Indiana

346 20 170 155 129

0 .6 0 0 0

311 23 26 309 2,350

33 0 0 0 0

657 43 196 464 2,480

33 .6 0 0 0

689 44 196 464 2,480

Iowa Kansas Kentucky Louisiana Maine

71 50 93 289 9.8

0 0 0 .6 0

148 3.8 220 2,070 244

0 0 0 67 0

219 53 313 2,360 254

0 0 0 67 0

219 53 313 2,430 254

621

Fresh

.5 0 0 0 0 0 .5 0 0 0

622

ESTIMATED USE OF WATER IN THE UNITED STATES IN 1990 INDUSTRIAL WATER USE Table 2. (continued) Self-Supplied Withdrawals By Source and Type Ground water

State Maryland Massachusetts Michigan Minnesota Mississippi Missouri Montana Nebraska Nevada New Hampshire New Jersey New Mexico New York North Carolina North Dakota

Surface water

Fresh

Saline

Fresh

Saline

Fresh

Saline

Total

21 65 175 65 144

0 0 3.7 0 0

49 22 1,510 89 126

379 0 0 0 0

70 87 1,680 154 269

379 0 3.7 0 0

450 87 1,690 154 269

53 30 39 9.4 .3 53 4.6 85 63 2.

0 0 0 0 0 0 0 0 0

32 27 2.4 .8 37 273 1.7 189 328 6.6

Ohio Oklahoma Oregon Pennsylvania Rhode Island

123 3.3 31 180 2.5

0 0 0 0 0

230 32 254 1,690 9.1

South Carolina South Dakota Tennessee Texas Utah

47 5.0 69 143 77

0 0 0 1.1 2.3

Vermont Virginia Washington West Virginia Wisconsin

1.0 195 104 106 58

0 0 0 0 0

Wyoming Puerto Rico Virgin Islands

6.0 11 .1

Total

Total

3,950

.2

0 0 1.2 9.7

0 0 0 0 0 1,020 0 0 5.5 0

85 57 41 10 37 326 6.3 274 390 8.8

0 0 0 0 0 1,020 0 0 5.5 0

0 0 0 0 0

354 35 284 1,870 12

585 10 813 741 29

0 0 0 1,460 0

632 15 882 884 106

0 0 0 1,460 2.3

43 300 397 26 409

0 66 36 0 0

44 495 501 132 468

0 0 50

16 11

3,260

19,300

9.9 0 0 15,400

freshwater, 5,190 Mgal/d of public-supplied freshwater, and an additional 3,270 Mgal/d of saline water. (See Table 1: water-resources regions and Table 2: States.) Industrial freshwater use during 1990 was 13 percent less than during 1985 and represents 7 percent of total freshwater use for all offstream categories. Surface water was the source for about 82 percent of self-supplied industrial withdrawals; ground water, 18 percent; and reclaimed wastewater, only a fraction of 1 percent. Publicsupplied deliveries to industries accounted for 13 percent of public-supply withdrawals. The source and disposition of water for industrial purposes are shown in the pie charts below (or as a GIF file or PostScript file (94 Kb)). The consumptive use of freshwater for industrial purposes during 1990 was

.1

0 0 0 0 0

Reclaimed Waste Water 63 0 0 0 0

85 57 41 10 37 1,340 6.3 274 396 8.8

0 0 0 0 0 0 0 0 0 0

354 35 284 1,870 12

0 0 1.6 0 0

632 15 882 2,340 108

0 0 0 22 0

0 66 36 0 0

44 561 536 132 468

0 0 0 0 0

0 0 51

16 11 51

0 0 0

3,270

22,600

90

3,330 Mgal/d, or 14 percent of freshwater withdrawals and deliveries; saline consumptive use was 913 Mgal/d, or about 28 percent of saline water withdrawals. In 1990, the Great Lakes and Mid Atlantic waterresources regions had the largest withdrawals for industrial purposes as shown in figure 20 (GIF file), or (PostScript file (620 Kb)). Indiana, Louisiana, Texas, Pennsylvania, and Michigan reported the largest state withdrawals for industries as shown in figure 21 (GIF file) or (PostScript file (508 Kb)). Indiana, Louisiana, Pennsylvania, and Michigan reported the largest freshwater use (figure 22 (GIF file)), or (PostScript file (508 Kb)), and Maryland and Texas reported the largest quantities of reclaimed wastewater used by industries.

WASTE WATER TREATMENT AERATION

transfer efficiency per meter submergence for a given aeration system under given ambient and operating conditions is relatively constant for installation depths in the technically most important range around 4 to 5 meters. It is, therefore, often quoted in percent per meter submergence to allow the user of this information to calculate the OTE or SOTE case-by-case for the respective aerator installation depths rather than providing a long list of values for different submergences. 6. Aeration efficiency (AE) is the OTR divided by the total power input for the aeration system, measured as either brake power or wire power. 7. Standard aeration efficiency (SAE) is the AE based on the SOTR.

ARNIM R.H. HERTLE GHD Pty Ltd Wembley, Washington Australia

OVERVIEW The aeration of wastewater and byproducts is a key process in the operation of most modern wastewater treatment plants (WWTP), which is reflected in the energy consumption of a WWTP’s aeration system, which can be up to 70% of the WWTP’s total energy consumption. As most wastewater treatment plants and many sludge treatment plants include aerobic biological processes, the transfer of oxygen into the wastewater (or sludge) is a key operation. As a result of the large quantities of CO2 produced in the course of the aerobic degradation of organic matter, the pH in such a reactor can be significantly lower than that of its feed. These CO2 -related effects can be amplified when nitrification occurs or when the liquor’s alkalinity is low. This drop in pH can lead to decreased performance of the plant as well as to impacts such as concrete and metal corrosion if not managed properly. The second key function of an aeration system is therefore to remove CO2 from the reactor (1). One of the important functions that an aeration system has to provide is mixing to prevent settling of the biomass in the reactors and optimize the contact of wastewater with the biomass.

Further explanations and definitions of further parameters and terms used to characterize aeration systems and their design are provided in Refs. 2 and 3. AERATION SYSTEMS Numerous aeration mechanisms are used in wastewater treatment plants; each has its particular advantages. The oxygen transfer from air (or pure oxygen where it is applied) into liquor can take place only through a common surface of the two media; so the common goal of all aeration mechanisms is to make as much active surface area available as possible. A classification of the technical systems used for aeration purposes can be based on the mechanical equipment that is used. Alternatively, the systems can be classified on the basis of where the oxygen transfer mainly takes place, which indicates how the design calculations are carried out. Aeration systems are therefore differentiated into

TERMS AND DEFINITIONS The following terms and definitions will be used herein:

— surface aeration and — submerged aeration.

1. α (alpha) factor is the ratio of the apparent volumetric mass transfer coefficient kL in wastewater to that in clean water. α is normally 23 >23 14

>23 >23 10.2

77

160

111

Chronic Acute

230 480 (24 h) 190 (48 h) 440 (48 h) Reproduction rate EC50 410

24 92

39 190

31 132

Acute

EC50 27

10

20

14.1

Acute 10 day NP in sediment Subacute, 14 day NP in sediment

260 mg/kg

Acute Chronic

2600 Growth and strength, 32 days

Acute

26.1 mg/kg

56

Reference 36.

The Soaps and Detergents Association found that the present use of NPES in detergents and cleaning products poses little or no risk to the environment in the United States. Environmental monitoring studies (39–41) demonstrate that APEs and their metabolite concentrations in rivers and lakes in the United States are below toxic levels. These studies also demonstrate that 92 to 99% of NPEs are effectively removed in wastewater treatment plants in the United States. Based on research conducted in the USA, APEs do not appear to pose a significant ecological risk. The Canadian Government is currently reviewing the environmental and health characteristics of NP and NPEs under its PSL2 program. European countries, however, have started phasing out the use of APEs in their products. In 1986, Germany instituted voluntary restrictions and Switzerland banned the use of surfactants in laundry detergents (39). A voluntary ban on APE use in household products began in 1995 throughout northern Europe (42). Denmark has introduced an environmental quality standard for NP and NPEs of 1 µg/L. Studies conducted in Japanese wastewater treatment plants (43) indicated widespread pollution by NPEs and their metabolites. CONCLUSIONS Literature indicates that the presence of detergents in wastewaters may pose problems, depending on the treatment methodology for the wastewater and sludge. Most

detergents are biodegradable under aerobic conditions. However, concerns with degradation products of APEs, particularly nonylphenol, remain. There is also concern over the resistance of all major classes of detergents to biodegradation in anaerobic environments, especially because this is the predominant method of sludge treatment from primary tanks. The amphiphilic nature of surfactants makes them prone to adsorption to the sludge during primary tank settling. Therefore a significant portion of the surfactants in wastewaters may be passed untreated into the sludge. Studies indicate that application of sludge in aerobic soil environments can lead to further biodegradation. However, more studies are required to determine the fate of toxic intermediates in soil. BIBLIOGRAPHY 1. Schroder, H.Fr. (2001). Tracing of surfactants in the biological wastewater treatment process and the identification of their metabolites. J. Chromatogr. 926: 127–150. 2. Swisher, R.D. (1987). Surfactant Biodegradation. Marcel Dekker, New York. 3. Ahel, M., Giger, W., and Koch, M. (1994). Behaviour of alkylphenol polyethoxylate surfactants in the aquatic environment-I. Occurrence and transformation in sewage treatment. Water Res. 28: 1131–1142. 4. Ahel, M., Giger, W., and Koch, M. (1994). Behaviour of alkylphenol polyethoxylate surfactants in the aquatic

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5.

6.

7.

8.

9. 10. 11.

12.

13.

14.

15.

16.

17.

18.

19.

20.

21.

22.

23.

DETERGENTS environment-II. Occurrence and transformation in rivers. Water Res. 28: 1143–1152. Banat, F., Prechtl, S., and Biscof, F. (2000). Aerobic thermophilic treatment of sewage sludge contaminated with 4-nonylphenol. Chemosphere 41: 297–302. Bennett, E.R. and Metcalfe, C.D. (1999). Distribution of degradation products of alkylphenol ethoxylates near sewage treatment plants in the lower great lakes, North America. Environ. Toxicol. Chem. 19(4): 784–792. Brunner, P.H., Capri, S., Marcomini, A., and Giger, W. (1988). Occurrence and behaviour of linear alkylbenzenesulphonates, nonylphenol, nonylphenol mono- and nonylphenol diethoxylates in sewage and sewage sludge treatment. Water Res. 22: 1465–1472. Lewis, M.A. (1992). The effects of mixtures and other environmental modifying factors on the toxicities of surfactants to freshwater and marine life. Water Res. 28(8): 1013– 1023. Rosen, M.J. (1989). Surfactants and interfacial phenomena, 2nd Edn. John Wiley & Sons, New York. AWWA, APHA, and WEF. (1995). Standard Methods for the Evaluation of Waste Water. Washington, DC. Wickbold, R. (1972). On the determination of nonionic surfactants in river and wastewater. Ten. Surf. Deterg. 9: 173–177. Riu, J. et al. (2000). Determination of LAS in WWTPs and coastal waters by automated solid phase extraction followed by capillary electrophoresis. J. Chromatogr. 889: 221–229. Stephanou, E. and Giger, W. (1982). Persistent organic chemicals in sewage effluents: quantitative determinations of NP and NPEs by glass capillary gas chromatography. Environ. Sci. Technol. 16: 800–805. McEvoy, J. and Giger, W. (1986). Determination of LAS in sewage sludge by high resolution gas chromatography/mass spectrometry. Environ. Sci. Technol. 20: 376–383. Matthijs, E. and De Hanau, H. (1987). Determination of LAS in aqueous samples, sediments, sludges and soils using HPLC. Tens. Surf. Deterg. 24: 193–199. Giger, W., Brunner, P.H., and Schaffner, C. (1984). 4nonylphenol in sewage sludge: accumulation of toxic metabolites from nonionic surfactants. Science 225: 623–625. Field, J.A. and Reed, R.L. (1996). NPEC metabolites of nonionic surfactants in US paper mill effluents. Environ. Sci. Technol. 30: 3544–3550. Lee, H.-B. and Peart, T.E. (1995). Determination of 4nonylphenol in effluent and sludge from sewage treatment plants. Anal. Chem. 67: 1976–1980. Scott, M.J. and Jones, M.N. (2000). The biodegradation of surfactants in the environment. Biochim. Biophys Acta 1508: 235–251. Dorn, P., Salanitro, J., Evans, H., and Kravetz, L. (1992) Assessing the aquatic hazard of some branched and linear non-ionic surfactants by biodegradation and toxicity. Environ. Toxicol. Chem. 12: 1751–1762. Hennes-Morgan, E.C. and Oude, N.T. (1993). Detergents In: Handbook of Ecotoxicology. P. Calow (Ed.). Blackwell Scientific, Oxford, UK, pp. 130–154. Dentel, S.K., Allen, H.E., Srinivasarao, C., and Divincenzo, J. (2002). Third Year Completion Report Project No. 6, Water Resources Center. University of Delaware, Newark, DE. Available: http://bluehen.ags.udel.edu/dewre/surfact.htm. Johnson, A.C., White, C., Bhardwaj, L., and Jurgens, M.D. (2000). Potential for octylphenol to biodegrade in some english rivers. Environ. Toxicol. Chem. 19(10): 2486–2492.

24. Jensen, J. (1999). Fate and effects of LAS in the terrestrial environment. Sci. Total Environ. 226: 93–111. 25. Marcomini, A., Capel, P.D., Goger, W., and Hani, H. (1988). Residues of detergent derived organic pollutants and PCBs in sludge amended soil. Naturwissenschaften 75: 460– 462. 26. Birch, R.R., Gledhill, W.E., Larson, R.J., and Nielsen, A.M. (1992). Role of Anaerobic Biodegradability in the Environmental Acceptability of Detergent Materials, Proc. 3rd CESIO. London, Int. Surf. Congr. Exhib., 26, pp. 26–33. 27. Gode, P., Guhl, W., and Steber, J. (1987). Environmental compatibility of fatty acid, alpha-sulfomethyl esters. Fat. Sci. Technol. 89: 548–552. 28. Maurer, E.W., Weil, J.K., and Linfield, W.M. (1965). The biodegradation of esters of α-sulfo fatty acids. J. Am. Oil Chem. Soc. 54: 582–584. 29. Steber, J. and Weirich, P. (1987). The anaerobic degradation of detergent range fatty alcohol ethoxylates-studies with 14 C labeled surfactants. Water Res. 21: 661–667. 30. Thomas, O.R.T. and White, G.F. (1989). Metabolic pathway for the biodegradation of sodium dodecyl sulphate by pseudomonas sp-c12b. Biotechnol. Appl. Biochem. 11: 318–327. 31. Holt, M.S., Mitchell, G.C., and Watkinsson, R.J. (1992). The environmental chemistry, fate and effects of nonionic surfactants. In: Anthropogenic Compounds. Vol. 3, Part F, Handbook of Experimental Chemistry, O. Hutzinger (Ed.). Springer, Berlin, pp. 89–144. 32. Ankley, G.T. and Burkhard, P. (1992). Identification of surfactants as toxicants in a primary effluent. Environ. Toxicol. Chem. 11: 1235–1248. 33. Argese, E. et al. (1994). Submitochondrial particle response to LAS, nonylphenol polyethoxylates and their biodegradation derivative. Environ. Toxicol. Chem. 13(5): 737–742. 34. Staples, C.A. (1998). An evaluation of the aquatic toxicity and bioaccumulation of C8- abd C9- alkylphenol ethoxylates. Environ. Toxicol. Chem. 17: 2470–2480. 35. Lewis, M.A. (1992). The effects of mixtures and other environmental modifying factors on the toxicities of surfactants to freshwater and marine life. Water Res. 26(8): 1013–1023. 36. Naylor, C.G. (1998). Environmental Fate and Safety of Nonylphenol Ethoxylates. The Alkylphenols and Alkylphenol Ethoxylates Review, APE Research Council, Washington, DC. 37. Renner, R. (1997). European bans on surfactants trigger transatlantic debate. Environ. Sci. Technol. 31: 316A–320A. 38. VandenHenvel, C. (2000). A more certain future for nonylphenol ethoxylates. Chem. Times Trends Spring: 38–39. 39. Hale, R.C. et al. (2000). Nonylphenols in sediments and effluents associated with diverse wastewater outfalls. Environ. Toxicol. Chem. 19(4): 946–952. 40. Bennett, E.R. and Metcalfe, C.D. (2000). Distribution of degradation products of APEs near sewage treatment plants in the lower great lakes, north america. Environ. Toxicol. Chem. 19(4): 784–792. 41. Naylor, C.G., Williams, J.B., Varineau, P.T., and Webb, D.A. (1998). Nonylphenol ethoxylates in an industrial river. Alkylphenols Alkylphenol Ethoxylates Rev.: 44–53. 42. Warhurst, A.A. (1985). Environmental and Human Safety of Major Surfactants. Friends of the Earth Scotland, Edinburgh, Scotland. 43. Fujita, M. et al. (2000). Behavior of nonylphenol ethoxylates in sewage treatment plants in Japan—biotransformation and ecotoxicity. Water Sci. Technol. 42: 23–30.

ECOLOGICAL WASTEWATER MANAGEMENT

ECOLOGICAL WASTEWATER MANAGEMENT GUENTER LANGERGRABER BOKU—University of Natural Resources and Applied Life Sciences Vienna, Austria

Conventional sanitation concepts are neither an ecological nor an economical solution for rural areas in both industrialized and developing countries. Ecological sanitation (EcoSan) represents a holistic approach toward ecologically and economically sound sanitation. The underlying aim is to close nutrient and water cycles with as small an expenditure on material and energy as possible to contribute to sustainable development. EcoSan is a systemic approach and an attitude. Single technologies are only a means to an end. Therefore, EcoSan-technologies may range from nearly natural wastewater treatment techniques to compost toilets, simple household installations, and to complex, mainly decentralized systems. These technologies are not ecological per se but only in relation to the observed environment. Promotion of EcoSan concepts, therefore, is the strategy for achieving the goal—closing the loop in wastewater management and sanitation.

disease organisms. The nutrients contained in excreta are then recycled by using them, for example, in agriculture. The basic motivation behind the need to reshape the management of nutrients and streams of organic residuals in society may be found in the so-called ‘‘basic system conditions for sustainable development’’ for water and sanitation management, formulated in Agenda 21 (1): 1. The withdrawal of finite natural resources should be minimized. 2. The release of nonbiodegradable substances to the environment must be stopped. 3. Physical conditions for circular flows of matter should be maintained. 4. The withdrawal of renewable resources should not exceed the pace of regenerating them. If a sanitation system shall contribute toward the goals of ecological sanitation, it has to meet or at least to be on the way toward meeting the following criteria, as given by Esrey et al. (2): 1. Prevent disease: A sanitation system must be capable of destroying or isolating pathogens. 2. Affordable: A sanitation system must be accessible to the world’s poorest people. 3. Protect the environment: A sanitation system must prevent pollution, return nutrients to the soil, and conserve valuable water resources. 4. Acceptable: A sanitation system must be aesthetically inoffensive and consistent with cultural and social values. 5. Simple: A sanitation system must be robust enough to be easily maintained within the limitations of

DEFINITION OF ECOLOGICAL SANITATION A sanitation system that provides ecological sanitation (EcoSan) is a cycle—a sustainable, closed-loop system (Fig. 1). The EcoSan approach is resource minded, not waste minded. Human excreta are treated as a resource and are usually processed on-site and then, if necessary, processed further off-site until they are completely free of

Soil Food

Potassium (K) Phosphorous (P) Nitrogen (N) Organic matter (C) Energy

Blackwater Treatment

675

Settlement

Other waste

Biological waste

Rainwater Graywater

Recycling Figure 1. An ecological sanitation system.

Reuse Recharge

Treatment (e.g., constructed wetland)

676

ECOLOGICAL WASTEWATER MANAGEMENT

the local technical capacity, institutional framework, and economic resources. 6. Comfortable: A sanitation system must be nearly as comfortable as a flush toilet. That means it has to be indoors or accessible under the same roof of the house. Successful implementation of sanitation systems requires understanding all components of the system. The components have to be considered together when designing and making sanitation systems work. Following are the main components (Fig. 2): • Nature: The most relevant natural variables are climate (humidity, temperature), water (amount available, groundwater level), and soil (stability, permeability). • Society: The factors that describe the society include the settlement pattern (concentrated/dispersed, low/high rise), attitude (fecophobic/fecophilic), habits, beliefs and taboos related to human excreta, and the economic status of the community. • Process: Physical, chemical, and biological processes turn human excreta into a nondangerous, inoffensive, useful product. Dehydration and decomposition are the principal processes.

• Device: The device is the on-site structure specifically built for defecation and urination. It is essential to sanitize human excreta before they are recovered and reused. Figure 3 shows the different approaches for handling human excreta: • Mix and drain: In conventional sanitation systems, urine and feces are mixed and flushed away with water. • Mix and evaporate: Excreta are mixed; however, they are not flushed away but treated on-site, for example, composted. • ‘‘Don’t mix’’ and dehydrate: Urine and feces are collected and treated separately. The principles underlying EcoSan are not novel. Sanitation systems based on ecological principles have been used in different cultures, for hundreds of years. EcoSan systems are still widely used in parts of East and Southeast Asia. In Western countries, this option was largely abandoned, as flush-and-discharge became the norm, but in recent years, there has been a revival of interest in EcoSan (2). WASTEWATER IS A RESOURCE

Nature climate water soil

Society settlement pattern economy habits and taboos

Device toilet latrine potty

Process physical chemical biological Figure 2. Main components of the ‘‘sanitation’’ system.

Figure 3. Different approaches for handling excreta: ‘‘Don’t mix’’ and dehydrate (left); mix and drain (middle); mix and evaporate (right).

Wastewater from households contains urine (yellowwater), feces (brownwater), and graywater. Graywater is the part of the wastewater which is not mixed with excreta (from kitchens, bathrooms, and laundries). If urine and feces are mixed, the resulting mixture is called blackwater. For a long time, wastewater has been regarded as a problem because wastewater involves hygienic hazards and contains eutrophying substances in the form of organic matter, nitrogen, and phosphorus. These substances cause problems in seas, lakes, and streams, but, on the other hand, they would be valuable to farmers. Nitrogen (N), phosphorus (P), and potassium (K) in wastewater can be used instead of artificial fertilizer, and the organic material increases the humus content. Recirculating nutrients from wastewater as fertilizer reduces the need for industrially produced fertilizer and also reduces

ECOLOGICAL WASTEWATER MANAGEMENT

discharges of nutrient-rich water from treatment plants into watercourses. Wastewater Characteristics One person produces about 500 liters of urine and 50 liters of feces per year. The same person, additionally, produces a range of 25,000 to more than 100,000 liters of graywater. Blackwater and graywater have very different characteristics (Table 1). Most of the nutrients essential in agriculture (N, P, K) occur in urine. Feces contain smaller amounts of nutrients, and the quantities in graywater are insignificant. If blackwater is collected separately at low dilution, it can be converted, for example, to a safe natural fertilizer that replaces synthetic products and prevents the spread of pathogens and other pollutants to receiving waters.

677

occurrence in infected persons or carriers connected to the system. Their concentrations also depend on the dilution of the water. Untreated wastewater should always be regarded as a potential carrier of pathogenic organisms. • Graywater normally contains small amounts of pathogenic organisms. However, due to a relatively high load of easily degradable organic substances, regrowth of indicator organisms of fecal pollution may occur. • Stormwater may have high loads of fecal contamination. This is of special concern in areas of the world where open-air defecation is practiced, because high loads of pathogens, as in wastewater, may occur. Stormwater may also contain high loads of zoonotic pathogens originating from animal or bird feces.

Sources of Hygienic Hazards

Resource Wastewater?

Watersheds may be affected directly by excreta that contain large amounts of pathogens from humans and animals, or indirectly through wastewater outlets, from large-scale wastewater treatment plants and from smaller units of wastewater or graywater on-site sanitation. Stormwater and runoff from agricultural lands may also carry large amounts of pathogens to watersheds emanating directly from excreta, or pathogens occurring in sludge, excreta, and manure applied to land. Stormwater and runoff water may also carry pathogens from domestic and wild animals and birds that may affect humans. Organic waste from human settlements and activities may also be a source of pathogenic organisms.

Summarizing the characteristics of wastewater and the sources of hygienic hazards, the following conclusions can be drawn:

• Human fecal excreta may be harmless but can contain large amounts of pathogenic organisms. The risk depends on the frequency of infected persons and symptomless carriers in the population. Anyway, human fecal excreta are responsible for the major part of hygienic hazards. • Human urine does not normally contain pathogenic organisms that will transmit enteric disease to other individuals. Fecal material is thus the main source of infectious organisms. Only in special cases, for example, a systemic infection with fever, will pathogenic organisms be present in urine. • All microorganisms in wastewater originating from human excreta occur in amounts reflecting their Table 1. Typical Characteristics of the Main Components of Household Wastewatera

Volume (l.p−1 .yr−1 )

Graywater

Urine

Feces

25,000–100,000

∼500

∼5

∼3% ∼10% (P-free detergents) ∼34% ∼41%

∼87% ∼50% ∼54% ∼12%

∼10% ∼40% ∼12% ∼47%

Yearly loads (kg.p−1 .yr−1 ) N P K COD a

∼4–5 ∼0.75 ∼1.8 ∼30

Reference 3.

• Most of the soluble nutrients are found in urine. If urine is segregated and converted to agricultural usage, the biggest step toward nutrient reuse and highly efficient water protection will have been taken. • The hygienic hazards of wastewater originate mainly from fecal matter. Segregation opens the way to sanitation and finally to an excellent end product. • Wastewater that is not mixed with feces and urine is a great resource for high quality reuse of water. • Source control should include evaluating all products that end up in the water. High quality reuse will be far easier when household chemicals are not only degradable (original substance disappears, even if metabolites do not degrade) but can be mineralized by the available technology. Additionally, pipes for drinking water should not emit pollutants (e.g., copper or zinc). • To reduce stormwater runoff, local infiltration and/or trenches to surface waters for relatively unpolluted rainwater can be used. Prevention of pollution includes avoiding copper or zinc gutters and roof materials that can cause heavy metal pollution of rainwater runoff. TREATMENT SYSTEMS Limitations of Conventional Sanitation In conventional sanitation systems, human excreta are mixed with water and flushed away by conventional flush toilets. The wastewater is then collected and transported in sewers and treated in a centralized plant. This results in high water demand, the spread of potentially dangerous pathogens and micropollutants (e.g., residues of pharmaceuticals) in a large volume of water and also the loss of the option to reuse graywater and to produce fertilizer. The initially small amount of feces could be hygienized easily by cheap methods.

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ECOLOGICAL WASTEWATER MANAGEMENT

For the strange mixture called ‘‘municipal wastewater,’’ hygienization is an expensive further treatment step. Hormones and medical residues reveal another weakness of sanitation systems. These substances reach receiving waters easily especially because of their polarity and usually very low degradation rates in treatment plants. Sewerage systems have a couple of severe disadvantages. In general, they are a very costly part of the infrastructure (if sufficient rehabilitation is done) and drain large amounts of water from the region. Combined systems emit raw wastewater into receiving waters with overflows. Storage tanks are very expensive if the number of overflows is low. Separate systems often have a large number of wrong connections (4).

the potential for homeowners, who have existing septic systems, to continue to benefit from their original investment and avoid the potentially large transfers of water from one watershed to another that can occur in centralized systems. In small communities of low population density, decentralized systems are the most cost-effective option. These systems are variable, can be designed for various sites that have, for example, shallow water tables, shallow bedrock, low-permeability soils, and small property lot size. Additionally, decentralized systems can provide cost-effective solutions for areas that require advanced treatment, such as nutrient removal and/or disinfection, while recharging local aquifers and providing reuse opportunities close to points of wastewater generation.

Centralized Versus Decentralized Wastewater Treatment

Technology Options

Sanitation systems can be classified as follows:

Decentralized wastewater treatment alternatives for small communities can be broadly placed in the following categories that represent the basic approaches to wastewater conveyance, treatment, and/or disposal (6, revised):

• Decentralized or ‘on-plot’ systems in which safe disposal of excreta takes place on or near a single household or a small settlement. • Centralized or ‘off-plot’ systems in which excreta are collected from individual houses and carried away from the plot to be treated off-site. The selection of the most appropriate sanitation system is influenced by ecological, technical, social, cultural, financial, and institutional factors. Proper decisions on where to connect houses to a sewerage system and where to build on-site facilities or small decentralized plants are the key issue for the economics of the whole wastewater infrastructure. Good regional planning can avoid the deadlock of very expensive sewerage systems that use all the money which could serve the environment in highly efficient decentralized treatment and collection systems. There are cost calculation procedures that include longterm development to balance operating- and investment costs and products (reuse water, fertilizer, soil improver). The price of secondary products can be very relevant in economically weak and water-scarce countries where water and industrial fertilizers are no longer subsidized. Reuseoriented sanitation can easily exceed the performance of the most advanced high-tech end-of-the-pipe plant often at much lower costs (4). Water and sanitation projects and programs will fail to be sustainable if they are not planned and designed to meet the needs of the end user (5). Decentralized Wastewater Treatment Systems The use of decentralized wastewater treatment systems offers these advantages (6). They save money, protect the homeowner’s investment, promote better watershed management, offer an appropriate solution for low density communities, provide a suitable alternative for varying site conditions, and furnish effective solutions for ecologically sensitive areas. Decentralized systems prevent unnecessary costs by focusing on preventive measures (assessment of community conditions/needs and maintenance of existing systems), instead of reacting to crises. They further maximize

• Natural systems use soil as a treatment and/or disposal medium, including land application, constructed wetlands, and subsurface infiltration. Some sludge and septage handling systems, such as sand drying beds, land spreading, and lagoons, are included. • Conventional treatment systems use a combination of biological, physical, and chemical processes and employ tanks, pumps, blowers, rotating mechanisms, and/or mechanical components as part of the overall system. These include suspended growth, fixed growth, and combinations of the two. This category also includes some sludge and septage management alternatives, such as digestion, dewatering and composting systems, and appropriate disposal. • Alternative collecting systems that use lightweight plastic pipe buried at shallow depths, have fewer pipe joints and less complex access structures than conventional gravity sewers. These include pressure, vacuum, and small-diameter gravity sewer systems. • Alternative treatment systems use source control and separating systems. An example of an alternative treatment system for ecological sanitation is given later. None of the described systems is a priori or not an EcoSan system. A number of criteria as given above have to be met for a sanitation system to be called an EcoSan system. Low Water Consumption and Water-Free Toilets and Urinals. A major part of alternative treatment systems are devices suitable for reducing water consumption and/or for separating feces and urine. Therefore, the available systems of low water consumption and water-free toilets and urinals are described: 1. Conventional water-flush toilet (listed as the reference system): The conventional water-flush toilet is

ECOLOGICAL WASTEWATER MANAGEMENT

2.

3.

4.

5.

standard in most industrialized countries. A precondition to the effective disposal of excrements in this manner is sufficient availability of flushing water (about 6–8 liters per flush) as well as a corresponding disposal system (sewer system and wastewater treatment plant or other disposal). At the high dilution of human excretions (approximately 100 to 250 g feces and 1 to 1.5 liters urine), the use of a conventional flush toilet requires about 40 liters of water per day, and the wastewater produced contains both nutrients and pathogens. Water-conserving toilet: The water-conserving toilet reduces the required consumption of water to 1.0 to 4.5 liters per flush. Together with a reduction in water costs and an increase in the available capacity of the existing wastewater collection system (e.g., collection pits), water-conserving toilets produce a more concentrated wastewater that can be used further more easily (e.g., biogas extraction, solid waste separation). Due to the smaller volume of flushing water, blockages in the pipes must be prevented by a flushing device, and certain structural requirements must be observed (e.g., minimum gradient in the downstream pipe). Water-conserving toilet with waste segregation: In addition to the reduction in water consumption, the water-conserving toilet with waste segregation allows the division of feces and urine. Therefore, urine can be used for, for example, agricultural purposes, and costs can be reduced by avoiding removal of nutirents in sewage treatment. The installation of a waste-segregation toilet is only possible in combination with a corresponding collection system (urine collection tank in the house, regular emptying of the tanks, agricultural application). Feces and urine are divided in a userfriendly manner by an adapted effluent vent in the toilet bowl. Vacuum toilets: Vacuum toilet systems dispose of flushing water and excrement by using a pipeline network under vacuum (approximately 0.5 bar) connected to a collection tank. Water is required only to rinse the toilet bowl, not for facilitating transport (approximately 0.7 to 1.0 liter per flush). Until now, the vacuum system has been used mostly in ships, trains, and aircraft (limited flush and wastewater capacities). The wastewater is very concentrated, so it is suitable for energy generation in a biogas plant. Drawbacks to the system are the technical requirements (operation and maintenance of the vacuum unit) and the associated financial costs (with susceptibility to breakdown, for example, blockages). Toilet systems without water (dry toilets): • Compost toilets: The compost in this system is formed from an aerated mixture of excrements and composting earth held in a container. For proper function, a minimum air temperature of 10 ◦ C is required as well as regular checks (approximately once/month) of the structural composition and equipment. Regular extraction of the humus is necessary.

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• Dehydration toilets are based on drying excrements, which are odorless and almost sterile in dehydrated form and can be used as a soil conditioner. It is therefore important to segregate urine and feces. Urine can then be collected for agricultural applications, drained away, or evaporated. The dehydration process can be accelerated by using moisture-absorbing material or heat, such as exposure to the sun. Dehydration toilets are ideal for dry and arid climates. • Water-free urinals are made of ceramics or plastics, that must be regularly impregnated with an antibacterial agent. Odor is contained by an insert with a special sealant that has a lower specific mass than urine and is buoyed upward or by using a float. As crystalline precipitation occurs only on contact of urine with water, such deposits cannot occur in the down pipe. A further advantage of the water-free urinal is the low installation and maintenance costs (no connection to a water supply and no mechanical or electrical flushing fixtures). It has been clearly demonstrated that urine sorting toilets are feasible (e.g., Sweden has more than 3000 installations; Ref. 7). Drawbacks have been observed from too small diameters of urine pipes that clog from scaling. For waterless separating toilets, one major problem remains: Men are often reluctant to sit down to urinate. This would cause a loss of urine and mix urine with feces. A luxury solution for this problem would be a private waterless urinal. Dry toilets will fail if those who plan/design do not understand the basics, wrong materials are used, or there is poor workmanship, and users are not involved and sufficiently instructed. ‘‘Fail’’ usually means that the content of the processing chambers turns wet, resulting in odors, fly breeding, and incomplete sanitation. Example of an Alternative Treatment System Figure 4 shows an alternative treatment system that is suitable for single houses and rural settlements and is based on sorting toilets (6). The example presented is more suitable for developed countries due to the number of technical systems used. Urine from separating toilets and waterless urinals flows into a storage tank where it stays until it is used mainly for agricultural purposes. The storage period should be at least half a year because this is an appropriate time for collection and part of the eventual medical residues can be destroyed during this period. Feces are flushed by an appropriate amount of water (e.g., 4 or 6 liters) and discharged into one chamber of a two-chamber composting tank (with filter floor or filter bag) where the solids are precomposted. After a 1-year collecting, dewatering, and composting period, the flow is directed to the second chamber; the first one is not fed for 1 year. This allows further dewatering and precomposting and makes removal from the tank safer (although the matter is not sanitized then). The products removed from the composting tank are brought further to full composting. They could be

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ECOLOGICAL WASTEWATER MANAGEMENT

Shower, washing

Kitchen

Graywater

Mechanical pretreatment

Separating toilet

Waterless urinal

Urine (yellowwater)

Feces (brownwater)

Two chamber composting tank

Storage

Solid bio-waste Constructed wetland

Composting

(Manure tank)

Water reuse tank

Garden (not directly on food)

Agricultural use

Receiving water Irrigation

Figure 4. Concept for an alternative treatment system.

mixed with solid kitchen and garden waste to decompose completely and allow further sanitizing. The filtrate from the composting tank is low in nutrients due to the previous separation of urine. Therefore, the filtrate can be treated together with the graywater unless high-quality reuse is planned. Graywater is pre-treated mechanically and afterward treated by a vertical-flow constructed wetland, intermittently loaded. The treated graywater is either stored and used in the garden and for irrigation or is discharged. SUMMARY The underlying aim of ecological sanitation (EcoSan) is to close nutrient and water cycles to contribute to sustainable development. Single technologies are only a means to an end to reach EcoSan goals. EcoSantechnologies, therefore, may range from nearly natural wastewater treatment techniques to separating toilets, simple household installations, and to complex, mainly decentralized systems. Technologies are not ecological per se but only in relation to the observed environment. The main objectives of sanitation systems are that they have to minimize health risks and protect the environment. Ecological sanitation systems, additionally, have to return nutrients to the soil and conserve valuable water resources. They have to be affordable and therefore accessible to the world’s poorest people, acceptable, aesthetically inoffensive, and consistent with cultural and social values, simple and robust in design and operation, and as comfortable as conventional systems. A sanitation system consists of the following components: nature (climate, water, and soil as main factors), society (including settlement pattern, attitudes, beliefs and taboos related to human excreta, and the economic status of the community), the processes occurring (that

convert human excreta into a nondangerous, inoffensive, useful product), and the device (the on-site structure specifically built for defecation and urination). These components have to be considered together when designing sanitation systems and making them work. EcoSan systems have a number of advantages that can be summarized as follows: • Advantages to the environment: If EcoSan systems could be adopted on a large scale, it would protect groundwater, streams, lakes, and the sea from fecal contamination. Less water would be consumed. Farmers would require smaller amounts of commercial fertilizers, much of which today washes out of the soil into water, thereby contributing to environmental degradation. • Advantages to households and neighborhoods: Urine separating systems, if properly managed and maintained, do not smell or produce flies and other insects. This is a great advantage over ordinary pit toilets. Urine and feces do not come into contact and produce odor. Moisture levels are too low for fly breeding. Over half the population of the developing world has no sanitary system for excreta disposal. The market for appropriate sanitation devices is enormous, and there is a big demand. The majority of separation toilets do not require expensive or high-tech equipment. Jobs could be created for builders and for collectors of urine and sanitized feces. These products can be sold to farmers or used in the garden. • Advantages to municipalities: Municipalities all over the world are experiencing greater and greater difficulty in supplying water to households and neighborhoods. EcoSan systems do not use these scarce water resources and may create therefore,

WASTE TREATMENT IN FISH FARMS

a more equitable allocation of water to rich and poor households. A wastewater infrastructure is usually built for extremely long service. The lifetimes of existing house installations, sewerage systems, and treatment facilities have to be considered. A change to EcoSan systems is easier for newly constructed settlements or rehabilitation of complete houses. The lifetime of a house installation is far shorter than that of sewerage systems. Components of source control sanitation could be installed in each renovated flat and be connected to conventional systems first. This can be economical based on the water saving from the beginning; later, after conversion of a group of houses, separate treatment can be implemented. Acknowledgments ¨ Special thanks to Elke Mullegger, Martin Edthofer, and Phil Russell for their work in proofreading.

BIBLIOGRAPHY 1. Agenda 21. (1992). The United Nations Program of Action from Rio. United Nations, New York. 2. Esrey, S., Gough, J., Rapaport, D., Sawyer, R., SimpsonH`ebert, M., Vargas, J., and Winblad, U. (1998). Ecological Sanitation. SIDA, Stockholm, Sweden. 3. Lange, J. and Otterpohl, R. (2000). Abwasser—Handbuch zu einer zukunftsfahigen ¨ Wasserwirtschaft (Wastewater—Manual for sustainable water management), 2nd Edn. MALLBETONVerlag, Donaueschingen-Pfohren, Germany [in German]. 4. Otterpohl, R. (2002). Resource efficient wastewater concepts—technical options and initial experience. Proceedings IFAT 2002. Munich, Germany. 5. Parry-Jones, S. (1999). Optimising the Selection of Demand Assessment Techniques for Water Supply and Sanitation Projects. Final Report Project/Task No: 207. Water and Environmental Health at London and Loughborough (WELL), UK. 6. Lens, P., Zeeman, G., and Lettinga, G. (Eds.). (2001). Decentralized Sanitation and Reuse—Concepts, Systems and Implementation. IWA Publishing, London, UK. 7. Johansson, M., J¨onsson, H., H¨oglund, C., Richert, S. A., and Rodhe, L. (2001). Urine Separation—Closing the Nutrient Cycle. Report, Stockholm Water Company, Stockholm, Sweden.

WASTE TREATMENT IN FISH FARMS ALEXANDER BRINKER Fischereiforschungsstelle des ¨ Landes Baden-Wurttemberg Langenargen, Germany

ASBJORN BERGHEIM RF-Rogaland Research Stavanger, Norway

WASTE TREATMENT: GENERAL ASPECTS With increasing aquacultural production, there is a greater need to reduce the amount of waste in fish

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farm effluents. These days, simple dilution is no longer considered an appropriate treatment (1). Generally, waste in fish farms is treated for two reasons: (1) in recirculating systems, it can negatively affect the fish and (2) in open systems, it can negatively affect the environment. Pollution control in general and in fish farming in particular should aim first at prevention rather than a cure, which can be achieved by using the best available technology (BAT) and best management practices (BMP). These have been reviewed and critically discussed in a recent comprehensive workshop (2). The use of highly digestible diets and stress-free husbandry have greatly reduced waste production per quantity of produced fish; for example, salmonid farms have achieved a reduction of about 80% in the last twenty years (3). However, despite BAT and BMP, the effluent often still needs to be treated and aquacultural waste management still relies heavily on end-of-pipe solutions. For treatment to be economically sustainable in terms of capital costs, running costs, and space requirements, some specific features of aquacultural waste loading must be taken into account. The waste loads in aquacultural facilities have some properties that make treatment difficult. These include very low but strongly fluctuating concentrations of solids and nutrients and high flow. Table 1 compares some important effluent load parameters in salmonid farms with those in municipal waste waters. (Fig. 1). The average effluent load from a land-based flow-through fish farm is normally much lower than that in the treated water leaving a sewage plant. However, the flow from fish farms can be very high—greater than 500 L·s−1 —and the total solid load can vary by two orders of magnitude in the course of a single day. Aquaculture treatment devices have to meet these diverse challenges but without the benefit of many of the technologies applied in sewage plants such as flocculation chemicals or biological treatments, which are often too expensive or otherwise impracticable under fish farm conditions. With respect to effluent treatment, it is important to differentiate between four types of fish culture systems: (1) flow-through fish farms, (2) recirculating aquaculture systems (RAS), (3) open net cages, and (4) pond or integrated aquaculture. In flow-through fish farms, the water quality for the fish is ensured by the steady discharge of wastes with the bulk flow. The relative loading is usually low, but due to the high flow rates, the total load in the recipient water body may become elevated, thus causing environmental problems. Waste treatment in flow-through systems is almost exclusively restricted to mechanical techniques. In RAS, where only compensatory water is added, wastes accumulate and remain within the system. The accumulated wastes must be removed to maintain physiologically adequate conditions in the culture water. In this case, waste control usually involves mechanical separation for the removal of particulate matter and the use of some kind of biofilter to remove dissolved compounds. In open net cages, waste control is very difficult as the waste can pass freely in nearly all directions. Particle traps provide the only means of collecting settled uneaten feed or fecal pellets. New

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WASTE TREATMENT IN FISH FARMS Table 1. Waste Loads of Flow-Through Fish Farms (Range of Values from Germany) Compared With Municipal Waste Loadings (11) Source

Total Suspended Solids, mg·L−1

Total Phosphorus, mg·L−1

Total Nitrogen, mg·L−1

BOD5 a , mg·L−1

0.5–10

dichlorophenol > trichlorophenol > tetrachlorophenol > pentachlorophenol The rate of biodegradation is inversely proportional to the number of chlorine atoms in the molecule for most chlorinated aromatic compounds. This is demonstrated by chlorobenzoic acids. Monochlorobenzoic acids are readily biodegradable, but most dichloro derivatives and trichlorobenzoic acids are much more stable. For polychlorinated biphenyl (PCB), there is also a relationship between the rate of biodegradation and the number and position of chlorine atoms in the molecule. The resistance of PCB in the environment is directly proportional to the number of chlorine atoms in the molecule. Chloro-substituted diphenylmethanes behave similarly. The rate of chlorinated aromatic compound biodegradation in the activated sludge process usually decreases in the order of mono-, di-, tri-, tetra-, and

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pentachloro derivatives, depending on the position of the chlorine in the molecule (3). Chloroanilines are biodegraded more slowly than chlorophenols because of the presence of both Cl and NH2 groups that retard the biodegradability of the compound. As for chlorophenols, an increased number of chlorine atoms in the chloroaniline molecule results in increased resistance to biological treatment by the activated sludge process. There is also a retardant effect due to the number, type, and position of substituents on the aromatic ring. In addition to the halogens, the substituents that retard biodegradation are NH2 , NO2 , and SO3 H. These substituents retard the rate of biodegradation by decreasing the electron density on the aromatic ring (3). Alternatively, substituents that facilitate biodegradation by increasing electron density on the aromatic ring are OH, COOH, CHO, and CH3 . Consequently, the susceptibility of chloroaniline to biodegradation depends on whether the NH2 group is readily transformed to OH by oxidative deamination. In certain compounds, a substituent (e.g., halogens) is not eliminated before the bacterium cleaves the aromatic ring. Chloroanilines are biodegraded most slowly when the Cl–NH2 group is meta-substituted, whereas para-substituted derivatives are degraded rapidly (4). The enzyme toluene dioxygenase catalyzes the hydroxylation of chloro-substituted benzenes (e.g., chlorobenzenes) as well as methyl- and chloro-substituted phenols. The dioxygenase system for initial attack, combined with the chlorocatechol degradation pathway, allows complete degradation of a range of industrial chemicals (1). Because of the nonspecific nature of the enzymes that transform benzoate to catechol, many aerobic bacteria can cometabolize chlorinated aromatic compounds. However, this biodegradation is not complete because chlorinated benzoates and catechols are the final end product of the oxidative biodegradation (5). The complete aerobic mineralization of chlorinated aromatic compounds is not typically seen, and the persistence of these compounds in the environment illustrates the ineffectiveness of bacteria for these degradations. It is known that bacteria can mediate the anaerobic dehalogenation of chlorobenzoate compounds. Reductive dehalogenation has been confirmed for a number of chlorinated aromatic acids, chlorobenzenes, chlorophenols, chlorophenoxyacetate, herbicides, and PCBs (6). During reductive dechlorination, the chlorinated aromatic compound serves as the electron acceptor. Theoretical calculations of the Gibbs free energy available from dehalogenation indicates that bacteria can benefit from the use of chlorinated aromatic compounds as electron acceptors under anaerobic conditions (7). Nevertheless, most anaerobic dehalogenation is probably the result of cometabolism. Methanogenic metabolism has been used successfully to dehalogenate a number of chlorinated aromatic compounds, including 2,4,5-trichlorophenoxyacetate, 3chlorobenzoate, 2,4-dichlorophenol, 4-chlorophenol, 2,3, 6-trichlorobenzoate, and 2,4-, 2,5-, 2,6-, 3,4-, and 3,5dichlorobenzoates (6,8). Haggblom et al. (9) found that methanogenic bacteria preferentially remove ortho-substituted chlorine and that the meta- and para-substituted

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chlorine atoms are removed at much slower rates. Methanogenic metabolism does not result in the final mineralization of chlorinated aromatic compounds. To achieve mineralization, a balanced microbial consortium is required to perform the sequential biodegradations so that the transformation products are not more toxic than the parent compound. Anaerobic dechlorination of aromatic compounds can be stimulated by adding electron donors. The addition of volatile fatty acids (i.e., acetate, butyrate, propionate) and ethanol increases the rate of dechlorination and the extent of biodegradation of many chlorinated aromatic compounds (8). Haggblom et al. (9) found that the addition of p-cresol and propionate enhances the methanogenic degradation of 2,4-dichlorophenol. The presence of electron acceptors also influences the extent of anaerobic dechlorination in the environment. Certain electron acceptors may block the desired reduction reactions through competitive inhibition of methanogenic metabolism. However, Haggblom et al. (9) report that sulfate-reducing conditions actually enhance the apparent rates of biodegradation of 4-chlorophenol, 3-chlorophenol, 2-chlorophenol, and 2,4-dichlorophenol. Chlorophenols act as a source of carbon and energy and are degraded under sulfidogenic conditions by sulfate-reducing bacteria. In groundwater environments that contain sulfate-reducing bacteria, sulfate may be the preferred electron acceptor, and methanogens and sulfate reducers probably compete for suitable electron donors. DEGRADATION OF CHLORINATED ALIPHATIC COMPOUNDS Chlorinated aliphatic compounds are typically used as solvents for cleaning and reagents for chemical synthesis. The uses include paint and ink formulations, dry cleaning, synthetic rubber production, fumigants, paint and varnish removers, degreasers, pesticide solvents, adhesives, photographic supplies, pharmaceutical products, and household and office supplies. Microbial Transformations Microbial degradation of chlorinated aliphatic compounds can use one of several metabolic processes. These include oxidation of chlorinated alkanes for an energy source, cometabolism under aerobic conditions, and reductive dechlorination under anaerobic conditions. The response of different chlorinated aliphatic compounds to these metabolic processes differs, depending on the nature of the contamination, the redox condition, and the available electron acceptors. When molecular oxygen is the electron acceptor, one- and three-atom substituted chlorinated aliphatic compounds are transformed by three types of enzymes: oxygenases, dehalogenases, and hydrolytic dehalogenases (10). The transformation products of oxygenases are alcohols, aldehydes, or epoxides. Dehalogenase transformation products are an aldehyde and glutathione. The glutathione is required as a cofactor for nucleophilic substitution by the dehalogenase enzyme. Hydrolytic dehalogenases hydrolyze the aliphatic compound and yield alcohols

as a transformation product. Higher chlorinated compounds, particularly when all available valences on carbon are substituted (e.g., tetrachloride or tetrachloroethylene), have not been transformed under aerobic conditions. They must be transformed by reductive dechlorination. Chlorinated aliphatic compounds may be either oxidized or reduced, depending on their structure and the redox potential of the aqueous environment where they are found. Reduction is possible because of their electron negative character. Consequently, polychlorinated aliphatic compounds often behave as an electron acceptor or the oxidant in a redox reaction. The more chlorinated a compound, the higher its oxidation state, and the more susceptible it is to reduction. This explains why the rate of dechlorination decreases under anaerobic metabolic conditions as, for example, tetrachloroethylene is dechlorinated to vinyl chloride. The reaction proceeds as follows: tetrachloroethylene >>>> trichloroethylene >>> 1,2-dichloroethylene >> vinyl chloride > ethylene Vinyl chloride (monochloroethylene) is more reduced, so the thermodynamic equilibrium tends to stabilize vinyl chloride as the typical end product of trichloroethylene degradation in aquifers that have negative redox potentials (i.e., an anaerobic environment). Aerobic Processes Biotransformation of some chlorinated aliphatic compounds has been demonstrated under aerobic conditions (11). Under aerobic conditions, many soil microorganisms can oxidize vinyl chloride. Most chlorinated aliphatic compounds are eventually mineralized to carbon dioxide. The aerobic degradation capabilities of these microorganisms for chlorinated aliphatic compounds have provided successful treatment processes in seeded (i.e., cultured) activated sludge reactors. Some chlorinated aliphatic compounds are degraded by cometabolism under conditions that support aerobic metabolism. These aerobic microbes generate oxygenase enzymes of broad-substrate specificity that oxidize chlorinated aliphatic compounds. These include microorganisms that belong to the genera Alcaligenes, Mycobacterium, Pseudomonas, Nitrosomonas, Xanthobacter, and Ancylobacter (12). For example, Nitrosomonas europaea catalyzes the aerobic transformation of vinyl chloride, cis- and trans-dichloroethylene, cis-dibromoethylene, and trichloroethylene (13). During the cometabolism of chlorinated alkenes, other microorganisms derive their energy from organic compounds such as methane, propane, phenol, and toluene. Several toluene-using microorganisms can degrade trichloroethylene by cometabolism. Even under varied pH and temperature, significant rates of microbial degradation are measured (14). Trichloroethylene degradation decreased by 30% at 4 ◦ C compared with that at 30 ◦ C (14). Phenol-oxidizing bacteria have a much higher capacity to degrade trichloroethylene than methane-oxidizing microorganisms (i.e., methanotrophs). Trichloroethylene degradation by phenol-oxidizing bacteria reportedly removed greater than 90% of TCE after

DEGRADATION OF CHLORO-ORGANICS AND HYDROCARBONS

phenol injection (15). In situ studies of trichloroethylene degradation have demonstrated that phenol-using microorganisms can be readily stimulated in the environment. Phenol addition is a good primary substrate; it achieves degradation of cis- and trans-dichloroethylene and trichloroethylene in situ. The apparent rate of chlorinated alkene transformation increases as phenol concentrations increase. Many investigators have confirmed the biotransformation of trichloroethylene by methanotrophic bacteria using methane as a primary substrate. These methanotrophs have the monooxygenase enzyme that will incorporate one oxygen atom from molecular oxygen into methane to produce methanol. The monooxygenase enzyme can hydroxylate many alkanes and aromatic compounds and form epoxides from chlorinated alkenes because it is not compound-specific (10). The products of these reactions are not further oxidized by methanotrophs, so a diverse community of microorganisms is needed to achieve complete mineralization of a given constituent. Trichloroethylene has been successfully degraded aerobically to carbon dioxide with methane in air, although the rate of transformation was less than that for dichloroethylenes (16). However, there are toxicity problems because some trichloroethylene oxidation products are toxic to many methanotrophic bacteria. It also appears that trichloroethylene concentrations greater than 50 mg/L inhibit methane use by methanotrophs (10). Another serious limitation of methane-oxidizing bacteria is that they cannot transform tetrachloroethylene (i.e., perchloroethylene) or higher chlorinated aliphatic compounds. The less chlorinated the compound, the greater the rate and extent of the transformation under aerobic conditions, as expected from thermodynamics. The lower the oxidation state of the compound, the less difficult it is to oxidize. Trichloroethylene is more oxidized than vinyl chloride, so it is more difficult to oxidize this compound further. As noted previously, the opposite is true when reducing an oxidized compound. The higher the degree of oxidation, the easier it is to reduce that compound. Anaerobic Processes Many chlorinated aliphatic compounds are transformed under anaerobic conditions. These compounds are mineralized in the presence of a diverse community of microorganisms. One of the predominant mechanisms for transforming chlorinated aliphatic compounds is reductive dechlorination. Reductive reactions result in replacing the chlorine atom by dihaloelimination. The reductive process usually occurs through cometabolism. Chlorinated aliphatic compounds are transformed by reductive dechlorination even at low concentrations of less than 200 parts per billion (ppb). During reductive dechlorination, the chlorinated organic compound serves as an electron acceptor. The rate of dechlorination under anaerobic conditions is linked to the rate of primary substrate oxidation. Electrons from the oxidation of a primary substrate carry out the dechlorination. The control of these reactions for bioremediation requires an understanding of the redox conditions and the influence and availability of specific electron acceptors and donors

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on the overall metabolic condition of the bacteria that perform the reduction. The availability of electron acceptors in anaerobic systems affects reductive dechlorination by competing with the chlorinated compounds for reducing potential. For example, nitrate and sulfate can inhibit the dechlorination of some chlorinated alkenes. It has been reported that the addition of nitrate to natural soil microcosms completely blocked the dechlorination of tetrachloroethylene (17). According to thermodynamic principles, microorganisms couple half reactions that yield the greatest free energy. As redox conditions become more reducing, more chlorinated compounds undergo transformation. Thus more compounds are transformed under methanogenic conditions than under other anaerobic respirations (e.g., sulfate or nitrate). Additionally, several biologically active donors (e.g., acetate and H2 ) and the ferrous ion have lower reduction potentials than most chlorinated aliphatic compounds. As a result, they can be involved in chlorine removal by reduction. Sulfate also influences these reactions. Using sulfate as the electron acceptor (e.g., sulfate reduction by Desulfovibrio sp.), the dechlorination of tetrachloroethylene proceeds at a slower rate than if carbon dioxide is the electron acceptor (e.g., methanogenesis). Sulfate may also block the dechlorination of trichloroethylene (18). The specific influence of sulfate and other electron acceptors cannot be generalized. Different microbial systems and different aquifer and groundwater chemical conditions shift the thermodynamic equilibrium. Therefore, treatability studies are always required to assess the particular situation. Chlorinated aromatic hydrocarbons can serve as both the electron acceptor and the electron donor in a reductive dechlorination reaction (7). However, chlorinated alkenes need an additional electron donor to support anaerobic dechlorination. Typical electron donors are the following: methanol, ethanol, glucose, sucrose, benzoate, lactate, formate, acetate, and butyrate. Volatile fatty acids produced under methanogenic conditions are generally considered the most effective electron donors for enhancing dechlorination. Because some methanogens consume hydrogen as an electron donor, microbial fermentations involve the reoxidation of a reduced electron carrier, such as nicotinamide adenine dinucleotide (NAD)H2 , as part of the metabolic reaction. This means that NADH2 is oxidized to NAD and H2 in the presence of methanogens. Usually, nonmethanogenic bacteria provide the hydrogen for methanogenesis (19). Abiotic Transformations Chlorinated aliphatic compounds include trichloroethylene, tetrachloroethylene, 1,1-dichloroethylene, 1, 2-dichloroethylene, carbon tetrachloride, chloroform, 1,1,1-trichloroethane, and chloroethylene (vinyl chloride). In addition to biological transformations, these compounds undergo abiotic transformations in the environment. The important abiotic transformations include substitution, dehydrohalogenation, and reduction in water (20). A typical substitution is the addition of water resulting in hydrolysis. The nucleophiles of OH− and H2 O are the principal species responsible for abiotic dehydrohalogenation

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DEGRADATION OF CHLORO-ORGANICS AND HYDROCARBONS Table 1. Ions Capable of Abiotic Displacement of Halogena Oxygenated Waters H2 O OH− Cl− Br− SO4 2− HCO3 − — a

Anaerobic Waters SO3 2− S2 O3 2− NH3 NO2 − Sn 2− R–C6 H13 S− C6 H5 S−

Reference 20.

of chlorinated aliphatic compounds in water. However, a variety of other species can displace the chlorine. These are presented in Table 1. Under anaerobic conditions, the sulfur nucleophiles are generally the most powerful. Sulfides react with chlorinated aliphatic compounds via substitution to produce mercaptans. These reactions proceed slowly in the absence of biological activity. The half-lives for the monochloroalkanes are approximately a month at 25 ◦ C (20). Polychlorinated species can have half-lives as long as 40 years. Stronger nucleophiles such as HS− can reduce the half-lives of the abiotic degradation of chlorinated aliphatic hydrocarbons. Microbial enzymes also catalyze these reactions and reduce half-lives significantly. A variety of transition metals, including nickel, iron, chromium, and cobalt, can reduce chlorinated aliphatic compounds. As a result of this oxidation–reduction (redox) reaction, the metals are oxidized. Vogel et al. (11) defined the reduction products and metals that mediate such reactions. The transition metal reduces a chlorinated compound, removing the chlorine and creating an alkyl radical that readily picks up a hydrogen atom from water, resulting in the formation of an alkane. The reduction of polychlorinated alkanes can result in both alkanes and alkenes. DEGRADATION OF PETROLEUM HYDROCARBONS Petroleum hydrocarbons include the common components of petroleum oils and products. They are found in wastewaters from petroleum refineries, petrochemical facilities, fuel storage and transportation facilities, and industrial organic chemical production facilities. Benzene and other single-ring aromatic hydrocarbons are used as solvents in industrial processes. Substituted forms of single-ring aromatic compounds are used in many industrial processes for preparing dyes, resins, antioxidants, polyurethane foams, fungicides, stabilizers, coatings, insulation materials, fabrics, and plastics. Polynuclear aromatic hydrocarbons (PAHs) are associated with petroleum refining and coal tar distillation. PAHs are also associated with waste by-products from coal gasification and coke production. Microbial Transformations Petroleum hydrocarbons can be classified into aliphatic, alicyclic, and aromatic hydrocarbon groups. Aliphatic hydrocarbons are divided into alkane, alkene, alkyne,

and unsaturated alkyl groups. Alicyclic hydrocarbons can be grouped into cycloalkanes and cycloalkenes. Aromatic hydrocarbons are comprised of benzene and its derivatives and polynuclear aromatic hydrocarbons. The presence of molecular oxygen as a terminal electron acceptor is required for successful microbial degradation of petroleum hydrocarbons. Nitrate and sulfate also serve as alternative electron acceptors during anaerobic respiration of hydrocarbons. Generally, the biodegradation rate decreases with decreasing redox potential. Only negligible biodegradation is observed under strictly anaerobic (i.e., fermentative) conditions. Consequently, hydrocarbons remain for a relatively long period of time in bottom sediments and other anaerobic portions of the aquatic environment. Petroleum hydrocarbons are generally hydrophobic compounds. Bacteria and fungi frequently attach to oil droplets because intimate contact between a microorganism and the surface of the petroleum hydrocarbon is necessary for biodegradation. Consequently, dispersing the petroleum oil in water makes it more susceptible to microbial attack. Bacteria often produce extracellular surfactants that aid in solubilizing petroleum hydrocarbons. These bacterial surfactants are complex mixtures of proteins, lipids (e.g., rhamnolipids, phospholipids), and carbohydrates (21). Aliphatic and Alicyclic Hydrocarbons The biodegradation potential of alkanes is a function of carbon chain length. Short carbon chains of less than 10 carbons are more difficult to biodegrade than longer chains. Because of their higher solubility, short-chain hydrocarbons also exhibit a high degree of toxicity in the aquatic environment (22). Longer chain aliphatic hydrocarbons are more easily biodegraded than the short-chain variety. A large number of facultative anaerobic bacteria are prevalent in the aquatic environment that can use aliphatic hydrocarbons as a source of carbon and energy. These bacteria include the genera Acinetobacter, Alcaligenes, Arthrobacter, Flavobacterium, Methylococcus, Mycobacterium, Nocardia, and Pseudomonas (21). Numerous fungi and yeast also biodegrade aliphatic hydrocarbons. Although fungi are more versatile than yeast in biodegrading short-chain hydrocarbons, both are effective in using long-chain alkanes. Aerobic biodegradation of a long-chain aliphatic hydrocarbon requires incorporating molecular oxygen into the compound. Oxygenase enzymes (i.e., monooxygenases and dioxygenases) mediate this degradation reaction (23). The pathway of alkane biodegradation is oxidation at the terminal methyl group of an alcohol and then of the corresponding fatty acid. The terminal oxidation proceeds by successive removal of two carbon units, termed the beta-oxidation sequence. Alkene biodegradation is more varied because bacteria attack at either the methyl group or the double bond. Unsaturated straight-chain hydrocarbons are usually biodegraded less easily than saturated compounds. Consequently, bacterial metabolism results in forming intermediates that consist of unsaturated alcohols and fatty acids, primary or secondary alcohols, methyl ketones,

DEGRADATION OF CHLORO-ORGANICS AND HYDROCARBONS

epoxides, and diols (21). The methyl group oxidation is the more likely biodegradation pathway. Petroleum hydrocarbons that have branch chains are less susceptible to biodegradation. Quaternary carbon and β-alkyl-branched compounds are generally considered recalcitrant and accumulate in the environment. However, combining chemical oxidative processes with biodegradation is effective for treating recalcitrant hydrocarbons. The biodegradation of cycloalkanes is usually by oxidation of the terminal methyl group and yields a primary alcohol. Hydroxylation must occur to initiate the biodegradation of cycloalkanes. The bacteria that can oxidize noncyclic alkanes can also hydroxylate cycloalkanes. Several alternate metabolic pathways exist for microbial attack on alicyclic hydrocarbons, and numerous intermediate compounds have been identified during their degradation. Substituted cycloalkyl compounds vary in their capacity for biodegradation, but those that contain carboxylic acid groups are readily biodegraded. Bacteria capable of degrading cycloalkyl carboxylic acids are numerous in the environment (24). Aromatic Hydrocarbons The nature and extent of the biodegradation of aromatic hydrocarbons depends on the number of rings in the structure, the number of substitutions, the type and position of the substituted groups, and the nature of the atoms in heterocyclic compounds. The solubility of the aromatic hydrocarbon greatly affects its potential for biodegradation under either aerobic or anaerobic conditions. Mixtures of aromatic compounds can also influence the rate of biodegradation. Aerobic microbial attack on single-, double-, and triplering aromatic compounds involves the foundation of a dihydrodiol compound. Oxidative attack on the dihydrodiol compound results in forming an alkyl catechol, a common intermediate formed during the oxidation of many aromatic hydrocarbons. Additional metabolic oxidation results in ring fission, forming either an aldehyde or an acid. This step results in the destruction of the aromatic ring leaving an oxidized aliphatic hydrocarbon, which is easily biodegraded, releasing hydrogen (25). Bacteria of the genus Nocardia can oxidize substituted aromatic hydrocarbons, such as p- and m-xylene, and use these compounds as a sole source of carbon and energy (4). A second metabolic pathway for degrading aromatic hydrocarbons involves oxidation of alkyl substitutes, which results in forming aromatic carboxylic acids that are then oxidized to dihydroxylated ring fission products (i.e., aldehydes and acids). The aldehydes and acids are then readily biodegraded by the beta-ketoadipic and meta fission pathways (26). In general, alkyl-substituted aromatic hydrocarbons are less biodegradable, the longer the chain length, or the more numerous the alkyl groups. Single-ring aromatic hydrocarbons can be transformed anaerobically by denitrifying, sulfate-reducing, ironreducing, and methanogenic bacteria (27). These are all anaerobic respirations where the nitrate, sulfate, ferric, and carbonate ions act as terminal electron acceptors (TEA) for energy metabolism. The use of these compounds as electron acceptors in microbial energy metabolism

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is called dissimilative metabolism. During dissimilative metabolism, a comparatively large amount of the TEA is reduced, and the reduced product is excreted into the aquatic environment. The possible end products of these reductions are HS− , N2 , NO2 − , N2 O, Fe2+ , and CH4 . The presence or absence of oxygen in the structure of an aromatic compound impacts both the degradation mechanism and the rate of biodegradation. The initial step in the degradation of aromatic hydrocarbons is conversion of the compound to an oxygenated form. Under anaerobic metabolic conditions, oxygen is incorporated from water into the aromatic structure by hydroxylation (28). The microbial fermentation of benzene and toluene is characterized by end products that are both partly oxidized and partly reduced. The oxidation might include both methyl group oxidation and ring oxidation. The reduction generally results in forming saturated cyclic rings (3). Grbic-Galic (29) observed that the biodegradation of benzene is initiated by ring oxidation, resulting in the formation of phenol. Three pathways are possible for toluene, starting with ring oxidation to p-cresol or o-cresol and methyl group oxidation to benzyl alcohol. Thereafter, the biodegradations proceed along pathways that are similar to the anaerobic transformations of oxygenated aromatic compounds (30). The biodegradation of multiring aromatic hydrocarbons (PAHs) is a function of the complexity of the chemical structure of the compound. In general, PAHs that contain four or more aromatic rings are much less biodegradable than compounds that contain only two to three rings (3). Several of the higher ring number PAHs and the intermediate products of their biodegradation are either toxic or carcinogenic. They are also strongly hydrophobic, which predicts that their concentration in the aqueous phase is always relatively low. Therefore, significant portions of the PAHs are found adsorbed on particles and possibly trapped in the micropores of these particles. Consequently, the rate of biodegradation is controlled by the sorption–desorption kinetics of the strongly sorbed PAH compounds. Enhancing solubilization by introducing bacteria that produce extracellular surfactants improves the in situ biodegradation of PAHs in groundwater. Biodegradation of the unsubstituted di- and tri-ring PAHs in marine and freshwater is well documented in the literature (25,31,32). The bacterial degradation rates for phenanthrene and anthracene appear to be related to the water solubilities of these compounds. As for benzene and its derivatives, the oxidation of diand tri-ring PAHs involves the formation of dihydrodiol intermediates. Catechol is the principal intermediate product of these microbial degradations (25). Bacteria that can degrade anthracene and phenanthrene include the genera Aeromonas, Beijerinckia, Flavobacterium, Nocardia, and Pseudomonas (31). Biodegradation of unsubstituted PAHs that contain four or more aromatic rings (e.g., fluorene, pyrene, benzo(a)pyrene, benzo(a)anthracene, dibenzo(a)anthracene) has been documented in the literature (33–35). The bacteria that can degrade the higher molecular weight PAHs were also identified in this literature; they include the genera Alcaligenes, Beijerinckia, Mycobacterium, and

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DEGRADATION OF CHLORO-ORGANICS AND HYDROCARBONS

Pseudomonas. These research studies have indicated that cometabolism coupled with analog substrate enrichment may be necessary to treat PAHs that contain four or more aromatic rings. Cometabolism appears to be the principal mechanism for biodegrading benzo(a)anthracene (35). The presence of biphenyl, m-xylene, and salicylate were necessary to induce oxidation of benzo(a)anthracene to carbon dioxide and a mixture of o-hydroxy polyaromatic compounds. Because the high molecular weight PAHs do not induce enzyme production in many bacteria, the addition of naphthalene as an analog substrate was necessary to biodegrade benzo(a)anthracene in this study. BIBLIOGRAPHY 1. Spain, J.C., Haigler, B.E., and Nishino, S. (1989). Development of bacteria for biodegradation of chloroaromatic compounds. In: Biotechnology Applications in Hazardous Waste Treatment. G. Lewandowski, P. Armenante, and B. Baltzis (Eds.). Engineering Foundation, New York, pp. 129–147. 2. Compeau, G.C., Mahaffey, W.D., and Patras, L. (1991). Fullscale bioremediation of contaminated soil and water. In: Environmental Biotechnology for Waste Treatment. G.S. Sayler, R. Fox, and J.W. Blackburn (Eds.). Plenum Press, New York, pp. 91–109. 3. Pitter, P. and Chudoba, J. (1990). Biodegradability of Organic Substances in the Aquatic Environment. CRC Press, Boca Raton, FL. 4. Gibson, D.T. (1989). Recent advances in the microbial degradation of aromatic hydrocarbons. In: Biotechnology Applications in Hazardous Waste Treatment. G. Lewandowski, P. Armenante, and B. Baltzis (Eds.). Engineering Foundation, New York, pp. 149–165. 5. Reineke, W. (1984). Microbial degradation of halogenated aromatic compounds. In: Microbial Degradation of Organic Compounds. D.T. Gibson (Ed.). Marcel Dekker, New York, pp. 319–360. 6. Young, L.Y. (1984). Anaerobic degradation of aromatic compounds. In: Microbial Degradation of Organic Compounds. D.T. Gibson (Ed.). Marcel Dekker, New York, pp. 487–523. 7. Dolfing, J. (1992). Gibbs free energy of formation of halogenated aromatic compounds and their potential role as electron acceptors in anaerobic environments. Environ. Sci. Technol. 26: 2213–2218. 8. Gibson, S.A. and Suflita, J.M. (1990). Anaerobic biodegradation of 2,4,5-Trichlorophenoxyacetic acid in samples from a methanogenic aquifer: stimulated by short chain organic acids and alcohols. Appl. Environ. Microbiol. 56: 1825– 1832. 9. Haggblom, M.M., Rivera, M.D., and Young, L.Y. (1991). Anaerobic degradation of chloroaromatic compounds under different reducing conditions. EPA Symposium on Bioremediation of Hazardous Wastes. Falls Church, VA. 10. Semprini, L., Grbic-Galic, D., McCarty, P.L., and Roberts, P.V. (1992). Methodologies for evaluating in-situ bioremediation of chlorinated solvents. U.S. Environmental Protection Agency, EPA/600/R-92/042. 11. Vogel, T.N., Criddle, G.S., and McCarty, P.L. (1987). Transformation of halogenated aliphatic compounds. Environ. Sci. Technol. 21(8): 722–736. 12. Rainwater, K. and Scholze, R.J. (1991). In-situ biodegradation for treatment of contaminated soil and groundwater. In: Biological Processes-Innovative Hazardous Waste Treatment

13.

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22. 23.

24.

25.

26.

27.

Technology Series. Vol. 3, H.M. Freeman and P.R. Sferra (Eds.). Technomic, Lancaster, PA, pp. 107–122. Vanelli, T., Logan, M., Arciero, D.M., and Hooper, A.B. (1990). Degradation of halogenated aliphatic compounds by the ammonia-oxidizing bacterium nitrosomonas europaea. Appl. Environ. Microbiol. 56(4): 1169–1171. Shields, M. (1991). Treatment of TCE and degradation products using Pseudomonas cepacia. Symposium on Bioremediation of Hazardous Wastes: EPA’s Biosystems Technology Development Program, Abstracts, Falls Church, VA., April 16–18. Hopkins, G.D., Semprini, L., and McCarty, P.L. (1993). Microcosm and in-situ field studies of enhanced biotransformation of trichloroethylene by phenol-utilizing microorganisms. Appl. Environ. Microbiol. 59: 2277–2285. Tsien, H.C., Bousseau, A., Hanson, R.S., and Wackett, L.P. (1989). Biodegradation of trichloroethylene by Methylosinus trichosporium. Appl. Environ. Microbiol. 55: 3155–3161. Sewell, G.W., Gibson, S.A., and Russell, H.H. (1990). Anaerobic in-situ treatment of chlorinated ethenes. Workshop on In-Situ Bioremediation of Groundwater and Contaminated Soils. Water Pollution Control Federation Hazardous Wastes Committee, WPCF, Annual Conference, October 7–11, Washington, DC, pp. 68–79. Bouwer, E.J. and Wright, J.P. (1988). Transformations of trace halogenated aliphatics in anoxic biofilm columns. J. Contam. Hydrol. 2: 155–169. Baek, N.H. and Jaffe, P.R. (1988). Anaerobic mineralization of trichloroethylene, presented at the International Conference on Physicochemical and Biological Detoxification of Hazardous Wastes. Vol. II. X.C. Wu (Eds.). Technomic, Lancaster, PA, pp. 772–782. Reinhard, M., Barbash, J.E., and Kunzle, J.M. (1988). Abiotic dehalogenation reactions of haloaliphatic compounds in aqueous solution, presented at the International Conference on Physicochemical and Biological Detoxification of Hazardous Wastes, Vol. II, X.C. Wu (Ed.). Technomic, Lancaster, PA, pp. 722–741. Britton, L.N. (1984). Microbial degradation of aliphatic hydrocarbons. In: Microbial Degradation of Organic Compounds. D.T. Gibson (Ed.). Marcel Dekker, New York, pp. 89–130. Bitton, G. and Gerba, C.P. (1984). Groundwater Pollution Microbiology. John Wiley & Sons, New York, pp. 23–24. Wackett, L.P., Brusseau, G.A., Householder, S.R., and Hansen, R.S. (1989). Survey of microbial oxygenases: trichloroethylene degradation by propane-oxidizing bacteria. Appl. Environ. Microbiol. 55: 2960–2964. Trudgill, P.W. (1984). Microbial degradation of the alicyclic ring: structural relationships and metabolic pathways. In: Microbial Degradation of Organic Compounds. D.T. Gibson (Ed.). Marcel Dekker, New York, pp. 131–180. Gibson, D.T. and Subramanian, V. (1984). Microbial degradation of aromatic hydrocarbons. In: Microbial Degradation of Organic Compounds. D.T. Gibson (Ed.). Marcel Dekker, New York, pp. 181–252. Arvin, E., Jensen, B., Godsy, E.M., and Grbic-Galic, D. (1988). Microbial degradation of oil and creosote related aromatic compounds under aerobic and anaerobic conditions, presented at the International Conference on Physicochemical and Biological Detoxification of Hazardous Wastes. Vol. II, X.C. Wu (Ed.). Technomic, Lancaster, PA, pp. 828–847. Huling, S.G., Pivetz, B., and Stransky, R. (2002). Terminal electron acceptor mass balance: light nonaqueous phase liquids and natural attenuation. J. Environ. Eng. 128: 246–252.

LANDFILL 28. Vogel, T.N. and Grbic-Galic, D. (1986). Incorporation of oxygen from water into toluene and benzene during anaerobic fermentative transformation. Appl. Environ. Microbiol. 52: 200–202. 29. Grbic-Galic, D. (1990). Microbial Degradation of Homocyclic and Heterocyclic Aromatic Hydrocarbons under Anaerobic Conditions. Department of Civil Engineering, Environmental Engineering and Science, Stanford University, Stanford, CA. 30. Grbic-Galic, D. and Vogel, T.M. (1987). Transformation of toluene and benzene by mixed methanogenic cultures. Appl. Environ. Microbiol. 53: 254–260. 31. Hogan, J.A., Toffoli, G.R., Miller, F.C., Hunter, J.V., and Finstein, M.S. (1988). Composting physical model demonstration: mass balance of hydrocarbons and PCBs, presented at the International Conference on Physicochemical and Biological Detoxification of Hazardous Wastes. Vol. II, X.C. Wu (Ed.). Technomic, Lancaster, PA, pp. 742–758. 32. Mahaffey, W.R., Compeau, G., Nelson, M., and Kinsella, J. (1990). Development of Strategies for Bioremediation of PAHs and TCE. In-Situ Bioremediation of Groundwater and Contaminated Soils. WPCF Annual Conference, Washington, DC, pp. 3–47. 33. Environmental Protection Agency. (1993). Pilot-Scale Demonstration of a Slurry-Phase Biological Reactor for CreosoteContaminated Soil—Applications Analysis Report. EPA/540/ A5-91/009. 34. Grifoll, M., Casellas, M., Bayona, J.M., and Solanas, A.M. (1992). Isolation and characterization of a fluorene-degrading bacterium: identification of ring oxidation and ring fission products. Appl. Environ. Microbiol. 58: 2910–2917. 35. Mahaffey, W.R., Gibson, D.T., and Cerniglia, C.E. (1988). Bacterial oxidation of chemical carcinogens: formation of polycyclic aromatic acids from benzo(a)anthracene. Appl. Environ. Microbiol. 54: 2415–2423.

LANDFILL JIM PHILP Napier University Edinburgh, Scotland, United Kingdom

INTRODUCTION All countries rely to a greater or lesser degree on landfilling to dispose of the huge quantities of municipal solid waste (MSW) generated. For example, the United Kingdom has traditionally relied very heavily on landfilling of MSW as it has a relatively poor recycling infrastructure compared with some other European countries. The United Kingdom produces roughly 28 million tons of MSW per annum, of which about 80% is landfilled. Only some 12% is recycled and about 8% incinerated with energy recovery. Moreover, the amount of waste being generated is growing some 3–4% per annum. Globally, landfilling has had a checkered history, and poor practice in design and operation of landfills has led to serious environmental problems. With regard to the threats to water bodies, the production of landfill leachate is by far the most significant. Liquid leachate develops at a site when its water holding capacity is exceeded. If the site is unlined, the leachate makes its way off-site to

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groundwater or to a surface water body and can cause drastic water pollution. The Landfill Directive of the EU (1) has the overall aim ‘‘to prevent or reduce as far as possible negative effects on the environment, in particular the pollution of surface water, groundwater (authors’ italics), soil and air, and on the global environment, including the greenhouse effect, as well as any resulting risk to human health, from the landfilling of waste, during the whole life-cycle of the landfill.’’ This succinctly crystallizes the objectives of modern landfill design, and the Directive has farreaching consequences for the way we handle and dispose of MSW. There are various stringent requirements of the Directive, the most important of which for preventing water pollution are: • Higher engineering and operating standards are to be followed. • Biodegradable waste has to be progressively diverted away from landfill. By 2020, the amount going to landfill will be 35% of that of 1995. • Banning of disposal of liquid wastes to landfill sites, along with certain hazardous and other wastes, will be implemented. THE MICROBIOLOGY OF REFUSE DECOMPOSITION One view of a landfill is as an enormous, solid-state fermenter in which naturally selecting microbial populations, usually bacteria, anaerobically decompose refuse components, ultimately to their mineral constituents. The process of anaerobic decomposition is microbiologically complicated, and a great many details remain to be elucidated. However, the overall process can be summarized as in Fig. 1. Although this image seems to separate the various processes, it should be kept in mind that all happen contemporaneously. As the landfill is microbiologically active over long periods, often decades, only the anaerobic processes (those in blue in Fig. 1) concern us here: hydrolysis and fermentation, acetogenesis, and methanogenesis. One of the reasons that these processes occur over decades is that

Process Hydrolysis/aerobic degradation

Products CO2 H2O

Hydrolysis and fermentation

Organic acids H2 CO2 H2O Ammoniacal nitrogen

Acetogenesis

Acetic acid H2 CO2

Methanogenesis

CH4 CO2

Oxidation

CO2

Aerobic

Anaerobic

Aerobic

Figure 1. The various microbiological processes in a landfill.

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the energy available from anaerobic processes is much lower than that from aerobic processes. For example, the aerobic mineralization of glucose to CO2 and H2 O liberates more than seven times as much energy as the anaerobic mineralization of glucose to CO2 and methane (CH4 ).

valerate by Syntrophomonas, produce acetate, hydrogen, and carbon dioxide, which are used by methanogens. The result is the overall conversion of fatty acids to methane. The conversion of fatty acids to acetate is energetically unfavorable, and the reaction depends on the removal of hydrogen by methanogens. Likewise, the methanogens need hydrogen for their metabolism.

Hydrolysis and Fermentation Long-chain, insoluble carbohydrates, lipids, and proteins are not in a form in which they are readily metabolized by microorganisms. Hydrolytic reactions break down these long-chain molecules to smaller, more water-soluble molecules that can then be metabolized. Fermentation reactions produce soluble fermentation end products such as short-chain volatile fatty acids (VFA) and gaseous products of variable water solubility such as CO2 and hydrogen. Acetogenesis The soluble acids from fermentation are converted to acetic acid, CO2 , and hydrogen by acetogenic bacteria. Other bacteria, acidogenic bacteria, convert carbohydrates, CO2 , and hydrogen to acetic acid. An important point is that the conversion of fermentation products to acetic acid occurs only at low concentrations of hydrogen. Hydrogen is produced at several stages, so it must be removed to prevent the inhibition of acidogenesis. If the partial pressure of hydrogen is too high, reduced organic acids, such as propionic, lactic, butyric, valeric, and caproic, start to accumulate, producing smells. Either from lowering the pH, mobilizing toxic metals, or inhibition due to the presence of high concentrations of these acids, there would be subsequent inhibition of methanogenesis. Methanogenesis Hydrogen concentration is kept low as a result of consumption by strictly anaerobic sulfate-reducing bacteria and methanogenic (methane-producing) bacteria. The methanogenic pathways of all species have the conversion of a methyl group to methane in common; however, the origin of the methyl group is variable. Although most isolated species can reduce CO2 , the majority of biological methanogenesis (about 70%) originates from conversion of the methyl group of acetate to methane, although in most cases acetate is not used as an energy source. Others acquire the methyl group directly from substrates such as methanol or methylamines. Syntrophism The low-energy yields involved in the anaerobic conversion of refuse to methane forces these different organisms into very efficient cooperation. Such cooperations are known as syntrophic relations. Syntrophism is a special case of symbiosis between two metabolically different types of bacteria, which depend on each other for degradation of a certain substrate, usually for reasons of energetics. There is a classic example in landfill microbiology. Metabolism of low molecular weight fatty acids, such as propionate by Syntrophobacter and caproate and

Landfill Microbiology Is Mass Transfer Limited Very often mass balance calculations of methane generation at landfills suggest that the overall process is not working nearly as efficiently as it might, even considering that several steps along the way have poor thermodynamics, which leads to very protracted timescales for return of the site to stability, and the consequent need is for long-term monitoring. If the bottleneck is not in thermodynamics, then where is it? The answer lies in a very common observation; that newspaper, although very rich in calories, can still be read after excavation from a landfill site decades after its disposal there. Cellulose, the most abundant polymer on the planet, represents about 50% of the organic material going to landfills. The rate limitation in a landfill is much more likely to be associated with solid substrates, such as paper. Here the limitation is not thermodynamic, but mass transfer. The substrate has to be converted from the solid form into an aqueous form before other metabolic associations can continue the degradation. The enzymatic hydrolysis of polymers to monomers has long been known as the rate-limiting step in the conversion of cellulose to methane and in the digestion of refuse. ENVIRONMENTAL IMPACTS OF LANDFILLS Landfilling domestic refuse creates a whole host of negative environmental impacts, summarized in Fig. 2. These impacts have variable magnitudes and can be categorized as localized (odor, noise, litter, transport risks, public health risks created by birds and vermin, explosive gas migration), diffuse (groundwater, surface water, drinking water contamination), and global (greenhouse gas generation). In the present context, the effects on water pollution are the primary concern. And the water pollution problems are a direct result of leachate generation and its off-site migration. Leachate Composition Leachate is the result of water infiltration to the site that exceeds the water holding capacity of the waste and other site materials. The water balance of a landfill site can be summarized as LC = PR + SRT − SRO − EP − ST where

LC = leachate PR = precipitation SRO = surface run-off SRT = surface run-to (zero on a well designed site) EP = evapotranspiration ST = change in water storage

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Gases, odours Evaporation Cap Noise, litter

Rain

Run-off River Waste

Liner Gas migration Leachate Water borehole Leachate seepage Groundwater

Figure 2. Environmental impacts of landfill practice.

Leachates vary in composition from site to site and also according to the age of the leachate. Leachates generated during the early stages of anaerobic decomposition are characterized by high concentrations of VFA, acidic pH, high BOD:COD ratio, and high levels of ammoniacal nitrogen and organic nitrogen. Ammonia is largely produced from the degradation of proteins. The low redox potential of such leachates facilitates the production of soluble, reduced-state metals, including chromium, iron, and manganese. These ‘‘young’’ leachates are much more environmentally damaging than mature leachates produced during the stable methanogenic phase. By the time that methanogenesis is occurring at a high rate, many of the fatty acids have been converted ultimately to methane and CO2 . Methanogenic leachates are more likely to have a higher pH (resulting in lower heavy metal concentrations due to precipitation), lower levels of ammoniacal nitrogen, and a lower BOD:COD ratio (from the biodegradation of fatty acids).

Longer than 18 months results in excessive leachate production in wet areas. Cells are subsections of phases, which vary in size according to operational exigencies. Cell sizes are minimized according to the surface area required for maneuvering large machinery on the site. It may also be possible to size each cell in which the rate of vertical filling exceeds the rainfall plus water holding capacity to minimize leachate generation. The practice of phasing (Fig. 3) has the objective of progressive excavation and filling of the site. As a result, at any one time, part of the site may be restored, part may be in the process of being capped, part is being prepared to receive waste, and only a relatively small part is being actively filled. When properly done, there will also be sufficient space for storage and protection of materials for subsequent restoration, and also coordination of haul roads and access routes. The environmental benefits of the phase/cell strategy are: • reduction of leachate generation;

LANDFILL DESIGN AND CONSTRUCTION IN RELATION TO ENVIRONMENTAL MITIGATION

• progressive installation of leachate and gas control systems;

It is not an objective to discuss detailed engineering design. Rather, in modern landfills, several design and construction measures can be taken to minimize the environmental effects of the site; and it is these that are summarized here. For example, a critical element of any landfill design is capacity, which is influenced by factors such as waste density, amount of daily cover, and the thickness of the final cap. However, for this discourse, capacity has little relevance to environmental effects.

• segregation of clean surface water run-off within and outside the site;

Phasing and Cell Construction A phase is a subsection of the total landfill to be filled; generally, it has an operational period of 12 to 18 months.

• protection of local amenity. Phases are generally filled from base to cap in a continuous operation, then capped and restored, leaving a temporary unrestored face sloping to the landfill base (Fig. 3). In deep landfills, such as those constructed in old quarries or opencast mining sites, the phases are vertically tiered so that, overall, the site is a three-dimensional honeycomb of cells.

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LANDFILL A: Phasing plan Most sensitive visual and 1 noise receptors

Direction of work

3 5 2

4

Haul road C: Phase 4 in operation

Under restoration

Under restoration 1

3

2

6

Site entrance

In preparation

B: Phase 2 in operation

Borrow pit/ stockpile location

In operation

In preparation

D: Phase 6 in operation

5

Restored

3 Restored

4

Under restoration

5

6 In operation

Figure 3. The phased approach to site management (adapted from Ref. 2).

In operation

Daily Cover

Synthetic liners, normally made of high-density (HDPE) or low-density (LDPE) polyethylene, are available in thicknesses of 0.5 to 2.0 mm. Although of very low permeability, their installation requires great care to prevent tearing. Geosynthetic clay liners (GCL) (Fig. 4) are relatively new products gaining acceptance as barrier systems in municipal waste landfills. These offer some advantages over traditional bottom liners and covers and retain low hydraulic conductivity. Advantages include

This practice involves covering refuse to a certain minimum depth daily with the following environmental benefits: • prevention of wind-blown litter; • suppression of odor; • deterrence to vermin and birds; • improvement of the site appearance. From the point of view of water (and therefore leachate) and gas management, the material used for daily cover should be sufficiently permeable that it does not impede water or gas flow. Impermeable material creates perched conditions and makes it difficult to extract leachate. Ideally, the material of daily cover is soil excavated from within the boundaries of the landfill to prevent net consumption of void. Liners A key design parameter for MSW is to attain an impermeability of 10−9 m/s and so prevent leachate breakthrough to the unsaturated zone. The objective should be to select a new site on soil with a hydraulic conductivity lower than this. If this is not possible, then a variety of materials can be used to line the new site, either as single or multiple layers. Clay liners varying in thickness from about 0.5 m (for imported clay) to 2 m (for in situ clay) are natural liners of high ion exchange capacity to retard the movement of toxic metals. Bentonite, the typical clay, is extremely absorbent. The hydraulic conductivity of dry, unconfined bentonite is 10−9 m/s. When saturated, however, it drops to less than 10−12 m/s−1 .

• fast and easy to install; • self healing of rips and tears due to the swelling property of bentonite; • cost-effective in regions where clay is not readily available; • their thin cross-section compared to a clay liner maximizes the capacity of a landfill and still protects groundwater. Tests show that holes up to 75 mm diameter will self-heal when the clay hydrates and swells. Stitchbonding or needle-punching creates small holes in the geotextile that heal due to swelling of the bentonite. The geotextile is often a blend of HDPE and very low-density polyethylene (VLDPE). Drainage and Leachate Collection/Recirculation/Treatment Leachate collected by the liner system must be removed to accelerate stabilization of the site and prevent liner damage, which is accomplished by drainage to collection sumps at low points via a granular layer containing perforated pipes of sufficient slope to allow gravity drainage. Leachate is then removed from the sumps, either by pumping in vertical wells or by gravity drains in a valley

LANDFILL LEACHATES, PART I: ORIGIN AND CHARACTERIZATION

699

Upper geotextile ~5 mm

Clay and adhesive

Clay bound with adhesives to upper and lower geotextiles

Lower geotextile

~5 mm

~4−6 mm

Clay and adhesive or clay

Clay stitch-bonded between upper and lower geotextiles

Clay

Clay needle-punched through upper and lower geotextiles

Figure 4. Geosynthetic landfill liner types.

site, where the leachate can either be treated on-site by a dedicated wastewater treatment plant or transported off-site for treatment. Using a proper leachate collection system, it is possible to spray leachate back onto emplaced waste, which effectively uses the landfill site as a flushing anaerobic bioreactor and can improve landfill gas generation by uniformly wetting the waste.

LANDFILL LEACHATES, PART I: ORIGIN AND CHARACTERIZATION A.I. ZOUBOULIS Aristotle University of Thessaloniki Thessaloniki, Greece

PETROS SAMARAS

Gas Abstraction and Use This is mentioned here for completeness, although methane is highly insoluble in water and therefore contributes very little to the water-related environmental problems of landfills. During the stable methanogenic phase of a landfill, by far the longest phase, landfill gas, consists mainly of methane and carbon dioxide, both of which are greenhouse gases; the former is highly explosive; the latter is relatively water-soluble and corrosive. Most of the landfill gas management systems for landfills are designed with the characteristics of methane in mind. By appropriate siting of vertical or horizontal gas abstraction wells, it is possible to collect the gas, flare it, or, if economically viable, burn it for energy generation. BIBLIOGRAPHY 1. Council Directive 1999/31/EC on the Landfill of Waste. (1999). 2. Department of the Environment. (1995). Landfill design, construction and operational practice. Waste Management Paper 26B. HMSO, London.

READING LIST Kiely, G. (1996). Environmental Engineering. McGraw-Hill, London. Kjeldsen, P., Barlaz, M.A., Rooker, A.P., Baun, A., Ledin, A., and Christensen, T.H. (2002). Present and long-term composition of MSW landfill leachate: A review. Crit. Rev. Environ. Sci. Technol. 32: 297–236. Qian, X., Koerner, R.M., and Gray, D.H. (2002). Geotechnical Aspects of Landfill Design and Construction. Pearson Education, Harlow. Senior, E. (1995). Microbiology of Landfill Sites, 2nd Edn. Lewis, Boca Raton, FL.

Chemical Process Engineering Research Institute Thermi-Thessaloniki, Greece

INTRODUCTION Sanitary landfilling is the most widely used method for disposing of urban solid wastes around the world. The extensive use and the public awareness of this disposal method have raised concerns, over the negative environmental impacts and the pollution potential that this practice creates, as well as by the by-products of landfills (e.g., leachates, biogas, odors, etc.). Among them, leachates are considered the most important environmental burden. Depending on the composition and extent of decomposition of the disposal of refuse, as well as on the hydrological parameters existing in the landfill site, leachates may become highly contaminated wastewaters. Landfill leachate, as defined in the U.S. Environmental Protection Agency Code of Federal Regulations (CFR) Title 40, Part 258.2, is the liquid that has passed through, or emerged from the disposal of solid wastes and contains soluble, suspended, or miscible materials from these wastes. Over time, the seepage of water through the landfill mainly from precipitation increases the mobility of pollutants and the potential for transferring them into the surrounding environment. As water passes through the layers of disposed solid wastes, it may ‘‘leach’’ pollutants from them, moving them deeper into the soil. The mobility of pollutants may present a potential hazard to public health, as well as to the environment, causing significant pollution problems in the groundwater aquifer and in

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adjacent surface waters. As a result, understanding and predicting leachate generation routes, as well as containing it and the subsequent appropriate treatment are required for environmentally proper handling of these heavily polluted wastewaters. A simple measure to prevent the movement of toxic and hazardous waste constituents from a landfill is a liner operated in conjunction with a leachate collection system. Leachates are typically collected from a collection system placed at the bottom of the landfill. Leachates may also be collected by using slurry walls, trenches, or other containment systems. The leachate generated may vary from landfill site to site, based on a number of factors, which include the types of waste accepted for disposal, the operating practices (such as shredding, daily cover with soil, or capping), the depth of fill, the applied compaction of wastes, the annual precipitation at the landfill site, and the landfill operational age. LEACHATE GENERATION Leachates are the combined wastewater, containing organic and inorganic constituents, produced when water and/or other liquids seep through solid wastes, deposited in urban or hazardous solid waste landfills. The quantity of leachates is influenced by several interacting factors, such as annual precipitation, runoff, infiltration, evaporation, transpiration, mean ambient temperature, waste composition, waste density, initial moisture content, and underlying soil conditions (depth) (1). A number of techniques have been reported, using the water budget analysis through a landfill site, to estimate the amount of leachate generated (2). The various components of moisture used in the water budget are shown in Fig. 1. According to this analysis, the primary source of moisture is precipitation over the landfill site. A part of this moisture results in surface runoff, another part is returned to the atmosphere in the form of evapotranspiration from the soil and the surfaces of plants, and the remainder is added to soil moisture storage. The maximum moisture that can be retained, without continuous downward percolation by gravity, is known as field capacity. Whenever the moisture content exceeds the field capacity of the soil, water percolates down into

(through) the solid waste. The addition of moisture to solid waste over a period of time saturates the solid waste to its field capacity, resulting thereafter in leachate generation. The various moisture components, which constitute the processes taking place in a landfill that produce leachate, are affected by several parameters, such as (3): 1. Precipitation, which varies geographically and seasonally. 2. Surface runoff and infiltration, which depend on the intensity and duration of storms, surface slope, permeability of soil cover, and amount and type of vegetation. 3. Evapotranspiration at a landfill site, which is affected by the type of the soil and vegetation. 4. Soil moisture storage capacity, which is continually changing; it increases due to infiltration and decreases due to evapotranspiration. Several methods used water balance calculations of these components to assess leachates generation rates. In general, following are the characteristics of the leachates produced (4): 1. A higher leachate generation rate is expected in humid than in dry areas. 2. Leachate generation follows a pattern similar to that of precipitation (rain), then remains at constant flow for a time period. 3. Production of leachates may be minimized by proper and efficient covering operations, careful drainage design, selection of vegetative cover, etc. 4. The quantity of leachate generated and its qualitative characteristics are significant for designing and constructing the most appropriate collection and treatment devices. In general, leachates are generated over a long time period, unless percolation is prevented by the closure procedures and the final land use (5). In this case, leachate generation will cease shortly after the completion and closure of the landfill. However, for the proper design of

Evapotranspiration Surface runoff Soil cover

Precipitation Irrigation

Vegetative cover Soil moisture storage

Percolation Solid wastes Solid waste moisture storage

Figure 1. The various moisture components in a sanitary landfill.

Leachate collection

Leachates Liner

LANDFILL LEACHATES, PART I: ORIGIN AND CHARACTERIZATION

a leachate treatment system, the characteristics of the leachate are necessary in addition to quantitative data. LEACHATE CHARACTERISTICS As leachates pass through or emerge from deposited solid wastes, they may contain soluble, suspended, or miscible materials from the wastes. Several factors may affect leachate quality, such as (6) • • • • • •

specific types of solid waste accepted/deposited operating practices (shredding, cover, or capping) amount of infiltration depth of fill compaction age

The specific waste types received for disposal are the most representative characteristic of a landfill and, therefore, of the respective wastewater generated because the main contaminants in this wastewater are derived from the materials deposited in the site. The amount of infiltration and the age of a landfill are the primary factors that affect the concentration of contaminants in the leachate produced. The remaining factors influence mainly the rate of infiltration. The highest concentrations of contaminants are typically present in leachates of new or very young landfills (7). However, the overall loads (i.e., the mass) of pollutants are generally not very large because new landfills typically generate low volumes of leachate. As the volume of waste approaches the capacity of the landfill and the production of leachate increases, both the pollutant loadings (i.e., flow × concentration) and the concentrations of certain contaminants, which are mainly organic pollutants, tend to increase. The increase of pollutant concentration is attributed to the onset of decomposition within the landfill and to the leachates that traverse the entire depth of the refuse. Therefore, large pollutant loadings from a typical landfill occur during the period of high leachate production and contain high levels of contaminants. The periods of varying leachate production cannot be quantified readily because they are site specific and depend on each of the aforementioned variables. Over a period of time (as the landfill ages and leaching continues), the concentration of contaminants in the leachate decreases. The landfill may continue to generate substantial quantities of leachate; however, gradually the load of pollutants become lower because of the lower concentrations of soluble, suspended, or miscible contaminants that remain in the landfill. As the decomposition process within the landfill continues, the landfill attains a stabilized state of equilibrium, where further leaching produces leachates with a pollutant load lower, than during the period of peak leachate production. This stabilized state is presumably the result of decomposition of landfill waste by indigenous microorganisms that remove (biodegrade) many of the organic contaminants usually susceptible to further leaching.

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Leachate characteristics change over time because there is a shift from the initial relatively short period of aerobic decomposition toward a longer period of anaerobic decomposition that has two distinct subphases (6, 8). The biological decomposition of landfilled municipal refuse is often based on the anaerobic breakdown of organic wastes. The biological activity occurs in a landfill shortly after the deposition of urban wastes, containing a large percentage of organic materials. Initially, the solid wastes, which contain high moisture content, can be decomposed rapidly under aerobic conditions, creating large amounts of heat. As oxygen is depleted, the intermediate anaerobic stage of decomposition begins. This change from aerobic to anaerobic conditions occurs unevenly through the landfill and depends upon the rate of oxygen diffusion into the fill layers. In the first stage of anaerobic decomposition, the socalled ‘‘acidic phase,’’ extra-cellular enzymes convert complex organic wastes, including carbohydrates, proteins, and fats, to more soluble organic molecules. Once the organic wastes are solubilized, their conversion to simpler organic molecules, such as acetic, propionic, butyric, isobutyric, valeric, isovaleric and hexanoic acids takes place; acetic acid is the main catabolic product of anaerobic fermentation. As a group, the low molecular weight, but highly polar, organic acids are termed volatile fatty acids (VFA). These soluble organic acids enter the leachate percolating through a landfill, resulting in a decreased pH of the leachate and increased oxygen demand. VFA impart to the leachate from this phase their characteristic ‘‘barnyard’’ odor and comprise the majority of its organic load. Anaerobic activity in the landfill can also lower the oxidation–reduction (redox) potential of the wastes, which under low pH conditions, can cause an increase in the concentration of dissolved inorganic contaminants. Eventually, in the second or ‘‘methanogenic’’ phase of anaerobic decomposition, methane gas-forming bacteria within the landfill begin to convert the organic acids to methane and carbon dioxide, reducing the organic strength. The fraction of organic carbon, remaining after this degradation process, tends to be more oxidized, but has a higher molecular weight, higher than 500 amu. The absence of organic acids in the landfill increases the pH of the leachate, toward neutral or alkaline, which can subsequently decrease the solubility of inorganic contaminants and lower their concentrations in the resulting leachate. The age or degree of decomposition of a landfill may be ascertained by observing the concentration of various leachate ‘‘gross’’ parameters, such as BOD5 , COD, TDS, or the organic nitrogen (Norg ) concentration (9). The values of these leachate parameters can vary over the decomposition life of a landfill, depending on the specific phase. Typically, leachates from the early, acidic phase of anaerobic decomposition may be up to 35 times stronger than domestic wastewater and can have a COD content of more than 20,000 mg/L, BOD5 greater than 12,000 mg/L, high volatile fatty acids concentration of about 6000 mg/L, and high content of inorganic compounds, such as 1300 mg/L chloride, but low phosphorous concentration (less than 1 mg/L) (10). Leachates from the older landfills have

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lower values of BOD5 and COD, as well as of most organic pollutants, indicating the presence of smaller amounts of degradable compounds, derived from the aged stabilized waste. The COD of leachates from the ‘‘methanogenic’’ phase tend to be lower, between 1500 and 4000 mg/L, and the significant decrease in the VFA concentration results in an increase in the pH to 7, or even higher. In addition, aged leachates can contain high levels of compounds existing in reduced form, such as ammonia (greater than 1000 mg/L), as well as a high concentration of chlorides because of the anaerobic environment of the landfill. Furthermore, certain metals such as iron, lead, and zinc tend to form stable complexes with the high molecular weight organic compounds (i.e., higher than 50,000), increasing their respective concentrations in the leachate (6). However, using only these parameters, other refusefilling variables, such as the processing of wastes prior to disposal or the fill depth, would not be taken into consideration. To compensate for these additional variables, several researchers have proposed examining certain ratios of leachate parameters over time (7). The most important (and widely used) such ratio is BOD5 :COD. Leachates from younger landfills typically exhibit BOD5 :COD ratios of approximately 0.8, whereas older landfills exhibit lower ratios (in many cases as low as 0.1). The decline in the BOD5 :COD ratio with age is due primarily to readily biodegradable material (e.g., phenols, alcohols, VFA) that degrade faster than the more recalcitrant compounds (such as the heavy molecular weight organic compounds, including humic and fulvic acids), which are much more difficult to treat biologically. As a result, as the landfill ages, the BOD5 of the leachate decreases faster than the COD. Other ratios that decrease over time include the volatile solids to fixed (inorganic) solids (VS:FS), volatile fatty acids to total organic carbon (VFA:TOC), and sulfate to chloride (SO4 :Cl), which is inversely related to the redox potential (ORP). As a result of the variation in leachate strength, leachates are commonly distinguished as young, acidphase leachates and old, methanogenic ones, as well as medium- and low-strength leachates (11). For ‘‘young’’ leachates, a typical ratio of BOD5 /COD around 0.7 has been suggested; the corresponding values for mature leachates are 0.5, for aging 0.3, and for ‘‘old’’ 0.1. A typical time period for the transition from young to older leachate types is between 3 to 10 years from the landfill start-up, but may be as short as, 2 years in specific cases. In summary, the following conclusions can be drawn regarding landfill leachate quality and their treatability (3): • Leachate characteristics are highly variable. • The quality of leachates changes with age, and therefore, the treatment facility should be flexible enough to handle/treat appropriately the changing leachate quality. • A reliable estimate of the chemical quality of leachates should include analytical experimental data measured under the particular conditions prevailing in the landfill area.

The characteristics of leachates are usually very different from those of domestic wastewaters and similar to heavily loaded industrial wastewaters, indicating the need to use advanced treatment methods for the effective removal of pollutants before leachates are discharged. BIBLIOGRAPHY 1. Leckie, J.O., Pacey, J.C., and Halvadakis, C. (1979). Landfill management with moisture control. J. Environ. Eng. Div., ASCE 110(4): 780–796. 2. Tchobanoglous, G., Theisen, H., and Vigil, S. (1993). Integrated Solid Waste Management: Engineering Principles Management Issues. McGraw-Hill, New York. 3. Qasim, S.R. and Chiang, W. (1994). Sanitary Landfill Leachate: Generation, Control and Treatment. Technomics, Lancaster, PA. 4. Tatsi, A. and Zouboulis, A.I. (2002). A field investigation of the quantity and quality of leachate from a municipal solid waste landfill in a Mediterranean climate (Thessaloniki, Greece). Adv. Environ. Res. 6(3): 207–219. 5. Barlaz, M.A. et al. (2002). A critical evaluation of factors required to terminate the post-closure monitoring period at solid waste landfills. Environ. Sci. Technol. 36(16): 3457–3464. 6. Forgie, D.J.L. (1988). Selection of the most appropriate leachate treatment methods. Part 1. A review of potential biological leachate treatment methods. Water Pollut. Res. J. Can. 20(2): 308–328. 7. Environmental Protection Agency. (2000). Development Document for Final Effluent Limitations Guidelines and Standards for the Landfills Point Source Category. EPA-821R-99-019, Washington, DC. 8. Kjeldsen, P.K. et al. (2003). Present and long term composition of MSW landfill leachate—a review. Crit. Rev. Environ. Sci. Technol. 32(4): 297–336. 9. Al-Yaqout, F. and Hamoda, M.F. (2003). Evaluation of landfill leachate in arid climate—a case study. Environ. Int. 29(5): 593–600. 10. Harrington, D.W. and Maris, P.J. (1986). The treatment of leachate: A UK perspective. Water Pollut. Control 85(1): 45–56. 11. Henry, J.G., Prasad, D., and Young, H. (1987). Removal of organics from leachates by anaerobic filter. Water Res. 21(11): 1395–1399.

LANDFILL LEACHATES: PART 2: TREATMENT PETROS SAMARAS Chemical Process Engineering Research Institute Thermi-Thessaloniki, Greece

A.I. ZOUBOULIS Aristotle University of Thessaloniki Thessaloniki, Greece

INTRODUCTION The characteristics (composition) of sanitary landfill leachates are very different from domestic wastewaters,

LANDFILL LEACHATES: PART 2: TREATMENT

and their quality varies from landfill to landfill, as well as with the particular landfill age. Hence, their treatment is based largely on industrial wastewater treatment processes. However, no single treatment method is considered efficient enough to achieve the high removal rates of pollutants, usually required; therefore, several treatment trains are currently used, including combinations of aerobic/anaerobic processes and several modes of physical–chemical treatment systems. The selection and design of leachate treatment facilities requires knowledge of several parameters, such as leachate quantity and quality, degree of necessary treatment, disposal methods, and effluents guidelines. These are the main problems that have to be considered: • Specific treatment schemes applied in a particular landfill may not be transferable to other sites. • Leachate quantity and quality vary seasonally, depending on climatic and hydrologic factors. • The composition of disposed of solid wastes greatly affects the composition of leachates. During plant design, the fluctuations in the leachate generation rate and its composition should be considered. Furthermore, the treatment system should also be flexible enough to treat the ‘‘young’’ leachates during the preliminary stages of landfill operation, as well as the ‘‘older’’ leachates produced during landfill aging. During the early development of appropriate methods for leachates, the main efforts were focused on the application of treatment processes commonly used for municipal wastewaters; they were based mainly on biological processes. Physical and chemical systems used were later complementary, aimed at the development of more efficient overall techniques (1). BIOLOGICAL PROCESSES FOR TREATING LEACHATES Both aerobic and anaerobic biological units have been used to treat landfill leachates. The number of landfill facilities that use variations of biological treatment as part of landfill wastewater treatment systems in the United States has been reviewed by the EPA (2), and it is shown in Table 1. According to this table, most of the biological treatment systems use aerobic processes, including suspended growth processes (i.e., activated sludge, sequencing batch reactors, lagoons, etc.), as well as fixed-film processes (i.e., trickling filters or rotating biological contactors). However, the use of anaerobic systems is rather limited; these systems are most effective for treating high strength leachates (i.e., whose COD values are over 4000 mg/L) and for wastewaters containing refractory (not easily biodegradable) contaminants because of the effectiveness of methanotropic microorganisms in metabolizing these compounds. An important disadvantage of anaerobic treatment systems is the sensitivity of the applied methanotropic microorganisms to certain toxic substances, commonly found in many leachates. The design of aerobic systems is based mainly on the requirements for removing organic loading in terms

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Table 1. Biological Treatment Facilities for Leachates in U.S. Landfillsa Type of Biological Treatment Activated sludge Aerobic lagoon systems Facultative lagoons Trickling filters Anaerobic systems Powdered activated carbon treatment (PACT)b Nitrification systems Rotating biological contactors (RBC) Sequencing batch reactors (SBR) Denitrification systems Other biological treatment systems a b

% of Nonhazardous Facilities

% of Hazardous Facilities

9 10 7 0 2 1

33 — — — — —

2 0

— —

1

33

1 13

— —

Reference 2. In combination with activated sludge.

of BOD5 and COD or on requirements for nitrogen removal. The selection of the design estimates depends on the effluents guidelines that have to be met and the problems anticipated from the existing high ammonia concentrations. A rule of thumb proposed for selecting the most appropriate design criteria is the ratio of BOD5 /Ncontent: when this ratio is less than 1, leachates are characterized by very high nitrogen (mainly ammonia) concentrations, and then nitrification criteria prevail; when this ratio is greater than 1, then the organic removal criteria dominate (3). Suspended growth biological treatment systems usually include mechanically based aerators to provide the required oxygen to the microbial population and for mixing the liquor components. The ranges of typical design and performance parameters of activated sludge systems are as follows (4): • hydraulic retention time: 1–10 days. • solids retention time between 1 and 10 days. • food to microorganism (F/M) ratio from 0.02–0.4 kg BOD5 /kg MLVSS/day. • average nutrient requirement ratio BOD5 : N:P = 100:3.2:0.5. • Removal efficiencies in terms of BOD5 and COD from 90–99%, depending on experimental conditions and the properties of raw leachates. Several measurements of operational parameters have indicated that a large part of the organic compounds in raw leachates are usually not readily biodegradable and require prolonged reaction times and extensive biological activity to oxidize them. Efforts to determine the removal rates of various compounds in an aerobic reactor resulted in discriminating four distinct and successive steps of substrate use by microorganisms, in order of gradually increased degree of difficulty: carbohydrates, fatty acids,

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amino acids, and humic substances of high molecular weight (4). Residual organics consist mainly of fulviclike materials with molecular weights in the range 500–10,000, which are not readily biodegradable. Introducing an anoxic stage may enhance the removal of nitrogen from leachates, achieving more than 70% total nitrogen removal. However, the highest removal rates may be achieved by adding an organic carbon source, such as methanol. High metal removal rates (up to 99%) have also been observed in activated sludge treatment systems of leachates; this was attributed to the oxidation of metals, forming insoluble compounds, and the incorporation of the respective precipitates into bacterial flocs (2). However, when metal content is high (e.g., 80 mg/L iron and 10 mg/L manganese), a pretreatment step may be necessary for their efficient reduction to prevent the resulting low MLVSS/MLSS ratios and certain mixing problems. Stabilization ponds and aerated lagoons have also been used as pretreatment steps for leachates, prior to disposal in municipal sewers or recycling into the landfill. Although extended aeration and lagoon-type systems are favored for treating leachates because of low manpower requirements and operational simplicity, however, this application has a primary drawback, the extended land required. Additional problems connected with suspended growth systems include the intense foaming of leachates, high power consumption, potential inhibition of biological activity by increased concentrations of metals, high sludge production rate, filamentous bulking of sludge, and decreased biological activity due to deficiency in certain nutrients (usually phosphorous) (5). Fixed-film aerobic systems provide an appropriate substrate for attaching and growing aerobic and facultative (anaerobic) bacteria. However, their use is currently limited for treating leachates. Typical loading rates for such systems are detention time about 10 h and loading rates from 2 to 5 g of NH3 -N/m2 /day. The systems achieve high removal efficiencies and result in effluents whose ammonia content is lower than 1 mg/L and BOD5 lower than 20 mg/L (2). Anaerobic biological systems used for treating leachates are often based on fixed-film type reactors using inert media, as well as on suspended growth systems, such as the upflow anaerobic sludge blanket (UASB) reactor. Detention times reported for these systems range from 1–15 days for anaerobic filters and from 1–6 days for UASB units; for the latter systems, removal capacities greater than 80% have been observed, and for anaerobic filters, the corresponding values range from 70–99% COD removal efficiency. The determination of BOD5 /COD and COD/TOC ratios in effluents from anaerobic treatment units showed that the composition of these effluents compared with those of leachates created from landfills of intermediate age. It may be concluded that a substantial part of biodegradable organics can be removed by the landfill itself, acting as an anaerobic bioreactor, and thus, the subsequent application of biological treatment methods and in particular of anaerobic processes are only moderately effective for the removing the remaining organic matter from ‘‘older’’ leachates (1).

In general, aerobic biological treatment of leachates is possible for ‘‘young’’ leachates that have high BOD5 /COD ratios, higher than 0.4, as well as low ammonia content. Activated sludge systems are the most common treatment techniques, and fixed-film systems are best used for nitrification of ‘‘older’’ leachates rich in ammonia. In this case, phosphorous addition may be necessary to provide the proper nutrient balance for sufficient cell growth. Anaerobic treatment systems may also be successful for treating leachates because of the advantages of relatively simple design; low capital, operating, and maintenance costs; and the ability to treat leachates with high BOD5 /COD ratios. However, anaerobic systems also present certain drawbacks that limit their application: slow biomass establishment, requirement for higher (at least mesophilic) temperatures (i.e., difficult use in cold climates); and poor solids separation (2,5). Combined treatment of leachates in an existing municipal wastewater treatment plant is considered convenient; it has been applied in several cases, and it is the preferred disposal method for leachates, when the principal following requirements are met: availability of a sewer system, wastewater treatment plant capacity high enough to accept the heavily loaded leachates, process compatibility with the specific (composition) characteristics of leachates, and a sludge treatment facility large enough to handle the increased sludge production rates (4). Several studies of the cotreatment of municipal wastewaters and landfill leachates concluded that the overall treatment process and the effluent quality are not seriously affected by the addition of leachates up to 10% by volume at the municipal sewage, although this depends mainly on the loading strength of the leachates (6,7). However, several problems may arise during the cotreatment that are connected to the possible negative effect and accumulation of heavy metals, the conversion of ammonia, the variations in temperature, the (much higher) sludge production, the (usually intense) foaming, and (poor) solids settleability. As a result, the cotreatment of leachates and municipal wastewaters has to be studied case-by-case, considering the significant problems that may appear due to the presence of toxic compounds, as well as the specific portion of leachate that has to be cotreated. The introduction of a pretreatment facility, such as a simple aerated lagoon, may satisfy the requirements for a preliminary polishing step in these cases. PHYSICAL–CHEMICAL TREATMENT OF LEACHATES Physical–chemical techniques are becoming increasingly common for treating industrial wastewater and for reclaiming municipal wastewaters, especially when intended for reuse. These techniques include mainly processes such as equalization, neutralization/pH adjustment, chemical precipitation and coagulation, chemical oxidation, activated carbon adsorption, air-stripping, ion exchange, and membrane separation. The application of physicochemical treatment methods to leachates offers the advantages of short start-up periods, relative stability to temperature variations, and the potential for automation.

LANDFILL LEACHATES: PART 2: TREATMENT

Equalization The composition and generation rates of leachates at landfills may vary widely due to their direct relationship to rainfall, storm water run-on and runoff, groundwater entering the waste-containing zone, and the moisture content and absorptive capability of disposed of wastes. To allow equalization of pollutant loadings and flow rates, the leachates are often collected, prior to treatment, in tanks or ponds that have sufficient capacity to hold the peak flows generated at the landfill facility (2). A constant flow is delivered to the treatment system from these holding tanks to dampen the variation in hydraulic and pollutant loadings to the wastewater treatment system. This reduction in hydraulic and pollutant variability increases the performance and reliability of treatment systems applied downstream and can reduce the size of subsequent treatment tanks as well as the chemical or polymer feed rates of supplementary reagents by reducing the maximum flow rates and the concentrations of pollutants to be removed. The equalization systems consist of steel or fiberglass holding tanks or lined ponds that can provide sufficient capacity to contain peak flow. Detention times determined by using a mass balance equation and depend on sitespecific generation rates and treatment design criteria. Detention times can range from less than a day up to 90 days; the median value is about 2 days. Equalization systems usually contain either mechanical mixing or aeration systems; they enhance the equalization process by keeping the tank contents well mixed and prohibiting settling of solids. pH Adjustment The pH of wastewater generated by landfills may have a wide range of values, depending on the specific types of wastes deposited in the landfill. In many instances, the raw wastewater may require neutralization to eliminate either high or low pH that may upset the treatment system subsequently applied, such as an activated sludge biological treatment. The landfill facilities may use neutralization systems in conjunction with chemical treatment processes, such as chemical precipitation, to adjust the pH of the wastewater and to remove metals to optimize process control. Acids, such as sulfuric acid or hydrochloric acid, are added to reduce pH, whereas alkalis, such as sodium hydroxide or lime, are added to raise the pH. Neutralization may be performed in a holding tank, in a rapid mixing tank, or in an equalization tank. Typically, the neutralization systems applied at the end of a treatment system are designed to control the pH of the final discharge between 6 and 9. Chemical Precipitation and Coagulation Suspended particulates and colloidal matter contained in surface waters or wastewaters can be removed by coagulation using multivalent cations, such as Ca2+ , Fe3+ or Al3+ . As a result, several investigations of the treatment of leachates dealt with the use of chemical precipitation and coagulation (8).

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In chemical precipitation, soluble metallic ions and certain anions, found in landfill wastewaters, are converted to insoluble forms that precipitate from the solution. Most metals are relatively insoluble as hydroxides, sulfides, or carbonates. Coagulation is used in conjunction with precipitation to facilitate their removal by agglomerating suspended and/or colloidal materials. The precipitated metals can be subsequently removed from the wastewater stream by filtration, settling clarification (sedimentation), or some other type of gravity-assisted separation. Other treatment processes such as equalization, chemical oxidation, or reduction (as in the case of hexavalent chromium) usually precede chemical precipitation. The performance of the chemical precipitation process is affected mainly by other chemical interactions, temperature, pH, the solubility of waste contaminants, and mixing effects (2). Common precipitating reagents used at landfills usually include lime, sodium hydroxide, soda ash, sodium sulfide, or alum. Other chemicals also used in precipitation and coagulation, as well as for pH adjustment, include sulfuric and phosphoric acids, ferric chloride, and polyelectrolytes (synthetic organic polymers). Often, landfills use an appropriate combination of these chemicals. Precipitation by sodium hydroxide or lime is the most conventional method for removing metals from leachates. Hydroxide and coagulant precipitation has proven effective for removing several metals, such as trivalent chromium, pentavalent arsenic, copper, lead, nickel, and zinc. However, sulfide precipitation may also be used, instead of hydroxide precipitation, to remove certain metal ions, such as mercury, lead and silver more effectively. Carbonate precipitation is another method of chemical precipitation; it is used primarily to remove antimony or lead. Use of alum as a precipitant/coagulant agent results in the formation of aluminum hydroxides in wastewaters, containing calcium or magnesium bicarbonate. Aluminum hydroxide is an insoluble gelatinous floc, which settles slowly and entraps suspended materials. It is considered particularly effective for removing certain metals, such as arsenic or cadmium. Lime is less expensive than sodium hydroxide, so it is used more frequently at landfills employing hydroxide precipitation. However, lime is more difficult to handle and feed, as it must be slaked, prepared in a slurry, and mixed intensively, often plugging the feed system lines. Lime precipitation also produces a larger volume of sludge. In addition to the type of chemical agent selected for treating leachates, another important design factor in the operation of chemical precipitation is the pH. Metal hydroxides are usually amphoteric, meaning they can react chemically both as acids or bases; as such, their solubilities increase at both lower (acidic) and higher (alkaline) pH levels. Therefore, there is an optimum pH value for the precipitation of each metal, which corresponds to its minimum solubility. Another key consideration in chemical precipitation is the necessary detention time during the sedimentation phase of the process. The optimal detention time depends on the wastewater being treated and on the desired effluent quality.

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LANDFILL LEACHATES: PART 2: TREATMENT

The first step in chemical precipitation is pH adjustment and the addition of coagulants. This process usually takes place in separate mixing and flocculation tanks. After mixing the wastewater with the appropriate chemical reagents, the resulting mixture agglomerates in the flocculation tank, and it is mixed slowly by mechanical means, such as mixers, or by recirculation pumping. The wastewater then undergoes further separation, by clarification (by settling) or filtration, where the precipitated metals are removed from the (cleared) solution. In a clarification system, an organic flocculant, such as a synthetic polymer (e.g., polyacrylamide), is sometimes added to help the settling. The resulting sludge from the clarifier or from the filter must be further treated, disposed of, or recycled. Several studies have been reported on the examination of coagulation–flocculation for the treating landfill leachates, aimed to optimize performance, by selection of the most appropriate coagulant, determination of experimental conditions, assessment of the pH effect, and investigation of flocculant addition (9,10). Aluminum sulfate (alum), ferrous sulfate, ferric chloride, and ferric chlorosulfate are commonly used as coagulants (3). Iron salts proved more efficient than aluminum salts, resulting in COD reductions of up to 56%, whereas the corresponding values for alum or lime were 39 and 18%, respectively (11). Additionally, high COD removal capacities have been observed during the combined action of alum and lime on stabilized leachates (12). Furthermore, the addition of flocculants together with coagulants may substantially enhance the floc settling rate (9). The coagulation–precipitation process has been investigated mainly by using stabilized or biologically pretreated landfill leachates, as a final polishing treatment stage. However, limited information exists on the efficiency of this physicochemical process, when used to remove pollutants from leachates, partially stabilized by recirculation or from recently produced (‘‘fresh’’) leachate. This technique may be important for enhancing leachate biodegradability prior to biological treatment. High COD removal capacities (about 80%) have been obtained during the addition of ferric chloride to partially stabilized leachates, whereas low COD reductions (lower than 35%) have been measured during the addition of coagulants to raw samples (13). In general, the coagulation and/or precipitation of raw leachates by the addition of lime resulted in the removal of multivalent cations, suspended solids, and color from raw leachates, but the effect on organic matter removal was rather negligible. In an early work of Slater et al. (14), it was found that only a small percentage (about 4–6%) of organic compounds with molecular weight of 10,000 or more was contained in an industrial raw leachate, whereas most of the organics had a molecular weight of 500 or less. However, after lime addition, the higher MW fraction disappeared, whereas the other fractions of MW 500 or less and between 500 and 10,000 remained almost untouched. As a result, lime addition to ‘‘young’’ leachates is not expected to be effective for removing organics because this type of leachate contains mainly high amounts of lower MW volatile fatty acids. Coagulation by lime may be an efficient method for treating ‘‘older’’ or biologically

treated leachates that contain a large fraction of high MW substances, such as humic and fulvic acids (1). However, the use of ferric salts or of alum has proven more effective than lime, possibly due to different optimum pH process conditions that range between 4.5 and 5.5. Chemical Oxidation Chemical oxidation processes can generally be used in wastewater treatment to remove ammonia, to oxidize cyanide, to reduce the concentration of residual organics, and to reduce the bacterial and viral content. Chemical oxidation for treating leachates has been successful, based on several oxidants, including chlorine gas, calcium hypochlorite, potassium permanganate, hydrogen peroxide, and ozone (2). Both chlorine and ozone are two chemicals that are commonly used to destroy residual organics in biologically pretreated wastewater. When these chemicals are used for leachate treatment, the resulting disinfection of the wastewater is usually an added benefit. Chemical oxidation is a potential treatment option for removing certain organic pollutants from leachates or groundwater. The amount of oxidant required in practice is generally greater than the theoretical mass calculated. The reasons for this are numerous and include incomplete oxidant consumption and oxidant demand caused by the simultaneous presence of other oxidizable species in solution. Oxidation reactions depend on the presence of appropriate catalysts, as well as on pH control, which is an important design variable. For many facilities using chemical oxidation, partial oxidation of organics, followed by additional treatment options, may be more efficient and cost-effective than using a complete oxidation treatment scheme alone. The use of chlorine gas in leachates has been tested and resulted in high color and iron removal rates, but in limited reduction of organic matter, which is possibly due to (1) the presence of ammonia, which has to be initially destroyed by break-point chlorination before any organic oxidation and (2) the presence of relatively difficult to oxidize organics. On the other hand, the use of hydrogen peroxide presents several benefits: control of odor from stored leachates, removal of sulfides discharged to municipal sewers, and growth control of undesirable microorganisms near discharge of leachates. In addition, the study of the MW distribution of organics in leachates treated with hydrogen peroxide showed that it enhances the percentage of compounds of MW less than 1000 and therefore, increases the possibility of further biological treatment. Ozone treatment of leachates provides several benefits, such as the removal of color, the degradation of particular organics (such as polyaromatic hydrocarbons), and the reduction of phenols and toxicity (15,16). Furthermore, ozone application may enhance the biodegradability of leachates by converting pollutants to end products or to intermediate products that are more readily biodegradable or can be more easily removed by adsorption (17). Ozone application proved very effective for removing color and iron, but was less efficient in removing COD. It was concluded that ozonation should not be used in leachates of high volatile fatty acid content, especially when acetic

LANDFILL LEACHATES: PART 2: TREATMENT

acid is present, due to their strong resistance to chemical oxidation (8). Activated Carbon Adsorption Activated carbon adsorption is a physical separation process, in which organic and inorganic materials are removed from wastewaters by sorption, attraction and accumulation of the contaminants on the surface of carbon granules. Most organic compounds and some metals typically found in landfill leachates can be effectively removed by using granular activated carbon (GAC). Although the primary removal mechanism is adsorption, biological degradation and filtration are additional pollutant removal mechanisms, also provided by an activated carbon filter. Adsorption capacities of 0.5 to 10% by weight are rather typical in many industrial applications. Spent carbon can be either regenerated on site by thermal processes, such as wet-air oxidation or steam stripping, or for smaller operations, it can be regenerated off site or sent directly for disposal in to hazardous waste landfills. Several studies have been presented concerning the use of activated carbon adsorption for treating landfill leachates. In general, this process is very effective in removing residual organics that remain after prior biological treatment of leachates, and thus it could be used as a final polishing step for biologically pretreated leachates and/or for well-stabilized ‘‘old’’ leachates (8). Furthermore, the combination of powdered activated carbon with an activated sludge system results in enhanced removal capacities of organic matter up to 98% of BOD5 (2). In conclusion, the adsorption of higher MW organic compounds is enhanced by the properties of these compounds. As the MW of organics increases, their polarity, solubility, and branching properties decrease, resulting in an increase in carbon adsorption. As a result, lower MW volatile fatty acids representative of ‘‘younger’’ leachates are poorly adsorbed on activated carbon particles, whereas higher MW compounds, such as fulvic acids, found in ‘‘old’’ leachates, are adsorbed on activated carbon to a greater extent. Both powdered and granular activated carbon may be used for leachate treatment, but special consideration should be given during the design period, because of the high cost of this material. Air Stripping Stripping is an effective method for removing dissolved volatile organic compounds from wastewater. Removal is accomplished by passing air or steam through the agitated waste stream. Air stripping is used to the treat leachates, mainly to remove ammonia. In this case, the pH must be increased to between 10.8 to 11.5, usually by adding NaOH or Ca(OH)2 solutions; ammonia stripping takes place by blowing large volumes of air upward through the leachate bulk volume. This process is carried out in towers, where leachates trickle down over some type of inert material or in a shallow aerated reaction vessel. Air stripping has been proved efficient for the extensive reduction of

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ammonia (up to 93%); the residual concentrations no longer inhibit nitrification (18). However, this method has two drawbacks; the cost of chemicals for pH adjustment and the problem of freezing as the air and the leachate temperature approach 0 ◦ C (2). Ion Exchange Ion exchange is an adsorption process that uses appropriate (usually synthetic organic) resins as media to remove charged contaminants from wastewater. Ion exchange is commonly used to remove heavy metals from relatively low-concentration waste streams. A key advantage of the ion exchange process is that it allows recovery and reuse of the removed metals. Ion exchange can also be designed to be selective for certain metals and can effectively remove them from wastewater that contains high concentrations of ‘‘background’’ metals, such as iron, magnesium, and/or calcium. A specific disadvantage of this treatment method is that the resins used are subjected to fouling by oils or other natural (high MW) polymers. However, the use of ion exchange to treat leachates is limited, by the high operating costs of this method, including the cost of the exchange media, the necessary chemical reagents used as regenerants, and the regenerants disposal costs. As a result, this method is appropriate for the supplementary removal of metals, as posttreatment final polishing step, that results in low residual ion concentrations of less than 1 mg/L (2,8). Membrane Filtration Membrane filtration systems employ a semipermeable polymeric membrane and a pressure differential to separate constituents of different size (from microparticles down to soluble ions) from an aqueous phase. Nanofiltration, ultrafiltration, and reverse osmosis are the most commonly used membrane filtration processes. Ultrafiltration uses a semipermeable microporous membrane, through which the wastewater is passed under pressure. Water and low molecular weight solutes, such as salts and surfactants, pass through the membrane and can be removed as permeate. Emulsified oils and suspended solids are rejected by the membrane and are removed with part of the wastewater as a more concentrated liquid. The concentrate is usually recirculated through the membrane unit, until the permeate flow drops substantially. Ultrafiltration is commonly used for removing substances whose molecular weights are greater than 500, including suspended solids, oil and grease, large organic molecules, and complexed heavy metals. Ultrafiltration is commonly used, when the solute molecules are greater than 10 times the size of the solvent molecules (usually water) and less than 0.5 µm. Reverse osmosis is a separation process that uses selective semipermeable membranes to remove dissolved solids, such as metal salts, from water. The respective membranes are more permeable to water than to contaminants or impurities. The wastewater is forced through the membrane at a pressure that exceeds the osmotic pressure caused by the dissolved solids. Molecules of water pass through the membrane as permeate, and the contaminants are rejected along the surface of the membrane and

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LANDFILL LEACHATES: PART 2: TREATMENT

exit as a concentrated stream. The concentrate (rejection) flow from a reverse osmosis system ranges from 10–50% of the feed flow; the concentrations of dissolved solids and of contaminants within this stream approach at least 10 times that of the feed (raw) wastewater. The percentage of permeate that passes through the membrane is a function of operating pressure, membrane type, and concentration of the contaminants in the feed. Cellulose acetate, aromatic polyamide, and thin-film composites are commonly used membrane materials. Membrane pore sizes for a typical reverse osmosis system range from 0.0005–0.002 µm, and pressures of 20–30 bars are usually required. Therefore, reverse osmosis feed water must be very low in turbidity to avoid direct blocking of the membrane. As a result, pretreatment of landfill wastewater prior to reverse osmosis may be necessary, including chemical addition and clarification or cartridge filtration (using 5-µm filters), to remove suspended particulates from the influent and protect pumps and membranes. Carbon adsorption is also recommended as a pretreatment for membranes sensitive to chlorine. Biofouling can be prevented by chlorination and dechlorination of the feed water. To maintain sufficient solubility of metals, such as calcium, magnesium, and iron, and avoid the formation of precipitates, the pH should be appropriately adjusted with acid. Aside from pH adjustment, chemical reagents may also be used as bactericides, as well as for dechlorination (2). Several reports on the use of ultrafiltration and especially on the application of reverse osmosis membranes for treating leachates showed that these methods are best used, following biological pretreatment, or for treating ‘‘older’’ leachates (19). In addition, during the treatment of leachates by membranes, compounds whose MW was lower than 200 were not rejected, in comparison with those having a MW higher than 200. The performance of reverse osmosis membranes may be further optimized, after careful adjustment of pH to values around 8 and by using polythylamine membranes that results in almost 94% reduction of total organic carbon (TOC) (8). However, severe membrane fouling was experienced in several cases, suggesting that reverse osmosis is most effective as a posttreatment step (following biological treatment) for removing residual COD and dissolved solids. Trebouet et al. (20) have shown that high pollutant removal can also be achieved by nanofiltration, especially for ‘‘older’’ leachates. Nanofiltration can be run at lower pressures than reverse osmosis, hence presenting lower operating costs and less membrane fouling. Recirculation of Leachates into the Landfill The recirculation of leachates is the redistribution of leachates that have been collected at the bottom of a landfill back to the top of it. The recirculation can usually be performed by spraying leachates onto the exposed surface of the landfill or by distribution through perforated pipes, located just beneath the surface of the landfill. As the recirculated leachate trickles downward through the fill, the disposed of solid waste materials in the landfill become an appropriate medium for developing anaerobic

microorganisms, and as a result, an anaerobic treatment process is initiated. Therefore, the landfill becomes an uncontrolled anaerobic digester and the biodegradable organics in the leachates are initially converted to volatile fatty acids and then to methane. Under these conditions, an initially low BOD5 leachate, which is similar in composition to that of ‘‘older’’ leachates, may be produced during periods up to 18 months of recirculation. It has been suggested that the recirculation of leachates can more rapidly develop an active anaerobic bacterial population of methane-forming bacteria within the landfill. The rate of removal of organic compounds may be further enhanced by the addition of excess sewage sludge, acting as additional carbon source, which is produced in biological wastewater treatment plants, as well as by appropriate pH control (8,21). The problems of recirculation usually include the development of odors, the high capital and maintenance cost of the recirculation system for leachates, and the precipitation of carbonates and iron oxides that may clog both the spraying equipment and the surface of the landfill and decrease the percolation rate. However, recirculation of leachates has a number of benefits, such as production of leachates with more uniform properties, the acceleration of overall landfill stabilization, the delay in the starting time for the application of other treatment systems, and reduction in the strength of the treated leachates. Nevertheless, the recirculation of leachates does not finally result in reduction of generated wastewater volumes and cannot provide a sufficient treatment process for leachates, because the treated leachates may have relatively high COD content (higher than 3000 mg/L), as well as high ammonia concentrations (2). In general, physical–chemical treatment methods are considered an effective means for treating leachates, which contain organic compounds of MW less than 500 and a BOD/COD ratio lower than 0.1. This is particularly important for the use of chemical coagulation and activated carbon adsorption because these methods are very sensitive to the MW distribution of organics. Chemical oxidation may be used for removing dissolved metals (mainly iron), but has little effect on COD, when applied separately. Air stripping can be highly effective in removing ammonia, but at highly alkaline pH. Finally, membrane separation processes may have some potential for the treating leachates, but they are subject to membrane fouling. In conclusion, neither physical–chemical treatment alone, nor biological methods may be able to treat leachates completely, when applied separately. As a result, the integration of several treatment processes is required to produce an effluent of acceptable quality; the selection of the most appropriate treatment train is very significant for an integrated leachate management system. SELECTING THE APPROPRIATE COMBINATION OF TREATMENT PROCESSES The characteristics of leachates may vary from place to place, as well as with time; thus, the construction of an appropriate treatment scheme is a difficult

LANDFILL LEACHATES: PART 2: TREATMENT

task, requiring the development/design of appropriate processes, consisting of the following subsequent steps: the first step involves assessment of leachate quality, an estimate of the (seasonal) quantity, and the type of treatment techniques available; the second step includes selection of the optimum biological treatment method for removing major pollutants from the leachate, followed by selection of the applicable polishing (final) stages as a third step (4,22). Selecting Treatment Techniques Based on Leachate Quality During this phase, the influent leachate quality is assessed, and the corresponding effluent characteristics are determined, based on the estimates of BOD5 , COD, nitrogen (as ammonia, i.e., NH3 -N), phosphorous, and metal content. Additional data that might be used in this phase include the concentration of volatile fatty acids, TOC, and total suspended solids. The most important parameter for the preliminary screening of treatment techniques for leachates is the BOD5 /COD ratio. Leachates from a relatively new landfill have high COD values, usually higher than 10,000 mg/L; low NH3 -N content; and BOD5 /COD ratios ranging from 0.4–0.8. These leachates are representative of ‘‘young’’ leachates, containing high amounts of easily biodegradable organic substances, which are amenable to both aerobic and anaerobic biological treatment. However, physical–chemical treatment may not be an appropriate method for such leachates due to the presence of low molecular weight volatile fatty acids, prevailing over higher molecular weight compounds. ‘‘Old’’ (mature) leachates are characterized by lower COD values, usually less than 3000 mg/L; higher NH3 N concentrations, due to the anaerobic decomposition of organic nitrogen content within the landfill; and BOD5 /COD ratios lower than 0.4. In this case, aerobic biological treatment is required because it can provide extensive nitrogen removal through nitrification, in addition to removing organic matter (COD values). However, for BOD5 /COD ratios lower than 0.1, the remaining organics consist mainly of nonbiodegradable materials, and physical–chemical treatment becomes a more attractive option. A second parameter, which can be used as an indicator for the preliminary selection of a treatment process, is the molecular weight partitioning of the organic content. In this case, biological treatment would be an effective technique for leachates that contain organic substances of MW lower than 500, whereas physical–chemical treatment is favored for removing organics, of molecular weights higher than 1000. However, this parameter is not easily measured in samples of leachates, and it is not used as a general index. Other less important parameters may include the BOD5 /NH3 -N ratio and the metal content, which are usually considered in the following steps of the design procedure. Selecting the Appropriate Biological Treatment Method To treat ‘‘young’’ leachates that have high BOD5 /COD ratios and high amounts of low MW (i.e., easily biodegradable) organics, biological treatment is the most efficient

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technique for reducing organic (carbonaceous) matter. The first option in this case is the examination of recirculation as a cost-effective method for preliminary reduction of leachate strength. However, when recirculation is not feasible, then aerobic or anaerobic biological treatment should be used. Anaerobic biological treatment is the most appropriate method for handling leachates of high BOD5 /COD ratio, low NH3 -N content, high temperature, and high VFA content; the optimum process in this case is the application of anaerobic filters. The effluents from anaerobic processes would be similar to ‘‘old’’ leachates that have COD values between 1000 to 3000 mg/L and require an additional posttreatment step to reduce residual organics. Aerobic biological treatment is a common method, which may be applied either for anaerobically pretreated or for raw leachates. In addition, biological denitrification may also be included in the aerobic process for effective removal of nitrates. Two significant problems are associated with the aerobic treatment of leachates; the first is the phosphorous deficiency in leachates that requires supplementary phosphorous, and the second is the high ammonia content that may inhibit the nitrification capacity of the system. In this case, the initial high ammonia concentration should be reduced by preliminary air stripping of the leachate. Additional problems include possible high metal content; long sludge retention time (up to 30 days), hence requiring the use of larger basins; foaming of leachates during aeration; and the potential for metal precipitate formation that affects the operation of aeration systems. The effluents from the aerobic processes usually have low BOD5 values, lower than 100 mg/L, but rather high COD values, up to 1500 mg/L, that require subsequent physical–chemical treatment (22). Selection of Physical–Chemical Treatment Method The third step in the treatment design procedure for leachates is selecting a process for posttreating the effluent. ‘‘Old’’ and biologically pretreated leachates may contain nonbiodegradable high MW organic compounds, particulates, and metal ions. Residual organics may be removed by chemical oxidation, using hydrogen peroxide and/or ozone, but these techniques are expensive, due to the high dosages of the necessary chemicals. Coagulation is an alternative method, which is effective for removing higher MW organics that represent about 50% of the residual organic matter. Activated carbon adsorption is also a viable process for removing lower MW organics. Membrane separations may be used for removing organics, but these methods are best applicable as final polishing stages, due to high costs and problems of membrane fouling. Chemical precipitation using lime or chemical oxidation followed by sedimentation and/or filtration has been suggested to decrease the metal content of leachates. Similar techniques can be also used to reduce particulates and metals content, such as coagulation followed by sedimentation/filtration, chemical oxidation, membrane separation, or ion exchange. Several alternative units for treating leachates are operating currently worldwide, consisting of a combination of subsequently applicable treatment steps. The

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LANDFILL LEACHATES: PART 2: TREATMENT

Sodium hydroxide

Anionic polymer

Weak hydrochloric acid and phosphate addition

02 Leachate Pumping stations

Leachate and gas condensate

Equalization tanks

Chemical Flocculation addition

01

Primary clarifier

Neutralization 04

Sludge

Heater Sludge thickener

Filtrate

Plate-and-frame press

Multimedia filters

Filter backwash

Wasted sludge

Sludge disposal in landfill

Effluent discharge to surface water

Sludge

Aerobic towers

Anaerobic towers

Chemical Flocculation addition

Secondary clarifier

Equalization tanks 07

Ferric chloride, sodium hypochlorite, and weak hydrochloric acid

Sampling location

Figure 1. Flow sheet of a typical leachate treatment plant (2).

U.S. EPA (2) has evaluated the performance of a number of treatment systems and representative results are presented in the following. A typical treatment system for leachates from a sanitary landfill is shown in Fig. 1 (2). The system employs equalization tanks, coagulation/sedimentation, pH adjustment, biological treatment consisting of anaerobic towers followed by aerobic ones, coagulation/sedimentation, and multimedia filtration; the

sludge treatment unit includes a sludge thickener and a plate-and-frame filter press. The results of the system operation are presented in Table 2. As shown in Table 2, the biological treatment unit experienced good overall removal for TOC (93%), COD (90.9%), and NH3 -N (99.1%). The biological unit operation alone did not demonstrate high removals of BOD5 (10.2%), TSS (9.3%), or for various metals (showing in general less

Table 2. Performance Data of a Typical Leachate Treatment Plant Biological Unit

Entire System

Parameter

Influent, Effluent, Removal, Influent, Effluent, Removal, mg/L mg/L % mg/L mg/L %

BOD5 Total suspended solids NH3 -N COD NO3 -N TDS TOC Total phenols Barium Boron Chromium Strontium Zinc

39.2 11.8 135 1742 1.5 5960 758 0.2 0.010 3 0.012 NDa NDa

a

ND = not detected.

35.2 10.7 1.1 159.4 130.5 5181 52.8 0.05 0.022 2.9 NDa 0.082 0.012

10.2 9.3 99.1 90.9 0.0 13.1 93.0 72.5 0.0 8.9 — 0.0 0.0

991 532.8 193.3 4028 0.693 5012 1316 1.2 2.43 4.33 0.036 2.9 0.144

35.2 10.7 1.1 159.4 130.5 5181 52.8 0.05 0.022 2.9 NDa 0.082 0.012

96.5 98.0 99.4 96.0 0.0 0.0 96.0 95.9 99.1 32.5 70.3 97.2 91.6

LANDFILL LEACHATES: PART 2: TREATMENT Leachate and maintenance facility sanitary wastewater

Holding tank

NaOH Lime

Equalization/ stripper

Grit chamber

Polymer and coagulant

Mix tank

Sulfuric acid

Primary clarifier

Flocculator

711

Phosphoric acid

Surge tank 03

01 Sludge

Plant sanitary wastewater Overflow, filtrate, and supernate Sequencing batch reactor (SBR)

Sodium hypochlorite

Sludge thickeners Perlite

06 Post-SBR tank

Granular activated carbon

Granular activated carbon

Multimedia filter

Chlorine contact chamber

Sludge holding tank

04 Sulfuric acid Sludge

Sampling location

Effluent dicharge to surface water

Plate -and- frame press

Sludge disposal in landfill

Figure 2. Flow diagram of an alternative leachate treatment plant using an SBR biological treatment unit.

than 10% removal) because the pollutants were generally not present in the biological treatment unit influent at treatable levels. The influent BOD5 in the treatment unit was rather low (39.2 mg/L), TSS was 11.8 mg/L, and most metals were not at detectable levels, even though the raw wastewater at this facility exhibited an initial BOD5 of 991 mg/L, TSS of 532.8 mg/L, and several metals at treatable levels. The biological treatment unit influent had low concentrations of pollutants because this facility employed large aerated equalization tanks and a chemical precipitation system prior to biological treatment. The equalization tanks had a retention time of approximately 15 days and were followed by a chemical precipitation system using sodium hydroxide. Due to the long retention time and the aeration of wastewater, significant biological activity also occurred in these tanks. The resulting insoluble pollutants were removed in the primary clarifier prior to entering the biological towers. The entire treatment system showed good removals for BOD5 , TSS, NH3 -N, COD, TOC, and total phenols. Most metals had good percentage removals or were removed to nondetectable levels. Another leachate treatment plant was evaluated by the EPA, including ammonia removal, hydroxide precipitation, biological treatment using a sequencing batch reactor, granular activated carbon adsorption, and multimedia filtration. A flow diagram of the landfill

wastewater treatment system is presented in Fig. 2 (2). The wastewater treatment process used at this (nonhazardous) facility was primarily treated landfill generated wastewater and a small amount of sanitary wastewater flow from the on-site maintenance facility. A summary of percentage removal data collected for the biological treatment unit operation (SBR) and for the entire treatment system is presented in Table 3. As shown in this table, the SBR treatment unit showed moderate overall removals for TOC (43.45%), COD (24.7%), and BOD5 (48.7%). Improved removal efficiencies were observed for TSS (82.9%), total phenols (74.2%), and NH3 -N (80.7%). Metals, such as barium, chromium, and zinc had low removal efficiencies. However, these metals in the influent of the biological treatment system were measured at low concentrations, often close to the detection limit. Other metals also had poor removal efficiencies, including boron and silicon. The entire treatment system showed good removals for BOD5 , TSS, NH3 -N, COD, TOC, and total phenols. Each of the metal parameters also experienced good removal rates through the treatment system. An alternative system for treating leachates from a nonhazardous facility was also constructed; it employed a two-stage reverse osmosis system and a multimedia filter. The flow diagram of this unit is shown in Fig. 3, and the corresponding performance data are given in Table 4.

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LANDFILL LEACHATES: PART 2: TREATMENT Table 3. The Performance Data of an Alternative Leachate Treatment System that Employed an SBR Biological Treatment Unit Biological Unit Parameter

Influent, mg/L

Effluent, mg/L

BOD5 TSS NH3 -N COD NO3 -N TDS TOC Total Phenols Barium Boron Chromium Strontium Zinc

232.6 59.6 134.8 635.0 14.4 4024 212.6 0.2 0.019 2.842 0.010 0.193 0.025

119.3 10.2 26.04 478.2 87.8 3987 120.4 0.052 0.032 2.483 0.011 0.237 0.058

a

Entire System

Removal, % 48.7 82.9 80.7 24.7 0.0 0.9 43.4 74.2 0.0 12.6 0.0 0.0 0.0

Influent, mg/L

Effluent, mg/L

1088 93.4 295.9 2932 0.494 6232 1098 0.940 0.283 6.7 0.0906 1.935 0.494

201 NDa 12.06 251 87 3834 82 NDa 0.0426 2.334 NDa 0.249 0.027

Removal, % 81.5 95.7 95.9 91.4 0.0 38.5 92.5 94.7 85.0 65.2 87.7 87.1 94.5

ND = not detected.

HCl (in-line addition) Raw leachate

Equalization tank

Multimedia filter

Cartridge filter

01

Feed pump

Second reverse osmosis unit

Equalization tank

to POTW

First reverse osmosis unit

03

02

Double pass permeate

Single pass permeate

Concentrate in landfill

Figure 3. Flow diagram of a leachate treatment plant using a two-stage reverse osmosis unit.

As shown in Table 4, the single-pass reverse osmosis treatment system demonstrated good overall removals for a number of parameters, including TSS, TOC, BOD5 , TDS, and COD. Total phenol and NH3 -N% removals were observed at 75.1 and 76.7%, respectively. Metals with quantitative percentage removals included arsenic (87.4%), boron (54.1%), and strontium (92.9%). The additional polishing reverse osmosis unit caused the removal efficiency of most parameters to increase further. These parameters include BOD5 , NH3 -N, COD, TDS, TOC, and total phenols. The percentage removal for boron also increased from 54.1% in the single-pass reverse osmosis system up to 94.4% in the two-stage reverse osmosis treatment system. In general, selecting the most appropriate treatment train should be based on leachate characteristics and cost estimates. For ‘‘young’’ leachates, which have BOD5 /COD

ratios higher than 0.4, biological treatment will prevail in the overall treatment system. For the highest BOD5 /COD ratios (in the range of 0.6 to 0.8), recirculation of leachates should be included as a preliminary treatment stage, followed by anaerobic or aerobic treatment. For BOD5 /COD ratios lower than 0.4, aerobic biological treatment becomes the most important method, especially when nitrification is required; in all cases, physical–chemical treatment is necessary, as a polishing step for biologically pretreated effluents. In some cases, physical–chemical treatment may precede aerobic treatment, aiming, for example, to reduce ammonia or metals. When the BOD5 /COD ratio is lower than 0.1, physical–chemical methods are the most appropriate, and aerobic biological treatment may be additionally used for nitrification/denitrification. During the design period, it is very important to consider the variation of

LANDFILL LEACHATES: PART 2: TREATMENT

713

Table 4. Performance Data of a Leachate Treatment Unit, Including a Two-Stage Reverse Osmosis System Single-stage Reverse Osmosisa System Parameter

Influent, mg/L

Effluent, mg/L

BOD5 TSS NH3 -N COD NO3 -N TDS TOC Total phenols Barium Boron Chromium Strontium Zinc

1182 171.8 58.48 1526 1.3 2478 642.6 1.26 0.28 1.808 ND 1.406 ND

54 ND 13.6 72.2 0.666 116.6 25 0.316 0.006 0.83 ND ND ND

a

Removal, % 95.4 97.7 76.7 95.3 48.8 95.3 96.1 75.0 98.0 54.1 — 92.9 —

Entire Treatment Systema Influent, mg/L

Effluent, mg/L

1182 171.8 58.48 1526 1.3 2478 642.6 1.26 0.28 1.808 ND 1.406 ND

5.4 ND 0.608 11.4 0.502 ND ND 0.063 0.001 0.10 ND ND ND

Removal, % 99.5 97.7 99.0 99.3 61.4 99.6 98.4 95.0 99.5 94.4 — 92.9 —

ND = not detected.

leachates properties over time that requires particular attention to the treatment plant design to achieve operating flexibility to cope with varying influent characteristics.

BIBLIOGRAPHY 1. Lema, J.M., Mendez, R., and Blazquez, R. (1988). Characteristics of landfill leachates and alternatives for their treatment: A review. Water Air Soil Pollut. 40: 223–250. 2. Environmental Protection Agency. (2000). Development document for final effluent limitations guidelines and standards for the landfills point source category. EPA-821-R99-019, Washington, DC. 3. Ehrig, H.J. (1984). Treatment of sanitary landfill leachate: biological treatment. Waste Manage. Res. 2: 131–152. 4. Qasim, S.R. and Chiang, W. (1994). Sanitary Landfill Leachate: Generation, Control and Treatment. Technomic, Lancaster, PA. 5. Forgie, D.J.L. (1988). Selection of the most appropriate leachate treatment methods. Part 1. A review of potential biological leachate treatment methods. Water Pollut. Res. J. Can. 20(2): 308–328. 6. Kelly, H.G. (1987). Pilot testing for combined treatment of leachate from a domestic waste landfill site. J. Water Pollut. Control Fed. 59(5): 254–261. 7. Diamadopoulos, E., Samaras, P., Dabou, X., and Sakellaropoulos, G.P. (1997). Combined treatment of landfill leachate and domestic sewage in a sequencing batch reactor. Water Sci. Technol. 36(2–3): 61–68. 8. Forgie, D.J.L. (1988). Selection of the most appropriate leachate treatment methods. Part 2. A review of recirculation, irrigation and potential physical-chemical treatment methods. Water Pollut. Res. J. Can. 20(2): 329–340. 9. Amokrane, A., Comel, C., and Veron, J. (1997). Landfill leachates pre-treatment by coagulation flocculation. Water Res. 31(11): 2775–2782. 10. Sletten, R.S., Benjamin, M.M., Horng, J.J., and Ferguson, J.F. (1995). Physical–chemical treatment of landfill leachate for metals removal. Water. Res. 29(10): 2376–2386.

11. Diamadopoulos, E. (1994). Characterization and treatment of recirculation-stabilized leachate. Water Res. 28(12): 2439–2445. 12. Loizidou, M., Kapetanios, E.G., and Papadopoulos, A. (1992). Assessment of leachate characteristics and its treatability. Fresenius Environ. Bull. 1: 748–753. 13. Tatsi, A., Zouboulis, A.I., Matis, K.A., and Samaras, P. (2003). Coagulation-filtration pretreatment of sanitary landfill leachates. Chemosphere 53(7): 737–744. 14. Slater, C.S., Uchrin, C.G., and Ahlert, R.C. (1985). Ultrafiltration processes for the characterization and separation of landfill leachates. J. Environ. Sci. Health A 20(1): 97– 111. 15. Wenzel, A., Gahr, A., and Niessner, R. (1999). TOC removal and degradation of pollutants in leachate using a thin-film photoreactor. Water Res. 33(4): 937–946. 16. Marttinen, S.K. et al. (2002). Screening of physical–chemical methods for the removal of organic material, nitrogen and toxicity from low strength landfill leachates. Chemosphere 46: 851–858. 17. Beaman, M.S., Lambert, S.D., Graham, N.J.D., and Anderson, R. (1998). Role of ozone and recirculation in the stabilization of landfills and leachates. Ozone Sci. Eng. 20: 121– 132. 18. Cheung, K.C., Chu, L.M., and Wong, M.H. (1997). Ammonia stripping as a pretreatment for landfill leachate. Water Air Soil Pollut. 94: 209–221. 19. Peters, T. (1999). Past and future of membrane filtration for the purification of landfill leachate. Proc. 7th Int. Waste Manage. Landfill Symp., Cagliari, Italy, pp. 335–344. 20. Trebouet, D. et al. (1999). Effect of operating conditions on the nanofiltration of landfill leachates: Pilot scale studies. Environ. Technol. 20: 587–596. 21. Morris, J.W.F., Vasuki, N.C., Baker, J.A., and Pendleton, C.H. (2003). Findings from long-term monitoring studies at MSW landfill facilities with leachate recirculation. Waste Manage. 23(7): 653–666. 22. Forgie, D.J.L. (1988). Selection of the most appropriate leachate treatment methods. Part 3. A decision model for the treatment train selection. Water Pollut. Res. J. Can. 20(2): 341–355.

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MACROPHYTES AS BIOMONITORS OF POLYCHLORINATED BIPHENYLS

MACROPHYTES AS BIOMONITORS OF POLYCHLORINATED BIPHENYLS KARINA S.B. MIGLIORANZA MIRTA L. MENONE Universidad Nacional de Mar del Plata Mar del Plata, Argentina and Consejo Nacional de Investigaciones Cient´ificas y T´ecnicas (CONICET), Buenos Aires, Argentina ´ DE MORENO JULIA E. AIZPUN ´ VICTOR J. MORENO

Universidad Nacional de Mar del Plata, Mar del Plata, Argentina

INTRODUCTION The freshwater, estuarine, and marine ecosystems around the world are now being increasingly subjected to greater stress from various human activities. As a result of several contaminants, significant changes in freshwater and marine plant communities have occurred (1). Polychlorinated biphenyls (PCBs) are one of the more ubiquitous and toxic organic pollutants. Altogether, some 209 different individual PCB compounds exist, although only about 130 of these are found in commercial mixtures. PCBs have a number of physical and chemical characteristics that have contributed to their persistence in the environment, such as low aqueous solubility, resistance to oxidation and hydrolysis, and high volatility (2). The use of PCBs is now restricted to nondispersive systems (e.g., as dielectric fluids in condensers and transformers), but until few years ago, they were important constituents of a number of industrial chemical formulations, including paints, inks, plastics, pesticides, and copying papers. In most industrialized countries, the use of PCBs has been banned since the late 1970s, resulting in a decline in PCB releases in the environment. However, in several aquatic ecosystems, the biota contamination remained relatively stable since the mid-1980s (3–5). The concern about the presence and effects of PCBs on the biota and, ultimately, on human health, have mobilized worldwide research in order to find remediating PCB levels in the environment and therefore prevent possible future undesirable consequences. Toxic effects in animals include reproductive impairment, mutagenesis, carcinogenesis, and teratogenesis (6). Recently, several studies have demonstrated the ability of PCBs to induce oxidative stress (a cellular situation characterized by an elevation in the steady-state concentration of reactive oxygen species) (7) in fishes, birds, reptiles (8), and dinoflagellates (9); but in plants, a lack of studies about this subject exist. PCBs are very hydrophobic contaminants and preferentially adsorb onto sediment particles. Particle size distribution and organic carbon content seem to be important factors in determining the

extent of partitioning of PCBs to natural sediment (10). In particular, the high octanol/water partition coefficient (Kow) indicates that PCBs have a high affinity for suspended solids, especially those rich in organic carbon. Mobilization of PCBs from sediments is one of the causes of the persistence of these contaminants (11,12), because sediments were, in the past, the most important sink of these pollutants. THE ROLE OF MACROPHYTES IN THE AQUATIC ECOSYSTEM Macrophytes include various emergent, submergent, and floating leaved species. Together with algae, periphyton, and phytoplankton, macrophytes form the base of most aquatic food chains. Macrophytes are important in nutrient cycling and respond rapidly to water quality changes. It is well known that aquatic vascular plants can enhance a freshwater ecosystem in many ways, such as by improving water clarity, contributing oxygen via photosynthesis, and providing spawning sites and protection for fish, mollusks, crustaceans, and other invertebrate species (13,14). Plants interact with their environment through processes that include contaminant bioconcentration and excretion, shading, and organic matter production and decomposition. As a result of these interactions, aquatic plants may significantly affect water and sediment quality (15). As plants are the base of most food chains, they will experience the effects of toxic compounds released into the aquatic media sooner than will organisms occupying higher trophic levels. In this respect, plants may be able to act as an ‘‘early warning signal’’ of impending contaminant impacts on other trophic levels of aquatic environments. Contaminants present in aquatic macrophytes may be transferred to higher levels of the food chain, when consumed as live plants by herbivores or as detritus by detritivores (16,17). Aquatic macrophytes in the littoral zones of lakes have two fundamental properties, which make them attractive as limnological indicators. First, they react slowly and progressively to changes in nutrient conditions, in contrast to bacteria and microalgae, i.e., over several years. Macrophytes, therefore, function as integrators of environmental conditions to which they are subjected, and thus can be used as long-term indicators with high spatial resolution. Second, the littoral zone may experience patterns of nutrients and pollutant concentrations (18,19), caused by natural or artificial inflows as well as by diffuse and nonpoint sources. The latter are difficult to localize and quantify, but even chemical analysis of point sources is often neglected in limnological routine work (20). Vascular plants like bulrushes (Schoenoplectus spp.), duckweeds (Lemna spp.), and pickerel weed (Pontederia spp.) are the more common species used in North American wetlands treatment systems to improve water quality in a variety of ways, including binding soil and reducing the resuspension of muds (15). Bioaccumulation by plants can remove substantial quantities of potential toxicants (like organochlorines) and nutrients of water entering

MACROPHYTES AS BIOMONITORS OF POLYCHLORINATED BIPHENYLS

and passing through wetlands. Thus, Menone et al. (21) calculated that about 75 µg of Heptachlor-epoxide per square meter of cordgrass marsh exists, or over the 32 km2 of cordgrass marsh, roots of Spartina densiflora concentrates 2 Kg of this contaminant. Total PCBs in the same estuary can be estimated as 22 µg per square meter or 0.7 Kg over the 32 km2 of cordgrass marsh. EXPOSURE AND UPTAKE OF MACROPHYTES TO CONTAMINANTS The uptake of contaminants depends on the physicochemical characteristics of the compounds and on the life form of the macrophyte (floating-leaved, free-floating, well-rooted, or rootless species). Free-floating plants, such as Lemna spp., Eichornia spp., and Pistia spp., take up contaminants from the water by roots and/or leaves. Likewise, the rootless Ceratophyllum spp. takes up contaminants mainly through its finely divided leaves. The situation of species with a well-developed rootrhizome system and totally submersed foliage, such as Myriophyllum spp., Potamogeton spp., and Vallisneria spp. species, is much more complex. Thus, these submersed aquatic macrophytes grow at the interface of two distinct environments, being their leaves exposed to the water column while their roots are in contact to the sediment. Both media (the water column and the sediment pore water) are potential sources of uptake of contamination (22,23). Indeed, the rooted aquatic macrophyte Hydrilla verticillata—common throughout the southern United States—has been shown to take up contaminants from both the sediment pore water and the overlying water column. It also has the ability to translocate sediment-incorporated contaminants from the roots into the vegetative portion of the plant (24). Other species, such as Schoenoplectus californicus, Spartina densiflora (mainly from the Southern Hemisphere), and Spartina alterniflora (mainly from the Northern Hemisphere), have the root-rhizome system well developed, but have not totally submersed foliage. The first two macrophytes form the bulk of the emergent vegetation in many of the shallow lakes and estuarine areas of the Southern America regions. The pollutant sequestration in these macrophytes occur across many interfaces, sediment, water, and air (21,25,26). Several environmental factors, including water and sediment pH, water current, sediment texture, organic carbon, and mineral composition, are known to influence contaminant adsorption by sediments and macrophytes growth, so they may contribute to overall site quality. BIOMONITORING During the past, direct chemical measurement of the concentrations of toxic and hazard substances have been carried out for the evaluation of water quality in contaminated natural environments. However, for environmental management, contaminant concentrations do not necessarily account for, or enable prediction of, the impairment of biota. If the objective is to

715

monitor and improve environmental quality of the ecosystems, biotic measurements of contamination are more useful than only measures of water or sediment contaminant concentrations. Furthermore, the cost of evaluation and measurement of persistent and toxic organochlorines, such as PCBs, is substantial. Therefore, for ecologic and economic reasons, it is important to develop biological monitoring that is useful and valuable. The use of biomonitoring to assess and control discharges of toxic chemicals into the environment has been promoted as being a desirable alternative to more expensive, less realistic, and time-consuming chemical analysis (27). Thus, the presence of indicator organisms (biomonitors) provides a measure of cumulative exposure to contaminants over time and avoids the need for frequent sampling. The use of biomonitors in situ to identify and quantify toxicants in an environment is referred to as biomonitoring (28). This technique takes advantage of the ability of organisms to accumulate contaminants in their tissues through bioaccumulation and bioconcentration. The primary objective of biomonitoring investigations is to assess the quality of water in an area by relating observed responses of organisms that live within a suspected polluted site to the concentrations of contaminants detected within their tissues. An additional and very significant advantage of biomonitoring is that the bioaccumulated sublethal levels of contaminants within the tissues of organisms indicate the net amount of contaminants that have been integrated over a period of time (27). The response of biota to pollution stress can be observed at the ecosystem, community, population, individual, and suborganismal levels of organization (29). Environmental assessments may be made by establishing quantitative relationships between (a) concentrations of pollutants that are accumulated within the tissues of organisms residing within a particular area and (b) manifest biological effects (30). In general, three historical stages exist in the biomonitoring with plants, based on the use of different parameters: 1. various physiological, morphological, and community parameters (pigment content, photosynthetic activity, diversity indices, biomass, etc.) 2. environmental concentrations of pollutants in plant tissues 3. early warning systems or biomarkers for assessing contaminant exposure and effects (histological, biochemical, or genetic) The presence of certain toxicants may induce a physiological response in an organism, often involving a heightened production of enzymes that are capable of metabolizing and/or degrading the toxicant in question. In this way, the quantity and activity of such xenobioticmetabolizing enzymes may be used as indicators of the bioavailability of a specific contaminant in the environment. An increasing number of studies have used biochemical as well as physiological endpoints for assessing toxic effects on plants. These effects are

716

MACROPHYTES AS BIOMONITORS OF POLYCHLORINATED BIPHENYLS

often more sensitive, but their environmental relevance and their relationship to the impact of toxic chemicals on biomass are not known. However, they can be combined with other parameters of chemical exposure to predict the ecological consequences of chemical-specific contamination (31). For a biomarker to be applied in the field, it should be correlated with a significant effect, such as survival, growth, or reproduction. Although many biomarkers have been validated for use in evaluating animal health and exposure to toxic substances, both in the laboratory and in natural ecosystems, very few biomarkers have been validated and used to assess plant exposure and health effects under field conditions (31). Padinha et al. (32) has shown that variations in some indices of physiological stress, like thiolic protein concentrations in Spartina maritima, could be used as a tool to monitor contamination by heavy metals. On the other hand, Wall et al. (33) have found no adverse influences of PCBs in terms of peroxidase activity (POD), glutathione concentration (tGSH), photosynthesis, and transpiration on S. alterniflora from a Superfund site contaminated with 46.0 ± 52.7 µg g−1 dry weight in sediments. Therefore, the utility of POD as well as other detoxication enzymes as biomarkers of PCB stress deserves attention but needs further study. In addition, laboratory research can complement and contribute to the best understanding of data from field studies (biomonitoring). Among laboratory experiments, we can mention: — Toxicity tests: Toxicity tests involve exposing a welldefined test organism to a dilution series of a suspected toxicant under controlled laboratory conditions. The goal of toxicity tests is to correlate the level of toxicants to observed organismal responses. Of the multitude of organismal responses that could be observed as endpoints in toxicity test, the one most frequently used is survivorship. In ecotoxicology, survivorship is usually expressed as LD50 , which is the dosage of the suspected toxicant that is seen to cause mortality in half of the individuals tested within a specified time period. — Bioassays: Bioassays are often used to assess the toxic effects of mixtures of compounds on biota by exposing test organisms to naturally contaminated water or sediment samples. When a pollutant enters an aquatic environment, it is expected that its initial effect on an exposed organism will be a suborganismal one—either biochemical or genetic. In this manner, biochemical and genetic indicators may be able to detect the presence of minimum contaminant concentrations compared with the levels of toxicants required to elicit a response at the level of the entire organism (e.g., death) (28). MACROPHYTES AS BIOMONITORS Plant biomonitors provide an integrated description of pollution within an ecosystem (34). Freshwater species were used as sentinels of contaminant stress for many

years, such as in the biomonitoring study of Wang and Williams (35), who used Lemna minor to examine the phytotoxicity of industrial effluents. Macrophytes have also proved to be useful for assessing organochlorine contamination in laboratory experiments (36,37) and in the field (27,28). The organochlorine bioaccumulation in submersed macrophytes may be very high; macrophytes may be three to four times more contaminated than sediment, and 6000 to 9000 times more contaminated than the water (38). Under field conditions, Schoenoplectus californicus has already shown potential as a biomonitor of organochlorine pesticides. Moreover, this species has demonstrated the ability to function as phytoremediator—plant use for the remediation of contaminated environments—of these pesticides, because, in combination with other aquatic biota, about 40% of the more hydrophobic pesticides have been retained in the lake environment, leading to their lower release through the effluent creek (25,26,39). Spartina densiflora grows abundantly and contributes significantly to the primary productivity of the estuarine ecosystems on the Southern America coast. In Mar Chiquita coastal lagoon (Argentina), it can bioaccumulate both organochlorine pesticides and PCBs, and the total amount of these compounds in cordgrass biomass may represent a significant proportion of the total PCBs burden in the estuary system (21). The same species is the dominant plant in Humboldt Bay salt marshes. S. densiflora’s ability to rapidly expand in bare areas has implications for marsh mitigation and restoration activities (40). Other known macrophytes, like Vallisneria americana, var. americana—one of the most abundant macrophyte in the Great Lakes (41,42)—Potamogeton spp., Najas spp., Myriophyllum spp., and Elodea spp., accumulate contaminants within its tissues and have also shown potential as biomonitors of organic contaminants in the field (27,28). MACROPHYTES AND PCBS Laboratory studies indicate that uptake of PCBs from contaminant water or sediments in aquatic macrophytes is expected (43). In addition, considering their limited mobility and their abundance in many aquatic systems, they could function as in situ biomonitors of water contaminants, like PCBs. However, the extent of sequestration in natural populations remains almost unknown. Thus, little impetus has developed for studying the role of submersed aquatic macrophytes in PCBs biogeochemical cycling in freshwater ecosystem, as evidenced by our literature survey, which revealed a lack of field data documenting the incidence of hydrophobic organic contaminants in feral aquatic macrophytes. Exposure to sediment-borne PCBs is particularly important for rooted macrophytes. As a result of the hydrophobicity of these compounds, sediments frequently contain higher concentrations of contaminants than the surrounding water (44,45). Thus, plants grown in contaminated environments typically have higher concentrations of hydrophofobic pollutants in below-ground tissues than

MACROPHYTES AS BIOMONITORS OF POLYCHLORINATED BIPHENYLS

in foliage and, because their mobility within the plant tissues is very limited (23), they tend to accumulate in roots (28). Macrophytes collected from different sites may contain different concentrations of contaminants within their tissues, reflecting sediments, water, and air loads at each location (26,28). In a shallow lake from Argentina, it has been demonstrated that S. californicus accumulate PCBs in direct relationship to the sediment PCB concentrations. Moreover, the higher PCB levels were found in root tissues revealing the high ability of this macrophyte to function as PCB biomonitor (Miglioranza, personal communication). The Bioaccumulation Factor (BF) is the ratio of the concentration in the biota and the concentration in the soil/sediment. It primarily depends on the properties of the soil/sediment and the biota, particularly the ratio of lipid in the biota and the organic carbon content of the soil/sediment (2). Root Bioaccumulation Factors (RBF) have been calculated for Schoenoplectus californicus and Spartina densiflora from Los Padres lake and Mar Chiquita coastal lagoon, respectively. The values in S. californicus ranged between 49 and 600 for different congeners of PCBs, being the highest RBF for the lower chlorinated congeners. In S. densiflora, the RBF ranged between 0.75 and 26.5. These results show the magnitude of bioaccumulation of PCBs in freshwater and estuarine macrophytes species and their importance in the ecology of these environments. The investigations of Butler et al. (46) demonstrated for the first time that plant cells are capable of hydroxylating and glycosylating a chlorinated biphenyl in a manner similar to what has been reported for animals (47). The metabolism of PCBs varies between the plant species and is affected by the substitution pattern and the degree of chlorination (48,49). Wilken et al. (49) analyzed 12 different terrestrial plant species and showed that lower chlorination grade is associated with higher metabolism rates. Recent studies (50,51) have shown the important role of plant cytochrome P450 in metabolism of different toxicants, but have admitted involvement of peroxidases too (52,53). Despite the scarce information about PCB concentrations in macrophytes under field conditions, we can conclude that macrophytes can play a crucial role in the biomonitoring and remotion of these toxic organic compounds from the environment, not only accumulating but also transforming them. For this reason, phytoremediation has been proposed as an alternative or complementary technique to treat sediment polluted by PCBs, but still needs much basic research. BIBLIOGRAPHY 1. Orth, R.J. (1994). Lake Reservoir Manage. 10: 49–52. 2. Connell, D.W. (1997). Lewis Publishers, New York, p. 505. 3. Borgmann, U. and Whittle, D.M. (1992). J. Great Lakes Res. 18: 298–308. 4. Picer, M. and Picer, N. (1995). Water Res. 29: 2707–2719. 5. Stow, C.A., Carpenter, S.R., Eby, L.A., Amrhein, J.F., and Hesselberg, R.J. (1995). Ecol. Appl. 5: 248–260.

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Quality index

40. Kittelson, P.M. and Boyd, M.J. (1997). Estuaries 20: 770–778.

th

P GD

41. Schloesser, D.W., Edsall, T.A., and Manny, B.A. (1985). Can. J. Bot. 63: 1061–1065.

w gro

42. Catling, P.M., Spicer, K.W., Biernacki, M., and LovettDoust, J. (1994). Can. J. Plant Sci. 74: 883–897. 43. Gobas, F.A.P.C., McNeil, E.J., Lovett-Doust, L., and Haffner, G.D. (1991). Environ. Sci. Technol. 25: 924–929.

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ut io

44. Di Toro, D. et al. (1991). Environ. Toxicol. Chem. 10: 1541–1583.

n

fr o m

45. Burton, A. (1992). In: A. Burton (Ed.). Lewis Publishers, Boca Raton, FL, Ch. 8, pp. 167–181.

wa st ew a

46. Butler, J.M., Groeger, A.W., and Fletcher, J.S. (1992). Bull. Environ. Contam. Toxicol. 49: 821–826. 47. Sandermann, H. (1982). In: Environmental Mutagenesis, Carcinogenesis, and Plant Biology. E.J. Klekowski (Ed.). Vol. 1, Praeger, New York, pp. 2–32. 48. Lee, I. and Fletcher, J.S. (1992). Plant Cell Reports 11: 97–100. 49. Wilken, A., Bock, C., Bokern, M., and Harms, H. (1995). Environ. Toxicol. Chem. 14: 2017–2022. 50. Pflugmacher, S., Geissler, K., and Steinberg, C. (1999). Ecotoxicol. Environ. Safety 42: 62–66. ´ M., Schmeiser, H.H., and Frei, E. (2000). Phyto51. Stiborova, chemistry 54: 353–362. 52. Koller, G., Moder, M., and Czihal, K. (2000). Chemosphere 41: 1827–1834. ´ L., Kucerova, ´ P., Macek, T., and Mackova, ´ M. 53. Chroma, (2002). Book of Abstracts, The Second PCB Workshop. I. Holoubek, and I. Holoubkova (Eds.). Masaryk University Brno, p. 153.

WASTEWATER MANAGEMENT FOR DEVELOPING COUNTRIES KONSTANTINOS P. TSAGARAKIS University of Crete Rethymno, Greece

INTRODUCTION A country is characterized as a developing one according to specific economic indicators. The World Bank has linked gross domestic product (GDP) to pollution for developing nations (1) (Fig. 1). This link breaks when incentives to protect the environment are introduced, followed by the adoption of cleaner and more efficient technologies, which can be adjusted to the case of water pollution originating from the disposal of untreated wastewater. Once increasing pollution has had negative effects on the wellbeing and economy of an area, incentives for wastewater treatment are induced. Legislation is introduced requiring polluters to pay for the treatment of wastewater at a certain level. Construction of wastewater treatment plants (WTPs) in compliance with legislation reduces pollution from wastewater and has a positive

te r

Time First incentives Legislation for wastewater treatment at polluted areas

Construction of WTPs in response

Figure 1. Breaking the link between economic growth and pollution from wastewater (1).

effect on a quality index based on a combination of treated effluent qualitative parameters. Wastewater management involves collection treatment and disposal/reuse. Potable water supply is inevitably related to them, hence some principles discussed below will include or directly apply to such issues. Ujang and Buckley (2) summarize sanitation problems for developing countries as lack of environmental awareness, insufficient expertise, inappropriate policies, insufficient funding, insufficient water resources, inappropriate management systems, and institutional support. Following this introduction, the application of existing technologies to developing countries (DC) is discussed, focusing on technology and knowledge transfer and the role of international experts. Some differences over wastewater production and its quality are stressed with special reference to the high standards often set by decision makers in DC. Finally, issues on sustainable technology selection and key points referring to methodologies and indexes are discussed. APPLICATION OF EXISTING TECHNOLOGIES TO DEVELOPING COUNTRIES Wastewater management must be considered as an integral part of the development process and national plans should be formulated (3). Conditions applicable when planning wastewater treatment facilities for DC are not identical to those prevalent during planning in developed countries in the past. Although some similarities exist, many differences also exist. One reason is that water pollution issues are not the main concern in DC because of other more pressing issues such as national or racial security, food availability, and epidemic control (2).

WASTEWATER MANAGEMENT FOR DEVELOPING COUNTRIES

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Technology Transfer

Knowledge Transfer

Today, technology is available for the treatment of wastewaters of any origin and strength. DC cannot be expected to play a key role in the development of cutting edge technology in wastewater management or water research. However, research is needed in order for available technology to be applied to specific country-region conditions. This kind of research is an investment for those countries concerned. The existing transfer of technology should be done wisely because today, developed countries are regarded as an area of investment for water companies. For example, the design approach in rural Egypt was to select wellestablished technologies; little stress was placed on selecting more innovative technologies. However, less sophisticated treatment technologies, such as stabilization ponds, have proven capable of meeting the required effluent standards (4). A similar situation was reported in Greece. From the 147 small (10 d) Fats, oils, and grease High pH (>8.0) Low DO and high MCRT Low F/M ( Aldrin > DDD > DDT > Endrin > Heptachlor > γ -HCH. Zitomer and Speece (43) reported that reductive dehalogenation takes place under anaerobic conditions, reducing the level of chlorination of organochlorine pesticides, thus making them more amenable to further degradation and in general rendering residues less toxic. Dechlorination of OPs is facilitated by the presence of suitable microorganisms in sewage sludge; for instance, the anaerobic bacterium Clostridium rectum is capable of degrading γ -HCH (44). In the case of PCDD/Fs, the dechlorination pathway is from the higher chlorinated 2,3,7,8-subtituted PCDD/Fs to the lower ones, which, in antithesis to PCBs and pesticides, have a higher toxic equivalence factor (TEF) rating, something that increases the overall toxicity of sewage sludge (45,46). Fu et al. (47) reported that abiotic dechlorination of PCDD/Fs in sludge leads to DiCDD/Fs as end products. This suggestion was also prompted by Stevens et al. (48), who found that DiCDFs were the predominant congeners in U.K. sewage sludges. However, this suggestion was not confirmed by Klimm et al. (49) and by Disse et al. (50), who under strictly anaerobic conditions did not observe any formation or destruction of PCDD/Fs. On the contrary, Klimm et al. (49) observed formation of hepta- and octaCDDs under semianaerobic conditions, something that led to a twofold increase in the concentration of these congeners. Under the same conditions, there was no formation of other PCDD/Fs or PCBs congeners. From the above-mentioned, it is difficult to decide if the POPs degrade or do not degrade during the anaerobic digestion of sewage sludge. The fact that POPs can be detected in almost all sewage sludges after treatment (46) suggests that they are rather resistant to this process. The concentration of POPs, especially of PCBs and of PCDD/Fs, in sewage sludges is of great importance, since sewage sludge is often used in agriculture for soil amendment. The European Union (51), in order to improve the long-term protection of soils, is working on a new directive, which includes maximum permissible concentrations for use of sludge in agriculture. According to this upcoming directive (51), the sum concentration of 7 PCB congeners (IUPAC −28, −52, −101, −118, −138, −153, −180) should not exceed 800 µg/kg (dry matter, dm), the concentration of PCDD/Fs should not exceed 100 ng TEQ/kg (dm), and the sum of organochlorine compounds should not exceed 500 mg/kg (dm). In studies dealing with the occurrence of POPs in sewage sludges, PCB levels have been found to vary dramatically from 22.7 to 8000 µg/kg (dm) (1,2,52–62). The reported PCDD/Fs concentrations are also greatly variable ranging between 0.7 and 4100 ng TEQ/kg (dm) (52,63–67).

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The Fate of POPs Throughout the Overall Treatment Process Studies have shown that POPs are recalcitrant to the conditions prevailing in typical biological WWTPs and persist in one or more phases within the treatment plant. Reported removal efficiencies throughout the overall treatment process range within 18–100% for PCBs and 75–90% for several OPs (2,6,31). There are no available data concerning the removal of PCDD/Fs in WWTPs. The distribution of POPs within the WWTP is dependent on the physicochemical properties of the chemicals and the operating conditions within the plant. The principal removal mechanism for the most hydrophobic POPs is through sorption to sludge particles and transfer to the sludge processing system. Advective transport into the final effluent, in association with suspended solids or in the dissolved phase, is also important for less hydrophobic compounds. It has been reported that from the 50 kg of all PCBs annually entering a WWTP, 60% is removed through the wasted sludge, 26% is discharged into the recipient with the final effluent, while 14% is lost due to volatilization or biotransformation (35). Therefore, the long-term ecotoxicological effects on both terrestrial and aquatic organisms need to be assessed for safe disposal of the products of the wastewater treatment process. BIBLIOGRAPHY 1. Blanchard, M. et al. (2001). Origin and distribution of polyaromatic hydrocarbons and polychlorobiphenyls in the urban effluents to wastewater treatment plants of the Paris area (France). Water Res. 35(15): 3679–3687. 2. Katsoyiannis, A. and Samara, C. (2004). Persistent organic pollutants (POPs) in the conventional activated sludge treatment process: occurrence and removal. Water Res. 38(11): 2685–2698. 3. Bedding, N.D., McIntyre, A.E., Perry, R., and Lester, J.N. (1982). Organic contaminants in the aquatic environment. I sources and occurrence. Sci. Total Environ. 25: 143–167. 4. Valls, M., Fernandez, P., and Albaiges, J. (1989). Broad spectrum analysis of organic contaminants in urban wastewaters and coastal receiving systems. In: Organic Contaminants in Wastewater, Sludge and Sediment: Occurrence, Fate and Disposal. D. Quagherbeur, I. Temmerman, and G. Angeletti (Eds.). Elsevier Applied Science, New York, pp. 19–34. 5. Paxeus, N., Robinson, P., and Balmer, P. (1992). Study of organic pollutants in municipal wastewater in Goeterborg, Sweden. Water. Sci. Technol. 25: 249–256. 6. Pham, T.T. and Proulx, S. (1997). PCBs and PAHs in the Montreal urban community (Quebec, Canada) wastewater treatment plant and in the effluent plume in the St. Lawrence River. Water Res. 31(8): 1887–1896. 7. Carballa, M. et al. (2004). Behavior of pharmaceuticals, cosmetics and hormones in a sewage treatment plant. Water Res. 38: 2918–2926. 8. Johnson, A.C. and Sumpter, J.P. (2002). Removal of endocrine disrupting chemicals in activated sludge treatment works. Environ. Sci. Technol. 256: 163–173. 9. Manoli, E. and Samara, C. (1999). Occurrence and mass balance of polycyclic aromatic hydrocarbons in the Thessaloniki sewage treatment plant. J. Environ. Qual. 28: 176–187. 10. Bergh, A. and Peoples, R. (1977). Distribution of polychlorinated biphenyls in a municipal wastewater treatment plant and environs. Sci. Total Environ. 8(3): 197–204.

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11. Shannon, E., Ludwig, F., and Waldmanis, I. (1977). Polychlorinated biphenyls (PCBs) in municipal waste waters. An assessment of the problem in the Canadian lower Great Lakes. Environmental Canada Research Report, No 49. 12. McIntyre, A., Perry, R., and Lester, J. (1981). The behaviour of polychlorinated biphenyls and organochlorine insecticides in primary mechanical waste water treatment. Environ. Poll. Ser. B 2: 223–233. 13. Barrick, R. (1982). Flux of aliphatic and polycyclic aromatic hydrocarbons to central Puget Sound from Seattle (Westpoint) primary sewage treatment effluent. Environ. Sci. Technol. 16: 682–692. 14. Eganhouse, R. and Kaplan, I. (1982). Extractable organic matter in municipal wastewater. 1. Petroleum hydrocarbons: temporal variations and mass emission rates to ocean. Environ. Sci. Technol. 16: 180–186. 15. Granier, L., Chesterikoff, A., Chevreuil, M., and Letolle, R. (1988). Essai de bilan des polychlorobiphenyles (PCB) a la station d’epuration de l’agglomeration parisienne (Acheres). C.R. Acad. Sci. Paris 306(Ser. II): 1175–1178. 16. Wania, F. and Mackay, D. (1996). Tracking the distribution of persistent organic pollutants. Environ. Sci. Technol. News 30(9): 390–396. 17. Pujadas, E. et al. (2001). Application of the new C18 speedisks to the analysis of polychlorinated dibenzo-pdioxins and dibenzofurans in water and effluent samples. Chemosphere 43: 449–454. 18. Peters, T., Nestrick, T., and Lamparski, L. (1984). The determination of 2,3,7,8-tetrachlorodibenzo-para-dioxin in treated wastewater. Water Res. 18(8): 1021–1024.

28. Wang, Y. and Wong, P. (2002). Mathematical relationships between vapor pressure, water solubility, Henry’s law constant, n-octanol–water partition coefficient and gas chromatographic retention index of polychlorinated-dibenzodioxins. Water Res. 36: 350–355. 29. Anonymous (www.es.lancs.ac.uk/ecerg/kcjgroup/11.pdf). 30. Byrns, G. (2001). The fate of xenobiotic organic compounds in wastewater treatment plants. Water Res. 35(10): 2523– 2533. 31. Chevreuil, M., Granier, L., Chesterikoff, A., and Letolle, R. (1990). Polychlorinated biphenyls partitioning in waters from river, filtration and wastewater plant: the case for Paris (France). Water Res. 24(11): 1325–1333. 32. Morris, S. and Lester, N. (1994). Behaviour and fate of polychlorinated biphenyls in a pilot wastewater treatment plant. Water Res. 28(7): 1553–1561. 33. Kipopoulou, A.M., Zouboulis, A., Samara, C., and Kouimtzis, Th. (2004). The fate of Lindane in the conventional activated sludge treatment process. Chemosphere 55: 81–91. 34. Rogers, H.R. (1996). Sources, behaviour and fate of organic contaminants during sewage treatment and in sewage sludges. Sci. Total Environ. 185: 3–26. 35. Katsoyiannis, A. and Samara, C. (2004). Persistent organic pollutants (POPs) in the conventional activated sludge treatment process: fate and mass balance. Environ. Res. in press.

19. De Alencastro, L.F. and Taradellas, J. (1983). Study of the concentration of PCBs in used water from a wastewater treatment plant. Gaz Eaux Eaux Usees 3: 113–122.

36. Pereira, W.E., Hostettler, F.D., and Rapp, J.B. (1996). Distributions and fate of chlorinated pesticides, biomarkers and polycyclic aromatic hydrocarbons in sediments along a contamination gradient from a point-source in San Francisco Bay, California. Mar. Pollut. Bull. 41(3): 299–314.

20. Van Luin, A.B. and Van Starkenburg, W. (1985). Hazardous substances in wastewater. Water Sci. Technol. 17: 843–853.

37. Petrasek, A.C. et al. (1983). Fate of toxic organic compounds in wastewater treatment plants. JWPCF 55(10): 1286–1296.

21. Saleh, F.Y., Lee, G.F., and Wolf, H.W. (1980). Selected organic pesticides, occurrence, transformation, and removal from domestic wastewater. JWPCF 52: 19–28.

38. Howard, P., Boethling, R., Jarvis, W., Meylan, W., and Michalenko, E. (1991). Handbook of Environmental Degradation Rates. Lewis Publishers, Chelsea, MI.

22. Hannah, S.A., Austern, B.M., Eralp, A.E., and Wise, R.H. (1986). Comparative removal of toxic pollutants by six wastewater treatment processes. JWPCF 58: 27–34.

39. Dorussen, H.L. and Wassenberg, W.B.A. (1997). Feasibility of treatment of low polluted waste water in municipal waste water treatment plants. Water Sci. Technol. 35(10): 73–78.

23. Nicoud, S., Humbert, B., De Alencastro, L.F., and Taradellas, J. (1988). Organic micropollutants at the effluents of the wastewater treatment plant, at the water of Rhone and the water of Leman-Campagne 1987. Rapport sur les e´ tudes et recherches entreprises dans le basin Lemanique, pp. 225–234.

40. Bamford, H.A., Ko, F.C., and Baker, J.E. (2002). Seasonal and annual air–water exchange of polychlorinated biphenyls across Baltimore harbor and the northern Chesapeake Bay. Environ. Sci. Technol. 36: 4245–4252.

24. Lorber, M., Barton, R., Winters, D., Bauer, K., Davis, M., and Palausky, J. (2002). Investigation of the potential release of polychlorinated dioxins and furans from PCP-treated utility poles. Sci. Total Environ. 290(1–3): 15–39. 25. Huckins, J.N., Petty, J.D., Prest, H.F., Clark, R.C., Alvarez, D.A., Orazio, C.E., Lebo, J.A., Cranor, W.L., and Johnson, B.T. (2000). A Guide for the Use of Semipermeable Membrane Devices (SPMDs) as Samples of Waterborne Hydrophobic Contaminants. Columbia Environmental Research Center, USGS, USDI and California Analytical Division, Agilent Technologies, Inc., Palo Alto, CA.

41. Ahmed, F.E. (2003). Analysis of polychlorinated biphenyls in food products. TRAC–Trends Anal Chem. 22: 170–185. 42. Hill, D. and McCarty, P. (1967). Anaerobic degradation of selected chlorinated hydrocarbon pesticides. J. Water Pollut. Cont. Fed. 39(8): 1259–1277. 43. Zitomer, D. and Speece, R. (1993). Sequential environments for enhanced biotransformation of aqueous contaminants. Environ. Sci. Technol. 27(2): 227–244. 44. Mogilevich, N. (1982). Microbial destruction of organohalogen compounds. Sov. J. Water Chem. Technol. 43(3): 98–109.

26. Eitzer, B. and Hitte, R. (1989). Atmospheric transport and deposition of polychlorinated dibenzo-p-dioxins and dibenzofurans. Environ. Sci. Technol. 23: 1396–1401.

45. Barkowskii, A.L. and Adriaens, P. (1996). Microbial dechlorination of historically present and freshly spiked chlorinated dioxins and diversity of dioxin-dechlorinating populations. Appl. Environ. Microbiol. 62: 675–681.

27. U.S. EPA. (1982). Fate of Priority Pollutants in Publicly Owned Treatment Works. Final Report, Vol. I. EPA 440/182/303. Effluent guidelines division WH-552. U.S. EPA, Washington, DC.

46. Stevens, J., Green, N., and Jones, K.C. (2003). Fate of 1,2,3,4,6,7,8-heptachlorodibenzofuran and pentachlorophenol during laboratory-scale anaerobic mesophilic sewage sludge digestion. Chemosphere 50: 1227–1233.

THE ROLE OF ORGANOCLAY IN WATER CLEANUP 47. Fu, Q.S., Barkowskii, A.L., and Adriaens, P. (1999). Reductive transformation of dioxins: an assessment of the contribution of dissolved organic matter to dechlorination reactions. Environ. Sci. Technol. 33: 3837–3842. 48. Stevens, J., Green, N.J.L., and Jones, K.C. (2001). Survey of PCDD/Fs and non-ortho PCBs in UK sewage sludges. Chemosphere 44: 1455–1462. 49. Klimm, C., Schramm, K.W., Henkelmann, B., Martens, D., and Kettrup, A. (1998). Formation of octa- and heptachlorodibenzo-p-dioxins during semi anaerobic digestion of sewage sludge. Chemosphere 37: 2003–2011. 50. Disse, G., Weber, H., Hamann, R., and Haupt, H.J. (1995). Comparison of PCDD and PCDF concentrations after aerobic and anaerobic digestion of sewage sludge. Chemosphere 31: 3617–3625. 51. EU. (2000). Working Document on Sludge. 3rd draft—unpublished. 52. Eljarrat, E., Caixach, J., and Rivera, J. (2003). A comparison of TEQ contributions from PCDDs, PCDFs and dioxin-like PCBs in sewage sludges from Catalonia, Spain. Chemosphere 51: 595–601. 53. Stevens, J., Northcott, G., Stern, G., Tomy, G., and Jones, K.C. (2003). PAHs, PCBs, PCNs, organochlorine pesticides, synthetic muscs, and polychlorinated n-alkanes in UK sewage sludge: survey results and implications. Environ. Sci. Technol. 37: 462–467. 54. Ottaviani, M., Crebelli, R., Fusselli, S., La Rocca, C., and Baldassari, L.T. (1993). Chemical and mutagenic evaluation of sludge from a large wastewater treatment plant. Ecotox. Environ. Safety 26: 18–32. 55. Sulkowski, W. and Rosinka, A. (1999). Comparison of the efficiency of extraction methods for polychlorinated biphenyls from environmental wastes. J. Chromatogr. A 845: 349–355. 56. Paulsrud, B., Wien, A., and Nedland, K.T. (2000). A Survey of Toxic Organics in Norwegian sewage Sludge, Compost and Manure. Aquateam, Norwegian Water Technology Centre, Oslo, Norway. 57. Alcock, R. and Jones, K.C. (1993). Polychlorinated biphenyls in digested UK sewage sludge. Chemosphere 25(12): 2199–2207. 58. Kedding, M., Langenohl, T., and Witte, H. (1989). PCB, Dioxine und Furane im Klarschlamm und deren Auswirkungen bei der landwirtschaftlichen Klarschlammverwertung. Korrespondenz Abwasser. 35: 19–26. 59. Lazzari, L., Sperni, L., Bertin, P., and Pavoni, B. (2000). Correlation between inorganic (heavy metals) and organic (PCBs and PAHs) micropollutant concentrations during sewage sludge composting processes. Chemosphere 41: 427–435. 60. Taradellas, J., Muntau, H., and Beck, H. (1985). Abundance and analysis of PCBs in sewage sludge. In: COST 681 Symposium on Processing and Use of Organic Sludge and Liquid Agricultural Wastes. October 8–11, Rome, Italy. 61. Blanchard, M., Teil, M.J., Ollivon, D., Legenti, L., and Chevreuil, M. (2003). Polycyclic aromatic hydrocarbons and polychlorobiphenyls in wastewaters and sewage sludges from the Paris area (France). Environ. Res. 95(2): 184–197. 62. McGrath, D., Postma, L., McCormack, R.J., and Dowdall, C. (2000). Analysis of Irish sewage sludges: suitability of sludge for use in agriculture. Irish J. Agric. Food Res. 39(1): 73–78. 63. Fabrellas, B., Sanz, P., Abad, E., Rivera, J., and Larrazabal, D. (2004). Analysis of dioxins and furans in environmental samples by GC-ion-trap MS/MS. Chemosphere 55(11): 1469–1475.

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64. Eljarrat, E., Caixach, J., and Rivera, J. (1999). Decline in PCDD and PCDF levels in sewage sludges from Catalonia (Spain). Environ. Sci. Technol. 33(15): 2493–2498. 65. Langenkamp, H., Part, P., Erhardt, W., and Pruess, A. (2001). Organic Contaminants in Sewage Sludge for Agricultural Use. A publication of the European Union. Available at http://europa.eu.int/comm/environment/waste/sludge/ organics in sludge.pdf. 66. Rappe, C., Andersson, R., Studer, C., and Karlaganis, G. (1997). Decrease in the concentrations of PCDDs and PCDFs in sewage sludge from Switzerland. Organohal. Comp. 33. 67. Rappe, C., Andersson, R., Karlaganis, G., and Bonjour, R. (1994). PCDDs and PCDFs in samples from various areas in Switzerland. Organohal. Comp. 20: 79–84.

THE ROLE OF ORGANOCLAY IN WATER CLEANUP GEORGE R. ALTHER Biomin, Inc. Ferndale, Michigan

Organically modified clays, also called organoclays, have been used to clean up water since 1985. Their prime function is as a prepolisher for activated carbon, ion exchange resins, and membranes. They are also used in a stand-alone mode after dissolved air flotation (DAF) and oil/water separation units. Their main use is to remove oils, greases, and other large hydrocarbons of low solubility from water. They are very adept at removing chlorinated hydrocarbons. In this application, organoclays are usually blended with anthracite in a ratio of 30% organoclay and 70% anthracite. The reason is that the organoclay, in its pure form, would collect so much oil in its interstitial pore spaces, and due to swelling of the clay, that it would last no longer than activated carbon, which removes 8–10% of oil based on its weight, before its pores are blinded. The organoclay blend, on the other hand, removes 50–70% of oil based on its weight, some seven times as much as activated carbon. The economic benefit for the end user, the one who pays for the cleanup, is a savings of 50% or more of operating costs. Organoclays can be called prepolishers to carbon, but it can also be said that carbon is a postpolisher to organoclay. The reason is that the organoclay also removes other compounds, such as PNAHs, BTEX, PCBs, and other hydrocarbons of low solubility, with extreme efficiency. This has been shown in many publications through the last 15 years (1–5). Organoclays are bentonites modified with quaternary amines. Bentonite is a volcanic, chemically altered rock that consists primarily of the clay mineral montmorillonite. Montmorillonite contains inorganic exchange ions, particularly sodium, calcium, and magnesium that hydrate in the presence of water and produce a hydrophilic environment on the surface of the clay. Mixing a cationic quaternary alkyl ammonium chloride or bromide compound with the bentonite makes the clay organophilic. Bentonite, which is hydrophilic in its natural state, becomes hydrophobic and organophilic when modified with

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THE ROLE OF ORGANOCLAY IN WATER CLEANUP

Oil

Oil

Oil

Pores

Pores

Figure 1. How activated carbon and organoclay remove oil from water.

Activated carbon granule pore spaces of activated carbon, blinded by emulsified oil.

quaternary amines (6,7). The positively charged end of the amine chain, which consists of a carboxylic head that includes a nitrogen ion, exchanges for a sodium or calcium ion on the surface of the bentonite. The cationic amine now becomes neutral, and the thus formed organoclay has turned into a nonionic surfactant that has a solid phase. The swelling capacity of sodium bentonite in water is up to 15 times its volume, but nearly zero in hydrocarbon fuels and solutions (8). After the organic modification, the organoclay swells some 15 times in a fuel such as gasoline and by some 20% when placed into a column and exposed to fuels such as gasoline (9–11). Its swelling capacity in water is very low. If alcohol is added to a fuel, the organoclay will swell even more and turn the system into a gel, or even grease. For this reason, organoclays have been used as thickeners for paints, drilling muds, greases, printing inks, and many other systems since the 1950s. Figure 1 shows how the two media differ in removing oil from water. The organoclay removes it by partition, whereas the activated carbon removes oil and other hydrocarbons by adsorption. Partition takes place outside the clay particle; adsorption takes place inside the pores, which is the reason that they become blinded by organic compounds equal to or exceeding the diameter of the pores. During the late 1950s, it was discovered that organoclays remove organic compounds from water, including benzene (12). It was found that organoclays can remove chlorinated organic hydrocarbons of low solubility efficiently (13,14). Excellent descriptions of the mechanisms of organoclay interactions with organic compounds are also given by Lagaly (15,16). The removal of these compounds from water by organoclays is the result of a partition mechanism similar to the process when immiscible organic compounds such as octanol are added to water contaminated with organic compounds (16,17–19). The interlayer phase of the organoclay acts as a partition medium for oils, greases, and other hydrocarbons (1,20–24). A portion of these organic compounds moves out of the water and into the organic compound where it is more soluble; like dissolves into like. The relative solubility of the contaminant determines the amount retained in each phase. The terminology can be extended to contaminants partitioning from the water phase into a solid phase such as organic cations sorbed to a clay surface. The higher the solution concentration of a compound and the lower its solubility, the larger the quantity removed by the organoclay by partitioning. The organic compounds are held closely by the quaternary

Clay platelets, modified with quaternary amine, removal of emulsified oil on the clay surface.

Activated carbon downstream of organoclay, ready to remove more soluble compounds.

amine by coulombic forces (25), and the contaminant is not easily leached off. If an amount of amine is exchanged into the bentonite that exceeds its stoichiometric capacity, the amine chains will still attach to the organoclay in a tail-totail interaction. The result is that this organoclay now has a positive charge and will remove organic and inorganic anions from water, such as humic acids and hexavalent chromium (Fig. 2) (26). Only part of the clay surfaces is covered with the quaternary amines, so a portion remains free, available for cation exchange with heavy metals such as lead, zinc, nickel, cadmium, and iron. The cation exchange capacity of bentonite ranges from 70–95 meq/100 gram. Column studies were conducted with pure organoclay and an organoclay/anthracite blend, testing for a number of metals simultaneously, and then testing for the removal capacity for single metals (27). The main application of organoclays has been in groundwater remediation. Removal of oil from water is the organoclay’s main function. Figure 3 shows a column test with organoclay/anthracite to determine the sorptive capacity of this medium for a vegetable oil. A 30-inch long (76.2 cm) by 3-inch diameter (7.62 cm) column was constructed from polyvinyl chloride (PVC) and filled with about 6 pounds of sorbent material to be studied. A peristaltic pump forced an aqueous solution containing 680 mg/L of vegetable oil through the column, after the column was backwashed with water to displace any air pockets. Samples were collected periodically at the outflow of the column and analyzed for their organic content using chemical oxygen demand (COD) analysis. The results are shown in Table 1 and Fig. 3. The percent removal capacity for a mineral oil such as Bunker C would be much higher because it is less soluble than vegetable oil. Figure 3 is a graphic description of these data. Figure 4 shows the results of a column study comparing the removal capacity for oil between organoclay/anthracite and bituminous activated carbon. To gain some background data, the ability of powdered, nonionic organoclay to remove a variety of oils from water was tested in jar tests (26,28). The results displayed in Figs. 5–8 show that organoclay removes all oils from water exceptionally well, as long as they are refined. It is far superior to activated carbon, which is why it is used as a prepolisher. If oil is not refined, which means its composition includes polar compounds, either a cationic organoclay has to be used, or the oil, which is now partially chemically emulsified, must be deemulsified (1,21,23,28–30). Figure 9 shows

THE ROLE OF ORGANOCLAY IN WATER CLEANUP Lypophile

Alkyl chains

CH3 N+

B. CH3

CrO42

Hydrophile

Oil

H H H H H H H H H H C H C H CH C H C H C H C H C H C CH CH C H C H C H C H C H C H C H H H H H H H H H H

A.

773

H H H H H H H H C H C H CH CH C H C H C H CH C H C H C H C H C H C H C H H H H H H H H H H H H H H H H H H C H C H CH C H C H C H C H C H C CH CH C H CH C H C H C H C H C H H H H H H H H H H H H H H H H H H H H C H C H CH CH C H C H C H C H C C H CH C H CH C H C H C H C H C H H H H H H H H H H H H H H H H H H C H C H CH C H C H C H C H CH CH C H C H C H C H C H C H H H H H H H H A. H H H H H H H H C H C H CH CH C H C H CH CH CH C H C H C H C H C H C H H H H H H H H

N+

H H C H C H C H H

CH3 Cl−

Na+

CH3

Cl− Na+ Cl− Ca++ Cl− Exchangeable cations nH2O

H H CH3 C H C Cl− H C H N+ H H H C H C CH3 H C H H

Montmorillonite

Oxygen Hydroxyls Aluminum, iron, magnesium and Silicon, occasionally aluminum

A. Partition reaction that removes non-polar organic compounds. B. Ionic bonding that removes anionic inorganic and organic compounds.

Figure 2. Model of an organoclay indicating partition and ion exchange adsorption mechanisms.

the effect of surfactants, which act as emulsifiers for oil, on the performance of organoclay. The effects are the same as on activated carbon; as solubility increases due to emulsification, the sorptive capacity decreases. Only the nonionic surfactant shows little effect on organoclay performance; probably it does not polarize the oils to any

significant extent. Tests have shown that organoclay is just as effective in the removing surfactants from water as activated carbon (30). However, organoclay always prefers oil to any other compound; thus once the emulsion is split, the organoclay is used to remove the oil, followed by carbon which removes the surfactants.

Oil on ogranoclay 1.0

Normalized concentration

0.9 0.8 0.7 0.6 0.5 0.4 0.3 Inflow 0.2

Oil

0.1

95% breakthrough

0.0 0

100

200

300

400

500

600

700 800 900 Pore volumes

1000 1100 1200 1300 1400 1500 1600

Figure 3. Breakthrough curve of oil on organoclay.

774

THE ROLE OF ORGANOCLAY IN WATER CLEANUP

Table 1a. Sorbent Mass, Porosity, Flow Rate, and Residence Time

kg 0.141

Porosity

Flow Rate,

90

Residence Time

80 70

lb

%

mL/min

gal/h

min

0.31

0.3

15.45

0.23

8

% removal

Mass Sorbent,

100

Organically modified clay

60 50 40 30

Table 1b. 95% Breakthrough for Organoclay/Anthracite Given in Pore Volumes and Minutes Along with Estimated Mass of Oil Sorbed Per Mass of Sorbent in mg/kg, lb/lb, and On a Percent Basis Breakthrough PV

Mass Sorbed

min

g/lb

20 10

10 20 30 40 50 60 70 80 90 100110120130140150 Gallons of effluent

Mass Sorbed per Mass Sorbent g/kg

lb/lb

% by sorbent

475

0.475

47.5

Activated charcoal

Figure 4. Removal of oil from water. 1150

9,200

65.8

0.14

RESULTS A minicolumn test was used to determine the ability of organoclay to remove such compounds as benzene, toluene, xylene, naphthalene, and PCB, and to be able to compare the data with those of bituminous activated carbon. The minicolumn method consists of spiking water with the compound to be evaluated and pumping that water through 1 gram of sorbent powder, which is tightly packed into a very small column. Pumping is performed until the influent concentration equals that of the effluent concentration (26). The sorbents are of 200 U.S. mesh size. This method is more comparable to large-scale, real situations than isotherms.

Figure 10 is a graphical illustration of minicolumn tests. When testing the removal capacity of the sorbents for benzene, toluene, o-xylene, and naphthalene, the organoclay performs similarly to carbon and performance improves as the solubility of the compounds decreases. Nonionic organoclay outperforms activated carbon with PCB 1260, as well as with motor oil, which is nearly insoluble in water. Surprisingly, organoclay removes methylene chloride much more effectively than activated carbon. The reasons are unclear, except that organoclay has an affinity for chlorinated compounds. Earlier results for vinyl chloride proved similar. It is theorized

No. 6 Bunker C No. 4 Oil Gear oil Terpentine Terpentine

Control test using activated carbon

Transformer oil Motor oil Lubricating oil Blando oil Hydraulic oil Crude oil Glycol oil Diesel fuel Kerosene Chlorinated paraphin klovo 40 sus. visc.135 CW8560 Sus. Visc. 4200 CW 170 Sus. Visc. 5600 10

20

30

40

50

60

70

% oil removal Figure 5. Removal of mineral oils from water by organoclay.

80

90

100

THE ROLE OF ORGANOCLAY IN WATER CLEANUP

Olive oil Olive oil Extra virgin olive oil Crude palm oil ∗RBD palm oil Refined palm oil Peanut oil Linseed oil Raw linseed oil Superb linseed oil Ropeseed oil Crude corn oil Refined corn oil Safflower oil Soy oil Crude sunflower oil Degummed sunflower oil *RBW sunflower oil Coconut oil Crude cottonseed oil Refined cottonseed oil AA standard castor oil No. 1 castor oil Canola oil ∗RBD

∗RBW

775

Control test using activated carbon

= Refined Bleached Deodorized = Refined Bleached Winterized

10

20

30

40

50 60 % oil removal

70

80

90

100

Figure 6. Removal of plant oils from water by organoclay.

and toluene. This allowed the observation of competition among these compounds for adsorption sites. Of each compound, 900 mL/gram were added to water. It was possible to add that much naphthalene because benzene and toluene helped dissolve it. Usually, its solubility is only 10 mg/L. This concentrate was pumped separately through a column of organoclay, powdered activated carbon, and organoclay/carbon combined. In that case, the bottom of the minicolumn contained 0.5 grams of activated carbon, and the upper half contained 0.5 grams of organoclay. This was done to determine if the organoclay/carbon combination is more effective than each sorbent alone.

that compounds such as methylene chloride have high electronegativity due to the presence of large amounts of halogens such as chlorides. The organoclay possesses positive charges on the surface due to the presence of inverted quaternary amine chains, so methylene chloride could chemically bond to these charges via their electronegativity. Therefore, two removal mechanisms, partitioning and ionic bonding, account for organoclay removal capacity for these compounds. These tests were followed by a set of tests using a ternary effluent, which is a wastewater containing three different organic hydrocarbons, naphthalene, benzene,

Lard oil Kidney rosin oil Herring oil Cod liver oil Tall oil Mineral oil Cold pressed valencia orange oil D-Limouene citrus based solvent Glycol based synthetic oil Glycol based synthetic oil

Control test using activated carbon 10

20

30

40

50

60

% oil removal

70

80

90

100

Figure 7. Removal of miscellaneous oils from water.

Crude cottonseed oil (1) (2) Boiled linseed oil

(1) (2)

Indonesian palm oil

(1) (2)

Herring oil

(1) (2)

Synthetic watch oil

(1) (2) 10

Figure 8. Removal of miscellaneous oils from water by different organoclays.

20

30

40 50 60 70 % oil removal

80

90

100

100% 95%

Organoclay (no surfactant added) ×

% removal capacity

90% 85% 80%

Triton sp-160 non-ionic surfactant

75% ×

70%

SLS 30 cationic surfactant

65% ×

60%

×

55%

L-60 anionic surfactant

50% 1 Drop 2 Drops 4 Drops 0.03 gram 0.06 grams 0.13 grams

Figure 9. Percent removal capacity of organoclay for oil after adding a surfactant.

8 Drops 0.24 grams

12 Drops 0.39 grams

Drops of surfactant added to oil contaminated water

Benzene Legends: Organoclay

Toluene

Activated carbon

Naphtalene

PCB-1260

Methylene chloride 3,500 mg/gram

Motor oil 100

200

300

400

500

mg/gram Sorbent loading at breakthrough

Figure 10. Minicolumn tests.

776

600

THE ROLE OF ORGANOCLAY IN WATER CLEANUP

The results in Fig. 11 illustrate that benzene breaks through first, followed by toluene, and lastly naphthalene. This was expected based on their solubilities in water. The competition, however, is not 100% proof. The total adsorbed amounts are higher than the individually adsorbed amounts of the three solvents at breakthrough, probably because the geometrically arranged packing of the solvents of different sizes, either within the carbon pores or around the amine chains, favors a higher packing density. The most important result is shown in the ‘‘total combined’’ graph. By placing the organoclay in front of the carbon, the removal capacity is doubled compared with the removal capacity of carbon and organoclay individually. This is also shown in a standard permeation column experiment with gasoline (Fig. 12). Again, the combination of organoclay/anthracite, followed by activated carbon, is much more effective in removing gasoline from water than either sorbent by itself, even though the amount of sorbent in each column is twice that in the combined column. Table 2 shows results from an actual groundwater cleanup project. The organoclay removes the oil completely and also a significant amount of other solvents and increases

Legend: Organoclay

Benzene

Activated Carbon Toluene

Organoclay/ Activated Carbon Combination

Naphtalene

the effectiveness of activated carbon in removing volatile organic compounds. Because of the roll-off phenomenon, where less soluble compounds such as toluene and xylene, knock benzene off sorption sites within activated carbon, thus recontaminating the effluent, it is important to prepolish the water with organoclay, followed by activated carbon, which these results clearly illustrate. The improved performance of an organoclay as the chlorination of the organic compound increases, along with a corresponding decrease in solubility in water, is shown in Figs. 13 and 14 (8). Organoclay is effective in removing phenol from water, but much more so, and far superior to activated carbon, for pentachlorophenol (PCP). Freundlich isotherms confirm the results previously discussed, illustrated in Figs. 15–18. Figure 19, which shows the effectiveness of organoclay in removing methylene chloride, is impressive. Figure 20 shows a time study, comparing how effectively nonionic organoclay and activated carbon remove turpentine from water. This test was conducted by pumping water spiked with turpentine through minicolumns and removing and analyzing a water sample every 50 minutes. Once the curve flattens out, saturation of the sample has occurred. These test results confirmed that the retention time of 6–10 minutes for water in an adsorber

Table 2. Organoclay/Carbon Sequence for Treating Contaminated Groundwater at an Abandoned Manufacturing Site

Organic Compound

Total combined 100

200

300

400

500

600

mg/gram Sorbent loading at breakthrough

777

Effluent Effluent After Solubility, After Influent, O. Clay, Carbon, mg/L at µg/L µg/L 20–25 ◦ C µg/L

Oil, mg/L 0.02 1,1,1-Trichloroethane 480–4000 Trichloroethane 110 1,1-Dichloroethene 335 Toluene 535 pH

5.0 42,622 688 285 967 8.64

ND 26,044 271 ND 242 8.01

ND ND ND ND ND 9.2

100 Granular activated carbon (2)

90

% breakthrough

80

EC-100 organoclay (1)

70 60 50 EC-100 organoclay + Granular activated carbon (3)

40 30 20 10 1

2

3

4

5

6

7

Gallons of water treated

Figure 12. Breakthrough curves for gasoline saturated water.

Concentration remaining in solution, ppm

Figure 11. Minicolumn tests. 700 600 500 PT-1E PC-1E AC

400 300 200 100 0 25

125

250

1000

Solution concentration, ppm

Figure 13. Adsorption of phenol by organoclay and activated carbon.

778

THE ROLE OF ORGANOCLAY IN WATER CLEANUP 1000 Adsorption, mg/g of adsorbent

Concentration remaining in solution, ppm

700 600 500

PT-1E PC-1E AC

400 300 200 100 0 −100

25

250

100

1 0.1000

1000

Solution concentration, ppm

filled with either medium as is standard operating procedure, is applicable to both media. Another test showed that the organoclay may need a slightly longer retention time (8–12 minutes?) for more soluble benzene. Bentonite is a natural ion exchange resin. Thus is an organoclay, even though a portion of its surface is covered with the chains of the quaternary amines. Some capacity for removing metals is still available in both media.

1.0000

10.0000

Concentration remaining in solution, mg/L

Figure 17. Isotherm showing a comparison of organoclay and activated carbon for the adsorption of naphthalene.

1000

Adsorption, mg/g

Figure 14. Adsorption of pentachlorophenol (PCP) by organoclay and activated carbon.

PT-1 AC

10

100

10

AC-PCB1221 AC-PCB1232 PT1-PCB1260

1

Adsorption, mg/g

1000 0.1 0.0001

100

1

Figure 18. Isotherms showing a comparison of organoclay and activated carbon for the adsorption of PCB.

PT-1E line AC line

10

0.001 0.01 0.1 Concentration remaining in solution, mg/L

4.5

1

10

100

Concentration remaining in solution, mg/L

Figure 15. Isotherm showing a comparison of organoclay and activated carbon for o-xylene adsorption.

Adsorption, mg/g of adsorbent

1000

Log [adsorption, mg/g]

4

1

3.5 3 2.5 2 1.5

Carbon line PT-1E line

1 0.5 0 0

100

0.5

1

1.5

2

2.5

Log [concentration, mg/L]

Figure 19. Isotherms showing a comparison of organoclay and activated carbon for the adsorption of turpentine.

PC-1 AC

10

Cation Exchange Capacity (CEC) Organoclay/anthracite; 0.04 meq/gram.

1 1

10

100

Concentration remaining in solution, mg/L

Figure 16. Isotherm showing a comparison of the removal capacity of organoclay and activated carbon for toluene.

0.04 meq/gram;

Organoclay:

Surface Area (m2 /gram) Organoclay/anthracite: 0.97 ± 0.05; Organoclay: 1.23 ± 0.69. A series of column tests were conducted in the

THE ROLE OF ORGANOCLAY IN WATER CLEANUP (a)

100

PT-1E eq AC eq

10

1 1

10 100 1000 10000 Concentration remaining in solution, mg/L

Comparsion of PT-1E and activated carbon turpentine adsorption 250 200 Activated carbon PT-1E

150

100 50

Multimetals 0.014 0.012 0.010 0.008 0.006 0.004 0.002 0.000 Cd

Cr

Cu

Ni

Zn

Pb

(b) Mass sorbed (lb)/mass sorbent (lb) × 100 percentage

Turpentine remaining in solution, mg/L, ppm

Figure 20. Isotherm showing a comparison of organoclay and activated carbon for the adsorption of methylene chloride.

Mass sorbed (lb)/mass sorbent (lb) × 100 percentage

Adsorption, mg/g

1000

779

Single metals 0.30 0.25 0.20 0.15 0.10 0.05 0.00 Cd

Cr

Cu

Ni

Zn

Pb

Fe

Figure 22. Adsorption of heavy metals from water by an organoclay/anthracite blend: Column study.

0 0

500

1000

1500

2000

2500

3000

Time, minutes

Figure 21. Time study showing a comparison of organoclay and activated carbon for the adsorption of turpentine from water.

same manner as that described for oil removal (Fig. 3), and the capacities of organoclay/anthracite and straight organoclay for removing various metals were determined. Bar charts were then constructed to illustrate the results, shown in Figs. 21 and 22 (24). The fact that the straight organoclay is not much better than organoclay/anthracite suggests that the diluting action of the anthracite results in improved access to sites on the organoclay. The U.S. Standard mesh size of the medium is 8 × 30 mesh. This capacity to remove small amounts of metals is not of great importance, but it can be a factor in calculating whether an ion exchange resin needs to be added to the treatment train if metal removal is required. Figure 23 shows the ability of a cationic organoclay to remove chlorine from water. This isotherm proves that a cationic organoclay is an excellent medium for removing humic and fulvic acids (natural organic matter) and is far superior to bituminous activated carbon (Fig. 24). A minicolumnn test compares the removal capacity of cationic organoclay, bituminous activated carbon, and an ‘‘organotrap’’ ion exchange resin, again showing superior results for the cationic organoclay (Fig. 24). These results are thoroughly discussed in Alther (26).

Table 3 shows the results of minicolumn tests to determine the removal capacity of cationic organoclay for negatively charged metals, including hexavalent chromium, selenite, arsenate, and fluoride. These results are in line with the capacity of this clay for chlorine removal. An anionic organoclay was developed and tested in a minicolumn to show its capacity to remove cations, including ethylenediaminetriacetate (EDTA) (Fig. 25). The results are that coal based activated carbon removes 21.8% EDTA from the spiked solution, nonionic organoclay removes 47.3%, and anionic organoclay removes 70% of the EDTA. Another set of minicolumn tests (Fig. 26) on water spiked with monoethanol amine revealed the following results: coal-based activated carbon removed 41.4%, nonionic organoclay removed 45.2%, anionic organoclay removed 29%, and cationic organoclay removed 79% of the monoethanolamine from the spiked water. These results reveal the extreme versatility of organoclays in removing a variety of contaminants. As a final test, the iodine numbers were determined for the organoclays, using the ASTM D-4607-94 testing method, to see if this test could be used to compare the two media. The results are as follows: activated carbon (coal based): 700–900 nonionic organoclay: 275 cationic organoclay: 190 anionic organoclay: 410

780

THE ROLE OF ORGANOCLAY IN WATER CLEANUP

Mass sorbed (lb)/mass sorbent (lb) × 100 percentage

(a) 0.030

Arsenate

0.025

Amount of cationic organoclay used: 16.5 grams (16 × 30 mesh) Amount of spiked water passed through column: 8.31 liters

0.020 0.015

Removal capacity of cationic organoclay for arsenate by weight: 0.3%, or 3 grams per 1000 grams cationic organoclay

0.010 0.005

Fluoride

0.000 Cd

Cr

(b) Mass sorbed (lb) / mass sorbent (lb) × 100 percentage

Table 3. Laboratory Column Study with Cationic Organoclay

Multimetals

0.035

Cu

Ni

Zn

Pb

Single metals

0.50 0.45 0.40 0.35 0.30 0.25 0.20 0.15 0.10 0.05 0.00

Amount of cationic organoclay used for fluoride removal: 16.5 grams (16 × 30 mesh) Amount of spiked water passed through column: 0.54 liters. Removal capacity of cationic organoclay by weight: 0.1% or 1 gram fluoride per 1000 gram of cationic organoclay Chromate Batch test: 100 mL water spiked with 5 mg/L hexavalent chromium, added 2 grams powdered cationic organoclay: Removed 96% of 5 mg/L. Conclusion

Cd

Cr

Cu

Ni

Zn

Pb

Fe

Figure 23. Adsorption of heavy metals from water by straight organoclay: Column study.

Aside from the extreme efficiency with which cationic organoclay removes humic acids from water, it also removes negatively charged metals, such as those shown above, and selenite.

Points Line estimate

100

1000

Adsorption, mg/g

Adsorption, mg/g

1000

10

1 1

10

100

1000

PC-1 AC

100

10

1

10000

Concentration remaining in solution, mg/L Figure 24. Isotherm showing the adsorption capacity of cationic organoclay for chlorine from water.

Performance of activated carbon in this test is superior because iodine molecules fit inside the pores and do not cause blinding and all adsorption on the organoclay takes place outside the clay platelets. Therefore, it is concluded that this test is of no use for organoclay, although the results are interesting. Below are several case histories, which give a practical appreciation of organoclays to the engineer who designs remediation systems. Case Histories 1. A creosote superfund site on the East Coast installed a pump and treat system consisting of two filter vessels, each containing 20,000 lb activated carbon. The flow rate was 170 gpm. The COD consisted of

0.1 1.00

10.00

100.00

1,000.00

10,000.00

Concentration remaining in solution, mg/L

Figure 25. Isotherm showing the adsorption capacity of cationic organoclay with that of activated carbon.

40–60 ppm, including benzene, VOCs, and phenols. The activated carbon lasted about 2 weeks with a breakthrough of 7–12 ppm COD; then it had to be replaced. After another vessel containing 19,000 lb of organoclay was installed, the effluent after the activated carbon was not detectable. Furthermore, there was a TSS content of 32–35 ppm (discharge limit is 40 ppm), primarily due to the presence of ferric iron. Once the organoclay was installed, the TSS content in the effluent was 3 ppm because the organoclay, a bentonite, also removes heavy metals by ion exchange.

Fulvic acid

THE ROLE OF ORGANOCLAY IN WATER CLEANUP

Legend: Organoclay Organotrap anion exchange resin Activated carbon 100

200

300 400 500 mg/gram Sorbent loading at breakthrough

600

Figure 26. Microcolumn study data comparing the adsorption capacity of cationic organoclay and activated carbon for fulvic acids.

2. An old wood-treating site in Colorado is situated above an aquifer, which had a concentration of 30 ppm of an oil and 25 ppm of PCP. The discharge limit for PCP is 50 ppb. When an activated carbon system was installed, replacement was required within 1 month. After 20,000 lb of organoclay was installed prior to the activated carbon, discharge limits were met, and replacement was required only after 12 to 15 months. 3. An old railroad site in southeastern United States, where railroad ties were once treated with creosote, required excavating the soil and thermally treating it to destroy the creosote. A condensate built up that contained PCP. Rather than accepting the high cost of incinerating the condensate water, it was passed through an organoclay/carbon system and discharged locally. This brief description of the use of organoclays for water treatment should convince the reader of their usefulness. Anyone who is interested in more detail should consult the references. BIBLIOGRAPHY 1. Alther, G.R. (2002). Removing oils from water with organoclays. J. Am. Water Works Assoc. 94(7): 115–121. 2. Alther, G.R. (1997). Working hand in hand with carbon for better results. Soil & Groundwater Cleanup May: 17–19. 3. Sheng, G., Xu, S., and Boyd, S.A. (1996). Co-sorption of organic contaminants from water by hexadecyltrimethylammonium exchanged clays. Water Res. 30(6): 1483–1489. 4. Smith, J.A. and Jaffe, P.R. (1994). Benzene transport through landfill liners containing organophilic bentonite. J. Environ. Eng. 120(6): 1559–1577. 5. Wolfe, T.A., Demiral, T., and Bowman, E.R. (1985). Interaction of aliphatic amines with montmorillonite to enhance adsorption of organic pollutants. Clays Clay Miner. 33(4): 301–311. 6. Jaynes, W.F. and Boyd, S. (1991). Clay mineral type and organic compound sorption by hexadecyltrimethylammonium-exchanged clays. Soil Sci. Soc. Am. J. 55(1): 43–48. 7. Jordan, J.W. (1949). Organophilic bentonites. I. Swelling in organic liquids. J. Phys. Colloid Chem. 53(2): 294–306. 8. Alther, G.R. (2004). A winning combination. Water Wastewater Prod. July/August: 22–27. Available at www.wwponline.com.

781

9. Alther, G.R., Evans, J.C., and Tarites, R. (1991). The use of organoclays for stabilization of hazardous wastes. Proc. 4th Annu. Hazardous Waste Manage. Conf./Central, pp. 547–552. 10. Alther, G.R., Evans, J.C., and Pancoski, S. (1989). A Composite Liner System to Retain Inorganic and Organic Contaminants. Hazardous Materials Research Institute, Superfund 89, Silver Spring, MD. 11. Alther, G.R., Evans, J.C., and Pancoski, E.S. (1988). Organically modified clays for stabilization of organic hazardous waste. Hazardous Materials Control Research Institute 9th Natl. Conf., Superfund 88, Silver Spring, MD, pp. 440–445. 12. Cowan, C.T. and White, D. (1962). Adsorption by organoclay complexes, I. Proc. 9th Natl. Clay Conf., pp. 4509–4517. 13. Slabough, W.H. and Hanson, D.B. (1969). Solvent selectivity by an organoclay complex. J. Colloid Interface Sci. 29(3): 460–463. 14. Street, G.B. and White, D. (1963). Adsorption by organoclay derivatives. J. Appl. Chem. 288–291. 15. Lagaly, G. (1981). Characterization of clays by organic compounds. Clay Miner. 16: 1–21. 16. Lagaly, G. (1979). The ‘‘layer charge’’ of regular interstratified 2:1 clay minerals. Clays Clay Miner. 27: 1–10. 17. Fashan, A., Tittlebaum, M., and Cartledge, F. (1993). Nonionic organic partitioning onto organoclays. Hazardous Waste Hazardous Mater. 10(3): 313–322. 18. Mortland, M.M. (1970). Clay–organic complexes and interactions. Adv. Agron. 22: 75–117. 19. Theng, B.K.G. (1974). The Chemistry of Clay-Organic Reactions. Halstead Press, John Wiley & Sons, New York. 20. Alther, G.R. (1999). Organoclays remove oil, grease, solvents and surfactants from water. CleanTech 99 Proc., Witter, Flemington, NJ, pp. 72–79. 21. Alther, G.R. (1998). Put the breaks on oil and grease. Chem. Eng. March: 82–88. 22. Alther, G.R. (1996). Organically modified clay removes oil from water. Waste Manage. 15(8): 623–628. 23. Alther, G.R. (1995). Organically modified clay removes oil from water. Waste Manage. 15(8): 641–650. 24. Jaynes, W.F. and Vance, G.F. (1996). BTEX sorption by organo-clays: Cosorptive enhancement and equivalence of interlayer complexes. Soil Sci. Soc. Am. J. 60: 1742–1749. 25. Mortland, M.M., Shaobai, S., and Boyd, S.A. (1986). Clay–organic complexes as adsorbents for phenols and chlorophenols. Clays Clay Miner. 34(5): 581–585. 26. Alther, G.R. (2001). Organoclays remove humic substances from water. In: Humic Substances. E.A. Ghabbour and G. Davis (Eds.). Royal Society of Chemistry, Cambridge, UK, pp. 277–288. 27. Tillman, F.D., Bartelt-Hunt, S.L., Craver, V.A., Smith, J.A., and Alther, G.R. (2004). Relative metal ion sorption on natural and engineered sorbents: Batch and column studies. Environ. Eng. Sci. (submitted). 28. Alther, G.R. (2002). Using Organoclays to Enhance Carbon Filtration. Waste Management 22. Pergamon Press, New York, pp. 507–513. 29. Alther, G.R. (2000). Maximize water cleanup performance. Environ. Prot. Mag. Feb.: 37–39. Available at www.wponline.com. 30. Alther, G.R. (1999). Organoclay filtration technology for oil removal. Fluid/Part. Sep. J. 12(2): 96–102.

782

COMBINED SEWER OVERFLOW TREATMENT

COMBINED SEWER OVERFLOW TREATMENT

include (1) the contributing drainage area (catchment) and wastewater sources, (2) the combined sewer pipe network and interceptor(s), (3) the regulator and diversion structures, and (4) the CSO outlets (Fig. 1). When an overflow occurs, the excess flows tend to be discharged into the neighboring receiving body of surface water. CSOs typically discharge a variable mixture of raw sewage, industrial/commercial wastewater, polluted runoff, and scoured materials that build up in the collection system during dry weather. These discharges contain a variety of pollutants that may adversely impact the receiving waterbody, including pathogenic microorganisms, viruses, cysts, and chemical and floatable materials. Health risks associated with bacteria-laden water may result from dermal contact with the discharge, from ingestion of contaminated water, as well as from consumption of fish or shellfish.

C.A. PROCHASKA A.I. ZOUBOULIS Aristotle University of Thessaloniki Thessaloniki, Greece

INTRODUCTION Combined sewer overflows (denoted hereafter as CSOs) occur when flows exceed the hydraulic capacity of either the wastewater treatment plant (denoted as WWTP) or the collection system that transports the combined flow of storm water and sanitary sewage to the WWTP. The principal components of a combined sewer system

Rainfall Subcatchment

Overland flow ce rfa ge u S ra sto

Industrial wastes on

ti ltra nfi

I

Sub s

Inlet (catch basin) CSO treatment facility

Pip e

Domestic wastewater Infiltration

urfa

flow

ow

fl er ath ) e F w ry (DW

ce

Regulator structure

D

er ew ds e in flow mb Co pipe

er ew ds e in w mb rflo Co ove

Inte for rcepto DW r F

Flo w

DWF treatment facility Treated effluent discharge

Figure 1. Schematic diagram of a combined sewer system (2).

COMBINED SEWER OVERFLOW TREATMENT

The methods used to treat CSOs can be classified as physical, chemical, and biological and methods that include a combination of some or all the above, such as treatment by constructed reed beds (1).

2.

PHYSICAL TREATMENT 3.

Physical treatment alternatives include sewer separation, retention basins, swirl/vortex technologies, screening, netting systems for floatable control, dissolved air flotation, and filtration. Most of these physical unit operations have been in use for many years and are considered reliable. Physical treatment operations are usually flexible enough to be readily automated and can operate over a wide range of flows. They can also stand idle for long periods of time without affecting treatment efficiency (1).

4.

5.

783

objectionable source of certain pollutants, such as TSS, sanitary floatables, and bacteria. Reduced volume of flow to be treated at the publicly owned treatment works (POTWs), thus reducing operating and maintenance (O & M) costs by eliminating surface runoff inflows during wet weather. Reduced infiltration and excess flow to a POTW for new sanitary sewer construction, replacing old combined sewers. Reducing upstream flooding, as well as overflows, when the existing combined sewers are undersized and back up frequently during storms. Being more effective and economical than treatment facilities for remote segments of a combined sewer system, serving relatively small areas.

Retention Basins Sewer Separation Separation is conversion of a combined sewer system into separate storm water and sanitary sewage collection systems. This alternative, historically considered the ultimate answer to CSO pollution control, has been reconsidered in recent years because of increased cost and major disruptions to traffic and other daily community activities from separated collection systems. Several potential benefits of sewer separation might warrant its consideration in specific cases: 1. Eliminating CSOs and preventing untreated sanitary sewage from entering the receiving waters during wet weather. Sanitary sewage is a more

CSO retention basins (RBs) capture and store some of the excess combined sewer flow that would otherwise be bypassed to receiving waters. Stored flows are subsequently returned to the sewer system during dry weather, when the in-line flows are reduced and more capacity is available at the treatment facility. RBs can be designed to control both flow rate and water quality. Figure 2 shows an example of a multistage CSO RB that has some treatment capabilities. This facility handles peak flows by routing them through a mechanical bar screen and then pumping them into the first compartment. The main function of the first compartment is to allow primary settling and grit removal. If the flows continue to rise, the

Flushing water pump station

East side interceptor

Grand river

Ex. market ave. pumping station

Disinfection building

se we r

Sanitary sewers

Effluent sewer

First compartment

Cascade Parshall flume

Inf

lue

nt

Return sewers

Pump station

Effluent troughs

Gallery

Second compartment

Third compartment (chlorine contact tank)

Figure 2. Multistage CSOs RBs (3).

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COMBINED SEWER OVERFLOW TREATMENT

first compartment fills and then spills over into the second compartment. This compartment is designed specifically to store most of the overflow from the first compartment. The second compartment is also equipped with a floor wash system that flushes all settled sediments into a collection trough. If the flows continue to rise, the water spills over into a series of troughs, where sodium hypochlorite is applied for disinfection. The flow is then routed to a contact tank (third compartment), which eventually returns the water to the nearby surface waters (2). These are the primary concerns in the operation of RBs: 1. Managing flows to and from the retention basin. 2. Preventing the combined sewage from becoming septic or handling the wastewater appropriately after it has become septic. 3. Removing accumulated solids and floatables. 4. Disinfecting basin overflows to receiving waters. Swirl/Vortex Technologies Solids separation devices, such as swirl concentrators and vortex separators, have been used in Europe and (to a lesser extent) in the United States. These devices are relatively small, compact solids separation units with no moving parts. A typical vortex-type CSO solids separation unit is illustrated in Fig. 3. During wet weather, the

outflow from the unit is throttled, causing the unit to fill and to self-include a swirling vortex-like flow regime. Secondary flow currents rapidly separate settleable grit, as well as floatable matter. The concentrated foul matter is intercepted for treatment, whereas the cleaner, treated flow can be discharged to receiving surface waters. These devices are usually intended to operate under very high hydraulic flow regimes. Screening Generally, there are two types of bar screens, coarse and fine. Both are used at CSOs control facilities; with each type provides a different level of solids removal efficiency. Although there is no industrial standard for classifying screens based on aperture size, coarse bar screens generally have a 0.04 to 0.08 m clear spacing between bars, whereas fine screens generally have rounded or slotted openings of 0.3 to 1.3 cm clear space. Coarse Screens. Coarse screens are constructed of parallel vertical bars and are often referred to as bar racks or bar screens. In CSO control and treatment facilities, coarse screens are usually the first unit of equipment in the system. These screens are usually set at 0 to 30◦ from the vertical and are cleaned by an electrically or hydraulically driven rake mechanism that removes the collected material from the screen continuously or

Effluent launder

Liquid flow pattern Effluent

Influent (enters tangentially)

Figure 3. Cross section through a typical vortex-type solids separation device (1).

Concentrated solids

COMBINED SEWER OVERFLOW TREATMENT

785

Removable section Plate, weld to rack, Cover plate cut to fit channel Clearance line for bar rake 3 Bars in × 2 in 8 Notch rack bars and weld cross bars in notches Flow

Drainage plate

Bar clamp

Clip angle, Buttonhead bolts bolt to channel with stainless steel bolts

Masonry expansion anchor, stainless steel bolt Channel

Channel invert

Concrete fill Curve this section

Drainage plate

Figure 4. Diagram of trash rack used for treating CSOs (2).

periodically. The most common type of bar screen used at CSO control facilities is a trash rack. Trash racks typically have 0.04 to 0.08 m clear spacing between bars. Figure 4 is a diagram of a typical trash rack.

pipe, where they can be directed to the sanitary sewer line for ultimate removal at the wastewater treatment plant. Screen blinding is prevented by a hydraulically driven rake assembly.

Fine Screens. Fine screens at CSO facilities typically follow coarse bar screening equipment and provide the next level of physical treatment in removing the smaller solid particles from the waste stream. Both fixed (static) and rotary screens have been used in CSO treatment facilities. Fixed fine screens are typically provided with horizontal or rounded slotted openings of 0.02 to 1.27 cm. The screens are usually constructed of stainless steel in a concave configuration, at a slope of approximately 30◦ from the vertical. Flow is discharged across the top of the screen. The flow then passes through the slotted openings, and solids are retained on the screen surface. Solids are discharged from the screen surface by gravity and by washing onto a conveyor belt or other collecting system. Rotary fine screens include externally and internally fed screens. Externally fed screens allow the wastewater to flow over the top of the drum mechanism and through the screens surfaces, while collecting solids onto the screen surface. As the screen rotates, a system of cleaning brushes or sprayed water removes debris from the drum. Internally fed systems discharge wastewater in the center of the drum, allowing the water to pass through the screen into a discharge channel, while solids are removed from the screen surface by cleaning brushes or a water spray. In response to the need for solids and floatables control during storms, proprietary screen products, such as the ROMAGTM screen, have been designed for wet weather applications (Fig. 5). The ROMAGTM screen partitions the flow, sending screened flow to the CSO discharge point, while keeping solids and floatables in the flow directed toward the sanitary sewer. This screen works as follows: excess flow enters the screening chamber, flows over a spill weir, and proceeds through the screen into a channel, which discharges flow to a neighboring receiving waterbody. Floatables trapped by the screen move laterally along the face of the screen via combs/separators to the transverse end section of the

Netting Systems for Floatables Control Floatables control technologies are designed to reduce or eliminate the visible solid waste that is often present in CSO discharges. The Netting Trash-TrapTM system is a modular floatables collection system, located at the CSO outfall. It uses the passive energy of the effluent stream to drive the floatable materials into disposable mesh bags. These bags are suspended horizontally in the CSO flow stream within a support structure. The construction methodology and method of installation at the outfall are determined site-by-site. Ever since, several other end-of-pipe, but also in-line configurations have been developed and implemented. The standard nets used in the system are designed to hold up to 0.7 m3 of floatables and a weight of 227 kg each. For the floating units, the effluent stream and the collected floatables are directed into the bags by two floating booms and curtains, which run from the front corners of the pontoon to either side of the outfall, where they attach to a vertical piling that has a roller

Romag screen

Figure 5. ROMAGTM ‘‘combing’’ mechanical screen (vertical) for CSO floatables control (4).

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COMBINED SEWER OVERFLOW TREATMENT

mechanism or a shoreline support. This design allows the boom to float and accommodate changes in the water level. The extended curtains are weighted to conform to the water bottom. The maximum high water level, expected at the site, determines the depth of the curtain. Certain modifications of the outfall design may include adding structural support, attaching structural struts and strut anchor support, and adding foundations. Dissolved Air Flotation Dissolved air flotation (DAF) removes solids by introducing fine air bubbles into the CSO stream. The air bubbles can attach to solid particles suspended in the liquid, causing the solids to float toward the surface, from where they can be skimmed off. This technology has been tested in several CSO applications. A major advantage of DAF is its relatively high overflow rate and short detention time, which results in reduced facility size, compared to conventional sedimentation. Oil and grease are also more readily removed by DAF. Operating costs for DAF are high, due to larger energy demand, and skilled operators are usually required for efficient operation. Filtration Dual media high rate filtration has been piloted for treating of CSO flows. A two-layer bed, consisting of coarse anthracite particles on top of less coarse sand, was used. After backwash, the less dense anthracite remains on top of the sand. Filtration rates of 16 gal/ft2 /min or more were used, resulting in substantially smaller space requirements, compared with sedimentation. Demonstration test systems included pretreatment by fine-mesh screens. The addition of chemical coagulant agents improved the performance considerably. Filtration is more appropriately applied after pretreatment by fine screening. Operation may be automated but tends to be rather O & M intensive. Intermittently operated sand filters can also be used for CSO treatment and show considerable promise for larger scale operations (3). CHEMICAL TREATMENT (MAINLY DISINFECTION)

5. CSO outfalls are often located in remote areas and thus, may require automated controls for the disinfection systems. In addition to these problems, increased health and safety concerns for using chlorine to disinfect CSOs have prompted the development of alternative disinfectant agents/methods, which often present fewer problems and health hazards. Alternative methods to chlorine addition have been developed and evaluated for the continuously disinfecting wastewater discharges to small streams or sensitive waterbodies and are now also being considered for treating CSOs and other episodic discharges. These include the use of chlorine dioxide, the application of ozonation or of ultraviolet radiation, the addition of peracetic acid, and electron beam irradiation (e-beam) (5). Ozone Ozone is a strong oxidizer and is applied to wastewater as a gas mixture with air. Its use in CSO treatment facilities for wastewater disinfection is relatively new; few facilities are currently using ozone for disinfection. This can be attributed to the higher initial capital costs of ozone generating equipment. Ozone is equal or superior to chlorine in ‘‘killing’’ power for pathogenic microorganisms, but it does not cause the formation of harmful by-products (halogentated organics), as does chlorination (5). Ultraviolet Radiation UV radiation is electromagnetic radiation used for disinfection. UV disinfection incorporates the spectrum of light between 40 and 400 nm. Germicidal properties range between 200 and 300 nm; 260 nm is the most lethal. The primary method for using UV disinfection is to expose wastewater to a UV lamp. UV radiation is not a chemical disinfection method; it avoids the addition of chemical reagents, and it disinfects without altering the physical or chemical properties of water. However, UV efficiency is affected by the presence of suspended solids in the CSOs, which scatter and absorb light and lower the method’s efficiency. Thus, UV disinfection is not very effective for CSOs that containing high TSS (5).

Chlorine Chlorine has long been the disinfectant of choice for most disinfection systems. It offers reliable reduction of pathogenic microorganisms at reasonable operating costs. Disinfection by chlorine is the most common method used to kill pathogenic microorganisms at WWTPs, but this methodology may not be feasible at all CSOs for several reasons: 1. CSOs occur intermittently, and their flow rate is highly variable, thus making it difficult to regulate the addition of disinfectant. 2. CSOs have high concentrations of suspended solids. 3. CSOs can vary widely in temperature and bacterial composition. 4. Disinfectant residuals, following the use of chlorine, may be prohibited from entering receiving waters.

Peracetic Acid Peracetic acid (CH3 COOOH, denoted PAA), also known as ethaneperoxoic acid, peroxyacetic acid, or acetyl hydroxide, is a very strong oxidant. Based on limited demonstration data for disinfecting secondary treatment plant effluents, peracetic acid appears to be an effective disinfectant and should be further evaluated for treating CSOs. The equilibrium mixture of hydrogen peroxide and acetic acid that produces PAA is too unstable and explosive to transport; therefore, PAA must be produced on site. The decomposition of PAA results in the formation of acetic acid, hydrogen peroxide, and oxygen (5). Electron Beam Irradiation Electron beam irradiation (e-beam) uses a stream of high energy electrons that is directed into a thin film of water

COMBINED SEWER OVERFLOW TREATMENT

or sludge. The electrons break water molecules apart and produce a large number of highly reactive chemical species (mainly radicals), including oxidizing hydroxyl radicals, reducing aqueous electrons, and hydrogen atoms. These are the main disadvantages of this method: 1. Increased safety considerations due to the use of high-voltage technology and the generation of Xray radiation. 2. There is no full-scale application experience for CSOs. 3. High capital costs. 4. High O & M costs. 5. Thin process flow stream. 6. Sufficient pretreatment straining of influent is also required to remove most of suspended solids for efficient application of this system (5). BIOLOGICAL TREATMENT AND COMBINED SYSTEMS The use of biological treatment, combined with certain of the aforementioned physical–chemical treatment processes, for treating CSO presents certain serious limitations: 1. The biomass used to assimilate the nutrients in the CSOs must also be kept alive during dry weather, which can be rather difficult, except at an existing treatment plant. 2. Biological processes are subject to upset under to erratic loading conditions. 3. The land requirements for this type of treatment plant can be excessive near an urban area. 4. Operation and maintenance can be costly, and the facilities require highly skilled operators. Some biological treatment technologies are used in CSO control as elementary parts of a WWTP. Pump-back or bleed-back flows from CSO storage facilities commonly receive secondary (biological) treatment in the treatment plant, once wet weather flows have subsided. In a WWTP, which has maximized the wet weather flows that are

787

accepted, the hydraulic flows are sometimes split; only a portion of the primary treated flows is subjected to secondary treatment to avoid process upset. The split flows are blended again before the exit and disinfected appropriately for final discharge (1). Constructed Wetlands Constructed wetlands are artificial wastewater treatment systems, consisting of shallow (usually 1 m deep) beds, that have been planted with aquatic plants, and which rely upon natural microbial, biological, physical, and chemical processes to treat CSOs. They typically have impervious clay or synthetic liners and engineered structures to control the flow direction, liquid detention time, and water level. Depending on the specific type of system, they may contain inert porous media, such as rock, gravel, or sand. Constructed wetlands have been classified in the literature into two types. Free water surface (FWS) wetlands (also known as surface flow wetlands) closely resemble natural wetlands in appearance; they contain aquatic plants, and water flows through the leaves and stems of the plants. Vegetated submerged bed (VSB) systems (also known as subsurface flow wetlands) do not resemble natural wetlands because they have no standing water. Wastewater (i.e., CSOs) stays beneath the surface of the medium, flows in contact with the roots and rhizomes of the plants, and is not visible or available to wildlife. Finally, the term vertical flow wetland is used to describe a typical vertical-flow sand or gravel filter, which has been planted with aquatic plants. Successful operation of this type of system depends mainly on its operation as a filter (i.e., frequently applying dosing and draining cycles) (7). Figure 6 shows a typical cross section of a horizontal subsurface flow wetland (also known as a reed bed). These systems are used to treat the excess combined sewer flow that would otherwise be bypassed to receiving waters, with good treatment efficiency that meets the tight permitting conditions. During exceptionally dry weather, secondary or tertiary treated effluent can be diverted to this system to conserve the plantings and the microorganism’s population in the system (6).

Phragmites

Design

Inlet Trough with adjustable weirs

Level device

Outlet

5−10 mm gravel

Impermeable membrane

Figure 6. Diagrammatic longitudinal section of a horizontal subsurface flow wetland (6).

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BIOLOGICAL PHOSPHORUS REMOVAL IN THE ACTIVATED SLUDGE PROCESS

BIBLIOGRAPHY 1. U.S. Environmental Protection Agency. (1993). Manual: Combined Sewer Overflow. Report EPA/625/R-93/007. Cincinnati, OH. 2. Metcalf & Eddy, Inc. (1991). Wastewater Engineering: Treatment, Disposal and Reuse. McGraw-Hill, Singapore. 3. U.S. Environmental Protection Agency. (1999). Combined Sewer Overflow Technology Fact Sheet Retention Basins. Report EPA/832/F-99/042. Washington, DC. 4. Pisano, W.C. (1995). Comparative assessment: vortex separators, rotary sieves, and ‘‘Combing’’ screens for CSO floatable control. Presented at the Water, Environ. Fed. Annu. Conf., Miami, FL. 5. U.S. Environmental Protection Agency. (1999). Combined Sewer Overflow Technology Fact Sheet Alternative Disinfection Methods. Report EPA/832/F-99/033, Washington, DC. 6. Green, M.B. and Upton, J. (1995). Constructed reed beds: Appropriate technology for small communities. Water Sci. Technol. 332: 339–348. 7. U.S. Environmental Protection Agency. (2000). Manual: Constructed Wetlands Treatment of Municipal Wastewaters. Report EPA/625/R-99/010, Cincinnati, OH. 8. Prochaska, C.A. and Zouboulis, A.I. (2003). Performance of intermittently operated sand filters: A comparable study, treating wastewaters of different origins. Water, Air, Soil Pollut. 147: 367–388.

BIOLOGICAL PHOSPHORUS REMOVAL IN THE ACTIVATED SLUDGE PROCESS MICHAEL H. GERARDI Linden, Pennsylvania

Phosphorus often is the limiting nutrient that promotes the excessive growth of aquatic plants, especially algae, and excess quantities of phosphorus often are present in the effluent of activated sludge processes in quantities greater than those required for the growth of aquatic plants. Excess phosphorus promotes not only the undesired growth of aquatic plants but also the undesired impacts of the death of the aquatic plants upon the receiving water. Therefore, state and federal regulatory agencies limit the quantity of phosphorus in the effluent. The requirement limiting the quantity of phosphorus in the effluent of activated sludge processes is becoming more and more stringent for municipal wastewater treatment plants. For example, discharge limits for total phosphorus of 2 mg/L or lower have been applied broadly to many plants in the lower Susquehanna River Basin. Several environmental concerns related to the excessive growth of aquatic plants include clogging of receiving waters and the production of color, odor, taste, and turbidity problems when the receiving waters are used for potable water supplies. The die-off of large numbers of aquatic plants results in oxygen depletion when the plants decompose. Those portions of the aquatic plants that do not decompose accumulate in the receiving waters and contribute to eutrophication or rapid aging of the receiving waters. Additionally, some algae release toxic compounds. Municipal wastewater contains 10–20 mg/L of total phosphorus. The total phosphorus consists of inorganic

and organic phosphorus. Inorganic phosphorus consists of phosphorus-containing compounds that do not contain carbon and hydrogen. Organic phosphorus-containing compounds do contain carbon and hydrogen. Significant and common inorganic phosphorus-containing compounds are orthophosphate (PO4 -P) and polyphosphates. Orthophosphate makes up approximately 50–70% of the total phosphorus in influent and approximately 90% of the phosphorus in the effluent of municipal wastewater treatment plants. Orthophosphate is the preferred phosphorus nutrient for aquatic plants. The forms of orthophosphate found in the influent and effluent of wastewater treatment plants are pH dependent. At pH values 7. The forms of orthophosphate in the influent and effluent are produced through dissociation (Eq. 1). 2− + −

H2 PO4 − −  −− − − HPO4 + H

(1)

Significant and common organic phosphorus-containing compounds include phytin, nucleic acids, and phospholipids. Phytin is an organic acid found in vegetables such as corn and soybean. Phytin is difficult to digest and is found in domestic wastewater. Nucleic acids are large complex molecules that contain genetic material. Phospholipids also are large and complex molecules that are used in the production of structural materials. Phytin, nucleic acids, and phospholipids degrade slowly in the activated sludge process. Their degradation results in the release of orthophosphate. It is the orthophosphate that is used as the phosphorus nutrient by bacteria and incorporated into cellular material or MLSS (mixed liquor suspended solids) in the activated sludge process. The degradation of organic compounds in activated sludge processes and the incorporation of phosphorus into new cellular material are achieved by a large and diverse population of bacteria. However, there are some bacteria that are capable of removing and storing phosphorus in quantities larger then their cellular needs. Bacteria that are capable of removing phosphorus in excess quantities are known as ‘‘phosphorus-accumulating organisms’’ (PAO) or ‘‘poly-P bacteria’’ (Table 1). Of all poly-P bacteria, Acinobacter is the most commonly known and studied. Phosphorus is removed from the wastewater in the orthophosphate form and stored by the poly-P bacteria as polyphosphate granules. Removal of phosphorus by poly-P bacteria often is termed ‘‘luxury uptake of phosphorus.’’ There are several operational measures that can be used in activated sludge processes to remove phosphorus

Table 1. Genera of Wastewater Bacteria that Contain Poly-P Species Acinobacter Aerobacter Aeromonas Arthrobacter Beggiatoa Enterobacter

Escherichia Klebsiella Moraxella Mycobacterium Pasteurella Pseudomonas

BIOLOGICAL PHOSPHORUS REMOVAL IN THE ACTIVATED SLUDGE PROCESS Table 2. Biological and Chemical Measures Available for Phosphorus Removal Operational Measure Chemical precipitation

Assimilation

Biological phosphorus removal

Biological/chemical techniques

Anaerobic tank

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Aerobic tank

Description Use of alum, ferric chloride, ferrous sulfate, or lime at a high pH; results in increased operational costs for chemical addition and disposal of chemical sludge Phosphorus incorporated into cellular material (MLSS) as 1–3% dry weight and wasted from the activated sludge process Phosphorus incorporated into cellular material (MLSS), especially poly-P bacteria, as 6–7% dry weight and wasted from the activated sludge process Precipitation of phosphorus released by poly-P bacteria under anaerobic condition and chemically precipitated with alum, ferric chloride, ferrous sulfate, or lime

from the effluent. These measures include biological and chemical techniques (Table 2). Phosphorus as orthophosphate can be removed from wastewater through its chemical precipitation as a metal salt at a high pH. Chemical compounds commonly used to precipitate orthophosphate include alum, ferric chloride, ferrous sulfate, and lime. Orthophosphate also can be removed through its incorporation as cellular material. By increasing the MLSS, more orthophosphate is incorporated into bacterial cells. However, the amount of phosphorus in the bacterial cells is only 1–3% by dry weight. Orthophosphate also can be removed by combined biological and chemical techniques and through biological phosphorus removal. Of the biological and chemical measures available for phosphorus removal, biological phosphorus removal offers several advantages when compared to other measures. For example, biological phosphorus removal is relatively inexpensive due to the reduction in chemical costs and sludge disposal costs associated with chemical addition to precipitate orthophosphate. Biological phosphorus removal is also capable of removing phosphorus to low effluent concentrations. Biological phosphorus removal or luxury uptake of phosphorus occurs when orthophosphate uptake by poly-P bacteria exceeds cellular requirements. If luxury uptake of phosphorus does not occur in an activated sludge process, the phosphorus content of the activated sludge is approximately 1–3% on a dry weight basis. If luxury uptake of phosphorus does occur, the phosphorus content of activated sludge is approximately 6–7%. Bacteria that are capable of luxury uptake of phosphorus enter activated sludge processes in fecal waste and through inflow and infiltration as soil and water bacteria. These bacteria are unique and remove phosphorus in excess of cellular needs in the presence of rapidly degradable organic compounds when transferred from an anaerobic (fermentative) tank to an aerobic tank.

Figure 1. Microbial activity in the anaerobic tank. In the anaerobic tank, soluble cBOD is fermented in the absence of free molecular oxygen and nitrate ions. The fermentation process produces a variety of fatty acids. The acids are rapidly absorbed by the poly-P bacteria and stored as an insoluble starch (PHB). In order to absorb the fatty acids and store them as starch, energy in the form of orthophosphate is released by the poly-P bacteria to the bulk solution.

Table 3. Soluble Fatty Acids Produced in the Anaerobic Tank Fatty Acid Formic acid Acetic acid Propionic acid Butyric acid Valeric acid Caproic acid

Formula HCOOH CH3 COOH CH3 CH2 COOH CH3 CH2 CH2 COOH CH3 CH2 CH2 CH2 COOH CH3 CH2 CH2 CH2 CH2 COOH

In the anaerobic tank (Fig. 1) that has a retention time of 1–2 h, fatty acids are produced in large quantities through fermentation (Table 3). Fermentation is the microbial degradation of organic compounds without the use of free molecular oxygen (O2 ) or nitrate ions (NO3 − ). Poly-P bacteria quickly absorb the fatty acids produced through fermentation. Although the fatty acids are absorbed, the acids are not degraded. Instead, the soluble fatty acids are stored in the form of insoluble starch granules (poly-β-hydroxybutyrate or PHB). The conversion of fatty acids to insoluble starch granules and the storage of the granules requires the expenditure of energy by the poly-P bacteria. The expenditure of energy results in a release of orthophosphate from the poly-P bacteria into the fermentative tank. With the production of PHB in the poly-P bacteria, the fermentative tank contains two ‘‘pools’’ of phosphorus—influent phosphorus and poly-P bacteria released orthophosphate. In the aerobic tank (Fig. 2) that has a retention time of 1–2 h, PHB granules are solubilized and degraded with the use of free molecular oxygen. The degradation of PHB granules results in the release of a large quantity of energy that is captured and stored by the poly-P bacteria. The energy is stored in the bacteria in the form of insoluble phosphate granules or voluntin. Phosphorus is removed from the activated sludge process when MLSS (bacteria) is wasted. There are several processes available for the biological removal of phosphorus. Often biological phosphorus removal is combined with nitrification and denitrification. Nitrification is the oxidation of ammonium ions (NH4 + ) to nitrate ions (NO3 − ), while denitrification is the reduction

790

BIOLOGICAL PHOSPHORUS REMOVAL IN THE ACTIVATED SLUDGE PROCESS Anaerobic tank

Aerobic tank

Anaerobic zone

Figure 2. Microbial activity in the aerobic tank. In the aerobic tank, PHB is solubilized and degraded in the presence of free molecular oxygen. The degradation of PHB results in the production of carbon dioxide, water, and new bacterial cells. Phosphorus released in the anaerobic tank as well as phosphorus present in the waste stream are absorbed by poly-P bacteria and stored in phosphorus granules.

of nitrate ions to molecular nitrogen (N2 ) and nitrous oxide (N2 O). Nitrification and denitrification are responsible for the biological removal of nitrogen from wastewater. When nitrification and denitrification are combined with biological phosphorus removal, these three biological processes are known as biological nutrient removal (BNR). There are five significant processes used for biological phosphorus removal or BNR. Several of these processes are proprietary. These processes are the A/O, Phostrip, A2O, Bardenpho, and UCT. Biological nutrient removal processes are either mainstream or sidestream (Table 4). A mainstream process contains an anaerobic tank within the major wastewater flow from influent to effluent. A sidestream process contains an anaerobic tank outside the major wastewater flow. Of the nutrient removal processes, two are designed to remove phosphorus only. These processes are the A/O and the Phostrip. The A/O (anaerobic/oxic) process is a mainstream process (Fig. 3). The A/O process is patented by Air Products and Chemicals, Incorporated and is similar to a conventional activated sludge process. The Phostrip process is a sidestream process and includes biological and chemical phosphorus removal (Fig. 4). A stripper tank is included in the Phostrip process. This tank has an anaerobic condition where phosphorus is released by poly-P bacteria from the return activated sludge (RAS). The released phosphorus is removed from the stripped tank by elutriation water. Lime is added to the elutriation water as it leaves the stripper tank. The addition of lime results in the precipitation of phosphorus as calcium phosphate.

Oxic zone

Figure 3. The A/O process.

Oxic zone

A/O Phostrip A2/O Bardenpho UCT

Secondary clarifier

Influent

Ras

Was

Overflow to additional treatment

Anaerobic stripper Stripped sludge

Figure 4. The Phostrip process.

The A2/O process consists of an anaerobic zone, an anoxic zone, and an oxic zone. As wastewater and bacteria move through these three zones, phosphorus is removed biologically and nitrogen is removed through nitrification and denitrification. Fermentation occurs in the anaerobic zone and phosphorus is released to the bulk solution by poly-P bacteria. In the anoxic zone nitrate ions are used (denitrified) by facultative anaerobic bacteria (denitrifying bacteria) to degrade soluble carbonaceous BOD. In the oxic zone ammonium ions (NH4 + ) in the wastewater and ammonium ions released from nitrogen-containing compounds are oxidized (nitrified) to nitrate ions (NO3 − ). The Bardenpho process is licensed and marketed in the United States by Eimco Process Equipment Company. The Bardenpho process includes five zones (anaerobic zone, anoxic zone, oxic zone, anoxic zone, and oxic zone). The University of Capetown or UCT process also contains anaerobic, anoxic, and oxic zones. However, the UCT

Table 4. Nutrient Removal Processes Process Name

Secondray clarifier

Nutrient Removed

Process

Nitrogen

Phosphorus

Mainstream X

X X X

X X X X X

X X X

Sidestream

Chemical Precipitation

X

X

PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION

process is designed to reduce the quantity of nitrate ions returned to the anaerobic zone in order to optimize the phosphorus removal. As phosphorus and nitrogen discharge limits become more stringent for activated sludge processes in the United States, these nutrient removal systems will become more popular. The choice of the biological nutrient removal system is based on cost, wastewater composition, and nutrient removal requirements.

PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION RAFFAELE MOLINARI Universita` della Calabria Rende, Italy

L. PALMISANO Universita` di Palermo Palermo, Italy

INTRODUCTION General Potable water, industrial water, and wastewater are often polluted by toxic organic species. Some classical methods (1,2) to cleanup waters, before sending them to rivers or to municipal drinking water supplies, transfer pollutants from one phase to another, hence creating further waste streams. New methods, such as those involving heterogeneous photocatalytic reactions, allow in many cases a complete degradation of organic pollutants to small and non-noxious species, without using chemicals, thus avoiding sludge production and its disposal. Membrane separation processes, thanks to the selective property of membranes, have already proved to be competitive with other separation processes (3–7). Photocatalytic membrane reactors (PMRs) have some advantages compared to conventional photoreactors. Indeed, confining the photocatalyst to the reaction environment as a result of the presence of the membrane enables operation with high amounts of catalyst, control of the residence time of the molecules in the reactor, and realization of a continuous process with simultaneous product(s) separation from the reaction environment. Some PMR configurations using membranes ranging from microfiltration (MF) to reverse osmosis (RO) have been investigated (8–13). The influence of various operating parameters on the photodegradation rate of pollutants present in aqueous effluents by means of discontinuous and continuous photocatalytic processes in the presence of NF membranes has been reported (14). Moreover, the possible use of solar radiation (15–18) in PMRs is of particular interest as the energy cost is one of the main drawbacks for industrial applications. Although many papers on photocatalysis have been published, the cases where membrane and photocatalyst are coupled are very few.

791

Historical Background The use of photocatalysis in waste treatment is usually based on the electronic excitation of a polycrystalline semiconductor caused by light absorption that drastically alters its ability to lose or gain electrons, promoting decomposition of pollutants to harmless by-products. Photocatalytic processes in liquid phase have been applied to the degradation of several organic compounds (19–24). Only a few compounds such as chlorofluorocarbons, trifluoroacetic acid, and 2,4,6-trihydroxy-1,3,5-triazabenzene (cyanuric acid) cannot be completely degraded by photocatalytic methods (24). The formation of transient by-products, more toxic than the starting substrates, could be observed in some cases. The research on types of reactors that can be used in photocatalysis is very active, and slurry photocatalytic reactors suitable for the requirement of continuous operation are described in the literature (25,26). Nevertheless, as the chemical industry is characterized by an almost exclusive use of continuous processes, a photocatalytic powder for potential application purposes should possess a suitable size and mechanical characteristics in addition to good catalytic properties. To date, only rare examples of pilot plant photoreactors have been reported, as the difficulty of making a clear assessment of the costs of the photocatalytic processes (which are typically efficient only in dilute systems) have prevented a wider development of them from an application perspective. Some authors have immobilized the semiconductor on Pyrex glass sheets (27,28), on aerogels (29), or on particles having high surface area (e.g., alumina or silica gel) (30,31). In all these methods, a drawback is the mass transfer resistance of the reacting species, which could control the reaction rate. Despite the potential advantages of using hybrid membrane photoreactors, the research on coupling photocatalysis and membranes is not yet sufficiently developed. Some papers reported in the literature concern the use of cellulose microporous membranes (32); porphyrin containing membranes (33); porphyrins immobilized as photosensitizers on sulfocationic membranes (34); filtration coupled to catalysis (35,36); and TiO2 immobilized inside the membrane (37), physically deposited on the membrane surface (38), or confined in suspension by means of the membrane (8,9,10,39). FUNDAMENTALS OF HETEROGENEOUS PHOTOCATALYSIS A simple definition of heterogeneous photocatalysis implies only the acceleration of a photoreaction by the action of a solid catalyst, which may interact with the species to be degraded and/or with the intermediates, depending on the reaction mechanism. The catalytic nature of the process should be demonstrated by checking that the turnover number (TON) is greater than 1. TON can be defined as the ratio of the number of photoinduced transformations for a given period of time to the number of photocatalytic sites. The total surface area of the solid photocatalyst can be considered when the number of sites is unknown, but figures of TON obtained in this way are lower limits. The most widely used semiconductor

792

PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION

Energy O2 UV e−

Excitation

Recombination

CB

VB

OH + organics

Photoreduction O2

TiO2

h+

H2O Photooxidation

OH + H+ CO2+ H2O

Figure 1. Basic photocatalytic mechanism occurring when a semiconductor particle (e.g., TiO2 ) is irradiated.

photocatalyst is polycrystalline TiO2 , in the allotropic phases of both anatase and rutile (the anatase phase appears generally more photoactive), although the use of many other solids (ZnO, WO3 , CdS, MoS2 , CdSe, Fe2 O3 , etc.) has also been reported (40–42). When a semiconductor is illuminated with light of suitable energy (i.e., greater than its bandgap), electrons are promoted from the valence band (vb) to the conduction band (cb), acquiring the reducing power of the cb energy; positive holes are created in the vb, acquiring the oxidation power of the vb energy. This photoproduced pair can give rise to recombination with emission of thermal energy and/or luminescence or to redox (reduction–oxidation) reactions involving electron acceptor or donor species adsorbed on the surface of the catalyst particles (Fig. 1). Consequently, the recombination rate of the photoproduced electron–hole pairs should be as low as possible in order to favor their availability on the surface of the catalyst particle. A photocatalyst should possess some essential characteristics: (1) light absorption should occur in the near UV and possibly in the visible wavelength ranges; (2) the stability should be such that its re-utilization is possible; and (3) some thermodynamic and kinetic constraints should be fulfilled (43). The photoreactivity depends not only on the intrinsic electronic characteristics of the photocatalysts, but also on their structural, textural, and surface physicochemical features. Among them we can cite the

vb and cb energies, the bandgap value, the lifetime of photogenerated electron–hole pairs, the crystallinity, the allotropic phase, the particle sizes, the presence of defects and dopants, the specific surface area, the porosity, the surface hydroxylation, and the surface acidity and basicity. The relative weight of their importance for the studied reaction determines the final level of the photoreactivity. For this reason, it is often difficult to explain and to discuss exhaustively the observed photoreactivity trends by considering only a few properties of the photocatalysts. The steps shown in Table 1 occur when polycrystalline TiO2 is used in aqueous medium in the presence of O2 and a generic substrate. The oxidant radicals along with the holes can give rise to oxidant attacks on a wide variety of substrates. The presence of O2 is essential in order to trap photoproduced electrons, improving the charge separation and consequently the availability of the holes. The application of heterogeneous photocatalysis to the purification of aqueous effluents containing dissolved organic and/or inorganic species (e.g., CN− ) has been widely studied. Nevertheless, it is worth noting that analytical studies to check if noxious intermediates are produced under irradiation are essential when the scale-up of photoreactors is proposed for application purposes. FUNDAMENTALS OF MEMBRANE PROCESSES Membrane processes are separation methods at the molecular level that have received more interest in recent years for their possibility of being used in many industrial applications. The main goals of such processes are concentration of a solute by removing the solvent, purification of a solution by removing nondesirable components, and fractionation of liquid or gaseous mixtures. Membrane separation processes offer interesting opportunities in pollution control, in the production of drinking water (44), and in the treatment of industrial wastewater (45). A membrane can be defined as a selective barrier between two phases (46,47); it can be thin or thick, natural or synthetic, neutral or charged, the structure can be homogeneous or heterogeneous; and the mass transport can be active or passive. In the last case the driving force can be due to a difference of pressure, concentration, or temperature. The driving force has the capacity to transport a component more rapidly

Table 1. Some Essential Steps When Polycrystalline TiO2 Is Used in Photodegradation Processes of Noxious Organic and Inorganic Substrates Role of TiO2 + TiO2 + hν → TiO2 (e− (cb) + h(vb) ) ž OH− + h+ → OH (vb)

Role of O2 − ž O 2 + e− (cb) → O2 ž O2 − + H+ → ž HO2 2ž HO2 → O2 + H2 O2 H2 O2 + ž O2 − → OH− + ž OH + O2

Substrate Degradation Substrate + ž HO2 → products Substrate + ž OH → products Substrate + h+ (vb) → products

PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION

than others owing to different physical and/or chemical properties between the membrane and the components. The membrane is assembled in a module whose geometry is generally plane or cylindrical and in which the feed is separated into two streams called retentate (the treated feed) and permeate. Each membrane process is characterized by the employement of a particular type of membrane. Some processes are microfiltration, ultrafiltration, nanofiltration, and reverse osmosis (48) in decreasing order of particles size and increasing pressures (0.1–0.2 to 50–100 bar) used as driving forces. Other membrane processes concern the separation of ionic species by applying an electrical potential (electrodialysis), the separation of mixtures of volatile liquids (pervaporation), and the separation of water from nonvolatile solutes (membrane distillation) by means of a temperature difference (49). SOME CASE STUDIES OF PHOTOCATALYTIC MEMBRANE REACTORS USED TO PHOTODEGRADE POLLUTANTS Preliminary Remarks The pioneering studies for coupling photocatalysis and membrane separations focused on the optimization of the system configuration (14). A continuous operation system (feed feeding and permeate withdrawing) was reported by Molinari et al. (39) involving the use of TiO2 particles in suspension (Fig. 2). This configuration seems appropriate for industrial applications, so experimental results deriving from this system will be mostly reported in the following. Membrane rejection of the target species

L M

M

M B

R F

H

F

A

P Sa

C

793

with and without photodegradation was measured to obtain information on the performance of the system. Degussa P25 TiO2 (specific surface area ∼ = 50 m2 /g; crystallographic phase, ∼80% anatase and 20% rutile) was used as the photocatalyst. Some of the molecules chosen as model pollutants were 4-nitrophenol (4-NP) and two dyes—patent blue (C27 H31 N2 NaO6 S2 ) and congo red (C32 H22 N6 Na2 O6 S2 ). The extent of the degradation of the contaminant was determined by UV-visible measurements and total carbon (TC), total organic carbon (TOC), and total inorganic carbon (TIC) determinations. 4-Nitrophenol Degradation in Batch and Continuous Membrane Photoreactors When NF-PES-010 or N30F nanofiltration membranes were tested (the first one was the most permeable), the permeate fluxes through the membranes in the presence of suspended catalyst were a little lower than those found in the absence of catalyst. Deposition of the catalyst on the membrane was mimimized by using a cell geometry that guaranteed turbulence and presence of vortexes in the bulk of the solution above the membrane. Photodegradation of 4-NP in batch (recycle) and continuous system configurations showed, for both types of membranes, bell-shaped curves of permeate concentration as a function of time. The concentration of 4-NP in the permeate was attributed to three factors: rejection, photocatalytic degradation, and adsorption. The increase of the initial concentration of permeate in the bell-shaped curves was lowered by the photocatalytic degradation that was responsible for the decrease of the concentration both in the retentate and in the permeate. Although 4-NP concentration in the retentate for the continuous configuration decreased less quickly than for the discontinuous one, the continuous system seems more promising for industrial application (39). In this system the optimum choice of the ratio between the irradiated volume and the total volume, Vi /Vt , was important. When the total suspension volume was increased from 400 to 700 mL, for instance, Vi /Vt increased owing to a constant recycle volume and, consequently, 4-NP abatement was higher due to an increased percentage of irradiated with respect to recycled suspension. The UV radiation mode was also important. The immersed lamp was found to be three times more efficient than the external lamp. Indeed, 99% w/w 4-nitrophenol degradation was achieved after about 1 h in the first case, whereas about 3 h were needed in the second case (39). Photodegradation of Other Pollutants

Sp P

Figure 2. Scheme of a continuous membrane photoreactor system with suspended catalyst. A, oxygen cylinder; B, recirculation reservoir (reactor); C, thermostatting water; L, UV lamp; M, pressure gauge; F, flowmeter; R, membrane cell; H, magnetic stirrer; P, peristaltic pump; Sa, feed reservoir; Sp, permeate reservoir (39).

NF-PES-010 and NTR-7410 nanofiltration membranes were tested in degradation runs after determination of their permeability and rejection for a variety of pollutants (50). It was found that membranes hold both catalyst and pollutants, but the NTR-7410 membrane tested at 8 bar gave a higher water permeate flux (105 L/h · m2 ) than the NF-PES-010 one (30 L/h · m2 ). The NTR-7410 membrane was also able to retain small molecules carrying negative charges (like the membrane)

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PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION

as a result of Donnan exclusion and, at 6 bar, yielded fluxes ranging between 20 and 40 L/h · m2 . The results of rejection tests and photodegradation studies of humic acids, patent blue dye, and 4-NP in a system utilizing this membrane are described in the following. The NTR-7410 membrane, not surprisingly, provided 100% retention of humic acids as they are made up of oligomers with molecular sizes greater than the membrane cutoff (600–800 g/mol). For patent blue (molar weight = 567 g/mol and initial concentration 10 mg/L), membrane rejection was about 78.6%. With respect to 4-NP, 0% rejection occurred at pH 6.75, while at pH 11 a rejection of about 77% and a negligible adsorption were observed. The pH dependence of rejection of 4-NP is to be expected by taking into account its acid–base properties and the resultant electrostatic interaction between 4-NP and the membrane (51,52): − + −

C6 H4 NO2 OH + H2 O −  −− − − C6 H4 NO2 O + H3 O −

+

−−

R—SO3 H + H2 O  −− − −

RSO− 3

+

+ H3 O

Humic acids are found in many natural waters (53) and often induce fouling problems when membranes are used. The interaction between humic acid and the particulate TiO2 photocatalyst has been explored by Lee et al. (54) who found that (1) humic solution or TiO2 suspension alone produced essentially a constant membrane flux of 160–170 L/h · m2 and (2) the mixture gave initial flux of 145 L/h · m2 that decreased to 101 L/h · m2 (a decrease of approximately 30%) in the first 30 min of operation. When the photoreactor was illuminated, a constant flux was measured during 6 h, consistent with destruction of humic acid. These authors proposed that the photocatalyzed conversion of humic acid into smaller and/or less absorptive species was occurring and concluded that ‘‘photocatalytic reactions appear to be attractive for the control of fouling materials such as natural organic matter.’’ Continuous Membrane Photoreactor at High Pollutant Concentrations The detrimental effect caused by a high pollutant concentration could be minimized by taking advantage of the ability of the membrane to retain both the catalyst and the pollutant. In order to investigate this, the photocatalytic system was tested in a continuous process with an initial concentration of pollutant in the photoreactor equal to zero, that is, in the presence of distilled water. A concentrated solution was continuously fed with a flow rate equal to that of the removed permeate. This solution was immediately diluted in the reactor, so photodegradation was effective at low pollutant concentration, although it was continuously fed at high concentration. Control of the residence time of the pollutant in the reactor was possible with the result that very low concentrations in the permeate could be obtained. This approach was tested by performing humic acid photodegradation studies (50) with an initial concentration in the photoreactor equal to zero and feeding a

200-mg/L solution. It was possible to maintain steadystate pollutant concentrations lower than 5 mg/L and 2 mg/L in the retentate and in the permeate, respectively. It was observed that the humic acid rejection was not 100% and this was mainly due to the lower size of the humic acid oligomers produced during the photodegradation process. The continuous process was also tested for high concentration (500 mg/L) feeds of patent blue dye and 4-NP. The photodegradation rate of patent blue was found to be lower than that obtained for other pollutants, possibly because of adsorption of the acid dye on the amphoteric catalytic surface, preventing UV light absorption. Indeed, at the end of the run, the catalyst was a dark blue color. Degradation of Dyes in the Continuous Membrane Photoreactor In order to achieve a better control of the residence time of pollutants, such as the dyes during the photodegradation process, a hybrid photoreactor was used in which the nanofiltration membrane was able to confine selectively dyes (congo red and patent blue) and catalyst in the reaction ambient while the permeate was withdrawn (55). The experimental results of runs carried out with patent blue in the membrane photoreactor with suspended catalyst showed that the photodegradation reaction followed pseudo-first-order kinetics (observed rate constant equal to 3.76 × 10−3 min−1 ). A similar run carried out in the absence of membrane showed an observed rate constant of 1.02 × 10−2 min−1 . The lower reaction rates for both dyes obtained by using the membrane with respect to that obtained in its absence were due to the smaller volume of irradiated suspension (320 against 500 mL) because a part (180 mL) circulated in the pipes of the plant and in the membrane cell. The possibility of successfully treating highly concentrated solutions of both dyes was examined, allowing the setup shown in Fig. 2 to work as a continuous system. In particular, the transient condition in the membrane photoreactor was studied by separating the effects of accumulation, adsorption, and photodegradation. Results of three runs with congo red for which the initial concentration of pollutant inside the photoreactor was zero are reported in Fig. 3. The first run, carried out in the absence of UV light and photocatalyst, indicated that the initial rate of dye accumulation in the photoreactor was 0.151 mg/min. The second run, carried out in the absence of UV light but in the presence of TiO2 , indicated that no increase of dye concentration occurred in the photoreactor during the first 45 min of continuous working of the plant because the congo red feed was adsorbed onto the catalyst surface. After the active sites of the catalyst were saturated, the concentration of the dye in the retentate increased linearly with an accumulation rate of 0.136 mg/min, very close to that determined for the first run. In the third run the continuous degradation of congo red in the presence of UV light and catalyst was performed and concentrations in the retentate and in the permeate versus irradiation time are reported. It can be noticed that, due to the concurrent

PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION

795

25

Concentration [mg/I]

20

(1) (3)

(2)

15

10

5 (4) 0

0

50

100

150

200 250 Time [min]

300

350

effect of dilution, adsorption, and photodegradation, the accumulation rate in the retentate was lower than that observed during the other two runs. The concentration of congo red in the permeate was virtually zero because the membrane maintained the substrate in the reacting ambient. The permeate flux (Jp ) throughout all the runs decreased from the initial value of 74.2 L/m2 · h to the value of 29.8 L/m2 · h and, consequently, the dye feeding rate also decreased. It is worth noting that the average photodegradation rate (0.274 mg/min) calculated for the overall run was lower than the average feeding rate (0.416 mg/min), calculated for the first 180 min, while it was 0.049 mg/min higher for longer time (average feeding rate 0.225 mg/min). Degradation of patent blue was also tested in the continuous system under the same experimental conditions used for congo red and the results are reported in Fig. 4. During the transient state (250–300 min in this specific case) the dye accumulated in the photoreactor because the photodegradation rate (0.570 mg/min) was lower than the feeding rate (0.863 mg/min). Subsequently, steady-state conditions were achieved, owing to the lower permeate flowrate, and no difference was observed between photodegradation and feeding rates. The rejection

400

450

Figure 3. Concentration of congo red in the retentate and in the permeate versus time for three different continuous runs. (1) Absence of UV light and TiO2 ; (2) absence of UV light and presence of TiO2 ; (3) presence of UV light and TiO2 for retentate; and (4) presence of UV light and TiO2 for permeate. V = 500 mL; T = 303 K; C0 = 0 mg/L; C(O2 ) = 22 ppm; TiO2 amount = 1 g/L; Cfeed = 500 mg/L; initial permeate flux, Jpin = 74.2 L/(m2 · h); final permeate flux, Jpfin = 29.8 L/(m2 · h); lamp, 125-W medium pressure Hg immersed lamp; initial pH = 6.42; membrane, NTR-7410; P = 3.5 bar.

of NTR-7410 membrane at steady-state conditions with respect to patent blue was 44.6%. Use of the membrane was beneficial because in addition to its role as a barrier for the catalyst, the product [i.e., cleaned up water (the permeate)] contained a very low concentration of dye with respect to the feed. It was approximately 1% in the case of congo red and approximately 11% in the case of patent blue with respect to 500 mg/L of the feed. It is worth noting that the concentration of the product corresponded to that of the retentate (approximately 3% for congo red and approximately 22% for patent blue) if the membrane was not used. CONCLUSIONS The continuous membrane photoreactor that combines both the advantages of classical photoreactors (catalyst in suspension) and membrane processes (separation at molecular level) appears very promising. Photocatalytic degradation can be carried out in reasonable times due to the high irradiated surface area of the suspended particles in the batch.

120 (1) Concentration [mg/l]

100 80 (2) 60 40 (1) Retentate (2) Permeate

20 0

0

50

100

150

200

250

Time [min]

300

350

400

450

Figure 4. Concentration of patent blue in the retentate and in the permeate versus time for a continuous photodegradation run carried out in the setup showed in Fig. 2. V = 500 mL; T = 303 K; C0 = 0 mg/L; C(O2 ) = 22 ppm; TiO2 amount = 1 g/L; Cfeed = 500 mg/L; regime permeate flux, Jpregime = 78.4 L/(m2 · h); lamp, 125-W medium pressure Hg immersed lamp; initial pH = 5.61; membrane, NTR-7410; P = 7.0 bar.

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PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION

Properly selected membranes should have both the capability to retain the catalyst and to partially reject organic species, enabling control of the residence time in the reacting system. In studies of 4-nitrophenol degradation, three factors—rejection, photocatalytic degradation, and adsorption—were found to contribute to maintain the steadystate 4-NP concentration in the permeate at very low values. The adsorption phenomenon is particularly important when oscillating concentrations of pollutant are fed to the membrane photoreactor, resulting in a negligible variation of concentration in the permeate. Photoreactors with an immersed lamp are generally more efficient than systems with external lamp. In addition, the pH of the polluted water, the molecular weight of the pollutant, and the type of pollutant and membrane are variables influencing pollutant rejection as a result of charge repulsion (Donnan exclusion) effects. In order to select a suitable membrane, rejection should be determined during operation of the photoreactor. The pressure in the membrane cell, the pH of the polluted water, the molecular size of the pollutants, and the photogenerated by-products and intermediate species can influence the permeate flux of the membrane and consequently its choice. High initial concentrations of the pollutants (e.g., 4-NP, patent blue, congo red) lower the photodegradation rate; however, this problem can be solved by diluting the feed in the reactor and by controlling the residence time of the pollutant by means of the membrane. The experimental results available in the literature indicate that the choice of a suitable membrane is essential for applying the photocatalytic membrane processes to the treatment of real effluents. Furthermore, use of photocatalysis combined with RO processes for drinking water production can eliminate the membrane fouling problem and the need for plant sanitizing. The hybrid continuous photoreactor where a nanofiltration membrane is used can give advantages over other approaches: simplification of clean-up or purification of various types of waters (for industrial, municipal/domestic, and agricultural uses), no sludge production, and saving in chemicals usage. It is expected that these hybrid processes will be considered particularly when plant upgrade is planned and, especially, if sunlight energy can be used for irradiation. Acknowledgment The authors thank MIUR (Italy) for financial support for this work.

BIBLIOGRAPHY 1. Gorenflo, A., Hesse, S., and Frimmel, H. (1998). A concept for advanced biodegradation. In: Proceedings of the European Research Conference on Natural Waters and Water Technology: Catalytic Science and Technology for Water. Acquafredda di Maratea, (I), October 3–8, p. 32. 2. Martin, R.J., Iwugo, G., and Pawlowski, L. (Eds.). (1989). Physicochemical Methods for Water and Wastewater Treatment. Elsevier, New York, p. 265.

3. Drioli, E. (Ed.). (1994). Proceedings of the International Symposium on Membrane and Membrane Processes. Hangzou, China, April 5–10. 4. Howell, A.J. and Noworyta, A. (Eds.). (1995). Towards Hybrid Membrane and Biotechnology Solutions for Polish Environmental Problems. Wroclaw Technical University Press, Wroclaw. 5. Caetano, A., De Pinho, M.N., Drioli, E., and Muntau, H. (Eds.). (1995). Membrane Technology: Applications to Industrial Wastewater Treatment. Kluwer Academic Publishers, Dordrecht, The Netherlands. 6. Pramauro, E. et al. (1993). Preconcentration of aniline derivatives from aqueous solutions using micellar-enhanced ultrafiltration. Analyst 118: 23. 7. Pramauro, E. and Bianco Prevot, E. (1998). Application of micellar ultrafiltration in environmental chemistry. Chim. Ind. Milan 80: 733. 8. Sopajaree, K., Qasim, S.A., Basak, S., and Rajeshwar, K. (1999). An integrated flow reactor–membrane filtration system for heterogeneous photocatalysis. Part I: Experimental and modelling of a batch-recirculated photoreactor. J. Appl. Electrochem. 29: 533. 9. Sopajaree, K., Qasim, S.A., Basak, S., and Rajeshwar, K. (1999). An integrated flow reactor–membrane filtration system for heterogeneous photocatalysis. Part II: Experiments on the ultrafiltration unit and combined operation. J. Appl. Electrochem. 29: 1111. 10. Molinari, R., Palmisano, L., Drioli, E., and Schiavello, M. (2002). Studies on various reactor configurations for coupling photocatalysis and membrane processes in water purification. J. Membr. Sci. 206: 399. 11. Ollis, D.F. (2002). Integrating photocatalysis and membrane technologies for water treatment. In: Proceedings of 2nd European Meeting on Solar-Chemistry and Photocatalysis: Environmental Applications. Saint Avold, France, May 29–31, Plenary Lecture 1. 12. Karakulski, K., Morawski, W.A., and Grzechulska, J. (1998). Purification of bilge water by hybrid ultrafiltration and photocatalytic processes. Sep. Purif. Technol. 14: 163. 13. Xi, W. and Geissen, S-U. (2001). Separation of titanium dioxide from photocatalytically treated water by crossflow microfiltration. Water Res. 35: 1256. 14. Molinari, R., Giorno, L., Drioli, E., Palmisano, L., and Schiavello, M. (2004). Photocatalytic nanofiltration reactors, Chap. 18. In: Nanofiltration: Principles and Applications. A.I. Schaefer, D. Waite, and A. Fane (Eds.). Elsevier Science, London. 15. Wyness, P., Klausner, J.F., Goswami, D.Y., and Schanze, K.S. (1994). Performance of nonconcentrating solar photocatalytic oxidation reactors, part I: flat-plate configuration. J. Solar Energy Eng. 116: 3. 16. Zhang, Y., Crittenden, J.C., Hand, D.W., and Perram, D.L. (1994). Fixed-bed photocatalysts for solar decontamination of water. Environ. Sci. Technol. 28: 435. 17. Liu, G. and Zhao, J. (2000). Photocatalytic degradation of dye sulforhodamine B: a comparative study of photocatalysis with photosensitization. New J. Chem. 24: 411. 18. Augugliaro, V. et al. (2004). Degradation of lincomycin in aqueous medium: coupling of solar photocatalysis and membrane separation. In: Serie Ponencias, The Improving Human Potential Programme. Access to research infrastructures activity. Research results at Plataforma Solar de Almeria within the year 2003 access campaign. CIEMAT, Madrid, pp. 43–52.

Next Page PHOTOCATALYTIC MEMBRANE REACTORS IN WATER PURIFICATION 19. Barbeni, M. et al. (1984). Photodegradation of 4-nitrophenol catalyzed by TiO2 particles. Nouv. J. Chim. 8: 547. 20. Okamoto, K. et al. (1985). Heterogeneous photocatalytic decomposition of phenol over TiO2 powder. Bull. Chem. Soc. Jpn. 58: 2015. 21. Augugliaro, V., Palmisano, L., Sclafani, A., Minero, C., and Pelizzetti, E. (1988). Photocatalytic degradation of phenol in aqueous titanium dioxide dispersion. Toxicol. Environ. Chem. 16: 89. 22. Augugliaro, V. et al. (1991). Photocatalytic degradation of nitrophenols in aqueous titanium dioxide dispersion. Appl. Catal. 69: 323. 23. Pramauro, E., Vincenti, M., Augugliaro, V., and Palmisano, L. (1993). Photocatalytic degradation of monuron in TiO2 aqueous dispersions. Environ. Sci. Technol. 27: 1790. 24. Pichat, P. and Enriquez, R.E. (2001). Interactions of humic acids, quinoline, and TiO2 in water in relation to quinoline photocatalytic removal. Langmuir 17: 6132. 25. Brandi, R.J., Alfano, O.M., and Cassano, A.E. (2000). Evaluation of radiation absorption in slurry photocatalytic reactors. 1. Assessment of methods in use and new proposal. Environ. Sci. Technol. 34: 2623. 26. Brandi, R.J., Alfano, O.M., and Cassano, A.E. (2000). Evaluation of radiation absorption in slurry photocatalytic reactors. 2. Experimental verification of the proposed method. Environ. Sci. Technol. 34: 2631. 27. Aguado, M.A., Anderson, M.A., and Hill, C.J., Jr. (1994). Influence of light intensity and membrane properties on the photocatalytic degradation of formic acid over TiO2 ceramic membranes. J. Mol. Catal. 89: 165. 28. Cheng, S., Tsa, S., and Lee, Y. (1995). Photocatalytic decomposition of phenol over titanium oxide of various structures. Catal. Today 26: 87. 29. Walker, S.A., Christensen, P.A., Shaw, K.E., and Walker, G.M. (1995). Photoelectrochemical oxidation of aqueous phenol using titanium dioxide aerogel. J. Electroanal. Chem. 393: 137. 30. Loddo, V., Marc`i, G., Palmisano, L., and Sclafani, A. (1998). Preparation and characterisation of Al2 O3 supported TiO2 catalysts employed for 4-nitrophenol photodegradation in aqueous medium. Mater. Chem. Phys. 53: 217. 31. Lepore, G.P., Persaud, L., and Langford, C.H. (1996). Supporting titanium dioxide photocatalysts on silica gel and hydro-physically unmodified silica gel. J. Photochem. Photobiol. A Chem. 98: 103. 32. Bellobono, I.R. (1995). Photosynthetic membranes in industrial waste minimization and recovery of valuable products. In: Membrane Technology Applications to Industrial Wastewater Treatment. A. Caetano et al. (Eds.). Kluwer Academic Publishers, Dordrecht, The Netherlands, p. 17. 33. Solovieva, A.B. et al. (2000). Porphyrin containing membrane photocatalytic systems for steroid olefine oxidation. Euromembrane: 345. 34. Kotova, S.L., Rumjantseva, T.N., Solovieva, A.B., Zav’jalov, S.A., and Glagolev, N.N. (2000). Porphyrins immobilized on sulfocationic membranes as photosensitizers in singlet oxygen generation. Euromembrane, 355. 35. De Smet, K., Vankelecom, I.F.J., and Jacobs, P.A. (2000). Filtration coupled to catalysis: a way to perform homogeneous reactions in a continuous mode. Euromembrane: 356. ¨ 36. Puhlfurß, P., Voigt, A., Weber, R., and Morb´e, M. (2000). Microporous TiO2 membranes with a cut off 99%

49 to 102,740

18 to 481

11

93 71 to >99 ∼88 86 to 94 ∼88 83 to 92 20 to >99

431 65 to 6540 1,025 1989 to 2453 5,450 54,504 11 to 981

312 11 to 433 5.3 5.3 to 124 7.9 42 77 to 365

1 8 1 6 1 1 5

RECLAIMED WATER A.S.C.E—A.W.W.A. (1990). Water Treatment Plant Design, 2nd Edn. McGraw-Hill, New York. 4. U.S. Environmental Protection Agency. Safe Drinking Water Act and Associated Final Rulings. Government Printing Office, Washington, DC. 5. Cox, C.R. (1964). Operation and Control of Water Treatment Processes. World Health Organization, Brussels, Belgium. 6. Tchobanoglous, G., Burton, F.L., and Stensel, D.H. (2003). Wastewater Engineering—Treatment, Disposal and Reuse, 4th Edn. Metcalf & Eddy, Inc., McGraw-Hill, New York, pp. 1104–1137, 1162–1196. Dean, J.A. (1985). Lange’s Handbook of Chemistry, 15th Edn. McGraw-Hill, New York. 7. Treybal, R.E. (1988). Mass Transfer Operations, 3rd Edn. McGraw-Hill Book Companies, New York, pp. 140–144, 158–211, 641–646. 8. Commission on Life Sciences. (1999). Risk Assessment of Radon in Drinking Water. Committee on Risk Assessment of Exposure to Radon in Drinking Water, Board on Radiation Effects Research, National Research Council, National Academy Press, Washington, DC. 9. Stocchi, E. (1990). Industrial Chemistry. Ellis Horwood, London, pp. 103–167. 10. Benefield, L.D., Judkins, J.F., and Weand, B.L. (1982). Process Chemistry for Water and Wastewater Treatment. Prentice-Hall, Englewood Cliffs, NJ, pp. 270–271, 293–294. Galperin, A.L. (1992). Nuclear Energy, Nuclear Waste. Chelsea House Publishers, New York. 11. Hafele, W. (1990). Energy from nuclear power. Scientif. Am. September. 12. U.S. Environmental Protection Agency. (1985). Guidance Manual For Compliance With the Filtration and Disinfection Requirements For Public Water Systems Using Surface Water Sources. Science and Technology Branch-Criteria and Standards Division-Office of Water, USEPA Publications, Washington, D.C., or NTIS, Springfield, VA.

READING LIST Montgomery, C.W. (1992). Environmental Geology, 3rd Edn. Wm. C. Brown Publishers, Chicago, IL. Hazardous Waste Cleanup Information, United States Environmental Protection Agency. Available: http://www.clu-in.com. Federal Remediation Technologies Roundtable. Available: http://www.frtr.gov.

RECLAIMED WATER ABSAR AHMAD KAZMI Nishihara Environment Technology, Inc. Tokyo, Japan

INTRODUCTION Reclaimed water is water from a wastewater treatment plant (WWTP) that has been treated and can be used for nonpotable uses such as landscape irrigation, cooling towers, industrial processes, toilet flushing, and fire protection. The inclusion of planned water reclamation, recycling, and reuse in water resource systems reflects the increasing scarcity of water resources to meet

805

societal demands, technological advancements, increased public acceptance, and improved understanding of public health risks. Per capita water use in the United States has quadrupled since the beginning of the twentieth century. Americans typically consume between 60 and 200 gallons (230 to 760 liters) per capita each day. The use of reclaimed water for nonpotable purposes can greatly reduce the demand on potable water sources—this use is encouraged by diverse organizations such as FEMP, EPA, and the American Water Works Association (AWWA). Municipal wastewater reuse amounts to about 4.8 billion gallons (18 million m3 ) per day (about 1% of all freshwater withdrawals). Industrial wastewater is far greater—about 865 billion gallons (3.2 billion m3 ) per day. Reclaimed water contains valuable nitrogen, phosphorus, and other nutrients, which promote plant growth. At the same time, the water meets stringent disinfection standards. Experience has shown that contact with reclaimed water does not promote waterborne disease transmission. In fact, reclaimed water quality standards are more stringent than those for surface streams, rivers, and irrigation channels. Reclaimed water is delivered through pipelines that are completely separate from the potable water system. In some areas of the United States, reclaimed water may be referred to as irrigation quality (‘‘IQ’’) water, but potential uses can extend well beyond irrigation. Using higher levels of treatment, such as reverse osmosis, reclaimed water as a potable source is technically and economically feasible. New technological breakthroughs in membrane filtration and combined biological and filtration treatment offer unprecedented opportunities for water recycling, especially in isolated locations and regions where the water supply is severely limited. BENEFITS OF USING RECLAIMED WATER • It saves millions of gallons of drinking water each day. • Its use for nonpotable (nondrinking) purposes is less expensive for the vast majority of of customers. • It delays the need for developing costly new water sources and building very expensive treatment plants. • There is no odor or staining from its use. • It allows a city to comply with permits relating to its water supply and wastewater treatment. • It minimizes negative effects around underground water sources, preserving the quality of life for plants and wildlife. • It reduces fertilizing costs because reclaimed water is rich in nitrogen and phosphorus. POTENTIAL USES OF RECLAIMED WATER Urban public water supplies are treated to satisfy requirements for potable use. However, potable use

806

RECLAIMED WATER Table 1. Summary of EPA Suggested Guidelines for Water Reusea

Levels of Treatment 1. Disinfected tertiaryb

Reclaimed Water Monitoring

Urban reusec

pH = 6–9

pH = weekly

Food crop irrigation

BOD5 ≤ 10 mg/L Turb. ≤ 2 NTU E. coli = none Res. Cl2 ≥ 1 mg/L pH = 6–9

BOD = weekly Turb. = cont. E. coli = daily Res. Cl2 = cont. pH = weekly

BOD5 = 30 mg/L TSS = 30 mg/L

BOD = weekly TSS = cont.

E. coli = 200/100 mL

E. coli = daily Res. Cl2 = cont.

Recreational impoundments 2. Disinfected secondary

Reclaimed Water Quality

Types of Reuse

Restricted-access-area irrigation

Food crop irrigation (commercially processed)

Nonfood crop irrigation Landscape impoundments (restricted access) Construction Wetlands habitat

Setback Distances 15 m (50 ft) to potable water supply wellsd

30 m (100 ft) to areas accessible to the public (if spray irrigation)

90 m (300 ft) to potable water supply well

Res. Cl2 ≥ 1 mg/L

a

From Reference 1. Filtration of secondary effluent. c Uses include landscape irrigation, vehicle washing, toilet flushing, fire protection, and commercial air conditioners. d Setback increases to 150 m (500 ft) if impoundment is not sealed. b

(drinking, cooking, bathing, laundry, and dishwashing) represents only a fraction of the total daily residential use of treated potable water. The remainder may not require water of potable quality. In many cases, water used for nonpotable purposes, such as irrigation, may be drawn from the same ground or surface source as municipal supplies, creating an indirect demand on potable supplies. There are opportunities for substituting reclaimed water for potable water or potable supplies for uses where potable water quality is not required. Specific water use where reuse opportunities exist include • • • • • •

urban industrial agricultural recreational habitat restoration/enhancement, and groundwater recharge

FEDERAL GUIDELINES FOR RECLAIMED WATER REUSE The U.S. Environmental Protection Agency (1) has suggested reclaimed water quality guidelines for the following reuse categories: • • • • • • •

urban reuse restricted-access-area irrigation agricultural reuse—food crops agricultural reuse—nonfood crops recreational impoundments construction uses industrial reuse

• groundwater recharge • indirect potable reuse Levels of treatment, minimum reclaimed water quality, reclaimed water monitoring, and setback distances are suggested for each reuse category (1). The guidelines are summarized in Table 1 for the two principal levels of treatment—disinfected tertiary (filtered secondary effluent) and disinfected secondary effluents. CONSTITUENTS OF RECLAIMED WATER The constituents of municipal wastewater subject to treatment may be classified as conventional, nonconventional, and emerging. Typical constituents included under each category are described in Table 2. The term conventional is used to define those constituents measured in mg/L that are the basis for designing most conventional wastewater treatment plants. Nonconventional applies to those constituents that may have to be removed or reduced using advanced wastewater treatment processes before the tank can be used beneficially. The term emerging is applied to those classes of compounds measured in the micro- or nanogram/L range that may pose long-term health concerns and environmental problems as more is known about the compounds. In some cases, these compounds cannot be removed effectively, even by advanced treatment processes. WATER RECLAMATION TECHNOLOGIES As noted in the previous section, the constituents of wastewater subject to treatment may be classified as conventional, nonconventional, and emerging. Conventional

RECLAIMED WATER Table 2. Classification of Typical Constituents in Wastewater Classification

Constituent

Conventional

Total suspended solids Colloidal solids Biochemical oxygen demand Chemical oxygen demand Total organic carbon Ammonia Nitrate Nitrite Total nitrogen Phosphorus Bacteria Protozoan cysts and oocystsa Virusesb Refractory organics Volatile organic compounds Surfactants Metals Total dissolved solids Prescription and nonprescription drugsc Home care products Veterinary and human antibiotics Industrial and household products Sex and steroidal hormones Other endocrine disrupters

Nonconventional

Emerging

a

Value per 100 mL. Plaque-forming units/100 mL. c Pharmaceutically active substances. b

807

constituents are removed by conventional treatment technologies. Advanced treatment technologies are used most commonly for removing nonconventional constituents. The removal of emerging constituents occurs in both conventional and advanced treatment processes, but the levels to which individual constituents are removed are not well defined. Typical performance data for selected treatment combinations are presented in Table 3. PLANNING FOR WASTEWATER RECLAMATION AND REUSE In effective planning for wastewater reclamation and reuse, the objectives and basis for conducting the planning should be defined clearly. The optimum water reclamation and reuse project is best achieved by integrating both wastewater treatment and water supply needs into one plan. This integrated approach is somewhat different from planning for conventional wastewater treatment facilities where planning is done only for conveyance, treatment, and disposal of municipal wastewater. Effective water reclamation and reuse facilities should include the following elements: (1) assessment of wastewater treatment and disposal needs, (2) assessment of water supply and demand, (3) assessment of water supply benefits based on water reuse potential, (4) analysis of reclaimed water market, (5) engineering and economic analyses of alternatives, (6) implementation plan

Table 3. Treatment Levels Achievable from Various Combinations of Unit Operations and Processes Used for Water Reclamation Typical Effluent Quality, mg/L, Except Turbidity, NTU

Activated sludge + granular medium filtration Activated sludge + granular medium filtration + carbon adsorption Activated sludge/nitrification single stage Activated sludge/nitrification denitrification separate stages Metal salt addition to activated sludge +nitrification/ denitrification separate stages Biological phosphorus removala Biological nitrogen and phosphorus removal + filtration Activated sludge + granular medium filtration + carbon adsorption +reverse osmosis Activated sludge/nitrificationdenitrification +granular medium filtration + carbon adsorption + reverse osmosis Activated sludge/nitrificationdenitrification and phosphorus removal + microfiltration +reverse osmosis a

TSS

BOD5

COD

Total N

NH3 -N

PO4 -P

Turbidity

4–6

Tsea ), z 620(Tair − Tsea ) = L (Tair + 273.2)Uz2

S.A. HSU Louisiana State University Baton Rouge, Louisiana

(2)

According to Hsu (13,14), B = 0.146(Tsea − Tair )0.49

INTRODUCTION Air–sea interaction is, according to Geer (1), the interchange of energy (e.g., heat and kinetic energy) and mass (e.g., moisture and particles) that takes place across the active surface interface between the top layer of the ocean and the layer of air in contact with it and vice versa. The fluxes of momentum, heat, moisture, gas, and particulate matter at the air–water interface play important roles, for example, in environmental hydraulics and water–environment–health interactions, during low wind speeds before the onset of wave breaking, the exchange of air bubbles is limited. If this situation persists for a long time, algal blooms may develop, ultimately affecting water quality. On the other hand, during typhoon/hurricane conditions, the storm surge may affect the sewerage outflow at a greater depth than normal because of shoaling. In the area on the right-hand side of the storm track (in the Northern Hemisphere), runoff may also be blocked due to the surge, thus increasing the flood potential and saltwater intrusion. Air–sea interaction encompasses vast scales in both spatial and temporal viewpoints, so only a few basic and applied topics are summarized here, including the parameterization of stability length, determination of friction velocity, wind–wave interaction, and the estimation of shoaling depth during storms. For more detailed laws and mechanisms in air–sea interaction, see Donelan (2) and recently Csanady (3); for air–sea exchange of gases and particles, see Liss (4) and most recently Donelan et al. (5); for the role of air–sea exchange in geochemical cycling, see Buat-Menard (6) and recently Liss (7); for larger scale air–sea interaction by La ˜ and its impacts, see Glantz (8); for more physics, Nina chemistry, and dynamics related to air–sea exchange, see Geernaert (9); and for wind–wave interaction, see Janssen (10).

where z is the height normally set to 10 m; Tair and Tsea stand for the air and sea temperatures, respectively; Uz is the wind speed at height z; and B is the Bowen ratio. For operational and engineering applications, z/L ≤ −0.4 is unstable, |z/L| < 0.4 is neutral, and z/L ≥ 0.4 is stable. PARAMETERIZATION OF THE ROUGHNESS LENGTH The roughness parameter Z0 can be computed based on the formula provided in Taylor and Yelland (15) that
4.5 Hs Z0 = 1200 Hs Lp

(4)

and, for deep water waves, Lp =

gT 2p 2π

(5)

where g is gravitational acceleration, Hs and Lp are the significant wave height and peak wavelength for the combined sea and swell spectrum, and Tp is its corresponding wave period. Note that Hs is defined as the average of the highest one-third of all wave heights during the 20-minute sampling period. ESTIMATION OF THE FRICTION VELOCITY The friction velocity (u∗ ) can be obtained as follows: 1/2

u∗ = U10 Cd

(6)

where U10 is the wind speed at 10 m and Cd is the drag coefficient. According to Amorocho and DeVries (16), the WAMDI Group (17), and Hsu (18), one may classify the air–sea interaction into three broad categories based on wave breaking conditions: In light winds, U10 < 7.5 m/s, prior to the onset of wave breakers,
 u∗ 2 = 1.2875 ∗ 10−3 (7) Cd = U10

PARAMETERIZATION OF THE STABILITY LENGTH In the atmospheric boundary layer, the buoyancy length scale, L, also known as the Obukhov (or Monin–Obukhov) length, is a fundamental parameter that characterizes the ‘‘stability’’ of the surface layer (11). L describes the relative importance between the buoyancy effect (or thermal turbulence) and wind shear (or mechanical turbulence). According to Hsu and Blanchard (12), L can be parameterized as follows. For unstable conditions (i.e., when Tsea > Tair ), 0.07 1000(Tsea − Tair )(1 + ) z B =− 2 L (Tair + 273.2)Uz

(3)

Both thermal and mechanical turbulence are important. In moderate winds, 7.5 ≤ U10 ≤ 20 m/s, the range after the onset but before the saturation of wave breakers,

(1)

Cd = (0.8 + 0.065U10 ) ∗ 10−3 1

(8)

2

AIR–SEA INTERACTION

Mechanical turbulence is more important than thermal effects. In strong winds, U10 > 20 m/s, after the saturation of wave breakers, (9) Cd = 2.5 ∗ 10−3 Mechanical turbulence dominates. ESTIMATING LATENT HEAT FLUX (OR EVAPORATION) Using the parameter of the Bowen ratio supplied by Hsu (13,14), as shown in Eq. 3, the latent heat flux (Hlatent ) can be estimated as 1 1 Hlatent (W m−2 ) = Hsensible = ρa Cp CT (Tsea − Tair )U10 B B (10) where Hsensible is the sensible heat flux, ρa (=1.2 kg m−3 ) is the air density, Cp (=1004 J kg−1 K−1 ) is the specific heat at constant pressure for dry air, CT (=1.1 ∗ 10−3 ) is the transfer coefficient for heat, (Tsea − Tair ) is in K, and U10 in m s−1 . A latent heat flux of 1 W m−2 is equivalent to an evaporation rate of 3.56 ∗ 10−3 cm day−1 , so Eq. 10 can be used to estimate the evaporation rate. ESTIMATING MAXIMUM SUSTAINED WIND SPEED DURING A HURRICANE Under hurricane/typhoon conditions, intense air–sea interaction occurs. Beach erosion, engineering structures, storm surge, and sewerage outflow can all be affected, so this topic should deserve more attention than the deepwater environment. The very first subject related to a tropical cyclone is to estimate its maximum sustained wind speed at the standard height of 10 m (i.e., U10 ). This is accomplished as follows. From the cyclostrophic equation (i.e., centrifugal force = pressure gradient) (11), 1 ∂P 1 P 1 Pn − P0 Ua2 = = = γ ρa ∂γ ρa γ ρa γ − 0

(11)

where Ua is the maximum sustained wind speed above the surface boundary layer, γ is the radius of the hurricane, ∂P/∂γ is the radial pressure gradient, Pn is the pressure outside the hurricane effect (1013 mb, the mean sea level pressure), and P0 is the hurricane’s minimum central pressure. Because ρa = 1.2 kg m−3 , P = (1013 − P0 ) mb, and 1 mb = 100 N m−2 = 100 kg m−1 s−2 , Eq. 11 becomes  Ua =

100 kg m−1 s−2 1.2 kg m

−3

1/2

√ √ P = 9 P

(12)

According to Powell (19), U10 = 0.7Ua ; therefore √ U10 = 6.3 P = 6.3(1013 − P0 )1/2

(13)

where U10 is in m s−1 and P in mb. Equation 13 has been verified by Hsu (18). In 1985, during Hurricane Kate over the Gulf of Mexico, the

U.S. National Data Buoy Center (NDBC) buoy #42003, located on the right-hand-side of the storm track near the radius of maximum wind, recorded a minimum sealevel pressure (P0 ) of 957.1 mb. Therefore, P = (1013 − 957.1) = 55.9 mb. Substituting this value in Eq. 13, U10 = 47.1 m s−1 which is in excellent agreement with the measured U10 = 47.3 m s−1 . Another verification is provided in Fig. 1. According to Anthes (20, p. 22 and Fig. 2.8),
U10γ = U10 max

R γ

0.5 (14)

where U10γ is the sustained wind speed at a distance 10 m away from the storm center and U10 max is the maximum sustained wind at 10 m at the radius of maximum wind, R. According to Hsu et al. (21), for operational applications, 
1013 − P0 R (15) = ln γ Pγ − P0 where Pγ is the pressure at a point located at a distance from the storm center and ln is the natural logarithm. Substituting Eq. 15 in Eq. 14, 
 1013 − P0 0.5 U10γ = U10 max ln Pγ − P0

(16)

During Hurricane Lili in 2002, the NDBC had two buoys, #42001 located near R, and #42003 located due east along 26 ◦ N, approximately 280 km from 42001. The wind speed measurement at both buoys was 10 m. From the NDBC website (www.ndbc.noaa.gov/), at 20Z 2 October 2002 at #42001, P0 = 956.1 mb. Substituting this P0 in Equation 13, U10 max = 47.5 m s−1 , in excellent agreement with the measured value of 47.2 m s−1 (=106 mph). Therefore, Eq. 13 is further verified. At the same time, Pγ = 1011.1 mb was measured at #42003. Substituting this Pγ in Eq. 16, we obtain U10γ = 8.8 m s−1 . The measured U10γ at 42003 was 9.2 m s−1 . The difference is only about 4%, so we conclude that Eqs. 13 and 16 can be used for nowcasting using the pressure measurements at Pγ and P0 which are normally available via the official ‘‘Advisory’’ during a hurricane. ESTIMATING MAXIMUM SIGNIFICANT WAVE HEIGHT DURING A HURRICANE According to the USACE (22, p. 3–85, Eq. 3–64),  Hs Tp = 12.1 g ∴

Hs gT 2p

= 0.0068

(17) (18)

More verification of Eq. 18 is provided in Hsu (23). According to Hsu et al. (21), based on the evaluation of nine fetch-limited wind–wave interaction formulas, that provided by Donelan et al. (24) ranked best as follows: 
gT p 1.65 gH s = 0.00958 2 U10 U10

(19)

AIR–SEA INTERACTION

3

Figure 1. Satellite data (visible channel from NOAA-16) received and processed at the Earth Scan Lab, Louisiana State University, during Hurricane Lili (2002) in the Gulf of Mexico. The solid line represents the storm track. Data from NDBC buoys 42001 and 42003 are employed in this study. Note that the anemometers for both bouys were located at the standard 10 m height.

ESTIMATING STORM SURGE AND SHOALING DEPTH

From Eqs. 18 and 19, 2 Hs = 0.00492 U10

(20)

Substituting Eq. 13 in Eq. 20, Hs

max

= 0.20 P

(21)

where Hs max is in meters and P is in mb. Equation 21 is verified in Fig. 1. Buoy 42001, located near the radius of maximum wind, measured P0 = 956.1 mb at 20Z 2 October 2002, so that P = (1013 − 956.1) = 56.9 mb. Substituting this value in Eq. 21, Hs max = 11.38 m, which is in excellent agreement with that of 11.22 m measured at 21Z 2 October 2002 (within 1 hour after the measured minimal P0 ).

To estimate a typhoon/hurricane—generated storm surge (S), and shoaling depth (Dshoaling ), the following formulas are useful operationally, provided that the storm’s minimum (or central) pressure near the sea surface (P0 ) is known. According to Hsu (23), for the storm surge in deep water before shoaling (i.e., when the waves feel the sea floor), SI = 0.070(1010 − P0 )

(22)

where SI is the initial peak storm surge before shoaling. For the peak surge at the coast, SP = 0.070 (1010 − P0 ) ∗ FS ∗ FM

(23)

4

NOAA’S ATLANTIC OCEANOGRAPHIC AND METEOROLOGICAL LABORATORY

where FS is a shoaling factor dependent on shelf topography and width and FM is a correction factor for storm motion. Both FS and FM for certain areas are included in Hsu (23). A verification of Eq. 23 during Hurricane Georges in 1998 is also available in Hsu (23). The shoaling depth is computed as follows: From Taylor and Yelland (15), Dshoaling = 0.2 Lp , and from Eq. 5,

12. Hsu, S.A. and Blanchard, B.W. (2004). On the estimation of overwater buoyancy length from routine measurements. Environ. Fluid Mech., in press.

gT 2p

15. Taylor, P.K. and Yelland, M.J. (2001). The dependence of sea surface roughness on the height and steepness of the waves. J. Phys. Oceanogr. 31: 572–590. 16. Amorocho, J. and DeVries, J.J. (1980). A new evaluation of the wind stress coefficient over water surfaces. J. Geophys. Res. 85(C1): 433–442.

Dshoaling = 0.2

2π

=

0.2 Hs  Hs 2π gT 2p

where Hs /gT p 2 is called wave steepness, a useful parameter in coastal engineering. From Hsu et al. (21) and under hurricane conditions from Eq. 18, Hs /gT p 2 = 0.0068. Thus, from Eq. 21, Dshoaling = 4.7 Hs = 4.7 ∗ 0.2(1013 − P0 ) ∴ shoaling depth (meters) ≈ (1013 − P0 )

(24)

CONCLUDING REMARKS Although all formulas presented in this article are based on the open literature, they may need some verification before being applied to site-specific conditions. For example, Eq. 22 for the storm surge is for an open coast before shoaling. It needs to be adjusted for flooding at the coast due to different storm speeds and local bathymetry, as needed in Eq. 23. BIBLIOGRAPHY 1. Geer, I.W. (Ed.). (1996). Glossary of Weather and Climate With Related Oceanic and Hydrologic Terms. American Meteorological Society, Boston, MA.

13. Hsu, S.A. (1998). A relationship between the Bowen ratio and sea-air temperature difference under unstable conditions at sea. J. Phys. Oceanogr. 28: 2222–2226. 14. Hsu, S.A. (1999). On the estimation of overwater Bowen ratio from sea-air temperature difference. J. Phys. Oceanogr. 29: 1372–1373.

17. The WAMDI Group. (1988). The WAM model—a third generation ocean wave prediction model. J. Phys. Oceanogr. 18: 1775–1810. 18. Hsu, S.A. (2003). Estimating overwater friction velocity and exponent of the power-law wind profile from gust factor during storms. J. Waterways Port Coastal Ocean Eng. 129(4): 174–177. 19. Powell, M.D. (1982). The transition of the Hurricane Frederic boundary-layer wind field from the open Gulf of Mexico to landfall. Mon. Weather Rev. 110: 1912–1932. 20. Anthes, R.A. (1982). Tropical Cyclones, Their Evolution, Structure, and Effects. Meteorological Monographs Number 41, American Meteorological Society, Boston, MA. 21. Hsu, S.A., Martin, M.F., Jr., and Blanchard, B.W. (2000). An evaluation of the USACE’s deepwater wave prediction techniques under hurricane conditions during Georges in 1998. J. Coastal Res. 16(3): 823–829. 22. U.S. Army Corps of Engineers. (1984). Shore Protection Manual. Vicksburg, MS. 23. Hsu, S.A. (2004). A wind-wave interaction explanation for Jelesnianski’s open-ocean storm surge estimation using Hurricane Georges (1998) measurements. Natl. Weather Dig., in press. 24. Donelan, M.A., Hamilton, J., and Hui, W.H. (1985). Directional spectra of wind-generated waves. Philoso. Transa. Roy. Soc. London, Ser. A 315: 509–562.

2. Donelan, M.A. (1990). Air-sea interaction. In: The Sea. Wiley, Vol. 9. Hoboken, NJ. 3. Csanady, G.T. (2001). Air-Sea Interaction: Laws and Mechanisms. Cambridge University Press, Cambridge, UK. 4. Liss, P.S. (1983). Air-Sea Exchange of Gases and Particles. Kluwer Academic, Springer, New York. 5. Donelan, M.A., Drennan, W.M., Saltman, E.S., and Wanninkhof, R. (Eds.). (2002). Gas Transfer at Water Surfaces. Geophysical Monograph #127, American Geophysical Society, Washington, DC. 6. Buat-Menard, P. (1986). Role of Air-Sea Exchange in Geochemical Cycling. Kluwer Academic, Springer, New York. 7. Liss, P.S. (1997). Sea Surface and Global Change. Cambridge University Press, Cambridge, UK. ˜ and Its Impacts: Facts and 8. Glantz, M.H. (2002). La Nina Speculation. Brookings Inst. Press, Washington, DC. 9. Geernaert, G.L. (1999). Air-Sea Exchange: Physics, Chemistry & Dynamics. Kluwer Academic, Springer, New York.

NOAA’S ATLANTIC OCEANOGRAPHIC AND METEOROLOGICAL LABORATORY National Oceanographic and Atmospheric Administration (NOAA)

June 18, 1999—The Atlantic Oceanographic and Meteorological Laboratory (AOML) in Miami, Florida, is one of 12 environmental research laboratories that work on environmental issues for NOAA’s Office of Oceanic and Atmospheric Research (OAR). OAR research advances NOAA’s ability to predict weather, helps monitor and provides understanding of climate and global change, as well as improve coastal ocean health.

10. Janssen, P. (2004). The Interaction of Ocean Waves and Wind. Cambridge University Press, Cambridge, UK. 11. Hsu, S.A. (1988). Coastal Meteorology. Academic Press, San Diego, CA.

This article is a US Government work and, as such, is in the public domain in the United States of America.

NOAA’S ATLANTIC OCEANOGRAPHIC AND METEOROLOGICAL LABORATORY

5

AOML’s mission is to conduct a basic and applied research program in oceanography, tropical meteorology, atmospheric and oceanic chemistry, and acoustics. The programs seek to understand the physical characteristics and processes of the ocean and the atmosphere, both individually and as a coupled system. The principal focus of these investigations is to provide knowledge that may ultimately lead to improved prediction and forecasting of severe storms, better use and management of marine resources, better understanding of the factors affecting both climate and environmental quality, and improved ocean and weather services for the nation.

Originally under the jurisdiction of the Environmental Science Services Administration (ESSA), the forerunner of NOAA, AOML was founded in Miami, Florida, in 1967. Several months after NOAA was established in 1970, groundbreaking began on a new 12-acre federally funded research facility on Virginia Key. AOML dedicated its new location on Feb. 9, 1973. It celebrated its 25th anniversary in 1998. AOML has four main research divisions: Hurricane Research, Ocean Acoustics, Ocean Chemistry, and Physical Oceanography. To learn more about AOML visit: http://www.aoml. noaa.gov/ HURRICANE RESEARCH DIVISION The Hurricane Research Division (HRD) is NOAA’s primary component for research on hurricanes. Its high-

est priority is improving the understanding and prediction of hurricane motion and intensity change. A key aspect of this work is the annual hurricane field program, supported by the NOAA Aircraft Operation’s Center research/reconnaissance aircraft. Research teams analyze data from field programs, develop numerical hurricane models, conduct theoretical studies of hurricanes, prepare storm surge atlases, and study the tropical climate. HRD works with the National Hurricane Center/Tropical Prediction Center in all phases of its research, the National Meteorological Center and the Geophysical Fluid Dynamics Laboratory—another of OAR’s research labs—in research related to numerical modeling of hurricanes, and the National Severe Storms Laboratory—yet another OAR lab—in the study of landfalling hurricanes,

6

LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS

as well as other NOAA groups, federal agencies, and universities in a variety of basic and applied research. OCEAN ACOUSTICS DIVISION The Ocean Acoustics Division (OAD) gathers, analyzes and reports coastal ocean data on human-related discharges and their potential environmental impacts. Additionally, OAD has an ongoing research program on the use of acoustics to measure coastal and deep ocean rainfall, an important element in calculating the global energy balance for climate monitoring and prediction. The Division works in cooperation with other federal, state, and local authorities to maximize research knowledge for use in economically and environmentally important projects in the coastal ocean.

OCEAN CHEMISTRY DIVISION With a diverse scientific staff of marine, atmospheric, and geological chemists, as well as chemical, biological, and geological oceanographers, the Ocean Chemistry Division (OCD) is able to use multidisciplinary approaches to solve scientific research questions. The Division’s work includes projects that are important in assessing the current and future effects of human activities on our coastal to deep ocean and atmospheric environments. PHYSICAL OCEANOGRAPHY DIVISION The Physical Oceanography Division (PhOD) provides and interprets oceanographic data and conducts research relevant to decadal climate change and coastal ecosystems. This research includes the dynamics of the ocean, its interaction with the atmosphere, and its role in climate and climate change. Data is collected from scientific expeditions, both in the deep ocean and in coastal regions. Satellite data is processed and incorporated into the analyses. PhOD manages the Global Ocean Observing (GOOS) Center, which manages the global collection, processing, and distribution of drifting buoy data and the information collected from ocean temperature profilers. This information is crucial to understanding and predicting shifts in weather patterns and the relationship of the ocean and the atmosphere as a coupled system.

LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS PAOLO MAGNI IMC—International Marine Centre Torregrande-Oristano, Italy

SHIGERU MONTANI Hokkaido University Hakodate, Japan

BACKGROUND Benthic nutrient regeneration may be referred to as a new availability to the water column of significant amounts of nitrogen, phosphorus, and other nutrients, as a consequence of the metabolism of organic matter by the benthos (1). The processes of benthic nutrient regeneration in coastal marine systems are strongly influenced by the presence of abundant macrofauna (2–5). Correct evaluation of the biogenic flux of nutrients due to the excretory activity of infaunal species is therefore an important background of information to investigate the cycling of biophilic elements (nitrogen, phosphorus, and silicon). In field studies, major drawbacks include the difficulty to discern between nutrient upward flux due to animal excretion and a number of local effects, such as microbial mineralization (4,6–8) and uptake (9–14), animal bioturbation and irrigation currents (15–19), tidal currents and

LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS

wind-generated waves (20–22). Laboratory experiments on the animal excretion rates of nutrients under more controlled conditions represent a useful tool for quantifying the actual biogenic contribution by macrofauna to the total upward flux of nutrients from sediments. Nevertheless, these experiments have often restricted their investigations to ammonium (23–28) or, in a few cases, to ammonium and phosphate (29). CASE STUDY This study was conducted within a multidisciplinary project on the cycling of nutrients and organic matter in a tidal estuary in the Seto Inland Sea (30–37). Laboratory experiments were carried out on the excretion rates of ammonium, phosphate, and silicate by different size classes of the bivalves, Ruditapes philippinarum and Musculista senhousia. These species were selected as they were dominant on a sandflat of the estuary under investigation. An extrapolation of these results to a field community is presented in TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT. Both studies will be the basis of a third companion paper on the relationship between the temporal scaling of bivalve nutrient excretion and the seasonal change of nutrient concentrations in the porewater (SEASONAL COUPLING BETWEEN INTERTIDAL MACROFAUNA AND SEDIMENT COLUMN POREWATER NUTRIENT CONCENTRATIONS). In these experiments, 2.5 L aquaria with and without (control) animals were employed and run on two different occasions for 24 hours. Each experiment consisted of a 10 h day (light) and a 10 h night (dark) treatment in which the changes in nutrient concentrations were measured every 2 h. Between the two (light and dark) treatments, a low tide lasting 2 h (like that approximately on the flat where animals were collected) was created, during which the experimental animals were not removed from sediments to keep the experiment continuous. The experimental setup and procedure are detailed in our associated paper

7

where the bivalve excretion rates of ammonium and phosphate have been presented and discussed (35). We will extend this study to silicate, a nutrient species whose regeneration through animal excretory activity has been less investigated, either in situ (38,39) or in the laboratory (5,40). Table 1 includes some characteristics of the experimental animals, as well as the field-relevant (33) amount of algal food (Thalassiosira sp.) offered in four spikes during each experiment (35). In all treatments, there was a marked increase in all three nutrient concentrations, in the control (no animals), the increase was much more limited (i.e., silicate) or not observed (i.e., ammonium) (Fig. 1). Based on the differences in nutrient concentrations between treatments and controls, relevant linear regression lines of five to six measurements were used to calculate the nutrient excretion rates for each size class of R. philippinarum and M. senhousia (Table 2). This approach may be a more reliable way to quantify the daily bivalve excretion, whereas previous similar experiments have been based on shorter incubations and/or the sole difference between initial and final values of nutrient concentrations (4,7,22,28,29). The data sets were subjected to ANOVA in a two-factor randomized complete block design, using the day/night variable as factor A, the time (hour) variable as factor B, and the size classes of each bivalve species as replicates (35). As found for ammonium in R. philippinarum, but not in M. senhousia, silicate excretion was significantly higher (57%, p < 0.001, n = 36) during the day than during the night, suggesting a possible effect of light on the excretory activity of this bivalve species. A comparison of nutrient excretion rates (µmol g−1 DW h−1 ) of bivalve species obtained through in situ or laboratory experiments is given in Table 3. According to the excretion rates of silicate found in our laboratory experiments, this study points to the importance of the excretory activity of these bivalve species to the biogenic regeneration of silicate. This aspect

Table 1. Animals Employed in the Laboratory Experiments and Experimental Conditions. Ind.: Number of Individualsa,b Species

Size, mm

Ind., n

9.4 ±1.4 15.5 ±1.0 18.9 ±0.8

12 – 15 – 9 –

TOT, mg

DW, mg

Temp, ◦ C

Chla, (µg L− 1)

Expt, date

19.6 ±1.5 19.6 ±1.5 21.6 ±0.3

26.3 ±8.7 24.6 ±8.8 38.9 ±12.8

May 1996 – May 1996 – Sep 1996 –

19.6 ±1.5 21.6 ±0.3

24.5 ±9.1 47.4 ±22.8

May 1996 – Sep 1996 –

Ruditapes philippinarum Size class I Size class II Size class III

AVG SD AVG SD AVG SD

197 ±79 830 ±174 1520 ±149

9.9 ±4.0 37.0 ±9.2 63.6 ±10.4

Musculista senhousia Size class I Size class II a

AVG SD AVG SD

16.7 ±1.3 23.5 ±1.7

14 – 8 –

431 ±117 1264 ±172

27.6 ±7.7 52.4 ±8.0

Reproduced from Reference 35. TOT: mean (live) weight for each size class of the experimental animals; DW: mean dry soft-body weight for each size class of the experimental animals; Temp: experimental temperature; Chl a (Chlorophyll a) is the mean (AVG) ± standard deviation (SD) of four spikes of cultures of Thalassiosira sp. (Chl a = 0.01 × Thalassiosirasp. cell + 3.6, r2 = 0.908; p < 0.001, n = 40) for each day/night treatment b

8

LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS

May 9–10, 1996

September 6–7, 1996

Control

Control

R. philippinarum I R. philippinarum II M. senhousia I

R. philippinarum III M. senhousia II 50

30 Dark

40 NH4+, µM

NH4+, µM

Light 20 10

Light

Dark

Light

Dark

Light

Dark

30 20 10

0

0 Light

PO43−, µM

3 2 1

Si(OH)4, µM

30 20 10

Time, hour

Nutrient Excretion Rate, µmol g−1 DW h−1

Species

Light

Ruditapes philippinarum Size class I 10.6 Size class II 9.6 Size class III 5.0 Light/dark mean 8.4b Total mean 7.1

PO4 3− Dark Light 7.9 5.8 3.8 5.8

Musculista senhousia Size class I 9.3 11.4 Size class II 16.9 9.7 Light/dark mean 13.1 10.6 Total mean 11.8

Si(OH)4

Dark Light

3.4 1.0 0.9 1.8

3.9 1.1 0.7 1.9

10.9 3.3 1.1 5.1 6.6

1.5 1.3 1.4 1.4

Dark

15.8 4.1 4.0 8.0b

1.9 1.2 1.6 1.4

2

Dark

10 12 14 16 18 20 22 24 2 4 6 8

Si(OH)4, µM

Light

Table 2. Nutrient Excretion Rate for Each Size class of Ruditapes philippinarum and Musculista senhousia During day/night Treatments (Experimental Temperature as in Table 1)a

NH4 +

4

0

0

0

6

14.5 4.2 9.4

4.8 5.5 5.2 7.3

a

Based on Reference 35. b Mean day (light) excretion significantly higher (ANOVA p < 0.001) than night (dark) excretion (based on Reference 35).

has been more controversial than the contribution of bivalves to the regeneration of ammonium and phosphate. The extent of silicate excretion also varies considerably, depending on the bivalve species investigated and the environmental/experimental conditions employed.

40 30 20 10 0

10 12 14 16 18 20 22 24 2 4 6 8

PO43−, µM Figure 1. Laboratory experiments: changes of nutrient concentrations [ammonium, NH4 + –N); phosphate, PO4 3− –P; and silicate Si(OH)4 –Si] during the day (light) and night (dark) treatments. Vertical dashed lines: left line (time 20:00) indicates the end of the day (light) treatment; right line (time 22:00) indicates the start of the night (dark) treatment; between lines: low tide between treatments (based on Ref. 35).

Dark

Time, hour

In in situ experiments, Prins & Small (4) found no significant excretion of silicate by Mytilus edulis beds on an intertidal zone of the Westerschelde (The Netherlands). The occurrence of silicate fluxes was attributed to the possible increased rate of dissolution of silicate at higher temperature. Asmus et al. (38) in the eastern Wadden Sea (Germany) and Dame et al. (39) in the Easterschelde and the western Wadden Sea (The Netherlands) found high fluxes of silicate from mussel beds of M. edulis. Although, in both studies, the actual excretion rate of silicate by M. edulis was not estimated per biomass unit (e.g., µmol g−1 DW h−1 ), Dame et al. (39) suggested that silicate release from mussel beds results from phytoplankton cells breaking down as they are metabolized by the mussels. Dame et al. (39) argued that the longer turnover time for silicate, compared to phosphate and ammonium, implies a lesser role for the mussel beds in recycling this nutrient species in the two estuaries under investigation. In contrast, Asmus et al. (38) found rapid silicate release in the Sylt-flume study and suggested that the mussels are an accelerator in recycling biogenic silica. Similarly, a study on the nutrient excretion of R. philippinarum in core incubation experiments found that silicate regeneration was, on average, 9.2 times faster in the site farmed with clams (5). In mesocosm experiments using large tanks, Doering et al. (40) also found that the level of flux was elevated in the presence of another clam, Mercenaria mercenaria, by 86% and 57% for silicate and ammonium, respectively. Our results indicate

LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS

9

Table 3. Comparison of Nutrient Excretion Rates (µmol g−1 DW hour−1 ) for Different Species of Mussels (m), Clams (c) and Oysters (o)a Species and Study Area Mytilus edulis (m) Narragansett Bay, USA Linher River, U.K. Sound, DK Western Scheldt, NL Musculista senhousia (m) Seto Inland Sea, JPN Seto Inland Sea, JPN Modiolus demissus (m) Narragansett Bay, USA Great Sippewissett, USA Donax serra (m) Maitland River, S. Africa Sundays River, S. Africa Donax sordidus (m) Sundays River, S. Africa Aspatharia wahlbergi (m) Lake Kariba, Zimbabwe Corbicula africana (c) Lake Kariba, Zimbabwe Corbicula japonica (c) Lake Shinji, JPN Mercenaria mercenaria (c) Delaware Bay, USA Macoma balthica (c) Wadden Sea, DK Ruditapes philippinarum (c) Virgin Islands, USA Moss Landing, USA Moss Landing, USA Marennes-Ol´eron, F Hatchery, Ireland Seto Inland Sea, JPN Seto Inland Sea, JPN Crassostrea virginica (o) Delaware Bay, USA Crassostrea gigas (o) North Brittany, F

Methodb

NH4 +

PO4 3−

Si(OH)4

T, ◦ C

Reference

Lab Lab In situ In situ

3.1 4.9–34.6 0.14–3.1 1.1

ndc nd 0.10–0.53 nd

ndc nd nd nd

15 11–21 0.7–18 12

41 23 42 43

Lab Lab

9.3–16.9 9.3–16.9

1.2–1.6 1.2–1.6

– 4.2–14.5

18–22 18–22

35 This study

Lab Lab

3.58 ± 1.73 2.5

nd nd

nd nd

21 annual

41 2

In situ Lab/In situ

0.35–8.1 2.2

nd nd

nd nd

15–25

44 26

Lab/In situ

2.9

nd

nd

15–25

26

Lab

6.1

0.48

nd

25.2

29

Lab

12.9

nd

nd

25.2

29

Lab

14.3

nd

nd

27

28

Lab

0.9–1.5

nd

nd

20

45

Lab

0.1d

nd

nd

13–15

16

Lab Lab Lab Lab Lab Lab Lab

1.9–4.9 1–2.3 0.6–0.9 0.5–13 0.16–1 3.8–10.6 3.8–10.6

nd nd nd nd nd 0.7–3.9 0.7–3.9

nd nd nd nd nd — 1.1–15.8

20.1 27.1 12, 15, 18 5–25 18.8 18–22 18–22

24 46 47 25 27 35 This study

Lab

0.5–0.9

nd

nd

20.1

24

In situ

0.28–6.6

nd

nd

27.1

48

a

Based on Reference 35. Lab: laboratory experiments. c nd: not determined. d Excretion rate calculated as a wet soft-body weight. b

well-balanced stoichiometric ratios among the nutrient species excreted by the bivalves (35), which can also be related to the use of the diatom Thalassiosira sp. as a food (35). This was aimed to approximate the actual field situation on the tidal flat, where abundant microalgal biomass, including resuspended benthic diatoms, is available to filter-feeders such as in R. philippinarum and M. senhousia (31,33). The high excretion rates of silicate of these two dominant bivalve species found in our laboratory experiments, together with those of ammonium and phosphate, suggest a major contribution of bivalve nutrient excretion to the upward flux of nutrients from sediments in our relevant study area. This will be the subject of the subsequent paper where we will apply these rates to the actual bivalve standing stock found in the field (TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT).

BIBLIOGRAPHY 1. Nixon, S.W., Oviatt, C.A., and Hale, S.S. (1976). Nitrogen Regeneration and the Metabolism of Coastal Marine Bottom Communities. The Role of Terrestrial and Aquatic Organisms in Decomposition Process. J. Anderson and A. MacFayden (Eds.). Blackwell, Oxford, pp. 269–283. 2. Jordan, T.E. and Valiela, I. (1982). A nitrogen budget of the ribbed mussel, Geukensia demissa and its significance in nitrogen flow in a New England salt marsh. Limnol. Oceanogr. 27: 75–90. 3. Murphy, R.C., and Kremer, J.N. (1985). Bivalve contribution to benthic metabolism in a California lagoon. Estuaries 8: 330–341. 4. Prins, T.C. and Smaal, A.C. (1994). The role of the blue mussel Mytilus edulis in the cycling of nutrients in Oosterschelde estuary (the Netherlands). Hydrobiol. 282/283: 413–429.

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LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS

5. Bartoli, M. et al. (2001). Impact of Tapes philippinarum on nutrient dynamics and benthic respiration in the Sacca di Goro. Hydrobiol. 455: 203–212.

23. Bayne, B.L. and Scullard, C. (1977). Rates of nitrogen excretion by species of Mytilus (Bivalvia: Mollusca). J. Mar. Biol. Ass. U.K. 57: 355–369.

6. Falcao, M. and Vale, C. (1990). Study of the Rio Formosa ecosystem: benthic nutrient remineralization and tidal variability of nutrients in the water. Hydrobiol. 207: 137–146.

24. Langton, R.W., Haines, K.C., and Lyon, R.E. (1977). Ammonia-nitrogen production by the bivalve mollusc Tapes japonica and its recovery by the red seaweed Hypnea musciformis in a tropical mariculture system. Helgol. Wiss Meeresunters 30: 217–229. 25. Goulletquer, P. et al. (1989). Ecophysiologie et bilan e´ nerg´etique de la palourde japonaise d’ e´ levage Ruditapes philippinarum. J. Exp. Mar. Biol. Ecol. 132: 85–108.

7. Gardner, W.S., Briones, E.E., Kaegi, E.C., and Rowe, G.T. (1993). Ammonium excretion by benthic invertebrates and sediment–water nitrogen flux in the Gulf of Mexico near the Mississipi river outflow. Estuaries 16: 799–808. 8. Smaal, A.C. and Zurburg, W. (1997). The uptake and release of suspended and dissolved material by oysters and mussels in Marennes-Ol´eron Bay. Aquat. Living Resour. 10: 23–30. 9. Rizzo, W.M. (1990). Nutrient exchanges between the water column and a subtidal benthic microalgal community. Estuaries 13: 219–226. ¨ 10. Sundback, K., Enoksson, V., Gran´eli, W., and Pettersson, K. (1991). Influence of sublittoral microphytobenthos on the oxygen and nutrient flux between sediment and water: a laboratory continuous-flow study. Mar. Ecol. Prog. Ser. 74: 263–279. 11. van Duyl, F.C., van Raaphorst, W., and Kop, A.J. (1993). Benthic bacterial production and nutrient sediment-water exchange in sandy North Sea sediments. Mar. Ecol. Prog. Ser. 100: 85–95. 12. Risgaard-Petersen, N., Rysgaard, S., Nielsen, L.P., and Revsbech, N.P. (1994). Diurnal variation of denitrification and nitrification in sediments colonized by benthic microphytes. Limnol. Oceanogr. 39: 573–579. 13. Feuillet-Girard, M., Gouleau, D., Blanchard, G., and Joassard, L. (1997). Nutrient fluxes on an intertidal mudflat in Marennes-Ol´eron Bay, and influence of the emersion period. Aquat. Living Resour. 10: 49–58. 14. Trimmer, M., Nedwell, D.B., Sivyer, D.B., and Malcolm, S.J. (1998). Nitrogen fluxes through the lower estuary of the river Great Ouse, England: the role of the bottom sediments. Mar. Ecol. Prog. Ser. 163: 109–124. 15. Callender, E. and Hammond, D.E. (1982). Nutrient exchange across the sediment–water interface in the Potomac river estuary. Estuar. Coast. Shelf Sci. 15: 395–413. 16. Henriksen, K., Rasmussen, M.B., and Jensen, A. (1983). Effect of bioturbation on microbial nitrogen transformations in the sediment and fluxes of ammonium and nitrate to the overlaying water. Environ. Biogeochem. 35: 193–205. 17. Yamada, H. and Kayama, M. (1987). Liberation of nitrogenous compounds from bottom sediments and effect of bioturbation by small bivalve, Theora lata (Hinds). Estuar. Coast. Shelf Sci. 24: 539–555. 18. Aller, R.C. (1988). Benthic fauna and biogeochemical processes: the role of borrow structures. In: Nitrogen Cycling in Coastal Marine Environments. T.H. Blackburn and J. Sørensen (Eds.). Wiley, pp. 301–338. 19. Luther, I.II. et al. (1998). Simultaneous measurements of O2 , Mn, Fe, I− , and S(-II) in marine pore waters with a solid-state voltammetric microelectrode. Limnol. Oceanogr. 43: 325–333. 20. Nakamura, Y. (1994). Effect of flow velocity on phosphate release from sediment. Water Sci. Technol. 30: 263–272. 21. Miller-Way, T. and Twilley, T.T. (1996). Theory and operation of continuous flow systems for the study of benthic–pelagic coupling. Mar. Ecol. Prog. Ser. 140: 257–269. 22. Asmus, R.M. et al. (1998). The role of water movement and spatial scaling for measurements of dissolved inorganic nitrogen fluxes in intertidal sediments. Estuar. Coast. Shelf Sci. 46: 221–232.

26. Cockcroft, A.C. (1990). Nitrogen excretion by the surf zone bivalves Donax serra and D. sordidus. Mar. Ecol. Prog. Ser. 60: 57–65. 27. Xie, Q. and Burnell, G.M. (1995). The effect of activity on the physiological rates of two clam species, Tapes philippinarum (Adams & Reeve) and Tapes decussatus (Linnaeus). Biol. Environ. Proc. Royal Irish Acad. 95B: 217–223. 28. Nakamura, M., Yamamuro, M., Ishikawa, M., and Nishimura, H. (1988). Role of the bivalve Corbicula japonica in the nitrogen cycle in a mesohaline lagoon. Mar. Biol. 99: 369–374. 29. Kiibus, M., and Kautsky, N. (1996). Respiration, nutrient excretion and filtratin rate of tropical freshwater mussels and their contribution to production and energy flow in Lake Kariba, Zimbabwe. Hydrobiol. 331: 25–32. 30. Magni, P. (2000). Distribution of organic matter in a tidal estuary of the Seto Inland Sea, Japan, and its relationship with the macrozoobenthic communities. In: Ad hoc Benthic Indicator Group—Results of Initial Planning Meeting. IOC Technical Series No. 57, UNESCO, pp. 20–25. 31. Magni, P. and Montani, S. (1997). Development of benthic microalgal assemblages on an intertidal flat in the Seto Inland Sea, Japan: effects of environmental variability. La mer 35: 137–148. 32. Magni, P. and Montani, S. (1998). Responses of intertidal and subtidal communities of the macrobenthos to organic load and oxygen depletion in the Seto Inland, Japan. J. Rech. Oc´eanogr. 23: 47–56. 33. Magni, P. and Montani, S. (2000). Physical and chemical variability in the lower intertidal zone of an estuary in the Seto Inland Sea, Japan: seasonal patterns of dissolved and particulate compounds. Hydrobiol. 432: 9–23. 34. Magni, P., Abe, N., and Montani, S. (2000). Quantification of microphytobenthos biomass in intertidal sediments: layerdependent variation of chlorophyll a content determined by spectrophotometric and HPLC methods. La mer 38: 57–63. 35. Magni, P., Montani, S., Takada, C., and Tsutsumi, H. (2000). Temporal scaling and relevance of bivalve nutrient excretion on a tidal flat of the Seto Inland Sea, Japan. Mar. Ecol. Progr. Ser. 198: 139–155. 36. Magni, P., Montani, S., and Tada, K. (2002). Semidiurnal dynamics of salinity, nutrients and suspended particulate matter in an estuary in the Seto Inland Sea, Japan, during a spring tide cycle. J. Oceanogr. 58: 389–402. 37. Montani, S. et al. (1998). The effect of a tidal cycle on the dynamics of nutrients in a tidal estuary in the Seto Inland Sea, Japan. J. Oceanogr. 54: 65–76. 38. Asmus, H., Asmus, R.M., and Reise, K. (1990). Exchange processes in an intertidal mussel bed: A Sylt-flume study in the Wadden Sea. Berichte Biolog. Anstalt Helgol. 6: 78. 39. Dame, R.F. et al. (1991). The influence of mussel beds on nutrients in the western Wadden Sea and Eastern Scheldt estuaries. Estuaries 14: 130–138.

TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT 40. Doering, P.H., Kelly, J.R., Oviatt, C.A., and Sowers, T. (1987). Effect of the hard clam Mercenaria mercenaria on benthic fluxes of inorganic nutrients and gases. Mar. Biol. 94: 377–383. 41. Nixon, S.W., Oviatt, C.A., Garber, J., and Lee, V. (1976). Diel metabolism and nutrient dynamics in a salt marsh embayment. Ecology 57: 740–750. ¨ 42. Schluter, L. and Josefsen, S.B. (1994). Annual variation in condition, respiration and remineralisation of Mytilus edulis L. in the Sound, Denmark. Helgol. Meeres. 48: 419–430. 43. Smaal, A.C., Vonck, A.P.M.A., and Bakker, M. (1997). Seasonal variation in physiological energetics of Mytilus edulis and Cerastoderma edule of different size classes. J. Mar. Biol. Ass. U.K. 77: 817–838. 44. Prosh, R.M. and McLachlan, A. (1984). The regeneration of surf-zone nutrients by the sand mussel, Donax Serra R¨oding. J. Exp. Mar. Biol. Ecol. 80: 221–233. 45. Snra, R.F. and Baggaley, A. (1976). Rate of excretion of ammonia by the hard clam Mercenaria mercenaria and the American oyster Crassostrea virginica. Mar. Biol. 36: 251–258. 46. Mann, R. and Glomb, S.J. (1978). The effect of temperature on growth and ammonia excretion of the Manila clam Tapes japonica. Estuar. Coast. Shelf Sci. 6: 335–339. 47. Mann, R. (1979). The effect of temperature on growth, physiology, and gametogenesis in the Manila clam Tapes philippinarum (Adams & Reeve, 1850). J. Exp. Mar. Biol. Ecol. 38: 121–133. 48. Boucher, G. and Boucher-Rodoni, R. (1988). In situ measurements of respiratory metabolism and nitrogen fluxes at the interface of oyster beds. Mar. Ecol. Prog. Ser. 44: 229–238.

TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT

11

dense assemblages of bivalves, it has been shown, play a major role in these processes (14–16). BACKGROUND The contribution of benthic macrofauna to the total upward flux of nutrients has been investigated (mostly for ammonium and, to a lesser extent, for phosphate) in many coastal and estuarine areas using several approaches. They include laboratory and mesocosm experiments (17–19). In situ benthic chambers and sediment core incubations (20–22), and open flow/tunnel systems (16,23–26). Measurements of macrofauna-influenced nutrient flux, however, often have temporal limitations as they are based on one or relatively few sampling occasions, and thus seasonal patterns are in most cases not known. In this article, we evaluate the magnitude and temporal scaling of biogenic flux of nutrients from intertidal sediments densely populated by bivalves, based on extrapolating nutrient excretion rates of dominant bivalves to a field community. In particular, we show that the seasonal pattern of nutrient fluxes can be strongly influenced by the animal standing stock and its temporal distribution. This is beside the effect and importance of variation in excretion rate due to animal physiological factors, such as seasonal cycles of gametogenesis, storage and use of body reserves, and water temperature (27,28). The nutrient species considered in this study include ammonium, phosphate, and silicate, for which we quantified in associated laboratory experiments LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS the excretion rates of different size classes of two dominant bivalve species. The relevance of macrofaunal excretion in regenerating the inorganic forms of three major bioelements such as N, P, and Si is discussed.

PAOLO MAGNI IMC—International Marine Centre Torregrande-Oristano, Italy

SHIGERU MONTANI Hokkaido University Hakodate, Japan

Beside light and temperature (1–5), nutrients such as ammonium (NH4 + -N), phosphate (PO4 3− -P), and silicate [Si(OH)4 -Si] are a key factor in controlling the growth, abundance, and structure of primary producers in the ocean (6,7). Hence, it is important to investigate the availability, sources, and distribution of these nutrient species, as well as their spatial and temporal scaling. Biological processes strongly influence nutrient regeneration in different marine systems. In the open ocean, an important portion of reduced N-forms (e.g., NH4 + -N), for instance, is made available in situ from waste products of plankton metabolism (8,9) and supports the socalled ‘‘regenerated’’ primary production (10). In coastal marine ecosystems, benthic nutrient regeneration is a major driving force in cycling biophilic elements (e.g., N, P, and Si) (11–13) and abundant macrofauna, for example,

MACROFAUNAL COMMUNITIES We present here the macrofaunal composition and distribution at an individual station (Stn B5) of a transect line selected in a sandflat of the Seto Inland Sea of Japan (29). At this station, the porewater nutrient concentrations (ammonium, phosphate, and silicate) in the uppermost 10 cm of sediments were also investigated in parallel from January 1995 to April 1996. They will be the subject of a subsequent associated paper focusing on the relationship between the seasonal pattern of bivalve nutrient excretion, described here, and the seasonal variation of porewater nutrient concentrations SEASONAL COUPLING BETWEEN INTERTIDAL MACROFAUNA AND SEDIMENT COLUMN POREWATER NUTRIENT CONCENTRATIONS. The total density and biomass of macrofauna varied from 7,400 (July 1995) to 22,050 ind.m−2 (October 1995), and from 70.9 (July 1995) to 244 g DW m−2 (September 7, 1995), respectively (Fig. 1). The bivalves Ruditapes philippinarum and Musculista senhousia and the polychaetes Ceratonereis erithraeensis and Cirriformia tentaculata were dominant; they accounted for 60.5% and 94.7% of the total density and biomass, respectively. Remarkably, R. philippinarum and M. senhousia alone accounted for up to 83.3 ± 6.7% of the total biomass when this exceeded

12

TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT

estimate the magnitude and temporal scaling of biogenic nutrient excretion, we used an indirect approach. The mean excretion rates of ammonium, phosphate, and silicate for the two dominant bivalves R. philippinarum and M. senhousia, which were obtained in laboratory experiments LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS, were applied to the relevant monthly biomass values found in the field. The ammonium and phosphate excretion rates of each bivalve species and their scaling to a field community have been extensively reported in Magni et al. (29). In this article, we applied these excretion rates to the bivalve biomass found at Stn B5 and extended this scaling to silicate, whose size-class dependent excretion rates are presented in LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS. For silicate, we adopted the same temperature-dependent excretion rate factors as those used for ammonium and phosphate (Table 1).

20000 15000 10000

0

Biomass, g DW m−2

(b) 250 200 150

20 Jan 17 Feb 17 Mar 15 Apr 16 May 30 May 14 Jun 12 Jul 10 Aug 7 Sep 29 Sep 30 Oct 27 Nov 26 Dec 24 Jan 22 Feb 21 Mar 17 Apr

5000

Others Cirriformia tentaculata C.nereis erithraeensis Ruditapes philippinarum Musculista senhousia

100

BIVALVE NUTRIENT EXCRETION

50 0

The highest excretion rates of nutrients were estimated in September 7, 1995, up to a total of 50.2, 7.5, and 34.1 mmol m−2 d−1 for ammonium, phosphate, and silicate, respectively (Fig. 2). This corresponded to the period of highest biomass of both R. philippinarum and M. senhousia, which also accounted for the highest bivalve percentage (91.8%) of the total macrofaunal biomass. The lowest excretion rates occurred in April 1995 for R. philippinarum (lowest biomass on the same occasion), in March 1995 for M. senhousia (lowest biomass on May 16, 1995), and in February 1995 as the sum of the two bivalve species excretion rates. These latter rates were 4.1, 0.64, and 2.9 mmol m−2 d−1 for ammonium, phosphate, and silicate, respectively. The upward flux rates of nutrients obtained through this extrapolation of laboratory experiments on bivalve nutrient excretion to a field community are comparable to the highest biogenic releases reported for dense assemblages of bivalves such as oyster reefs (34) and mussel beds (16,35). This study also points to the importance of bivalve excretion to the biogenic regeneration of silicate, as previously suggested by field measurements indicating evidence of increased levels of silicate flux in the

20 Jan 17 Feb 17 Mar 15 Apr 16 May 30 May 14 Jun 12 Jul 10 Aug 7 Sep 29 Sep 30 Oct 27 Nov 26 Dec 24 Jan 22 Feb 21 Mar 17 Apr

Density, ind m−2

(a) 25000

1995

1996 Time, month

Figure 1. Seasonal variation of density (a) and biomass (b) of dominant macrozoobenthic species at Stn B5 (29). Note that in May 1995 and September 1995, sampling was carried out fortnightly.

120 g DW m−2 , from August 1995 till the end of the investigations. The high values of total macrofaunal and bivalve biomass may be regarded as a typical feature of many estuarine and intertidal areas, which are amongst the most productive systems in the ocean (30). In addition, biomass was markedly lower during the first half of the year (January 1995 to July 1995) than between late summer and winter; yet values progressively increased from early spring (March 1995) to early summer (June 1995). These marked temporal changes of macrofaunal communities reflect the high variability of these ecosystems. To

Table 1. Adopted Temperature-Dependent Excretion Rates of Ammonium (NH4 + -N), Phosphate (PO4 3− -P), and Silicate [Si(OH)4 -Si] for Ruditapes philippinarum and Musculista senhousia Period

Temperature, ◦ C

Month

Station

Dec, Jan, Feb Nov, Mar – Apr, May, Jun, Oct Jul, Aug, Sep a

AVG SD AVG SD AVG SD AVG SD

Excretion Rate, µmol g−1 DW h−1 R. philippinarum

B5-B1

H1

Y3

4.7 ±1.8 11.7 ±0.9 19.6 ±2.8 27.7 ±2.3

5.9 ±1.8 11.9 ±0.2 19.9 ±4.0 28.0 ±3.1

10.4 ±2.1 15.2 ±3.0 21.7 ±3.6 26.9 ±1.1

f

a

0.6 – 0.9 – 1 – 0.9 –

NH4 4.3 – 6.4 – 7.1 – 6.4 –

+

PO4

3−

1.1 – 1.7 – 1.9 – 1.7 –

M. senhousia Si(OH)4 4.0 – 5.9 – 6.6 – 5.9 –

a

NH4 +

PO4 3−

Si(OH)4

0.5 – 0.7 – 1 – 1.2 –

5.9 – 8.3 – 11.8 – 14.2 –

0.7 – 1.0 – 1.4 – 1.7 –

3.7 – 5.1 – 7.3 – 8.8 –

f

A factor (f ) of 1 is used for the mean of the excretion rates obtained in laboratory experiments (see LABORATORY EXPERIMENTS ON BIVALVE EXCRETION RATES OF NUTRIENTS and based on Ref. 29).

NH4+-N, mmol m−2 day−1

TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT

60 50

Ruditapes philippinarum Musculista senhousia

40 30 20 10

20

Ruditapes philippinarum Musculista senhousia

15

BIBLIOGRAPHY 5 0

40

Ruditapes philippinarum Musculista senhousia

30

1. Admiraal, W. and Peletier, H. (1980). Influence of seasonal variations of temperature and light on the growth rate of culture and natural population of intertidal diatoms. Mar. Ecol. Progr. Ser. 2: 35–43. 2. Harrison, W.G. and Platt, T. (1980). Variations in assimilation number of coastal marine phytoplankton: effects of environmental co-variates. J. Plank. Res. 2: 249–260. 3. Verhagen, J.H.G. and Nienhuis, P.H. (1983). A simulation model of production, seasonal changes in biomass and distribution of eelgrass (Zoostera marina) in Lake Grevelingen. Mar. Ecol. Progr. Ser. 10: 187–195.

20 10 0 20 Jan 17 Feb 17 Mar 15 Apr 16 May 30 May 14 Jun 12 Jul 10 Aug 7 Sep 29 Sep 30 Oct 27 Nov 26 Dec 24 Jan 22 Feb 21 Mar 17 Apr

Si(OH)4-Si, mmol m−2 day−1

by a comparison with the extent of benthic nutrient regeneration through diffusive flux. In particular, nutrient flux measured from nutrient concentrations in the porewater in adjacent intertidal and coastal areas was more than one order of magnitude lower; it varied from 0.2 to 1.5 mmol NH4 + -N m−2 d−1 and from 0.01 to −2 d−1 . It can be inferred that a 0.05 mmol PO3− 4 -P m marked increase in biogenic nutrient regeneration is importantly controlled by the animal biomass increase (36) and has a major impact, acting as a positive feedback, on primary producers (41). These results indicate that abundant macrofauna and its excretory products play a primary role in benthic nutrient regeneration, are well balanced in their stoichiometric ratios, and thus act as a major factor to support primary production within the intertidal zone.

10

20 Jan 17 Feb 17 Mar 15 Apr 16 May 30 May 14 Jun 12 Jul 10 Aug 7 Sep 29 Sep 30 Oct 27 Nov 26 Dec 24 Jan 22 Feb 21 Mar 17 Apr

PO43−-P, mmol m−2 day−1

20 Jan 17 Feb 17 Mar 15 Apr 16 May 30 May 14 Jun 12 Jul 10 Aug 7 Sep 29 Sep 30 Oct 27 Nov 26 Dec 24 Jan 22 Feb 21 Mar 17 Apr

0

13

1996

1995 Time, month

Figure 2. Magnitude and temporal scaling of ammonium (NH4 + -N), phosphate (PO4 3− -P), and silicate [Si(OH)4 -Si] excretion by Ruditapes philippinarum and Musculista senhousia in a field community.

presence of bivalves (17,24,36). The temporal scaling of bivalve nutrient excretion showed a marked seasonal pattern of large variations of nutrient flux, up to ca. 10-fold (R. philppinarum) and 20-fold (M. senhousia) between March–April 1995 and September 7, 1995, and a progressive decrease from late summer through winter. This approach may involve some limitations, such as the effect of differences between the bivalve performance in controlled laboratory experiments and that in the field and a relative approximation in adopting different excretion rates at temperatures other than those actually employed in the laboratory experiments (29). However, it indicates the strong influence of animal distribution on the magnitude and temporal scaling of biogenic nutrient regeneration due to bivalve excretion. The great potential of this biogenic source of nutrients in cycling biophilic elements can also be highlighted

4. Giesen, W.B.J.T., van Katwijk, M.M., and den Hartog, C. (1990). Eelgrass condition and turbidity in the Dutch Wadden Sea. Aquat. Bot. 37: 71–85. 5. Agust´ı, S. et al. (1994). Light harvesting among photosynthetic organisms. Funct. Ecol. 8: 273–279. 6. Justic, D., Rabalais, N.N., and Turner, T. (1995). Stoichiometric nutrient balance and origin of coastal eutrophication. Mar. Poll. Bull. 30: 41–46. 7. Valiela, I. et al. (1997). Macroalgal blooms in shallow estuaries: controls and ecophysiological and ecosystem consequences. Limnol. Oceanogr. 5: 1105–1118. 8. Park, Y.C. and Carpenter, E.J. (1987). Ammonium regeneration and biomass of macrozooplankton and Ctenophores in Great South Bay, New York. Estuaries 4: 316–320. 9. Wen, Y.H. and Peters, R.H. (1994). Empirical models of phosphorus and nitrogen excretion rates by zooplankton. Limnol. Oceanogr. 39: 1669–1679. 10. Harrison, W.G. et al. (1992). Nitrogen dynamics at the VERTEX time-series site. Deep-Sea Res. 39: 1535–1552. 11. Rowe, G.T., Clifford, C.H., Smith, K.L., and Hamilton, P.L. (1975). Benthic nutrient regeneration and its coupling to primary productivity in coastal waters. Nature 225: 215–217. 12. Nixon, S.W. (1981). Remineralization and nutrient cycling in coastal marine ecosystems. In: Nutrient Enrichment in Estuaries. B. Nelson and L.E. Cronbin (Eds.). Humana Press, Clifton, NJ, pp. 111–138. 13. Klump, J.V. and Martens, C.S. (1983). Benthic nitrogen regeneration. In: Nitrogen in the Marine Environment. E.J. Carpenter and D.G. Capone (Eds.). Academic Press, New York, pp. 411–457. 14. Jordan, T.E. and Valiela, I. (1982). A nitrogen budget of the ribbed mussel, Geukensia demissa, and its significance in

14

BREAKWATERS nitrogen flow in a New England salt marsh. Limnol. Oceanogr. 27: 75–90.

15. Murphy, R.C. and Kremer, J.N. (1985). Bivalve contribution to benthic metabolism in a California lagoon. Estuaries 8: 330–341. 16. Prins, T.C. and Smaal, A.C. (1994). The role of the blue mussel Mytilus edulis in the cycling of nutrients in Oosterschelde estuary (the Netherlands). Hydrobiol. 282/283: 413–429. 17. Doering, P.H., Kelly, J.R., Oviatt, C.A., and Sowers, T. (1987). Effect of the hard clam Mercenaria mercenaria on benthic fluxes of inorganic nutrients and gases. Mar. Biol. 94: 377–383. 18. Nakamura, M., Yamamuro, M., Ishikawa, M., and Nishimura, H. (1988). Role of the bivalve Corbicula japonica in the nitrogen cycle in a mesohaline lagoon. Mar. Biol. 99: 369–374. 19. Kiibus, M. and Kautsky, N. (1996). Respiration, nutrient excretion and filtration rate of tropical freshwater mussels and their contribution to production and energy flow in Lake Kariba, Zimbabwe. Hydrobiol. 331: 25–32. 20. Yamada, H. and Kayama, M. (1987). Liberation of nitrogenous compounds from bottom sediments and effect of bioturbation by small bivalve, Theora lata (Hinds). Estuar. Coast. Shelf Sci. 24: 539–555. 21. G´omez-Parra, A. and Forja, J.M. (1993). Benthic fluxes in Cadiz Bay (SW Spain). Hydrobiol. 252: 23–34. 22. Yamamuro, M. and Koike, I. (1993). Nitrogen metabolism of the filter-feeding bivalve Corbicula japonica and its significance in primary production of a brackish lake in Japan. Limnol. Oceanogr. 35: 997–1007. 23. Dame, R.F., Zingmark, R.G., and Haskin, E. (1984). Oyster reefs as processors of estuarine materials. J. Exp. Mar. Biol. Ecol. 83: 239–247. 24. Asmus, H., Asmus, R.M., and Reise, K. (1990). Exchange processes in an intertidal mussel bed: a Sylt-flume study in the Wadden Sea. Berichte Biolog. Anstalt Helgol. 6: 78. 25. Prins, T.C. and Smaal, A.C. (1990). Benthic-pelagic coupling: the release of inorganic nutrients by an intertidal bed of Mytilus edulis. In: Trophic relationships in the marine environment. Proc. 24th Europ. Mar. Biol. Symp., Aberdeen Univ. Press, pp. 89–103. 26. Asmus, R.M., Asmus, H., Wille, A., Zubillaga, G.F., and Reise, K. (1994). Complementary oxygen and nutrient fluxes in seagrass beds and mussels banks? In: Changes in Fluxes in Estuaries: Implications from Science to Management. K.R. Dyer and R.J. Orth (Eds.). Olsen & Olsen, Fredensborg, pp. 227–238. 27. Bayne, B.L. and Scullard, C. (1977). Rates of nitrogen excretion by species of Mytilus (Bivalvia: Mollusca). J. Mar. Biol. Ass. U.K. 57: 355–369. ¨ 28. Schluter, L. and Josefsen, S.B. (1994). Annual variation in condition, respiration and remineralisation of Mytilus edulis L. in the Sound, Denmark. Helgol. Meeres. 48: 419–430. 29. Magni, P., Montani, S., Takada, C., and Tsutsumi, H. (2000). Temporal scaling and relevance of bivalve nutrient excretion on a tidal flat of the Seto Inland Sea, Japan. Mar. Ecol. Progr. Ser. 198: 139–155. 30. Heip, C.H.R. et al. (1995). Production and consumption of biological particles in temperate tidal estuaries. Oceanogr. Mar. Biol. Ann. Rev. 33: 1–149. 31. Magni, P. and Montani, S. (1997). Development of benthic microalgal assemblages on a tidal flat in the Seto Inland Sea, Japan: effects of environmental variability. La mer 35: 109–120.

32. Magni, P. and Montani, S. (1998). Responses of intertidal and subtidal communities of the macrobenthos to organic load and oxygen depletion in the Seto Inland Sea, Japan. J. Rech. Oc´eanogr. 23: 47–56. 33. Magni, P. and Montani, S. (2000). Physical and chemical variability in the lower intertidal zone of an estuary in the Seto Inland Sea, Japan: seasonal patterns of dissolved and particulate compounds. Hydrobiol. 432: 9–23. 34. Dame, R.F., Wolaver, T.G., and Libes, S.M. (1985). The summer uptake and release of nitrogen by an intertidal oyster reef. Neth. J. Sea Res. 19: 265–268. 35. Dame, R.F. et al. (1991). The influence of mussel beds on nutrients in the western Wadden Sea and Eastern Scheldt estuaries. Estuaries 14: 130–138. 36. Bartoli, M. et al. (2001). Impact of Tapes philippinarum on nutrient dynamics and benthic respiration in the Sacca di Goro. Hydrobiol. 455: 203–212. 37. Kuwae, T., Hosokawa, Y., and Eguchi, N. (1998). Dissolved inorganic nitrogen cycling in Banzu intertidal sand-flat, Japan. Mangroves Salt Marshes 2: 167–175. 38. Matsukawa, Y., Sato, Y., and Sasaki, K. (1987). Benthic flux of nutrient salts on an intertidal flat. Nippon Suisan Gakkaishi 53: 985–989. 39. Takayanagi, K. and Yamada, H. (1999). Effects of benthic flux on short term variations of nutrients in Aburatsubo Bay. J. Oceanogr. 55: 463–469. 40. Yamamoto, T., Matsuda, O., Hashimoto, T., Imose, H., and Kitamura, T. (1998). Estimation of benthic fluxes of dissolved inorganic nitrogen and phosphorus from sediments of the Seto Inland Sea. Umi to Kenkyu (Oceanogr. Soc. Japan) 7: 151–158 (in Japanese). 41. Peterson, B.J. and Heck, K.L. Jr. (2001). Positive interaction between suspension-feeding bivalves and seagrass—a facultative mutualism. Mar. Ecol. Progr. Ser. 213: 143–155.

BREAKWATERS STEFANO PAGLIARA PIETRO CHIAVACCINI Universita` di Pisa Pisa, Italy

Breakwaters are coastal structures used to protect harbor and shore areas by dissipating and reflecting wave energy. They are built to — reduce wave disturbance in coastal and harbor areas and preserve related activities; — protect ships and boats from wave forces; — when located near shore, in the same direction as the coastline, they can stabilize the coastline, modifying cross-shore and long-shore sediment transport. The choice of the type of structure depends on the availability of materials used, the characteristics of the incident wave, the bottom morphology, the geotechnical parameters of the soil, and the necessity of obtaining a flexible or rigid structure. Breakwaters can be classified as rubble-mound structures, vertical breakwaters, and floating breakwaters.

BREAKWATERS

RUBBLE-MOUND STRUCTURES The typical cross section of a rubble-mound breakwater is sketched in Fig. 1. It consists of different layers of stones. The center core is made up of quarry run. The external layer (armor) consists of large armor units, that can be either rock or specially designed concrete units (cubes, tetrapods, dolos). The breakwater crest is generally 1–2 m over the still water level (SWL). The crest width should be large enough to allow transport and installation of material during construction and when a repair is made (1). Because of the relative dimensions of the units of the armor and those of the core, in some cases, it is necessary to build the breakwater as a filter of three or four layers (underlayers), so that the finer material of the core cannot be removed by the waves through the voids of the armor layer. To prevent removal of finer material, the filter must satisfy the following relations: D15 (upper layer) < 4 ÷ 5D85 (lower layer)

(1)

D15 (upper layer) < 20 ÷ 25D15 (lower layer)

(2)

Armor Stability. The stability formulas are based on experiments carried out on hydraulic models. One of the most used stability formulas is Hudson’s (3) determined for a nonovertopping structure: H = (K cot α)1/3 D2n50 where

Stability The rubble-mound breakwater causes the dissipation of wave energy by generating eddies due to the breaking. The voids and the roughness of the structural material, as well as the permeability of the structure, are very important in the dissipation process. The wave energy entering through the structure creates shear stresses that can move the masses causing loss of stability.

Seaward

Crest width Breakwater crest

Crown wall

(3)

H = characteristic height of the wave (Hs to H1/10 ); Dn50 = equivalent cube length of median rock; α = slope angle;  = (ρs /ρw − 1) where ρs and ρw are rock density and water density, respectively; K = stability coefficient (Tables 1 and 2).

The damage D represents the measure of the modification of the structure’s profile under wave action. The damage can be defined by counting the number of rocks moved or by measuring the variation of the armor layer Table 1. Values of the Stability Coefficient K for H = Hs a Damage D 0–5%

where D15 = nominal size that is exceeded by the 85% of the sample D85 = nominal size that is exceeded by the 15% of the sample A toe filter is necessary if the breakwater is built on erodible material. The toe filter prevents breaking waves from removing material from the base of the structure. If the breakwater is located in shallow water, the filter toe is exposed to extreme wave action. To avoid, or just limit, wave overtopping, it is possible to use a concrete structure (crown wall) located over the crest (Fig. 1). When the breakwater is small and not high, it is possible to avoid using a center core. This kind of structure, of single sized stones, is called a ‘‘reef breakwater,’’ and it is normally used for small submerged breakwaters (2).

15

Stone Shape

Placement

Breaking Waves

Nonbreaking Waves

Smooth, rounded Rough, angular Rough, angular

Random Random Special

2.1 3.5 4.8

2.4 4.0 5.5

a

Slope 1.5 ≤ cot α ≤ 3.0.

Table 2. Values of the Stability Coefficient K for H = H1/10 Damage D 0–5% Breaking Waves Non Breaking Waves Stone shape Placement Trunk Smooth, rounded Rough, angular Rough, angular Tetrapods a b

Head

Trunk

Head

Random

1.2

1.1

2.4

1.9

Random

2.0

1.3a –1.9b

4.0

2.3a –3.2b

Special

5.8

5.3

7.0

6.4

Random

7.0

3.5a –5b

8.0

4.0a –6.0b

cot α = 3. cot α = 1.5.

Landward

SWL

r

Berm

mo

Ar

Underlayer

TOE filter core Figure 1. Typical section of a rubble-mound breakwater.

16

BREAKWATERS

area (eroded area A of the damaged section). For this second case, Broderick (4) introduced a parameter (relative eroded area) defined as S=

A D2n50

(4)

where Dn50 is the nominal diameter, corresponding to 50% of the weight of the sample. The damage can be considered the number of masses of dimension equal to Dn50 eroded in a strip of section of the same length. Zero damage means that there is nominally no removal of the armor units from the breakwater face. The K value of Hudson’s formula is different for the trunk and the head of the structure. The stones will be less stable on the head than on the trunk. In this case, K must be decreased by about 20%. Van der Meer (5) derives expressions that include some additional parameters for an incident wave: Hs −0.5 = 6.2S0.2 P0.18 Nz−0.1 ξm Dn50 plunging waves ξm < ξcm Hs P = 1.0S0.2 P−0.13 Nz−0.1 (cot α)0.5 ξm Dn50 surging waves ξm > ξcm ξm =

−0.5 sm

(5)

(6)

tan α

ξmc = [6.2P0.31 (tan α)0.5 ]1/(P+0.5) where

S = relative eroded area (normally equal to 2); P = notional permeability; Nz = number of waves; sm = wave steepness sm = Hs /Lom ; Lom = deepwater wavelength corresponding to the mean period.

For a homogeneous structure (no core, no filter, and stones of the same size), P = 0.6; a rock armor layer with a permeable core gives P = 0.5; an armor layer with filter on a permeable core gives P = 0.4. For a breakwater with an impermeable core, P = 0.1. For overtopped and low crested structures, Van der Meer suggests multiplying Dn50 by a reduction factor fi defined by


Rc sop (7) fi = 1.25 − 4.8 Hs 2π Rc is the freeboard, and sop = Hs /Lop (Lop is the deep water wavelength referred to the peak  Expression (7)
period). Rc sop < 0.052. can be used in the range 0 < Hs 2π For submerged breakwaters, the following expression can be used (6): hc = (2.1 + 0.1S) exp(0.14Ns∗ ) h where h = water depth; hc = height of the structure from the base; S = relative eroded area; Hs −1/3 Ns∗ = sp = spectral stability number. Dn50

(8)

Run Up and Overtopping Run up is a phenomenon in which the incident crest wave runs up along a sloping structure to a level higher then the original wave crest. Together with overtopping, it plays, a very important role in the design of a rubble-mound structure because it depends on the characteristics of the structure (slope roughness, berm length, permeability). Run up is expressed by Ru,x% that represents the level reached by the wave exceeded in x% of the cases by the incident wave. The run up level is referred to the SWL. For rubble-mound structures, Van der Meer’s (6) formula is used: Ru,x% = aξm , Hs Ru,x% c = bξm , Hs

ξm < 1.5

(9)

ξm > 1.5

(10)

These formulas are valid for breakwaters that have an impermeable or almost impermeable core (P < 0.1). If the breakwaters have a permeable core (0.1 < P < 0.4) Equations 9 and 10 become Ru,x% =d Hs

(11)

It is usual in breakwater design to consider that x% = 2; this means that Ru,2% is the run up exceeded by 2% of the waves. In this case, the values of the parameters of Equations (9,10), and 11 are a = 0.96, b = 1.17, c = 0.46, and d = 1.97. ξm is the breaker parameter for deep water, correspond−0.5 tan α where the symbols ing to the mean period (ξm = sm are explained in Equations 5 and 6). In a low crest elevation, overtopping is allowed. Overtopping is the quantity of water passing over the crest of a structure per unit time, and it has the same dimensions of a discharge Q(m3 /s), often expressed for unit length q[m3 /(sm)]. A knowledge of overtopping is important in defining the necessary protection of the splash area and in assessing the risk to people or installations behind the breakwater. The amount of overtopping varies considerably from wave to wave; the overtopping discharge changes in time and space, and the greatest quantity is due to a small number of the incident waves. Wave overtopping for an impermeable rock armored slope structure with a crown wall can be expressed by the equation of Bradbury and Allsop (7), using the parameters of Aminti and Franco (8): 


−b Rc 2 som q =a (12) gHs Tom Hs 2π where Som = deepwater wave steepness, based on mean period; Hs = significant wave height; Rc = crest freeboard relative to SWL in m; Tom = deepwater wave mean period; a, b = parameters as specified in Table 3. G is the width (seaward) of the armor crest till the crown wall, and α is the slope of the armor layer.

BREAKWATERS Table 3. Coefficients for Equation 15 from Experimental Results Armor Units

Cotα

G/Hs

a

b

Rock

2 2 2 1.33 1.33 1.33 2 2 2 1.33 1.33 1.33

1.10 1.85 2.60 1.10 1.85 2.60 1.10 1.65 2.6 1.10 1.85 2.60

17 × 10−8 19 × 10−8 2.3 × 10−8 5.0 × 10−8 6.8 × 10−8 3.1 × 10−8 1.9 × 10−8 1.3 × 10−8 1.1 × 10−8 5.6 × 10−8 1.7 × 10−8 0.92 × 10−8

2.41 2.30 2.68 3.10 2.65 2.69 3.08 3.80 2.86 2.81 3.02 2.98

Tetrapods

transmitted to incident characteristic wave height or the ratio of the square of transmitted mean energy to incident mean wave energy: Kt =

Hs Rc

3

Hs2 Ac B cot α

(13)

where Ac = level of the berm from SWL; B = the width of the berm; Rc = the level of the crest of the crown wall from SWL; α = slope of armor. Wave Reflection Each coastal structure causes a wave reflection. Reflection plays a very important role because of the interaction between reflected and incident waves that can create a very confused sea, increasing the wave steepness. It is a problem especially at the entrance of an harbor because the steepness makes ship and boat maneuver very difficult. Besides, strong reflection increases the erosive force in front of the structure. Rubble-mound breakwaters, which are permeable, rough, and sloping structures and structure of limited crest level, absorb a significant portion of the wave energy. For these structures, the reflection coefficient is small. For nonovertopped structures, that have a theoretical permeability P, we can use the following equation (10): −0.46 Kr = 0.071P−0.082 (cot α)−0.62 Sop

(14)

where Kr is the ratio of the reflected wave height and the incident wave height. Wave Transmission When energy passes over and through a breakwater, there is a wave transmission. The wave action in the landward side of the structure is smaller than in the seaward side. A wave is transmitted when a considerable amount of water overtops the structure and when the breakwater is very permeable and the wave period is relatively long. We define the coefficient of transmission as the ratio of




Est Es

0.5 (15)


 Hs Rc − 0.24 +b Kt = 0.031 Dn50 Dn50

For rock armored permeable slopes, that have a theoretical permeability P = 0.4 and a berm in front of a crown wall, we can use the Pedersen and Burcharth formula (9):


Hst = Hs

For rock armored, low crested, submerged and reef breakwaters, we can use the Van der Meer and d’Angremond formula (11):

where

qTom = 3.2 × 10−5 L2om

17

(16)


Hs B 1.84 − 0.0017 Dn50 Dn50 + 0.51 for a conventional structure, Hs b = −2.6sop − 0.05 + 0.85 for a reef type Dn50 structure, Hs = significant wave height; som = deepwater wave steepness, based on peak period; Rc = crest freeboard relative to SWL, negative for submerged breakwaters; B = width of crest; Dn50 = median of nominal diameter of rock for design conditions. b = −5.42sop + 0.0323

For conventional structures, Kt has a maximum of 0.75 and a minimum of 0.075, and for reef type structures, Kt varies between 0.15 and 0.6. This formula can be used in the following range: 1
30fold variability between layers and months (Fig. 2). Ammonium varied from 75.1 (0–0.5 cm, January 1995) to

Jan

PAOLO MAGNI

SEASONALITY OF POREWATER NUTRIENT CONCENTRATIONS

Feb

SEASONAL COUPLING BETWEEN INTERTIDAL MACROFAUNA AND SEDIMENT COLUMN POREWATER NUTRIENT CONCENTRATIONS

still a noticeable lack of knowledge about the relationship between macrofauna and the temporal distribution and spatial variability of nutrients in sediments. Some works have conducted a seasonal study of porewater chemistry focusing on nitrogen organic compounds, such as dissolved free amino acids (13), and ammonium profiles and production (13–15). These studies have indicated that temperature has a pronounced effect on the porewater ammonium production rate and seasonal variation. In addition to the effect of environmental parameters that exhibit seasonal patterns and mineralization processes in sediments by bacteria, macrofauna also contribute to the total benthic metabolism by feeding, assimilation, and respiration. Recently, we have shown that the excretory activity of macrofauna strongly influences the magnitude and seasonal variability of the biogenic upward flux of nutrients (16; TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVE-DOMINATED TIDAL FLAT). However, field studies on the relationship between porewater nutrient concentrations and macrofaunal communities come often from isolated surveys (17) or from transplantation/manipulation experiments (8,9). In particular, evidence of a coupling between the seasonal variability of nutrient concentrations in porewater and macrofauna-influenced upward flux of nutrients is still lacking.

Temperature, °C

the internet through real-time observations. The goal of this project is to put in place the infrastructure needed to simplify sensor deployment and data acquisition to allow information access by scientific researchers, educators and the public. This is an important contribution to GLERL’s leadership in supporting and promoting observation system development among Great Lakes universities and non-governmental organizations. The environmental observatory consists of an offshore buoy connected to a hub that receives data from various environmental sensors such as an acoustic doppler current profiler. The data are then sent through a wireless link to an onshore receiver connected to the internet.

73

’96

Figure 1. Seasonal variation of sediment temperature.

SEASONAL COUPLING NH4+-N

600 500 400 300 200

17 Apr

21 Mar

24 Jan

22 Feb

26 Dec

27 Nov

3 Oct

30 Oct

7 Sep

12 Jul

10 Aug

14 Jun

16 May

12 Apr

6 Feb

1 Mar

100

PO43−-P

µM

0–0.5 0.5–2 2–3 3–4 4–5 5–6 6–7 7–8 8–9 9–10

30 25 20 15 10

17 Apr

21 Mar

22 Feb

24 Jan

26 Dec

27 Nov

30 Oct

3 Oct

7 Sep

10 Aug

12 Jul

14 Jun

16 May

12 Apr

1 Mar

Si(OH)4-Si

µM

0–0.5 0.5–2 2–3 3–4 4–5 5–6 6–7 7–8 8–9 9–10

300 250 200 150 100

1995

908 µM (0–0.5 cm, 30 October 1995), phosphate from 0.9 (0–0.5 cm, January 1995) to 36.9 µM (4–5 cm, September 1995), and silicate from 17.1 (0–0.5 cm, February 1996) to 379 µM (5–6 cm, 30 September 1995). The spatial and temporal distribution of ammonium, phosphate, and silicate concentrations were consistent with each other. They were lowest in winter, progressively increased through spring and summer in the uppermost layers, and were highest between September and October 1995; a major increase occurred in intermediate layers (i.e., between 3 and 8 cm). Subsequently, minor but noticeable peaks of ammonium and phosphate concentrations were also found in March 1996, up to 518 and 32.7 µM at 6–7 cm, respectively. For each sampling occasion and nutrient species, we summed the concentrations measured in each layer of the sediment column to be representative of an alllayer monthly pool expressed on a square meter basis.

Time, month

17 Apr

21 Mar

22 Feb

24 Jan

26 Dec

27 Nov

30 Oct

3 Oct

7 Sep

10 Aug

12 Jul

14 Jun

16 May

12 Apr

1 Mar

50 13 Jan

Sediment depth, cm

(c)

6 Feb

5 13 Jan

Sediment depth, cm

(b)

Figure 2. Spatial and seasonal variation of porewater ammonium (NH4 + -N) (a), phosphate (PO4 3− -P) (b), and silicate [Si(OH)4 -Si] (c) concentrations.

µM

0–0.5 0.5–2 2–3 3–4 4–5 5–6 6–7 7–8 8–9 9–10 13 Jan

Sediment depth, cm

(a)

6 Feb

74

1996

This was obtained by calculating the sediment porosity of each layer from the water content (weight loss on drying at 105◦ C for 20 h), assuming the bulk density of sediment particles as 2.5 g cm−3 . Each volume of porewater was subsequently multiplied by the relevant nutrient concentrations, which were finally expressed as areal depth-integrated values (mmol m−2 ). Ammonium, phosphate, and silicate concentrations showed a strong correlation with each other (Fig. 3). These results suggest that similar and/or coincident processes may govern the spatial and seasonal variability of major inorganic forms of N, P, and Si in sediment porewater. We were thus interested in assessing the existence of common environmental factors (i.e., temperature) and/or biological (i.e., macrofauna-influenced) processes that influence the variability of ammonium, phosphate, and silicate in porewater. Ammonium and silicate were highly correlated with temperature (Fig. 4). This could

SEASONAL COUPLING

35

r2 = 0.600 (p < 0.001)

25

Temperature, °C

Porew. NH4+-N, mmol m−2

30

20 15 10

30 25 20 15 10 5

5

0 0.0

0.3

0.6

0.9

1.2

Porewater PO43−-P, mmol m−2 30

5 10 15 20 25 Porewater NH4+-N, mmol m−2

30

35

r2 = 0.853 (p < 0.001)

Temperature, °C

Porew. NH4+-N, mmol m−2

r2 = 0.684 (p < 0.001)

0

0

25 20 15

30 25 20 15 10 r2 = 0.331 (p < 0.05)

5

10

0

5

0.0 3 6 9 12 Porewater Si(OH)4-Si, mmol m−2

15

15 r2

0.6

0.9

1.2

35 Temperature, °C

0

0.3

Porewater PO43−-P, mmol m−2

0

Porew. Si(OH)4-Si, mmol m−2

75

= 0.460 (p < 0.01)

12 9

30 25 20 15 10 r2 = 0.679 (p < 0.001)

5

6

0 0

3

3

6

9

12

15

Porewater Si(OH)4-Si, mmol m−2

0 0.0

0.3

0.6

0.9

1.2

Porewater PO43−-P, mmol m−2 Figure 3. Relationship between porewater ammonium (NH4 + -N), phosphate (PO4 3− -P), and silicate [Si(OH)4 -Si] concentrations in the uppermost 10 cm of sediments.

be consistent with previous studies that have focused on the distribution of ammonium and found that it was strongly dependent on seasonal variations in temperature (13,24,25), whereas little combined information is available on the distribution of ammonium, phosphate, and silicate in the sediment column and relevant influencing factors. In our study, for instance, the correlation between temperature and phosphate was significant, yet rather weak. We then tested the hypothesis that the seasonal variation of all three nutrient species could be related to the activity of in situ benthic macrofauna. COUPLING WITH THE EXCRETORY ACTIVITY OF DOMINANT BIVALVES We based our considerations on previous physiological measurements (LABORATORY EXPERIMENTS ON BIVALVE

Figure 4. Relationship between porewater ammonium (NH4 + -N), phosphate (PO4 3− -P), and silicate [Si(OH)4 -Si] concentrations in the uppermost 10 cm sediments and temperature.

EXCRETION RATES OF NUTRIENTS) and scaling up (TEMPORAL SCALING OF BENTHIC NUTRIENT REGENERATION IN BIVALVEDOMINATED TIDAL FLAT) of nutrient excretion rates by two bivalve species dominant on this flat, Ruditapes philippinarum and Musculista senhousia. The plots of bivalve excretion rates of ammonium, phosphate, and silicate versus their relevant pool in the porewater showed a highly significant positive correlation in all cases (Fig. 5). These results suggest the importance of the physiological activity of the benthos on the seasonal variability of porewater chemistry. We conclude that the seasonal patterns of nutrient concentrations in the porewater are strongly coupled with the extent of biogenic regeneration of nutrients due to bivalve excretory activity. This study thus provides evidence of the influence of biological processes on the seasonal patterns of porewater nutrient distribution, suggesting a major role of macrofauna not only at the sediment–water interface, but also in the year-round processes that occur within sediments.

SEASONAL COUPLING

Porew. NH4+-N, mmol m−2

76

Mytilus edulis. In: Trophic Relationships in the Marine Environment. Proc. 24th Europ. Mar. Biol. Symp. Aberdeen, Univ. Press, pp. 89–103.

30 25

7. Yamamuro, M. and Koike, I. (1993). Nitrogen metabolism of the filter-feeding bivalve Corbicula japonica and its significance in primary production of a brackish lake in Japan. Limnol. Oceanogr. 35: 997–1007.

20 15 10 5 0 0

Porew. PO43−-P, mmol m−2

8. Reusch, T.B.H. and Williams, S.L. (1998). Variable responses of native eelgrass Zostera marina to a non-indigenous bivalve Musculista senhousia. Oecologia 113: 428–441.

y = 0.34x + 10.3 r 2 = 0.584 (p < 0.001)

1.2

10 20 30 40 50 NH4+-N excretion, mmol m−2 day−1

60

y = 0.10x + 0.15 r 2 = 0.719 (p < 0.001)

1.0

10. Uthike, S. (2001). Interaction between sediment-feeders and microalgae on coral reefs: grazing losses versus production enhancement. Mar. Ecol. Progr. Ser. 210: 125–138.

0.8 0.6

11. Uthike, S. (1999). Sediment bioturbation and impact of feeding activity of Holothuria (Halodeima) atra and Stichopus chloronotus, two sediment feeding holothurians, at Lizard Island, Great Barrier Reef. Bull. Mar. Sci. 64: 129–141.

0.4 0.2 0.0 0

2

4

6

8

PO43−-P excretion, mmol m−2 day−1 Porew. Si(OH)4-Si, mmol m−2

9. Peterson, B.J. and Heck, K.L., Jr. (2001). Positive interaction between suspension-feeding bivalves and seagrass—a facultative mutualism. Mar. Ecol. Progr. Ser. 213: 143–155.

15

12. Christensen, B., Vedel, A., and Kristensen, E. (2000). Carbon and nitrogen fluxes in sediments inhabited by suspension-feeding (Nereis diversicolor) and non-suspensionfeeding (N. virens) polychaetes. Mar. Ecol. Prog. Ser. 192: 203–217. 13. Land´en, A. and Hall, P.O.J. (1998). Seasonal variation of dissolved and adsorbed amino acids and ammonium in a nearshore marine sediment. Mar. Ecol. Prog. Ser. 170: 67–84.

y = 0.24x + 3.2 r 2 = 0.495 (p < 0.01)

12 9

14. Blackburn, T.H. (1980). Seasonal variation in the rate of organic-N mineralization in oxic marine sediments. In: Biogechimie de la mati`ere organique a` l’interface eau-sediment ´ marine. Edition du CNRS, Paris, pp. 173–183.

6 3 0 0

10

20

30

40

Si(OH)4-Si excretion, mmol m−2 day−1 Figure 5. Relationship between porewater ammonium (NH4 + -N), phosphate (PO4 3− -P), and silicate [Si(OH)4 -Si] concentrations in the uppermost 10 cm of sediments and bivalve-influenced upward flux of those nutrients.

BIBLIOGRAPHY 1. Jordan, T.E. and Valiela, I. (1982). A nitrogen budget of the ribbed mussel, Geukensia demissa, and its significance in nitrogen flow in a New England salt marsh. Limnol. Oceanogr. 27: 75–90. 2. Dame, R.F., Zingmark, R.G., and Haskin, E. (1984). Oyster reefs as processors of estuarine materials. J. Exp. Mar. Biol. Ecol. 83: 239–247. 3. Murphy, R.C. and Kremer, J.N. (1985). Bivalve contribution to benthic metabolism in a California lagoon. Estuaries 8: 330–341. 4. Boucher, G. and Boucher-Rodoni, R. (1988). In situ measurements of respiratory metabolism and nitrogen fluxes at the interface of oyster beds. Mar. Ecol. Prog. Ser. 44: 229–238. 5. Nakamura, M., Yamamuro, M., Ishikawa, M., and Nishimura, H. (1988). Role of the bivalve Corbicula japonica in the nitrogen cycle in a mesohaline lagoon. Mar. Biol. 99: 369–374. 6. Prins, T.C. and Smaal, A.C. (1990). Benthic-pelagic coupling: the release of inorganic nutrients by an intertidal bed of

15. Laima, M.J.C. (1992). Extraction and seasonal variation of NH4 + pools in different types of coastal marine sediments. Mar. Ecol. Prog. Ser. 82: 75–84. 16. Magni, P., Montani, S., Takada, C., and Tsutsumi, H. (2000). Temporal scaling and relevance of bivalve nutrient excretion on a tidal flat of the Seto Inland Sea, Japan. Mar. Ecol. Progr. Ser. 198: 139–155. 17. Lomstein, B.A., Blackburn, T.H., and Henriksen, K. (1989). Aspects of nitrogen and carbon cycling in the northern Bering Shelf sediment: I. The significance of urea turnover in the mineralization of ammonium ion. Mar. Ecol. Prog. Ser. 57: 237–248. 18. Magni, P. (1998). A multidisciplinary study on the dynamics of biophilic elements (C, N, P, Si) in a tidal estuary of the Seto Inland Sea, Japan: physico-chemical variability and macrozoobenthic communities. Ph.D. Thesis, The United Graduate School of Ehime University, Japan, p. 258. 19. Montani, S. et al. (1998). The effect of a tidal cycle on the dynamics of nutrients in a tidal estuary in the Seto Inland Sea, Japan. J. Oceanogr. 54: 65–76. 20. Magni, P., Montani, S., and Tada, K. (2002). Semidiurnal dynamics of salinity, nutrients and suspended particulate matter in an estuary in the Seto Inland Sea, Japan, during a spring tide cycle. J. Oceanogr. 58: 389–402. 21. Magni, P. and Montani, S. (1997). Development of benthic microalgal assemblages on an intertidal flat in the Seto Inland Sea, Japan: effects of environmental variability. La mer. 35: 137–148. 22. Magni, P. and Montani, S. (1998). Responses of intertidal and subtidal communities of the macrobenthos to organic

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load and oxygen depletion in the Seto Inland Sea, Japan. J. Rech. Oc´eanogr. 23: 47–56. 23. Magni, P. and Montani, S. (2000). Physical and chemical variability in the lower intertidal zone of an estuary in the Seto Inland Sea, Japan: seasonal patterns of dissolved and particulate compounds. Hydrobiol. 432: 9–23. 24. Blackburn, T.H. and Henriksen, K. (1983). Nitrogen cycling in different types of sediments from Danish waters. Limnol. Oceanogr. 28: 477–493. 25. Klump, J.V. and Martens, C.S. (1989). Seasonality of nutrient regeneration in an organic-rich coastal sediment: kinetic modeling of changing pore-water nutrient and sulfate distribution. Limnol. Oceanogr. 34: 559–577.

MAPPING THE SEA FLOOR OF THE HISTORIC AREA REMEDIATION SITE (HARS) OFFSHORE OF NEW YORK CITY BRADFORD BUTMAN U.S. Geological Survey

Figure 1. Map showing the area offshore of New York and New Jersey that has been used for the disposal of dredged materials and other wastes since the late 1800’s. The Historic Area Remediation Site (HARS) is outlined in red.

Repeated surveys using a multibeam mapping system document changes in the topography and distribution of sediments on the sea floor caused by placement of dredged material, remedial capping, and natural processes. INTRODUCTION The area offshore of New York City has been used for the disposal of dredged material for over a century. The area has also been used for the disposal of other materials such as acid waste, industrial waste, municipal sewage sludge, cellar dirt, and wood. Between 1976 and 1995, the New York Bight Dredged Material Disposal Site, also known as the Mud Dump Site (MDS), received on average about 6 million cubic yards of dredged material annually. In September 1997 the MDS was closed as a disposal site, and it and the surrounding area were designated as the Historic Area Remediation Site (HARS) (Figs. 1 and 2). The sea floor of the HARS, approximately 9 square nautical miles in area, currently is being remediated by placing a minimum 1-m-thick cap of clean dredged material on top of the surficial sediments that are contaminated from previous disposal of dredged and other materials. The U.S. Geological Survey (USGS) is working cooperatively with the U.S. Army Corps of Engineers (USACE) to map the sea floor geology of the HARS and changes in the characteristics of the surficial sediments over time. HIGH-RESOLUTION SURVEYS OF THE SEA FLOOR OF THE HARS Surveys of the HARS were conducted in November 1996 (prior to the closing of the Mud Dump Site), November 1998 (during early remediation of the HARS),

This article is a US Government work and, as such, is in the public domain in the United States of America.

Figure 2. Shaded relief image of the Historic Area Remediation Site (HARS) in April 2000 showing the Primary Remediation Area (PRA, divided into nine cells), the no discharge zone (ND), the former Mud Dump Site (MDS), and the discontinued Cellar Dirt Disposal Site (CDDS). Companion images are shown in Figures 4 and 5. See text for a description of this image and major features. Bathymetric contour interval is 5 m (red lines).

and April 2000 (during continued remediation of the HARS) using a Simrad EM1000 multibeam mapping system (Fig. 3). Survey lines were run approximately 100 m apart to provide full coverage of the sea floor. The EM1000 measured the depth of water (to an accuracy of

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MAPPING THE SEA FLOOR OF THE HISTORIC AREA REMEDIATION SITE (HARS) OFFSHORE OF NEW YORK CITY

Figure 3. High-resolution multibeam mapping systems use sound from an array of transducers to measure water depth and sediment characteristics of the sea floor. The horizontal resolution of the maps is a few meters, providing an image of the sea floor topography and sediment properties somewhat comparable to an aerial photograph.

about 30 cm) as well as the intensity of sound reflected from the sea floor, which is referred to as backscatter intensity. High backscatter intensity generally indicates the presence of rocks and coarse-grained sediments, while low backscatter intensity indicates the presence of finer grained sediments. Direct observations using bottom photographs, video, and grab samples are needed to verify interpretations of the sea floor geology based on backscatter intensity. IMAGES OF THE HARS SEA FLOOR In this fact sheet, the topography and backscatter intensity data measured by the multibeam mapping system are presented in three types of images. Each of these images highlights different features and characteristics of the sea floor. (1) A shaded relief image (Fig. 2) visually shows small topographic features (with relief of a few meters) that could not be effectively shown by contours alone at this scale. The image was created by vertically exaggerating the topography four times and then artificially illuminating the relief by a light source positioned 45◦ above the horizon from the north. In this image, topographic features are enhanced by strong illumination on the north-facing slopes and by shadows cast on the south-facing slopes. (2) A shaded relief image, colored by backscatter intensity, combines the high-resolution view of topography with a measure of sediment characteristics (Figs. 4, 5B, and 6). In these images, the backscatter intensity is represented by a suite of eight colors ranging from blue, which represents low intensity (fine-grained sediments), to red, which represents

Figure 4. Pseudo-colored backscatter intensity and shaded relief map of the entire HARS in April 2000. The faint north-trending stripes run parallel to the survey tracklines and are artifacts of data collection and environmental conditions. The pink, green, and yellow boxes outline areas shown in Figure 6 to illustrate changes in backscatter intensity between 1996, 1998, and 2000. See text for a description of this image and major features. Bathymetric contour interval is 5 m (red lines).

high intensity (rock outcrops and coarse-grained sediments). These data are draped over the shaded relief image. The resultant image displays light and dark intensities within each color band that result from a feature’s position with respect to the light source. For example, north-facing slopes, receiving strong illumination, show as a light intensity within a color band, whereas south-facing slopes, being in shadow, show as a dark intensity within a color band. (3) A shaded relief image, colored by bathymetry, combines the high-resolution view of topography with color to show water depth (Fig. 5A). THE SEA FLOOR OF THE HARS Within the HARS, one of the most striking aspects of the sea floor is the variability in backscatter intensity and bottom morphology over distances of a few kilometers or less, caused by both natural and anthropogenic processes. This fact sheet presents companion images showing the sea floor of the HARS as mapped in April 2000 in plan view (Figs. 2, 4, and 6) and in perspective view (Fig. 5). Images of selected areas in 1996, 1998, and 2000 illustrate changes over time (Fig. 6). Major features of the sea floor of the HARS shown in these images include two

MAPPING THE SEA FLOOR OF THE HISTORIC AREA REMEDIATION SITE (HARS) OFFSHORE OF NEW YORK CITY

Figure 5. Perspective view of the Historic Area Remediation Site, looking from south to north, based on the multibeam survey carried out in April 2000. A, Shaded relief map with color-coded bathymetry. B, Backscatter intensity draped over shaded relief (see text for a description of the color scheme). The north-trending stripes, running parallel to the survey tracklines, are artifacts of data collection and environmental conditions. The topography, surface features, and the surficial sediments of the HARS have been heavily influenced by the disposal of dredged and other material in this region over the last century, and by recent remedial capping. See text for a description of these images and major features.

relatively smooth topographic highs composed of material dumped in the late 1800’s and early 1900’s (‘‘Topographic highs’’ in Fig. 5A); mounds of material in the Mud Dump Site (‘‘Disposal mounds in MDS’’ in Fig. 5A); two circular features where contaminated sediments were placed and then capped with sand, one in the late 1980’s, and the other in 1997 (‘‘Sand capping’’ in Figs. 4 and 5B and ‘‘Previous capping’’ and ‘‘New sand capping’’ in Fig. 6); material deposited between the November 1996 and November 1998 survey (‘‘Recent placement’’ in Figs. 4 and 5A); many features about 50 m in size interpreted to be individual dumps of material (‘‘Historical dumps’’ in Figs. 4 and 5B); and material placed as part of remediation activities (‘‘Remedial capping’’ in Figs. 4, 5B, and 6). CHANGES IN SURFICIAL PROPERTIES BETWEEN 1996, 1998, AND 2000 Comparison of the topography and backscatter intensity from the three multibeam surveys show how the area changed as a result of dredged material placed before the Mud Dump Site was closed and ongoing remediation of the HARS (see Fig. 2 for locations of placed material from USACE records).

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Figure 6. Pseudo-colored backscatter intensity and shaded relief map of parts of the HARS in 1996, 1998, and 2000 (see Fig. 4 for location). These images illustrate changes in the sea floor topography and backscatter intensity that occurred between 1996 and 1998 and between 1998 and 2000 caused by placement of dredged material and by remedial capping. See text for a description of these images and major features.

Between 1996 and 1998, changes include (1) mounds of medium backscatter intensity dredged material in the northeastern corner of the MDS, some as high as 6 m, placed between November 1996 and September 1997 (compare panels A and B, Fig. 6); (2) a circular area of low-backscatter intensity material about 1 km in diameter and 2 m thick in the southern part of the MDS associated with sand capping (compare panels C and D, Fig. 6); and (3) a circular area of low backscatter intensity material in PRA1 associated with remedial capping (compare panels E and F, Fig. 6). Between 1998 and 2000, changes include (compare panels F and G, Fig. 6) (1) increased backscatter intensity in PRA1 due to additional placement of material and consolidation, de-watering, and possible winnowing of the previous cover; (2) a series of crater-like features in the western part of PRA2, 30 to 70 m long and on the order of 20 m wide with elevated rims and central depressions, that were apparently formed as remedial material impacted the soft sediments on the sea floor; and (3) an area of reduced backscatter intensity in the northeastern corner of PRA2 caused by the placement of remedial material.

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NOAA AND UNIVERSITY SCIENTISTS STUDY METHYL BROMIDE CYCLING IN THE NORTH PACIFIC

Resolution limits of the multibeam system, and the amount of material placed over a relatively large area, preclude using the repeated topographic surveys for determining the amount of material placed on the sea floor. However these multibeam data clearly show the overall regional geology and, through comparison of topography and backscatter intensity, document the location of placed material and changes in sediment properties over time.

NOAA AND UNIVERSITY SCIENTISTS STUDY METHYL BROMIDE CYCLING IN THE NORTH PACIFIC SHARI YVON-LEWIS Atlantic Oceanographic and Meteorological Laboratory

Figure 1. Cruise Track.

KELLY GOODWIN SARA COTTON University of Miami

JAMES BUTLER Climate Monitoring and Diagnostics Laboratory

DANIEL KING University of Colorado

ERIC SALTZMAN RYSZARD TOKARCZYK University of Miami

PATRICIA MATRAI BRIAN YOCIS EILEEN LOISEAU Bigelow Laboratory for Ocean Sciences

GEORGINA STURROCK Commonwealth Scientific and Industrial Research Organization—Australia

As part of a study supported by both NASA and NOAA, scientists from two NOAA laboratories, three universities and CSIRO participated in a research cruise aboard the R/V Ronald H. Brown. The cruise departed Kwajalein, Republic of the Marshall Islands on 14 September 1999 and arrived in Seattle, Washington on 23 October 1999 with stops in Honolulu, Hawaii, Dutch Harbor, Alaska, and Kodiak, Alaska. The objective of this research effort was to obtain reliable measurements of the uptake and emission of methyl bromide and other climatically important halocarbons in tropical to temperate regions of the North Pacific Ocean. Atmospheric methyl bromide (CH3 Br), which is of both natural and anthropogenic origin, has been identified as This article is a US Government work and, as such, is in the public domain in the United States of America.

Figure 2. Scientists collecting water samples for production and degradation incubations.

a Class I ozone-depleting substance in the amended and adjusted Montreal Protocol on Substances that Deplete Stratospheric Ozone. The role of the ocean in regulating the atmospheric burden of this gas is still somewhat uncertain. Methyl bromide is both produced and destroyed in the ocean through chemical and biological processes. The organisms or reactions that produce CH3 Br at rates sufficient to explain its observed concentrations are not known. Degradation has been shown to occur at rates that are faster than can be explained by known chemical degradation reactions, and evidence suggests that this additional degradation is bacterial consumption of CH3 Br. While recent measurements have shown that, on the whole, the ocean is a net sink for CH3 Br, measurement coverage to date has been limited and sporadic, which restricts our ability to map the spatial and temporal variations that are necessary for understanding how the system will respond to perturbations (e.g., Global Warming). The measurements made during this cruise are designed to help improve our understanding of the role that the oceans play in the cycling of CH3 Br. The program involved instrumentation from two NOAA

Next Page TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS

laboratories and two universities. Measurements were made of the concentrations of CH3 Br and a suite of natural and anthropogenic halocarbons in the air and surface water, degradation rates of CH3 Br in the surface water, production rates of CH3 Br and other natural halocarbons in the surface water, and depth profiles of CH3 Br and other halocarbons. The combined results from these measurements will be used to constrain the budget of CH3 Br in these waters at this time of year. The relative importance of the biological and chemical processes will be examined for tropical and high latitudes. Attempts will also be made to extract relationships between the rates and concentrations measured and satellite measurements in order to develop proxies that can provide global coverage on shorter time scales. At this time, there is insufficient data to examine seasonal and long-term trends in net flux, production, or degradation. Until satellite measurable proxies can be found, additional research cruises are needed to reduce the uncertainty in the global net flux estimate and to map the spatial and temporal variations in the net fluxes, production rates, and degradation rates of CH3 Br and other climatically important halocarbons.

TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS PAOLO MAGNI IMC—International Marine Centre Torregrande-Oristano, Italy

SHIGERU MONTANI Hokkaido University Hakodate, Japan

Freshwater runoff during ebb flow and salt water intrusion during the flood may have a major effect on short-term changes in nutrient (ammonium, nitrate+nitrite, phosphate, and silicate) concentrations along an estuary. Time series hourly measurements conducted in a mixedsemidiurnal type estuary (i.e., characterized by two major lower and higher tidal levels) show that these changes are a strong function of both tidal state (e.g., low vs. high tide) and amplitude (e.g., neap vs. spring tide). In particular, the changes in nutrient concentrations are higher during ebb than during flood tide and largest between the lower low tide and the higher high tide of a spring tide. Finally, the importance of investigating simultaneously different stations along the estuarine spine is highlighted, in addition to studying the nutrient distribution based on selected salinity intervals which may reflect only the conditions at a particular tidal state. BACKGROUND An important aspect of the high variability of tidal estuaries is related to the effect of the tidal cycle on the physical and chemical characteristics of the water. In particular, on a timescale of hours, freshwater runoff

81

during ebb flow and salt water intrusion during the flood may determine strong changes in salinity and dissolved and particulate compounds. Several studies in riverine and estuarine waters have investigated the distribution of nutrient (e.g., ammonium, nitrate, phosphate, and silicate) concentrations, based on the salinity gradient in the system at a particular tidal state (1–3). Accordingly, plots of nutrients versus salinity are often used to assess the source of different nutrient species, whether from inland, outside the estuary, or within it (3–7). An evaluation of the distributional pattern of nutrients along an estuary has important ecological implications in relation to the cycling of biophilic elements, such as N (nitrogen), P (phosphorous), and Si (silicon). It is known that numerous processes influence the behavior of nutrients, whether they show conservative mixing or reflect removal or addition along an estuary. Ammonium consumption and ammonium oxidation, for instance, are predominant in the water column, whereas denitrification in sediments is responsible for nitrate removal from the water column (8–10). By contrast, bioturbation and excretion by abundant benthic animals may greatly contribute to the upward flux of regenerated nutrients, such as ammonium and phosphate, which in turn enhance primary production (11,12). Accordingly, it has been shown that regeneration processes within an estuary are consistent with often encountered nonconservative mixing of ammonium (4,8,13,14). This corresponds to the tendency of ammonium concentration to be high at midsalinity ranges, resulting in a poor correlation with salinity. In contrast, nitrate tends to show conservative behavior, as evidence of its riverine origin (2,14,15), although addition (15) or removal (7) is also found. Moreover, it must be considered that in some cases, nitrate versus salinity plots may fail to unravel active nitrate turnover, leading to an approximate balance of sources and sinks (16). As for silicate, a general pattern indicates that estuarine mixing of this nutrient species tends to be conservative (2,7,13). Yet, either silicate removal (7,17) or addition (4) occurs in relation to the development of an algal population in rivers or to a closer interaction with estuarine sediments, respectively, and varies with season (14). A major upward flux of silicate from sediments might also be related to the biological activity and excretory processes of abundant macrofaunal assemblages (18,19). In addition to these general considerations, the distribution and cycling of nutrients depend strongly on the specific characteristics of each estuary, including water residence time and water depth, nutrient levels, and the extent of salt-water intrusion. Uncles and Stephens (3) showed that saline intrusion was a strong function of the tidal state and a weaker function of freshwater inflow. Accordingly, Balls (2) indicated that conservative mixing of phosphate, is a function of estuarine flushing time, as related to particle–water interaction and chemical speciation (20). In particular, phosphate removal at low salinities may be due to adsorption to iron and aluminium colloidal oxyhydroxides that aggregate and undergo sedimentation (7).

Previous Page TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS

laboratories and two universities. Measurements were made of the concentrations of CH3 Br and a suite of natural and anthropogenic halocarbons in the air and surface water, degradation rates of CH3 Br in the surface water, production rates of CH3 Br and other natural halocarbons in the surface water, and depth profiles of CH3 Br and other halocarbons. The combined results from these measurements will be used to constrain the budget of CH3 Br in these waters at this time of year. The relative importance of the biological and chemical processes will be examined for tropical and high latitudes. Attempts will also be made to extract relationships between the rates and concentrations measured and satellite measurements in order to develop proxies that can provide global coverage on shorter time scales. At this time, there is insufficient data to examine seasonal and long-term trends in net flux, production, or degradation. Until satellite measurable proxies can be found, additional research cruises are needed to reduce the uncertainty in the global net flux estimate and to map the spatial and temporal variations in the net fluxes, production rates, and degradation rates of CH3 Br and other climatically important halocarbons.

TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS PAOLO MAGNI IMC—International Marine Centre Torregrande-Oristano, Italy

SHIGERU MONTANI Hokkaido University Hakodate, Japan

Freshwater runoff during ebb flow and salt water intrusion during the flood may have a major effect on short-term changes in nutrient (ammonium, nitrate+nitrite, phosphate, and silicate) concentrations along an estuary. Time series hourly measurements conducted in a mixedsemidiurnal type estuary (i.e., characterized by two major lower and higher tidal levels) show that these changes are a strong function of both tidal state (e.g., low vs. high tide) and amplitude (e.g., neap vs. spring tide). In particular, the changes in nutrient concentrations are higher during ebb than during flood tide and largest between the lower low tide and the higher high tide of a spring tide. Finally, the importance of investigating simultaneously different stations along the estuarine spine is highlighted, in addition to studying the nutrient distribution based on selected salinity intervals which may reflect only the conditions at a particular tidal state. BACKGROUND An important aspect of the high variability of tidal estuaries is related to the effect of the tidal cycle on the physical and chemical characteristics of the water. In particular, on a timescale of hours, freshwater runoff

81

during ebb flow and salt water intrusion during the flood may determine strong changes in salinity and dissolved and particulate compounds. Several studies in riverine and estuarine waters have investigated the distribution of nutrient (e.g., ammonium, nitrate, phosphate, and silicate) concentrations, based on the salinity gradient in the system at a particular tidal state (1–3). Accordingly, plots of nutrients versus salinity are often used to assess the source of different nutrient species, whether from inland, outside the estuary, or within it (3–7). An evaluation of the distributional pattern of nutrients along an estuary has important ecological implications in relation to the cycling of biophilic elements, such as N (nitrogen), P (phosphorous), and Si (silicon). It is known that numerous processes influence the behavior of nutrients, whether they show conservative mixing or reflect removal or addition along an estuary. Ammonium consumption and ammonium oxidation, for instance, are predominant in the water column, whereas denitrification in sediments is responsible for nitrate removal from the water column (8–10). By contrast, bioturbation and excretion by abundant benthic animals may greatly contribute to the upward flux of regenerated nutrients, such as ammonium and phosphate, which in turn enhance primary production (11,12). Accordingly, it has been shown that regeneration processes within an estuary are consistent with often encountered nonconservative mixing of ammonium (4,8,13,14). This corresponds to the tendency of ammonium concentration to be high at midsalinity ranges, resulting in a poor correlation with salinity. In contrast, nitrate tends to show conservative behavior, as evidence of its riverine origin (2,14,15), although addition (15) or removal (7) is also found. Moreover, it must be considered that in some cases, nitrate versus salinity plots may fail to unravel active nitrate turnover, leading to an approximate balance of sources and sinks (16). As for silicate, a general pattern indicates that estuarine mixing of this nutrient species tends to be conservative (2,7,13). Yet, either silicate removal (7,17) or addition (4) occurs in relation to the development of an algal population in rivers or to a closer interaction with estuarine sediments, respectively, and varies with season (14). A major upward flux of silicate from sediments might also be related to the biological activity and excretory processes of abundant macrofaunal assemblages (18,19). In addition to these general considerations, the distribution and cycling of nutrients depend strongly on the specific characteristics of each estuary, including water residence time and water depth, nutrient levels, and the extent of salt-water intrusion. Uncles and Stephens (3) showed that saline intrusion was a strong function of the tidal state and a weaker function of freshwater inflow. Accordingly, Balls (2) indicated that conservative mixing of phosphate, is a function of estuarine flushing time, as related to particle–water interaction and chemical speciation (20). In particular, phosphate removal at low salinities may be due to adsorption to iron and aluminium colloidal oxyhydroxides that aggregate and undergo sedimentation (7).

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TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS

Therefore, it is important that investigations of the distribution of nutrients as a function of salinity are conducted on proper spatial and temporal scales that take into account the extent and variability of salt-water intrusion. Time series surveys represent a valuable approach to quantifying such variability in estuarine waters (21,22). In particular, short-term monitoring surveys during a complete tidal cycle and simultaneous information at different stations can be a powerful tool for evaluating the extent of salt-water intrusion along the estuary and its direct impact on the water chemistry of an estuary. This approach may overcome the limitations of evaluating the distribution and behavior of nutrients based on a particular tidal state at arbitrarily selected salinity intervals. Major drawbacks when working across long distances to cover different salinity ranges include work time to travel between samplings, tidal-water displacements, variations in tidal velocities, and a possible wide range of river discharge. Simultaneous observations at different stations are expected to track changes in nutrient concentrations that occur along the estuarine spine in relation to the extent of salt-water intrusion over time. Case study This article reviews the results of 24-hour surveys in a tidal estuary of the Seto Inland Sea, Japan, during a spring tide of May 1995 (13,23). It is aimed to give an example of the effect of freshwater runoff and salt-water intrusion on the spatial and temporal variability of nutrient concentrations during a tidal cycle. The effect of a tide is also evaluated in relation to the mixed-semidiurnal behavior (i.e., with pronounced differences between two successive low and high tides) of the estuary. The fieldwork was conducted along a transect line of approximately 1.4 km linking the river to the rear to the subtidal zone. Multiprobe casts were used for hydrologic measurements. Nutrient concentrations were determined every hour in surface waters, simultaneously at a riverine, an intertidal, and a subtidal station, and every two hours at additional depths through the water column at the subtidal station. Details of the sampling scheme and analysis are given in Montani et al. (13) and Magni et al. (23). At the beginning of the survey, there was a marked salinity gradient along the estuary; surface water salinity was 2.0, 18.5, and 31.3 psu at the riverine, intertidal, and subtidal stations, respectively (Fig. 1). As the lower low tide approached, salinity remained low at the riverine station (Fig. 1a), sharply decreased also at the intertidal station (Fig. 1b) and, subsequently, at the subtidal station (Fig. 1c). By contrast, soon after the lower low tide, saltwater intrusion rapidly caused an abrupt increase in salinity at both the intertidal and the riverine stations, up to >30 psu (Figs. 1a,b). During the flood, there was a homogeneous distribution of high-salinity water along the transect line. At the subtidal station, a major change in salinity as a function of the tidal cycle was also apparent, but restricted to the surface layer, whereas salinity remained constantly >31 psu below the surface (Figs. 1c,d).

The nutrient concentrations were also markedly affected by the tidal cycle. At the beginning of the survey, silicate and nitrate+nitrite concentrations were markedly higher at the riverine station than at the intertidal and subtidal stations, whereas the ammonium concentration was relatively higher at the intertidal station (Figs. 1a, b, c). Approaching the lower low tide, the nutrient concentrations in surface water increased rapidly, especially at the intertidal and subtidal stations. Differently at the riverine station, the ammonium concentrations remained low, suggesting no significant import of this nutrient species through freshwater inflow (Fig. 1a). At the subtidal station, the nutrient concentrations also showed a relatively consistent increase below the surface, yet progressively less noticeable with depth (Fig. 1d). By contrast, during the flood, as high-salinity water flushed backward into the estuary, the nutrient concentrations dropped to the lowest values at all stations and depths (Figs. 1a, b, c, d); a 7.5-fold and 8.8-fold decrease of silicate and nitrate+nitrite concentrations, respectively, occurred at the riverine station. During the second part of the survey, after the higher high tide, both salinity and nutrient changes were less marked. Figure 2 summarizes the relationships between salinity and nutrient concentrations. Salinity versus nutrient plots demonstrate that the distributional pattern of nutrients largely varied with station, depth, and the different nutrient species. In particular at the riverine station, silicate and nitrate+nitrite were negatively correlated with salinity; r2 explained a large portion of total variance (i.e., r2 = 0.879 and 0.796, respectively). By contrast, at this station, ammonium showed a positive correlation with salinity, whereas phosphate did not significantly correlate with salinity. Differently at the intertidal and subtidal stations, the concentrations of all nutrient species in surface waters were negatively and significantly correlated with salinity; levels of confidence varied from p < .05 (ammonium) to p < .001 (phosphate and silicate). The variability of both salinity and nutrient concentrations was lowest at the subtidal station below the surface. This test data set (Fig. 2d) indicated that all nutrient species were correlated positively with salinity at a high level of significance (p < .001). Relevant plots also highlighted that, within such a restricted variability of salinity, silicate and nitrate+nitrite concentrations comprised narrower values than those of phosphate and ammonium (Fig. 2d). Accordingly, the model equation for the former two nutrient species explained a higher portion of the total variance (r2 = 0.423, and r2 = 0.457, respectively) than that explained by the latter (r2 = 0.221, and r2 = 0.245, respectively). These results showed that the riverine input was a major source of silicate and, partially, nitrate+nitrite and phosphate. It was also apparent that the increase in nutrient concentrations at the intertidal station and subsequently at the subtidal station, was largest during the first part of the survey (Figs. 1b, c). Companion papers demonstrated that the intertidal zone also plays a major role in nutrient cycling, as a major site of nutrient regeneration within the estuary (12,14,24).

TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS

Subtidal station (surf. & water column)

(d) (a)

Riverine station (surface)

(b)

Intertidal station (surface)

(c)

Subtidal station (surface)

0

Depth, m

Salinity

25 20

Tidal range (2 m)

15

31

L 17.5

31 2

35 30

4 H

H

6 8

Salinity

5

0

0

Depth, m

Si(OH)4

H 76.7 50 40 10 L 30

2

Higher high tide

100 80 60 40

4 20 8

Lower low tide

10 L

L

Si(OH )4 50.3 H 25 25 20 15 20 15 10 5 10

0 5 2 Depth, m

60 40 20

15 10

4

47.4 20 25 5

5 6

50 40 40 30 20 30 20

10 L

6

20

NO3− + NO2−

27

31 29

31-23

10

0

83

L

L 5 L

8

5

NO3− + NO2−

0 0

H 4.9

2

4 Depth, m

PO43−

5

3 2 1

1

4

2

L

6 8

•

2 Depth, m

NH4+

0

30 20 10 0

4

May 30

May 31

May 30

May 31

May 30

May 31

Time, h

2 1

2 L

3 2 L

42.6 H 35.7 25 >30 >30 20 25 25 20 20 5 15 15 5 10 L 10 L 10 15

6 8

10 12 14 16 18 20 22 24 2 4 6 8 10 12 14 16 18 20 22 24 2 4 6 8 10 12 14 16 18 20 22 24 2 4 6 8 10

3

PO43−

0 40

2

3

NH4 +

10

10 10 12 14 16 18 20 22 24 2 4 6 8 10 May 30 May 31

Time, h

Figure 1. Time series of salinity (psu, practical salinity unit) and nutrient concentration (µM) in surface water at a riverine, intertidal, and subtidal station (Fig. 1a, b, c) and through the water column at the subtidal station (Fig. 1d) during a tidal cycle of a spring tide in a mixed-semidiurnal type estuary (Seto Inland Sea, Japan). Data sources: Fig. 1a, b, c from Montani et al. (13) (redrawn from Figs. 4 & 5); Fig. 1d: from Magni et al. (23) (adapted from Figs. 2 & 3).

CONCLUDING REMARKS This article showed that the effect of salt-water intrusion on the dynamics of nutrients varies strongly both spatially (on relatively short distances) and temporally (on an hour timescale), and that this is much dependent on the tidal state. In particular, it was shown that the effect of tidal amplitude is important in determining the extent of the variations in nutrient concentrations, which were stronger between the lower low tide and the higher high tide. It also indicated that nutrient concentrations were higher during the ebb than during the flood and highest at the surface layer, as strongly correlated inversely with salinity. Finally, this work highlighted the importance of considering simultaneous investigations at different stations along the estuarine spine during a tidal cycle, especially on short distances, besides studying the nutrient

dynamics based on selected salinity intervals that may reflect only the situation at a particular tidal state. BIBLIOGRAPHY ´ 1. Hernandez-Ay´ on, J.M., Galindo-Bect, M.S., Flores-Baez, B.P., and Alvarez-Borrego, S. (1993). Nutrient concentrations are high in the turbid waters of the Colorado River delta. Estuar. Coast. Shelf Sci. 37: 593–602. 2. Balls, P.W. (1994). Nutrient inputs to estuaries from nine Scottish east coast rivers; influence of estuarine processes on inputs to the North Sea. Estuar. Coast. Shelf Sci. 39: 329–352. 3. Uncles, R.J. and Stephens, J.A. (1996). Salt intrusion in the Tweed Estuary. Estuar. Coast. Shelf Sci. 43: 271–293. 4. Balls, P.W. (1992). Nutrient behaviour in two contrasting Scottish estuaries, the Forth and the Tay. Oceanol. Acta 15: 261–277.

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TIDALLY MEDIATED CHANGES IN NUTRIENT CONCENTRATIONS

Riverine station (surface )

PO43−

NO3− + NO2−

Si(OH)4

(a) 120 100 80 60 40 20 0 60

(c) Subtidal station (surface )

y = −2.9x + 110.7; r 2 = 0.879 (n = 25, p < .001)

y = −2.6x + 100.7; r 2 = 0.736 (n = 25, p < .001)

y = −1.25x + 48.8; r 2 = 0.796 (n = 25, p < .001)

y = −2.07x + 74.6; r 2 = 0.744 (n = 25, p < .001)

Subtidal station (d) (water column - no surface)

y = −3.6x + 133.7; r 2 = 0.624 (n = 25, p < .001)

y = −22.1x + 719; r 2 = 0.423 (n = 50, p < .001)

y = −2.10x + 80.7; r 2 = 0.381 (n = 25, p < .01)

y = −17.7x + 571; r 2 = 0.457 (n = 50, p < .001)

45 30 15 0 5 4 3 2 1 0

y = −1.79x + 58; r 2 = 0.221 (n = 50, p < .001)

y = −0.01x + 3.9; r 2 = 0.011 (n = 25, ns)

y = −0.19x + 8.0; r 2 = 0.428 (n = 25, p < .01)

y = −0.09x + 4.8; r 2 = 0.404 (n = 25, p < .001)

40 NH4 +

Intertidal station (surface )

(b)

y = 0.68x + 3.9; r 2 = 0.546 (n = 25, p < .001)

30

y = −14.5x + 472; r 2 = 0.245 (n = 50, p < .001)

20 y = −1.29x + 58.2; r 2 = 0.254 (n = 25, p < .05)

y = −0.63x + 35.2; r 2 = 0.337 (n = 25, p < .01)

10 0 0

5

10 15 20 25 30

0

5

10 15 20 25 30

0

5

10 15 20 25 30

0

5 10 15 20 25 30 35

Salinity, psu Figure 2. Plots of salinity (psu) versus nutrient concentration (µM) originated from the time series in Fig. 1. Data sources: Fig. 1a, b, c from Montani et al. (13) (redrawn from Figs. 8 to 11); Fig. 1d: after Magni et al. (23).

5. Clark, J.F., Simpson, H.J., Bopp, R.F., and Deck, B. (1992). Geochemistry and loading history of phosphate and silicate in the Hudson estuary. Estuar. Coast. Shelf Sci. 34: 213–233. 6. Page, H.M., Petty, R.L., and Meade, D.E. (1995). Influence of watershed runoff on nutrient dynamics in a southern California salt marsh. Estuar. Coast. Shelf Sci. 41: 163–180. 7. Eyre, B. and Twigg, C. (1997). Nutrient behaviour during post-flood recovery of the Richmond river estuary northern NSW, Australia. Estuar. Coast. Shelf Sci. 44: 311–326. 8. Soetart, K. and Herman, P.M.J. (1995). Nitrogen dynamics in the westerschelde (the Netherlands) using a box model with fixed dispersion coefficients. Hydrobiologia 311: 215–224. 9. Nixon, S.W. et al. (1996). The fate of nitrogen and phosphorus at the land–sea margin of the North Atlantic Ocean. Biogeochem. 35: 141–180. 10. Nedwell, D.B., Jickells, T.D., Trimmer, M., and Sanders, R. (1999). Nutrients in estuaries. Adv. Ecol. Res. 29: 43–92. 11. Herman, P.M.J., Middelburg, J.J., van de Koppel, J., and Heip, C.H.R. (1999). Ecology of estuarine macrobenthos. Adv. Ecol. Res. 29: 195–240. 12. Magni, P., Montani, S., Takada, C., and Tsutsumi, H. (2000). Temporal scaling and relevance of bivalve nutrient excretion on a tidal flat of the Seto Inland Sea, Japan. Mar. Ecol. Prog. Ser. 198: 139–155. 13. Montani, S. et al. (1998). The effect of a tidal cycle on the dynamics of nutrients in a tidal estuary in the Seto Inland Sea, Japan. J. Oceanogr. 54: 65–76. 14. Magni, P. and Montani, S. (2000). Water chemistry variability in the lower intertidal zone of an estuary in the Seto Inland

Sea, Japan: Seasonal patterns of nutrients and particulate compounds. Hydrobiologia 432: 9–23. 15. Middelburg, J.J. and Nieuwenhuize, J. (2000). Uptake of dissolved inorganic nitrogen in turbid, tidal estuaries. Mar. Ecol. Prog. Ser. 192: 79–88. 16. Middelburg, J.J. and Nieuwenhuize, J. (2001). Nitrogen isotope tracing of dissolved inorganic nitrogen behaviour in tidal estuaries. Estuar. Coast. Shelf Sci. 53: 385–391. 17. Liss, P.S. and Pointon, M.J. (1973). Removal of dissolved boron and silicon during estuarine mixing of sea and river waters. Geochim. Cosmochim. Acta 37: 1493–1498. 18. Bartoli, M. et al. (2001). Impact of Tapes philippinarum on nutrient dynamics and benthic respiration in the Sacca di Goro. Hydrobiologia 455: 203–212. 19. Magni, P. and Montani, S. (2004). Magnitude and temporal scaling of biogenic nutrient regeneration in a bivalvedominated tidal flat. In: Encyclopedia of Water. John Wiley & Sons, New York. 20. Froelich, P.N. (1988). Kinetic control of dissolved phosphate in natural rivers and estuaries: A primer on the phosphate buffer mechanism. Limnol. Oceanogr. 33: 649–668. 21. Yin, K., Harrison, P.J., Pond, S., and Beamish, R.J. (1995). Entrainment of nitrate in the Fraser river estuary and its biological implications. II. Effects of spring vs. neap tide and river discharge. Estuar. Coast. Shelf Sci. 40: 529–544. 22. Yin, K., Harrison, P.J., Pond, S., and Beamish, R.J. (1995). Entrainment of nitrate in the Fraser River estuary and its biological implications. III. Effects of winds. Estuar. Coast. Shelf Sci. 40: 545–558.

THE ROLE OF OCEANS IN THE GLOBAL CYCLES OF CLIMATICALLY-ACTIVE TRACE-GASES 23. Magni, P., Montani, S., and Tada, K. (2002). Semidiurnal dynamics of salinity, nutrients and suspended particulate matter in an estuary in the Seto Inland Sea, Japan, during a spring tide cycle. J. Oceanogr. 58: 389–402. 24. Montani, S., Magni, P., and Abe, N. (2003). Seasonal and interannual patterns of intertidal microphytobenthos in combination with laboratory and areal production estimates. Mar. Ecol. Prog. Ser. 249: 79–91.

THE ROLE OF OCEANS IN THE GLOBAL CYCLES OF CLIMATICALLY-ACTIVE TRACE-GASES ROBERT C. UPSTILL-GODDARD University of Newcastle upon Tyne Newcastle upon Tyne, United Kingdom

The oceans are important in the global cycles of a range of trace gases that influence atmospheric chemistry and climate. For some of these, the oceans are a net source to the atmosphere, whereas for others, they are a net sink. Major gases of interest are summarized in Table 1, along with their net flux directions and their principal roles in the troposphere and stratosphere. Carbon dioxide (CO2 ), methane (CH4 ), and nitrous oxide (N2 O) are major greenhouse gases: CO2 currently accounts for more than one-half of enhanced global warming, whereas CH4 and N2 O, respectively, account for about 15% and 6% (1). Volatile sulphurs are also implicated in climate forcing and they play important roles in atmospheric chemistry. As a result of length restrictions, our discussion focuses only on these gases. GLOBAL PARTITIONING OF ANTHROPOGENIC CO2 Several mechanisms contribute to natural CO2 cycling between the atmosphere and the Earth’s surface. The largest natural exchanges occur through respiration and photosynthesis on land and in the oceans, and by solubility-driven uptake in the oceans, and the net result of these exchanges is a natural carbon cycle in overall balance. By comparison with these natural fluxes, the flux

85

of anthropogenic CO2 , which is derived primarily from the burning of fossil fuels, cement production, deforestation, and other land-use changes, is rather small, ∼7 ± 1 Gt C per year (1) (1 Gt = 1 gigatonne = 1015 g), which is nevertheless large enough to significantly disturb the natural CO2 cycle. As a result of its unreactive nature, the residence time of CO2 in the troposphere is of the order of 50–200 years, hence anthropogenic CO2 tends to accumulate there. In fact, tropospheric CO2 has risen from about 280 ppmv (parts per million by volume) preindustrially, as determined from ice core studies (2), to approaching 380 ppmv by mid-2004 (http://www.cmdl.noaa.gov/). However, this corresponds to somewhat less than one-half of the known anthropogenic release (1). During the 1980s (the latest decade for which estimates of all carbon sources and sinks are available), the mean rate of tropospheric CO2 growth was only about 3.3 ± 0.1 Gt C per year, the remainder, about 3.7 ± 1.0 Gt C per year, having been absorbed by ‘‘sinks’’ located in the oceans and within the terrestrial biosphere (1). The fraction of anthropogenic CO2 absorbed by these sinks is, however, not constant. Large fluctuations are evident in the continuous tropospheric records that date back to the late 1950s (1), and these fluctuations are believed to relate directly to short-term variations in global climate. For example, increased atmospheric growth rates correlate ˜ climate warming events, whereas cooling with El Nino periods, such as that which followed the eruption of Mount Pinatubo in the early 1990s, seem to be associated with reduced atmospheric growth. These variations are thought principally to reflect changes in the balance of terrestrial primary production (photosynthesis) versus respiration, decomposition, and the combustion of organic material (3). High background variability has precluded directly measuring the relative magnitudes of the oceanic and terrestrial CO2 sinks, hence they have hitherto been estimated using models, often with conflicting results (4). However, recent techniques based on measuring carbon and oxygen isotopes in air (5) now enable the partitioning of anthropogenic CO2 between these reservoirs to be determined with greater certainty, and it is now generally agreed that the ocean and land sinks are of about the same magnitude (although the uncertainties are large), i.e., about 1.9 ± 0.6 Gt C per year (6).

Table 1. Some Atmospherically Active Trace Gases with Important Marine Sources or Sinks Gas

Net Flux

Effect in the Atmosphere Infra-red activity Infra-red activity Atmospheric redox Ozone regulation Infra-red activity Ozone regulation Atmospheric redox Acidity, Cloud formation Sulphate aerosol (cooling) Source of COS Atmospheric redox Atmospheric redox

Carbon Dioxide Methane

(CO2 ) (CH4 )

Into Ocean Out of Ocean

Nitrous Oxide

(N2 O)

Out of Ocean

Carbon Monoxide Dimethyl sulphide Carbonyl Sulphide Carbon Disulphide Organohalogens (Natural) Nonmethane hydrocarbons

(CO) (DMS) (COS) (CS2 )

Out of Ocean Out of Ocean Out of Ocean Out of Ocean Out of Ocean Out of Ocean

Troposphere √ √

Stratosphere

√ √ √ √ √ √

√ √ √

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MECHANISMS OF OCEAN CO2 UPTAKE Compared with many other atmospheric trace gases, CO2 is rather soluble in seawater, where it occurs as three principal dissolved species that together comprise seawater dissolved inorganic carbon (DIC), which are bicarbonate (HCO3 − , 91% of DIC), carbonate (CO3 2− , 8% of DIC), and dissolved CO2 (1% of DIC). The capacity of the oceans to absorb atmospheric CO2 is ultimately buffered by the DIC system. CO2 uptake results in an increase both in the partial pressure of CO2 (pCO2 ) and in the concentration of HCO3 − , which is produced through the reaction of CO2 with CO3 2− . With further CO2 uptake, its conversion to HCO3 − becomes limited by the decreasing CO3 2− availability, the result being that more CO2 remains in solution, further decreasing the capacity for CO2 uptake at increasing tropospheric levels. As an illustration, for atmospheric CO2 to rise by about 100 ppmv relative to today, the increase in seawater DIC would only be about 60% of that which has accompanied the approximately 100 ppmv rise in tropospheric CO2 since the industrial revolution (1,6). Studies based on the downward penetration of chemical tracers, such as 14 C, coupled with simple box models of ocean mixing (7), or more sophisticated General Circulation Models (8), show that tropospheric CO2 equilibrates with the surface ocean mixed layer rather rapidly, on the order of a few years. In contrast, CO2 penetration into the deep ocean interior is constrained by relatively slow vertical transport; the whole ocean equilibrates with the atmosphere on a timescale of more than 1000 years. As a result of this slow rate of downward mixing, with few exceptions, anthropogenic CO2 has yet to penetrate below about 1000 m depth. Transport of CO2 into the ocean interior occurs via the thermohaline circulation, in which cool surface waters in the high latitude North Atlantic (Greenland Sea) and Southern Oceans sequester CO2 from the troposphere and sink because of their higher density. This downward transport is balanced in regions where deep water with a high capacity for CO2 uptake wells up to the surface, such as occurs in parts of the tropics and the Southern Ocean. The sinking waters eventually spread laterally toward the subtropics, giving rise to a relatively uniform distribution of anthropogenic CO2 . Eventually, on the 1000-plus year timescale of deep ocean mixing, the deep CO2 -rich waters formed in this way will again well up to the surface, the principal site for this being the Equatorial Pacific. As this water warms during upwelling, its pCO2 will increase, resulting in CO2 loss to the atmosphere by out gassing. One other potential removal process for tropospheric CO2 is through reaction with CaCO3 contained in deep sea sediments; however, their response time to changes in tropospheric CO2 is several thousand years (9). According to some coupled ocean-atmosphere models, one possible consequence of global warming is an increase in the intensity of vertical stratification in the oceans (1), which would reduce the rate of surface to deep water mixing and, consequently, the uptake rate of tropospheric CO2 . In addition to chemically and physically driven uptake, tropospheric CO2 is also processed through the so-called

‘‘biological pump,’’ in which organic matter produced via photosynthesis and cycled through the upper ocean food-web is ultimately transported downward via sinking organic particles or through vertical biomass migrations. This sinking flux of organic carbon is remineralized or respired back to inorganic carbon at depth with an accompanying release of dissolved inorganic nutrients (principally nitrate and phosphate). Model simulations suggest that without the biological pump, tropospheric CO2 could be about 150 ppmv higher than at present (1). However, the likely response of the biological pump to increasing tropospheric CO2 is uncertain. Most current evidence tends to discount increased productivity on the grounds of limitation by the supply of nutrients, which are seasonally depleted in most surface waters. However, extensive regions exist of the subarctic Pacific, equatorial Pacific, and Southern Ocean with abundant nitrate and phosphate throughout the year but with very low phytoplankton productivity, the so-called high-nutrientlow-chlorophyll (HNLC) regions. Their low productivity likely reflects a deficiency in the supply of a minor nutrient such as iron (10). Although predicting climate-induced changes in the supply rate of iron to the oceans is far from straightforward, these regions could conceivably play a significant role in the future ocean uptake of anthropogenic CO2 . It is also conceivable that global warming-induced changes in stratification described earlier could also modify the ocean’s biological carbon cycle; however, the consequences of this are difficult to predict given the intrinsic complexity of the system (11). THE GLOBAL ROLE OF CH4 AND N2 O As well as being important greenhouse gases, CH4 and N2 O play important roles in atmospheric chemistry; N2 O is involved in the stratospheric cycling of NOx (reactive nitrogen oxides) and ozone (12), and CH4 takes part in reactions that govern levels of tropospheric ozone and hydroxyl radical (· OH) and stratospheric H2 O (13). Like CO2 , N2 O and CH4 are currently increasing in the troposphere, both by about 0.3% per year (1). However, the role of the oceans in their global budgets differs from that for anthropogenic CO2 in two fundamental ways. First, the oceans are a net source of tropospheric N2 O and CH4 , and second, for both, the marine source is one of several global sources whose relative magnitudes remain rather uncertain. GLOBAL SOURCE UNCERTAINTIES The marine system is one of two major N2 O sources, the other being terrestrial soils. Combustion, biomass burning, and fertilizers make additional minor contributions. Although the uncertainties are large, the oceans are thought to contribute around 20% of the natural global source (1); however, the contribution is larger with anthropogenic sources included (see below). For CH4 , many more global sources exist, in descending order of magnitude, these are: agriculture, wetlands, fossil fuels, biomass burning, termites, oceans, CH4 hydrates, and

THE ROLE OF OCEANS IN THE GLOBAL CYCLES OF CLIMATICALLY-ACTIVE TRACE-GASES

landfills. However, the range of uncertainties is no better than it is for N2 O, ranging from ±100% (e.g., wetlands), to in excess of ±2000% (e.g., CH4 hydrates) (1). According to these data, marine waters contribute only about 3% of the total CH4 source to the troposphere, but this may be an underestimate (see below).

SOURCES OF SEAWATER N2 O In seawater, N2 O develops as a byproduct during microbial nitrification (the conversion of dissolved ammonium, NH4 + , to nitrate, NO3 − ) and as a reactive intermediate during microbial denitrification (the reduction of NO3 − to gaseous nitrogen). Nitrification principally occurs in oxygenated waters, although it is inhibited by light, whereas denitrification is restricted to anoxic sediments and O2 -deficient waters. For these reasons, N2 O production is insignificant in most open ocean surface waters. In extremely O2 -deficient waters, N2 O can be consumed during denitrification because of its use as an electron acceptor by denitrifying bacteria (14). Coupling of the two processes, in which the NO3 − developing from nitrification is consumed by denitrification, occurs both in marine sediments (15) and around the fringes of O2 depleted waters in the open ocean (16). The net rate of N2 O production by nitrification and denitrification is influenced by several factors in addition to dissolved O2 availability, including the supply rates of NO3 − and NH4 + , the composition of the microbial ecosystem, and in sediments, physicochemical aspects such as porosity and grain size. Consequently, both processes show pronounced seasonal variability. These aspects, coupled with a nonuniform distribution of N2 O source regions, makes the total marine source of N2 O difficult to quantify. Current data indicate around two-thirds of the marine source of tropospheric N2 O to derive from the open ocean, with the remainder coming from coastal waters. However, these estimates contain uncertainties resulting from incomplete spatial and seasonal sampling and difficulties related to estimating sea-to-air fluxes. Open ocean emissions are approximately equally distributed between the northern and southern hemispheres. Most of the N2 O is located below the surface-mixed layer and highest concentrations occur at around 500–1000 m depth, where a high O2 demand results from the bacterial decomposition of sinking organic particles. Exchange of this water with the atmosphere is usually slow, except during wintertime when the surface-mixed layer deepens because of cooling and wind-driven mixing, entraining waters from below. Regions experiencing strong seasonal upwelling are especially strong sources of tropospheric N2 O, which include the Tropical North Pacific, the Arabian Sea/northwestern Indian Ocean, the equatorial upwelling, and along the coasts of northwest Africa and western central and South America. In these areas, the upwellings bring N2 O-rich waters to the surface along with a plentiful supply of nutrients that fuels high primary productivity. The resultant large downward flux of organic particles gives rise to strongly O2 -deficient waters that replenish

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the deep water inventory of N2 O through nitrificationdenitrification coupling (17). Some areas of the subtropical gyres and the North Atlantic seem to be weak sinks for tropospheric N2 O in winter and weak sources in summer. The major coastal N2 O source regions are estuaries, open coastal shelf waters being generally at or close to equilibrium with tropospheric N2 O levels. Both water column nitrification associated with high suspended particle populations (18) and sediment denitrification (19) have been identified in estuaries. Whereas oceanic N2 O emissions are considered to be wholly natural, significant N2 O in coastal waters appears to derive indirectly from anthropogenic activity. In particular, the use of fertilizers is reflected in enhanced transport of nitrogen to coastal waters and a consequent increase in N2 O production and tropospheric flux. As much as 90% of current estuarine N2 O emissions and 25% of continental shelf emissions may be anthropogenic and consistent with the geographic distribution of fertilizer use, human population, and atmospheric nitrogen deposition, more than 80% of these anthropogenic sources are located in the Northern Hemisphere mid-latitudes between 20◦ N and 66◦ N (20). SOURCES OF SEAWATER CH4 Like N2 O, seawater CH4 has a microbial source (microbial methanogenesis). Methanogenesis is inhibited by dissolved O2 and therefore usually occurs in anoxic sediments or in waters that are strongly O2 -depleted. Even so, CH4 concentrations in the oxygenated surface ocean are on average about 30% above the tropospheric equilibrium value, most likely reflecting methanogenesis by O2 -tolerant methanogens inside bacterially maintained ‘‘anoxic microniches’’ in the guts of zooplankton and/or in particles (21). Consequently, the open ocean represents a small CH4 source. In addition, regions of much higher CH4 concentration occur in upwelling areas, associated with enhanced primary productivity as for N2 O (16,22). Based on the available open ocean data, the marine contribution to tropospheric CH4 is about 10 Tg CH4 per year, which is equivalent to about 3% of the total global source (1). Coastal waters have much higher CH4 concentrations than the open ocean but have, until recently, been excluded from global CH4 source estimates because of a lack of data. For estuaries, values of 100 to 200 times the background equilibrium value are common. Such CH4 levels reflect direct inputs from rivers, coastal seawater, underlying sediments, and in situ production/consumption from water column methanogenesis and microbial CH4 oxidation. Recent work suggests that correctly accounting for these regions could increase the estimated marine CH4 source by around 50% (23). A potentially even larger CH4 source may be geologically sourced CH4 from natural marine seeps, which are most common on shallow continental shelves (24). Seeps are episodic in nature, and the CH4 fluxes developing from them are predominantly by bubbles, which complicates making accurate measurements. As a result of CH4 losses because of bubble dissolution and subsequent CH4 oxidation in the water column, shallow water seeps are

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much more effective than deep water seeps at contributing CH4 to the troposphere. Revised source estimates that take account of seep occurrences suggest that the total marine source of tropospheric CH4 could be as much as 40 Tg yr−1 (25). If so, the oceans could be a much more important source of tropospheric CH4 than previously thought. VOLATILE SULPHUR COMPOUNDS The marine system is an important source of sulphur globally; in particular, it is the principal source of biogenic atmospheric sulphate aerosol, which plays important roles in atmospheric chemistry and climate. The predominant volatile sulphur in surface seawater is dimethyl sulphide (DMS) [(CH3 )2 S], which is a byproduct of algal metabolism and accounts for about 95% of marine sulphur emissions and 20% of total global sulphur emissions (1). DMS is rapidly oxidized by free radicals in the lower troposphere. Sulphur dioxide (SO2 ) is the major reaction product and is subsequently transformed to sulphate aerosol through gas-to-particle reactions. This process is hypothesized to impact directly the radiative forcing of global climate, primarily through changes to cloud albedo (26). Carbonyl sulphide (COS) has also been implicated in global climate forcing. Compared with other volatile sulphurs, COS has a long tropospheric residence time of around 2–6 years. Consequently, it is transported into the stratosphere, where its photo-oxidation is believed to be an important source of sulphate aerosol, which is thought to impact Earth’s radiation balance (27) and stratospheric ozone levels (28). DIMETHYL SULPHIDE A global database of more than 15,000 measurements of surface seawater (29) revealed distinct annual DMS cycles in the open ocean at mid to high latitudes. In the northern hemisphere, open ocean DMS increases during spring–summer, whereas in the southern hemisphere, concentrations peak six months later. These patterns relate to the timing of phytoplankton blooms and seasonal changes in mixed layer depth. In contrast, tropical regions show weak seasonality; DMS is elevated in the upwelling regions off western Africa and South America, but these concentrations are lower than those at high latitudes during summer. On coastal shelves, DMS is spatially and temporally variable, broadly correlating with seasonal primary productivity and the presence of algal blooms associated with upwelling at water mass boundaries (hydrographic fronts) (29). Similar concentrations are found in many estuaries, notable exceptions being those with high concentrations of suspended particles. Seasonal patterns are, however, rather different, with maximum concentrations occurring during late winter/early spring. The available data indicate that coastal regions may be larger emitters of DMS per unit area than the global mean, however, because of its much larger surface area, the open ocean is the most important source of tropospheric DMS (29).

CARBONYL SULPHIDE Direct marine emissions are thought to account for about 20% of the global COS source (30), although the tropospheric oxidation of marine-derived DMS and carbon disulphide (CS2 ) may increase this to as much as 55% (31). COS in seawater primarily develops from the photodecomposition of humic-like colored dissolved organic matter (CDOM) (32), although a small nonphotochemical source has also been inferred. The distribution of sea surface COS, therefore, corresponds closely to the concentration and reactivity of CDOM, which primarily derives from terrestrial sources. Consequently, COS concentrations are highest in estuaries, which are about an order of magnitude higher than those of adjacent coastal waters (32). COS undergoes hydrolysis removal in seawater at a similar rate to that for its photo-production, hence COS shows a pronounced diel cycle in surface seawater with concentrations peaking in the early afternoon and declining to a minimum just before sunrise (33). Strong seasonal variation also exists. In the mid to high latitude open ocean the balance of photo-chemical production versus hydrolysis removal can lead to these regions becoming a seasonal sink for tropospheric COS (34). Net ocean uptake of COS has also been found in the subtropical ocean gyres (35). Taking account of these findings, the marine COS source is most likely dominated by the contribution from coastal and shelf areas. BIBLIOGRAPHY 1. Houghton, J.T., Ding, Y., Griggs, D.J., Noguer, M., van der Linden, P.J., Dai, X., Maskell, K., and Johnson, C.A. (Eds.). (2001). Climate Change 2001: The Scientific Basis. Cambridge University Press. 2. Barnola, J.M. et al. (1995). CO2 evolution during the last millennium as recorded by Antarctic and Greenland ice. Tellus 47B: 264–272. 3. Le Qu´er´e, C., Orr, J., Monfray, P., and Aumont, O. (2000). Interannual variability in the oceanic sink of CO2 from 1979 through 1997. Glob Biogeochem Cyc. 14: 1247–1265. 4. Siengenthaler, U. and Sarmiento, J.L. (1993). Atmospheric carbon dioxide and the ocean. Nature 365: 119–125. 5. Keeling, R.F. and Shertz, S.R. (1992). Seasonal and interannual variations in atmospheric oxygen and implications for the global carbon cycle. Nature 358: 723. 6. Sarmiento, J.L. and Gruber, N. (2002). Sinks for anthropogenic carbon. Physics Today 55: 30–36. 7. Broecker, W.S., Takahashi, T., Simpson, H.J., and Peng, TH. (1979). Fate of fossil fuel carbon dioxide and the global carbon budget. Science 206: 409–418. 8. Cox, P.M., Betts, R.A., Jones, C.D., Spall, S.A., and Totterdell, I.J. (2000). Acceleration of global warming due to carboncycle feedbacks in a coupled climate model. Nature 408: 184. 9. Archer, D.E., Kheshgi, H., and Maier-Reimer, E. (1997). Multiple timescales for neutralization of fossil fuel CO2 . Geophys Res. Lett. 24: 405–408. 10. Martin, J.H. (1992). Iron as a limiting factor. In: Primary Productivity and Biogeochemical Cycles in the Sea. P.G. Falkowski and A. Woodhead (Eds.). Plenum Press, New York, pp. 123–137.

PACIFIC MARINE ENVIRONMENTAL LABORATORY—30 YEARS OF OBSERVING THE OCEAN 11. Sarmiento, J.L., Hughes, T.M.C., Stouffer, R.J., and Manabe, S. (1998). Simulated response of the ocean carbon cycle to anthropogenic climate warming. Nature 393: 245–249. 12. Nevison, C. and Holland, E. (1997). A re-examination of the role of anthropogenically fixed nitrogen on atmospheric N2 O and the stratospheric O3 layer. J. Geophys Res. 10: 15809–15820. 13. Crutzen, P.J. (1990). Methane’s sinks and sources. Nature 350: 390–381. 14. Cohen, Y. and Gordon, L.I. (1979). Nitrous oxide in the oxygen minimum of the of the eastern tropical north Pacific: Evidence for its consumption during denitrification and possible mechanisms for its production. Deep Sea Res. 25: 509–524.

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predict sea surface DMS as a function of latitude, longitude, and month. Glob. Biogeochem. Cyc. 13: 399–444. 30. Watts, S.F. (2000). The mass budgets of carbonyl sulfide, dimethyl sulfide, carbon disulfide and hydrogen sulfide. Atmos. Environ. 34: 761–779. 31. Kettle, A.J. et al. (2001). Assessing the flux of different volatile sulfur gases from the ocean to the atmosphere. J. Geophys. Res. 106: 12193–12209. 32. Uher, G. and Andreae, M.O. (1997). Photochemical production of carbonyl sulfide in North Sea water: A process study. Limnol. Oceanogr. 42: 432–442. 33. Ulsh¨ofer, V.S., Fl¨ock, O.R., Uher, G., and Andreae, M.O. (1996). Photochemical production and air-sea exchange of carbonyl sulfide in the eastern Mediterranean Sea. Mar. Chem. 53: 25–39.

15. Vanraaphorst, W. et al. (1992). Nitrogen cycling in 2 types of sediments of the southern North-Sea (frisian front, broad fourteens)—field data and mesocosm results. Neth. J. Sea Res. 28: 293–316.

34. Ulshofer, V.S., Uher, G., and Andreae, M.O. (1995). Evidence for a winter sink of atmospheric carbonyl sulfide in the northeast Atlantic Ocean. Geophys. Res. Lett. 22: 2601–2604.

16. Naqvi, S.W.A. et al. (1998). Budgetary and biogeochemical implications of N2 O isotope signatures in the Arabian Sea. Nature 394: 462–464.

35. Weiss, P.S., Johnson, J.E., Gammon, R.H., and Bates, T.S. (1995). A reevaluation of the open ocean source of carbonyl sulfide to the atmosphere. J. Geophys. Res. 100: 23083–23092.

17. Suntharalingam, P. and Sarmiento, J.L. (2000). Factors governing the oceanic nitrous oxide distribution: Simulations with an ocean general circulation model. Glob. Biogeochem. Cyc. 14: 429–454. 18. Barnes, J. and Owens, N.J.P. (1998). Denitrification and nitrous oxide concentrations in the Humber estuary, UK, and adjacent coastal zones. Mar. Poll. Bul. 37: 247–260. 19. Dong, L.F., Nedwell, D.B., Underwood, G.J.C., Thornton, D.C.O., and Rusmana, I. (2002). Nitrous oxide formation in the Colne estuary, England: The central role of nitrite. Appl. Environ. Microbiol. 68: 1240–1249. 20. Seitzinger, S.P. and Kroeze, C. (2000). Global distribution of N2 O emissions from aquatic systems: natural emissions and anthropogenic effects. Chenmosphere 2: 267–279. 21. Lamontagne, R.A., Swinnerton, J.W., Linnenbom, V.J., and Smith, W.D. (1973). Methane concentrations in various marine environments. J. Geophys. Res. 78: 5317–5324. 22. Upstill-Goddard, R.C., Barnes, J., and Owens, N.J.P. (1994). Nitrous oxide and methane during the 1994 SW monsoon in the Arabian Sea/noerthwestern Indian Ocean. J. Geophys. Res. 104: 30,067–30,084. 23. Upstill-Goddard, R.C., Barnes, J., Frost, T., Punshon, S., and Owens, N.J.P. (2000). Methane in the southern North Sea: Low-salinity inputs, estuarine removal, and atmospheric flux. Glob. Biogeochem. Cyc. 14: 1205–1217. 24. Judd, A.G. (2003). The global importance and context of methane escape from the seabed. Geo-Marine Lett. 23: 147–154. 25. Kvenvolden, K.A., Lorenson, T.D., and Reeburgh, W.S. (2001). Attention turns to naturally occurring methane seepage. EOS 82: 457. 26. Charlson, R.J., Lovelock, J.E., Andreae, M.O., and Warren, S.G. (1987). Oceanic phytoplankton, atmospheric sulphur, cloud albedo and climate. Nature 326: 655–661. 27. Turco, R.P., Whitten, R.C., Toon, O.B., Pollack, J.B., and Hamill, P. (1980). OCS, stratospheric aerosols and climate. Nature 283: 283–286. 28. Chin, M. and Davies, D.D. (1995). A reanalysis of carbonyl sulphide as asource of stratospheric background sulphur aerosol. J. Geophys. Res. 100: 8993–9005. 29. Kettle, A.J. et al. (1999). A global database of sea surface dimethylsulfide (DMS) measurements and a procedure to

PACIFIC MARINE ENVIRONMENTAL LABORATORY—30 YEARS OF OBSERVING THE OCEAN NOAA—Pacific Marine Environmental Laboratory

Sept. 29, 2003—Although the NOAA Pacific Marine Environmental Laboratory in Seattle, Wash., celebrates its 30th anniversary this year, its staff has spent 43 years at sea. The figure of 15,654 days at sea was one of the many facts presented during the lab’s anniversary celebration in August. That, along with 1,290 published journal articles and 352,000,000 hits on the PMEL Web page indicate that there’s a lot going on out on Sand Point. For two-thirds of its life, the lab has been under the direction of Eddie Bernard. An oceanographer by training, Bernard became director in 1983, a decade after the former Pacific Oceanographic Laboratory became PMEL. ‘‘We have dedicated people at PMEL who devote a lot of energy and creativity to the work we do,’’ he said. NATIONAL TSUNAMI MITIGATION PROGRAM Some of that creativity and energy became evident when in 1994 the U.S. Senate asked NOAA to come up with a plan to reduce the risk of tsunamis to coastal residents. What resulted was the National Tsunami Hazard Mitigation Program, chaired by Bernard and composed of representatives from federal, state and local agencies from West Coast states, Alaska and Hawaii, working to save lives and property. ‘‘The National Tsunami Mitigation Program initiated by PMEL is a unique and effective partnership,’’ said Rich Eisner of the California Governor’s This article is a US Government work and, as such, is in the public domain in the United States of America.

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Office of Emergency Services, one of the program members. ‘‘The integration of science and mitigation policy, the warning centers, and local emergency management, and the application of new technologies fostered by PMEL have been successful beyond expectations.’’ Among the technological innovations is the system of buoys along the West Coast that serve as warning devices, or, as Bernard calls them, ‘‘tsunameters.’’ The tsunami program also includes a public education component that teaches coastal residents what to do in case of a possible tsunami. Signs now indicate evacuation routes and some coastal communities have been designated ‘‘TsunamiReady’’ for their efforts to educate and protect their residents. UNDERWATER VOLCANOES AND VENTS As one of NOAA’s ‘‘wet’’ labs, PMEL focuses on a variety of ocean issues. When underwater volcanoes or vents, were first discovered in the Galapagos Islands 26 years ago, PMEL was among the first to start investigating these unusual underwater communities, where unique marine life thrive on the chemical soup spewed from the sea floor. ‘‘We may be taking drugs in the future made of enzymes that are more compatible with our bodies than synthetic compounds, which may have side effects,’’ Bernard said.

‘‘What’s spewing from the ocean floor could someday give us resistance to some new strains of infection.’’ FISHERIES OCEANOGRAPHY COORDINATED INVESTIGATIONS PROGRAM PMEL began as a ‘‘small research laboratory with emphasis on water quality and environmental impact issues’’ in the waters off the West Coast extending to the equatorial Pacific Ocean. It now has an international reputation in many areas, especially its ability to collect ocean data and to work collaboratively in projects that cover many disciplines. One example is the Fisheries Oceanography Coordinated Investigations program that assists in forecasting fish stocks to help ensure a reliable supply and lower costs to consumers. ‘‘In 1985, Eddie Bernard took a big risk,’’ said Doug DeMaster of the NOAA Marine Fisheries Service. That risk was offering to establish with the Alaska Fisheries Science Center and his counterpart, William Aron, a cutting-edge, applied

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relaying surface wind, sea surface temperature, upper ocean temperatures and currents, air temperature and relative humidity in real-time via satellite. ‘‘We knew we were onto something when we linked data from the TAO buoys to the Internet and attracted millions of hits from all over the world,’’ Bernard said. ‘‘All it takes is time, money and commitment.’’ UNDERWATER ACOUSTICAL MONITORING

science program across NOAA line offices. Eighteen years after its inception, NOAA’s FOCI has published more than 450 scientific articles and was awarded the Department of Commerce Bronze Medal in 2002 for ‘‘scientific achievements that have advanced fisheries oceanography and marine ecology and have contributed to building sustainable fisheries in the North Pacific.’’ DeMaster noted, ‘‘Today, if you attend a FOCI meeting, you cannot tell which scientists are from NOAA Research and which are from NOAA Fisheries. In 1985, it took vision and courage to blur the lines between line offices. Today, it seems only natural.’’ PACIFIC TROPICAL ATMOSPHERE OCEAN ARRAY Understanding the natural systems is a key element of the lab. ‘‘The ocean is dynamic, it moves all of the time,’’ Bernard said. ‘‘We’re now in the third generation of observing systems. In the equatorial Pacific, we have the world’s longest continuous time series of open ocean data—25 years.’’ The equatorial Pacific also proved to be the place to be if humans wanted early warning ˜ events. El Nino ˜ is a disruption of the of El Nino ocean-atmosphere system in the tropical Pacific having important consequences for weather around the globe. ˜ considered the most intense After the 1982–83 El Nino, in the 20th century, the challenge was given to develop some sort of early warning device so people could prepare ˜ for the devastating and beneficial aspects of an El Nino ˜ Once again, in 1994, PMEL and its counterpart, La Nina. harnessed the creativity and talent of dedicated scientists and came up with the Pacific Tropical Atmosphere Ocean (TAO) array, the world’s largest ocean observing system. Bobbing in the Pacific are 70 buoys measuring and

Always eager to hear what the Earth has to say, PMEL scientists also listen to the planet via underwater acoustical monitoring. Using a variety of methods, including underwater hydrophones, PMEL listens for seismic activity, marine mammals and ship traffic. The systems also have picked up some so-far unidentified sounds. ‘‘People tend to think the ocean is quiet beneath the surface,’’ said Christopher Fox, director of the ocean acoustics project. ‘‘But it’s pretty noisy down there.’’ Some things are easy to identify, Fox said, such as whales and ship traffic. But visitors to the ocean acoustics Web site can listen to such unidentified sounds that the lab has dubbed ‘‘Train,’’ ‘‘Upsweep,’’ ‘‘Whistle’’ and ‘‘Bloop.’’ After 30 years, PMEL knows that the Earth still holds countless tantalizing secrets. And PMEL scientists and staff are eager to unlock those secrets. ‘‘As the planet aspirates, it provides new opportunities,’’ Bernard

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said. ‘‘It’s an ongoing science experiment with enormous challenges and rewards.’’ RELEVANT WEB SITES NOAA Pacific Marine Environmental Laboratory Dr. Eddie Bernard, Director, Pacific Marine Environmental Laboratory NOAA National Tsunami Hazard Mitigation Program NOAA Vents Program Fisheries Oceanography Coordinated Investigations program NOAA Marine Fisheries Service NOAA Alaska Fisheries Science Center ˜ Theme Page NOAA El Nino ˜ Theme Page NOAA La Nina NOAA Pacific Tropical Atmosphere Ocean (TAO) Project NOAA Vents Program: Underwater Acoustical Monitoring NOAA’s ALASKA FISHERIES SCIENCE CENTER Ocean Explorer: Sounds in the Sea

MEDIA CONTACT Jana Goldman, NOAA Research, (301) 713–2483

SEAWATER TEMPERATURE ESTIMATES IN PALEOCEANOGRAPHY GIUSEPPE CORTESE Alfred Wegener Institute for Polar and Marine Research (AWI) Bremerhaven, Germany

BACKGROUND Climatic issues, such as global warming, greenhouse gases release into the atmosphere, and the role mankind plays in affecting the climate of our planet have recently gained international interest. As a result, political panels and decision-makers are starting to look with growing interest at the results of committees established with the purpose of analyzing past, present, and future climate change. One of these panels [Intergovernmental Panel on Climate Change (IPCC)], established by the World Meteorological Organization (WMO) and the United Nations Environment Programme (UNEP), regularly publishes reports (available also online at http://www.grida.no/climate/ipcc tar/) intended for scientists, politicians, and anybody interested in and/or concerned by climate change. The most powerful tools currently available to the scientific community for the analysis and prediction of climate are computer models, which simulate the functioning of Earth as a complex system. These models differ based on which components of the climate system (atmosphere, ocean, biosphere, cryosphere, etc.) they include and account for, their spatial scale and degree of detail, and on the different sets of physicochemical equations they

use to describe and parameterize phenomena taking place at a subgrid spatial scale. Regardless of setup and characteristics, though, models should be benchmarked against an observational base (e.g., to test how well the model reproduces a known distribution of a climatic variable). Models are moreover usually initialized by a set of observations (to let the incorporated physics modify the starting distribution). Some of these datasets are readily available in the form of measurements performed over historic timescales: One of the best known examples is the atmospheric CO2 concentration curve measured at Mauna Loa, Hawaii, since the late 1950s (1). Other important climatic variables cannot be observed or measured directly, and indirect methods, so-called ‘‘proxies,’’ have been developed to obtain information on them. Some of the methods that estimate seawater temperature on geological timescales are described in the following. INTRODUCTION The heat balance of Earth is strongly affected, to a first approximation, by the amount of solar radiation reaching the top of the atmosphere. The motion of currents in the atmosphere and in the ocean redistributes this heat between low and high latitudes, as the low latitudes receive a higher amount of heat compared with the higher latitudes. The ocean can be regarded as the ‘‘thermostat’’ of such a heat machine, because water has a much higher heat capacity compared with air, and because the turnover time for the ocean (thousands of years) is several orders of magnitude higher than for the atmosphere (days). One of the main features of the large-scale circulation in the ocean is the formation of deep waters at high latitudes, in the Weddell and Ross Seas (around Antarctica), and in the Labrador and Greenland/Iceland Seas (in the northern North Atlantic), because of the cooling of enormous amounts of warm surface waters advected to these locations by ocean currents (e.g., the Gulf Stream in the North Atlantic). When these waters move to higher latitudes, they cool, become denser, and sink to the ocean depths, where they start a reversed journey, from high to low latitudes. During this cooling, they also release latent heat and moisture, which help to mitigate the climate of the coastal areas they fringe, including the whole of western Europe. Indications exist, from geological records and model simulations (2), that during glacial times, this oceanic overturning did not operate with the same intensity as today, and/or that the places where deep waters formed were displaced compared with today. As seawater temperature, together with salinity, determines the density of seawater (and therefore it strongly influences its circulation and sinking characteristics), a variety of methods have been developed to determine this important climatic variable in the distant, geological past. METHODS Sediment cores collected from the bottom of the World Ocean represent an ideal archive to trace the history of

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the ocean. Some of these cores can be accurately dated, and often they contain the fossilized remains of organisms that lived in past oceans, and reacted to changes in their characteristics. The most useful of these organisms, because of their high fossilization potential, species diversity, and sensitivity to a range of climatic variables, are micro- and nannofossils. As their name implies, these organisms range from nanometer to millimeter size and leave fossilizable body parts, with the most commonly used groups belonging to algae (diatoms, nannoplankton) and protozoans (radiolarians, foraminifera). Most species in these groups live close to the sea surface, and each of them is adapted to a particular set of sea surface water variables, including temperature, salinity, and macro- and micro-nutrient content. Several techniques have been developed to exploit this close link between species occurrences/abundances and past environmental variables. Our first global view of surface water temperatures in the ocean during the Last Glacial Maximum (ca. 18,000 years Before Present) came from one such technique, namely, Q-Mode Principal Component Analysis applied to foraminifer and radiolarian abundance data in sediment cores (3). Semiquantitative Floral/Faunal Estimates Some of the first attempts to obtain seawater temperature information from microfossil assemblages tried to derive this value directly from a formula. The latter was based on the presence/absence, in the fossil samples, of species having a well-known distribution in the modern ocean, and being representative of different climatic zones (e.g., equatorial, subtropical, temperate, etc.). This approach, initially developed for diatoms (4), has also been applied to radiolarians (5). Information on species diversities of modern planktonic foraminifera in the Indian Ocean have also been used, in the late 1970s, to estimate ocean paleotemperatures (6). Transfer Functions With the expansion of the knowledge about the biogeography of most plankton groups (and the ecological and environmental significance of many species), and the development of computers (allowing the implementation of more sophisticated algorithms and techniques), transfer functions made their breakthrough in paleoceanography. They were first described and applied to planktonic foraminifera by Imbrie and Kipp (7), who illustrated their utilization in paleoclimatology. The species assemblage present in a collection of modern sediment samples (calibration dataset), containing up to several hundreds of samples, is ‘‘simplified’’ into a limited number of faunal factors/assemblages. A multivariate regression is then used to calibrate these simplified assemblages to the desired environmental variable (e.g., surface seawater temperature) measured, in the water column, at the same locations where the surface sediment samples were recovered. The value of the environmental variable at a certain time in the past is estimated by calculating ‘‘pseudo-factors’’ for each of the past samples (i.e., the value the calibrated assemblages have in these samples) and replacing them in the calibration equation. One main

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assumption of this method is that the considered environmental variable (in our case, surface temperature) plays a major role in influencing the biogeographic distribution of the considered microfossil species, and that it is therefore responsible for most of the variance observed in the calibration dataset. Transfer functions have been successfully applied to past ocean surface temperature reconstructions by using planktonic foraminifera (8–14), radiolarians (15–22), and diatoms (23,24). Geochemical Methods Several methods have been recently developed to estimate past ocean temperatures by measuring chemical elements or organic compounds, which are either incorporated in the organism shell or produced during its life cycle. The general approach is to develop a calibration equation by growing the desired species in culture under a variety of environmental conditions, and measuring how its geochemical composition changes. Additionally, a calibration equation can also be obtained by analyzing extensive collections of recent sediments, covering areas where the environmental variables display a wide range. The resulting equation is then applied to fossil samples to derive the past value of the desired variable. Mg/Ca The relative amount of magnesium compared with calcium a certain planktonic foraminifera species incorporates in its test depends on temperature, and the Mg/Ca ratio can therefore also be used as a paleo-thermometer. The calibration equation is derived from laboratory culture studies, based on different species, including Globigerinoides sacculifer (25), Globigerina bulloides, and Orbulina universa (26). Revised calibrations and applications to geological records are being continuously developed (27–31). Another paleo-thermometer derived from elemental ratios in foraminifera tests is the Sr/Ca ratio (32). Alkenones Several species of nannoplankton, living in the shallower layer of the ocean (less than ca. 50 m depth), produce organic compounds named long-chain (C37 -C39 ) unsaturated ketones, also known as alkenones. The two species being responsible for most of the alkenone production are Emiliania huxleyi and Gephyrocapsa oceanica. Although it is still not clear what the function of such compounds is, they have become a very useful tool for the estimation of past seawater temperature. It has been demonstrated (33,34) that the three different varieties of the C37 unsaturated compounds (C37:2 , C37:3 , C37:4 ), characterized by 2, 3, or 4 double bonds, display abundance variations related to seawater temperature. The alkenone unsatu ration ratio (35), Uk 37 = [C37:2 ]/([C37:2 ] + [C37:3 ]), was calibrated to temperature (36) by growing E. huxleyi in culture under a range of temperatures. Since then, alkenones have been widely applied in paleoclimatology (37–45). Other Methods Other methods commonly used to estimate past ocean temperatures are as follows the modern analogue technique and its variations (46–51); artificial neural

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networks (52,53); and the δ 18 O isotopic composition of planktonic foraminifera (54–58). The study of the isotopic signal stored in corals, although covering shorter time intervals than marine sediment cores, provides excellent temporal resolution archives (down to a few weeks), which allow to analyze seasonal climate variability (59,60). BIBLIOGRAPHY 1. Keeling, C.D., Whorf, T.P., Wahlen, M., and van der Plicht, J. (1995). Interannual extremes in the rate of rise of atmospheric carbon dioxide since 1980. Nature 375: 666–670. 2. Ganopolski, A. and Rahmstorf, S. (2001). Rapid changes of glacial climate simulated in a coupled climate model. Nature 409: 153–158. 3. CLIMAP. (1976). The surface of the ice-age Earth. Science 191: 1131–1137. 4. Kanaya, T. and Koizumi, I. (1966). Interpretation of diatom thanatocoenoses from the North Pacific applied to a study of core V20-130 (studies of a deep-sea core V20-120, part IV). Sci. Rep. Tohoku Univ. Ser. 2. 37: 89–130. 5. Nigrini, C. (1970). Radiolarian assemblages in the North Pacific and their application to a study of Quaternary sediments in core V20-130. Memoir of the Geological Society of America 126: 139–183. 6. Williams, D.F. and Johnson, W. (1975). Diversity of recent planktonic foraminifera in the southern Indian Ocean and late Pleistocene paleotemperatures. Quaternary Res. 5: 237–250. 7. Imbrie, J. and Kipp, N.G. (1971). A new micropaleontological method for quantitative paleoclimatology: Application to a late Pleistocene Caribbean core. In: Late Cenozoic Glacial Ages. K. Turekian (Ed.). Yale University Press, New Haven, CT, pp. 71–181. 8. CLIMAP. (1981). Seasonal Reconstructions of the Earth’s Surface at the Last Glacial Maximum. Geological Society of America, Map and Chart Series MC-36. 9. CLIMAP. (1984). The Last Interglacial ocean. Quaternary Research 21: 123–224. 10. Mix, A.C. and Morey, A.E. (1996). Climate feedback and pleistocene variations in the atlantic south equatorial current. In: The South Atlantic: Present and Past Circulation. G. Wefer, W.H. Berger, G. Siedler, and D.J. Webb (Eds.). Springer-Verlag, Berlin, pp. 503–525. 11. Ortiz, J.D. and Mix, A.C. (1997). Comparison of Imbrie-Kipp transfer function and modern analog temperature estimates using sediment trap and core top foraminiferal faunas. Paleoceanography 12(2): 175–190. 12. Mix, A.C., Morey, A.E., Pisias, N.G., and Hostetler, S.W. (1999). Foraminiferal faunal estimates of paleotemperature: Circumventing the no-analog problem yields cool ice age tropics. Paleoceanography 14: 350–359. 13. Feldberg, M.J. and Mix, A.C. (2002). Sea-surface temperature estimates in the Southeast Pacific based on planktonic foraminiferal species; modern calibration and Last Glacial Maximum. Marine Micropaleontol. 44: 1–29. 14. Niebler, H.S. et al. (2003). Sea-surface temperatures in the Equatorial and South Atlantic Ocean during the Last Glacial Maximum (23-19 ka). Paleoceanography 18: 1069. 15. Moore, T.C., Jr. (1973). Late Pleistocene-Holocene oceanographic changes in the northeastern Pacific. Quaternary Res. 3(1): 99–109. 16. Morley, J.J. (1979). A transfer function for estimating paleoceanographic conditions, based on deep-sea surface

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48. Pflaumann, U., Duprat, J., Pujol, C., and Labeyrie, L. (1996). SIMMAX: A modern analog technique to deduce Atlantic sea surface temperatures from planktonic foraminifera in deep-sea sediments. Paleoceanography 11: 15–35. 49. Waelbroeck, C. et al. (1998). Improving past sea surface temperature estimates based on planktonic fossil faunas. Paleoceanography 13(3): 272–283. 50. Hale, W. and Pflaumann, U. (1999). Sea-surface temperature estimations using a Modern Analog Technique with foraminiferal assemblages from Western Atlantic Quaternary sediments. In: Use of Proxies in Paleoceanography—Examples from the South Atlantic. G. Fischer and G. Wefer (Eds.). Springer, Berlin, pp. 69–90. 51. Hendy, I.L. and Kennett, J.P. (2000). Dansgaard-Oeschger cycles and the California Current System: Planktonic foraminiferal response to rapid climate change in Santa Barbara Basin, Ocean Drilling Program Hole 893A. Paleoceanography 15: 30–42. 52. Malmgren, B.A. and Nordlund, U. (1997). Application of artificial neural networks to paleoceanographic data. Palaeogeography, Palaeoclimatology, Palaeoecology 136: 359–373. 53. Malmgren, B.A., Kucera, M., Waelbroeck, C., and Nyberg, J. (2001). Comparison of statistical and artificial neural network techniques for estimating past sea-surface temperatures from planktonic foraminifer census data. Paleoceanography 16(5): 520–530. 54. Emiliani, C. (1955). Pleistocene temperatures. J. Geol. 63: 538–578. 55. Shackleton, N.J. and Opdyke, N.D. (1973). Oxygen isotope and palaeomagnetic stratigraphy of equatorial Pacific Core V28-238: Oxygen isotope temperatures and ice volumes on a 105 year and 106 year scale. Quaternary Res. 3: 39–55. 56. Broecker, W.S. (1986). Oxygen isotope constraints on surface ocean temperatures. Quaternary Research 26: 121–134. ¨ 57. Mulitza, S., Durkoop, A., Hale, W., Wefer, G., and Niebler, H.S. (1997). Planktonic foraminifera as recorders of past surface-water stratification. Geology 25: 335–338. 58. Bemis, B.E., Spero, H.J., Bijma, J., and Lea, D.W. (1998). Reevaluation of the oxygen isotopic composition of planktonic foraminifera: Experimental results and revised paleotemperature equations. Paleoceanography 13(2): 150–160. 59. Felis, T. et al. (2000). A coral oxygen isotope record from the northern Red Sea documenting NAO, ENSO, and North Pacific teleconnections on Middle East climate variability since the year 1750. Paleoceanography 15: 679–694. 60. Nyberg, J., Winter, A., Malmgren, B., and Christy, J. (2001). Surface temperatures in the eastern Caribbean during the 7th century AD average up to 4 ◦ C cooler than present. Eos Trans. AGU 82(47).

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¨ ¨ 45. Ruhlemann, C., Mulitza, S., Muller, P.J., Wefer, G., and Zahn, R. (1999). Warming of the tropical Atlantic Ocean and slowdown of thermohaline circulation during the last deglaciation. Nature 402: 511–514.

Pacific Fisheries Environmental Laboratory, NOAA

46. Hutson, W.H. (1980). The Agulhas Current during the late Pleistocene: analysis of modern faunal analogs. Science 207: 64–66.

A great deal of the patterns and fluctuations observed in our living marine resources are attributable to the impact of physical processes in the environment on marine

47. Prell, W.L. (1985). The Stability of Low Latitude Sea Surface Temperatures: An Evaluation of the CLIMAP Reconstruction with Emphasis on the Positive SST Anomalies. Rep. TR 025 U.S. Dept. of Energy, Washington, DC, pp. 1–60.

This article is a US Government work and, as such, is in the public domain in the United States of America.

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ecosystems and their components. For this reason, PFEL places a strong emphasis on research that examines the role of the physical environmental variability on marine ecosystems in general and commercially important fish stocks specifically. The objectives of the physical oceanography task are: • perform research on the temporal and spatial scales of environmental variability in eastern boundary current systems in relation to other marine ecosystems • provide environmental input to SWFSC research programs, particularly the coastal groundfish program • provide high quality marine information to the research community. Research is performed at PFEL which integrates environmental and biological data sets, investigating the linkages between environmental variability and fluctuations in the abundance and distribution of marine populations on a continuum of scales (from global, basin-wide spatial scales to the scale of local upwelling centersand on time scales from decades down to days). Physical oceanography research is directed to: • large-scale climatic variability • environment/recruitment relationships in eastern boundary current ecosystems • mesoscale (smaller scale) processes affecting coastal circulation and fisheries recruitment Examples of research on the large-scale variability include studies of recurring temperature changes off the west coast of the U. S. and their effects on groundfish recruitment, and the investigation of environmental changes in the California Current region associated with recent ENSO events. Much of the mesoscale research focuses on relating environmental variability on day-to-year and 5–100 nm scales to patterns and events in the life history of groundfish (e.g., recruitment success). The physical oceanography research program is linked closely to those of the other tasks at PFEG and to research programs at the other SWFSC labs. PFEG scientists also are involved in numerous cooperative studies with oceanographers and

fisheries scientists at many federal and state government, academic and privately supported research institutions. Expertise in physical oceanography at PFEG and the linkages to the Navy’s Fleet Numerical Meteorology and Oceanography Center (FNMOC) and Naval Postgraduate School, as well as numerous other government, academic, and private research facilities, has historically meant that this task serves regionally, nationally and internationally as a resource to other ocean scientists. Within the SWFSC, many cooperative research programs have been developed and planned. As an example, the task provides physical oceanographic expertise to the Tiburon Laboratory Rockfish Recruitment surveys each spring, to relate ocean variability off central California to rockfish recruitment. PFEG physical oceanographers are asked frequently to attend workshop and present seminars as experts on environmental-fishery linkages, and represent SWFSC, NMFS and NOAA on numerous committees and working groups.

COASTAL WATER POLLUTANTS UPADHYAYULA V. K. KUMAR Choa Chu kang Ave-4 Singapore

POLLUTANT INPUT INTO THE MARINE ENVIRONMENT Among all the diversity of human activities and sources of pollution, we can distinguish three main ways that pollutants enter the marine environment: • direct discharge of effluents and solid wastes into the seas and oceans (industrial discharge, municipal waste discharge, coastal sewage, and others); • land runoff into the coastal zone, mainly from rivers; • atmospheric fallout of pollutants transferred by the air mass onto the sea surface. Certainly, the relative contribution of each of these channels to the combined pollution input into the sea

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will be different for different substances and in different situations. At present, the signs and consequences of human activity can be found everywhere on Earth. One of the typical features of marine pollution is global spreading of a number of contaminants. Numerous data undoubtedly indicate the existence of a large-scale (global) field of background contamination of the hydrosphere. Another important feature of marine pollution is the existence of increased pollution levels in the enclosed seas and coastal waters, compared with the open ocean. Contamination levels also increase during the transition from the southern parts of all oceans to the north, where the main industrial centers and main pollution sources are concentrated. Besides the general distribution pattern of pollution sources, there are two other factors explaining the relative stability of global pollutant distribution in the world ocean: the relative confinement of large-scale water circulation within the limits of each hemisphere and the predominance of the zonal transport of the traces in the atmosphere. Another distinctive and repeatedly registered feature of the general picture of contaminant distribution in the marine environment is their localization at the water–atmosphere and water–bottom sediment boundaries. Practically everywhere and for all trace components (primarily for oil hydrocarbons), their concentrations are considerably (usually hundreds and thousands of times) higher in the surface microlayer of water and in the upper layer of bottom sediments. These boundaries provide the biotopes for the communities of hyponeuston and benthos, respectively. The existence of elevated levels of contaminants in zones of high bioproductivity is extremely alarming ecologically. These zones include the water layer up to 100 m from the water surface (photic layer) and boundaries of natural environments (water–atmosphere and water–bottom sediment, as previously mentioned), as well as enclosed seas, estuaries, and coastal and shelf waters. In particular, in shelf and coastal zones, which occupy 10% of the world ocean surface and less than 3% of its volume, the most intense processes of bioproduction, including the self-reproduction of the main living resources of the sea, take place. The main press of anthropogenic impact is also concentrated here. The number and diversity of pollution components is growing as well. Contaminants that are globally distributed are combined here with hundreds and thousands of ingredients of local and regional distribution. Most of these substances are wastes and discharges from different local industries and activities. Based on the extreme diversity of marine pollution components, the variety of their sources, the scales of distribution, and the degree of hazards, these pollutants can be classified in different ways, depending on their composition, toxicity, persistence, sources, volumes, and so on. TYPES AND FORMS OF WATER POLLUTION Water pollution is attributed to various sources and can be broadly divided into three categories: domestic, industrial, and radioactive wastes, which can be categorized in

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the following forms: (1) thermal pollution; (2) the addition of pathogenic organisms, leading to public health hazard; (3) oil pollution; (4) the addition of inert, insoluble mineral material; (5) the addition of biodegradable organic material that will result in the depletion or complete removal of dissolved oxygen; (6) toxicity due to the presence of synthetic organic compounds and salts of heavy metals; (7) enhanced eutrophication; (8) acid deposition or discharges; and (9) radioactivity. Domestic Sewage The objectionable features of domestic sewage are its population of pathogenic bacteria; its high content of organic matter, which gives it a high oxygen demand; its nutrient content, which gives it the potential of supporting large populations of algae and other plants, which in turn may be as objectionable as the sewage itself; and the obvious aesthetics. The first and last of these can be overcome by proper treatment in biological sewage treatment plants, but the effluent from these plants is generally rich in nutrients. So, the effluent discharged into nearby inshore waters may result in large objectionable weed crops and other plants. An additional difficulty of disposing of domestic sewage and even the effluents from sewage treatment plants in the sea is that the density of the sewage is invariably less than that of seawater. Thus, the sewage tends to float on the surface unless introduced in a region of strong currents, will mix with seawater, and is diluted only slowly. Industrial Wastes These can be highly varied in composition and present a variety of special problems. The wastes may be toxic to plants or animals. They may be highly acidic or basic. If they contain large quantities of organic matter, their BOD may be objectionably high. Surface-active ingredients such as detergents may cause objectionable foaming or disrupt normal bacterial populations. The settable chemicals tend to settle to the seabed, react with mineral content there, and change the entire infrastructure of the seabed itself, which is causing accumulation of heavy metals in the sediments of several seas. Pathogens A variety of pathogenic organisms, including viruses, bacteria, protozoa, and parasitic worms, exist in seawater and can cause diseases in plants, animals, and people. Impacts include human illness, seafood contamination, and recreational beach closures. Pathogens are discharged to coastal waters through both point and nonpoint sources, especially from insufficiently treated sewage that is released from septic systems on land and on ships and from agriculture and storm water runoff. Higher concentrations tend to occur after storms and related overflow of sewer systems, making it difficult to interpret trends and temporal fluctuations. Nutrients Important parameters for monitoring nutrient pollution in coastal waters include the following: nitrogen and

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phosphorus concentrations, maximum bottom dissolved oxygen levels, extent and duration of anoxic and hypoxic conditions, extent of SAV, chlorophyll-a concentrations, turbidity, and duration and extent of algal blooms (by type). Some parameters are important in assessing the vulnerability of an area to pollutants, such as nitrogen and phosphorus, or in determining baseline conditions of the area. Oil Petroleum residues can contaminate marine and coastal waters through various routes: accidental oil spills from tankers, pipelines, and exploration sites; regular shipping and exploration operations, such as exchange of ballast water; runoff from land; and municipal and industrial wastes. Although the main global impact is due to tar balls that interfere with recreational activities at beaches, the impact of petroleum hydrocarbon concentrations on marine organisms in the neuston zone in the ocean—particularly fish eggs and larvae—requires more attention. Large-scale oil spills from tankers often make the headlines; yet, nonpoint sources, such

as regular maritime transportation operations and runoff from land, are actually considered the main contributors to the total oil discharge into the ocean, although conclusive statistics are lacking. Runoff and routine maintenance of the oil infrastructure, it is estimated, account for more than 70% of the total annual oil discharge into the ocean. Both the number and amount of accidental oil spills have been monitored and seem to have declined in the past decade. A single catastrophic event can, however, influence the statistics significantly and have a localized, yet tremendous impact on the ecosystem. Tables 1 and 2 show the variety of oil pollution sources and give expert estimates of the scales of distribution and impact of each of these sources on the marine environment. Even though these estimates can vary up to one to two orders of magnitude (especially for natural oil sources, atmospheric input, and river runoff), the main anthropogenic flows of oil pollution into the marine environment come from land-based sources (refineries, municipal wastes, and river runoff) and transportation activity (tanker oil transportation and shipping). Polycyclic aromatic hydrocarbons (PAHs),

Table 1. Sources and Scale of Oil Pollution Input Into the Marine Environment Environment Types and Source of Input

Scale of Distribution and Impact

Hydrosphere

Atmosphere

Local

Regional

Global

+ +

− −

+ +

? +

− +

+

−

+

+

?

+

−

+

?

−

+ + + + −

+ − − − +

+ + + + +

? + − + +

− ? − ? ?

Natural: Natural seeps and erosion of bottom sediments Biosynthesis by marine organisms Anthropogenic: Marine oil transportation (accidents, operational discharges from tankers, etc.) Marine nontanker shipping (operational, accidental, and illegal discharges) Offshore oil production (drilling discharges, accidents, etc.) Onland sources: sewage waters Onland sources: oil terminals Onland sources: rivers, land runoff Incomplete fuel combustion

Note: +, −, and ? mean, respectively, presence, absence, and uncertainty of corresponding parameters.

Table 2. Estimates of Global Inputs of Oil Pollution Into the Marine Environment (Thousands Tons/Year of Oil Hydrocarbon) 1981∗

1985∗ ∗ ∗

1990∗ ∗ ∗

Source

1973∗

1979∗∗

Land-based sources: Urban runoff and discharges Coastal refineries Other coastal effluents

2,500 200 –

2,100 60 150

1,080 (500–1,250) 100 (60–600) 50 (50–200)

34% – –

1,175 (50%) – –

Oil transportation and shipping: Operational discharges from tankers Tanker accidents Losses from nontanker shipping

1,080 300 750

600 300 200

700 (400-1,500) 400 (300–400) 320 (200–600)

45% – –

564 (24%) – –

Offshore production discharges

80

60

50 (40–60)

2%

47 (2%)

Atmospheric fallout

600

600

300 (50–500)

10%

306 (13%)

Natural seeps

600

600

200 (20–2,000)

8%

259 (11%)

Total discharges

6,110

4,670

3,200

100%

2,351

COASTAL WATER POLLUTANTS

especially benzo(a)pyrene, enter the marine environment mostly from atmospheric deposition. Table 2 illustrates the general trend of declining total input of oil pollution into the world ocean over the years. The global situation reflected in this table certainly may differ at the regional level, which depends on natural conditions, the degree of coastal urbanization, the population density, industrial development, navigation, oil and gas production, and other activities. For example, in the North Sea, offshore production input reached 28% of the total input of oil pollution in last decade, instead of the ‘‘modest’’ 2% on the world scale shown in Table 2, which equaled the annual input of more than 23,000 tons of oil products at the background of their general changeable flow of 120,000–200,000 tons a year in the North Sea. One can expect similar situations in other regions of intensive offshore oil and gas development, for example, in the Gulf of Mexico, Red Sea, Persian Gulf, and Caspian Sea. The persistent pollution in oil production areas in the Caspian Sea or the amounts of annual discharges (about 40 million tons of produced waters polluted by oil products) during offshore drilling in the Gulf of Mexico. At the same time, no reliable balance estimates exist for these regions. The continental shelf of the Gulf of Mexico is also distinctive for intense seepage of natural liquid and gaseous hydrocarbons, which can lead to formation of oil slicks and tar balls on the sea surface, which makes assessing and identifying anthropogenic oil pollution more difficult. In any case, the input of oil hydrocarbons from natural sources into the Gulf of Mexico is larger than in many other areas. In the Baltic Sea, the Sea of Azov, and the Black Sea, the leading role in oil input most likely belongs to landbased sources, which are dominated by river inflow. The Danube River alone annually brings about 50,000 tons of oil to the Black Sea, half of the total oil input of about 100,000 tons. Observations in the Caribbean basin, where annually up to 1 million tons of oil enter the marine environment, showed that about 50% of this amount came from tankers and other ships. In the Bay of Bengal and the Arabian Sea, oil pollution inputs from tanker and other ship discharges equal, respectively, 400,000 tons and 5 million tons of oil a year. The most intense tanker traffic exists in the Atlantic Ocean and its seas, which accounts for 38% of international maritime oil transportation. In the Indian and Pacific Oceans, this portion is, respectively, 34% and 28%. Enforcing stricter requirements on activities accompanied by oil discharges led to a global decline of oil pollution inputs in the marine environment mentioned before. A number of dramatic events show the vulnerability of making an optimistic prognosis about decreasing oil pollution at the regional and global levels. For instance, catastrophic large-scale events took place in the Persian Gulf during and after the 1991 Gulf War. Between 0.5 and 1 million tons of oil were released into coastal waters. Besides, products of combustion of more than 70 million tons of oil and oil products were emitted into the atmosphere. Another large-scale accident occurred in Russia in September–November 1994. About 100,000 tons of oil

99

were spilled on the territory of the Komi Republic, which threatened to cause severe oil pollution for the Pechora River basin and, possibly, Pechora Bay. It must be remembered that catastrophes, in spite of the obvious consequences and all the attention they attract, are inferior to other sources of oil pollution in their scales and degree of environmental hazard. Land-based, oil-containing discharges and atmospheric deposition of products of incomplete combustion can accordingly give 50% and 13% of the total volume of oil pollution input into the world ocean (see Table 2). These diffuse sources continuously create relatively low but persistent chronic contamination across huge areas. Many aspects of the chemical composition and biological impacts of these contaminants remain unknown. Persistent Organic Pollutants Persistent organic pollutants (POPs) consist of a number of synthetic compounds, including industrial polychlorinated biphenyl’s (PCBs); polychlorinated dioxins and furans; and pesticides, such as DDT, chlordane, and heptachlor, that do not exist naturally in the environment. A number of POPs often persist in the environment and accumulate through the food chain or in the sediment to a toxic level that is directly harmful to aquatic organisms and humans. The marine environment collects contaminants from the air, but also from ocean currents, rivers discharging into the ocean, and sea ice that transports POPladen particles. Figure 1 presents some examples of contaminant levels in seawater. Hexachlorocyclohexane dominates the picture, except for Russian waters where PCB levels are high, up to 15 nanograms per liter in the Kara Sea. These high levels seem to mirror the high input of PCBs from Russian rivers. Levels in seawater can also be used to shed light on the mechanisms that transport contaminants to the Arctic. Detailed measurements in the Bering and Chukchi Seas show that hexachlorocyclohexane levels in the water increase along a south–north gradient, which has been suggested as evidence for a cold-condensation theory; those semivolatile contaminants condense out of the atmosphere as temperatures drop. Less volatile contaminants, such as PCBs, DDT, and chlordane, were present at lower levels in the Bering and Chukchi Seas than in more temperate latitudes. Concentrations of organic contaminants in Arctic marine sediments are, in general, extremely low compared with freshwater sediments, and are ten to a hundred times lower than in the Baltic Sea. The most apparent geographic trends are that concentrations of PCBs, hexachlorocyclohexane, and hexachlorobenzene are higher closer to the shore along the Norwegian coast than in the open sea. They are also higher in gulfs and river mouths along the Russian coast and around Svalbard. Heavy Metals Heavy metals exist naturally in the environment, and it is sometimes difficult to distinguish variations developing from anthropogenic inputs and those from the

100

COASTAL WATER POLLUTANTS Annual nitrogen load to the Baltic Sea 11%

Annual phosphorus load to the Baltic Sea 11%

12%

10%

Sweden

2% 6%

6%

Finland

Finland Baltic States*

12%

Sweden

8%

Baltic States*

18%

Poland

13%

Germany

Germany Denmark

5%

Denmark Atmosphere

Atmosphere

33% 18%

N2 Fixation *Belarus, Estonia Latvia, Lithuania

Source : Helcom 1993

Poland

35%

*Belarus, Estonia Latvia, Lithuania

Helcom. 1993. First assessment of the state of the coastal waters of the Baltic sea. Baltic sea Environ. Proc. No. 54, 160 pp.

Source : Helcom 1993

Figure 1. Nutrient release to coastal waters taking the Baltic Sea as an example.

natural hydrologic cycle and the atmosphere. Among the trace metals commonly monitored are cadmium, copper, mercury, lead, nickel, and zinc. When they accumulate through the food chain at moderate to high concentrations, some of these metals can affect the human nervous system. The marine environment receives heavy metals from atmospheric deposition, river runoff, and local pollution. The relative importance of these sources will differ between regions. For example, rivers carrying metal-laden sediments deposit almost their entire load in the shelf seas, and only a minor portion reaches the deep ocean. Natural sources of metals are important and, in many cases, it is found that they are the main source to the marine environment. Mining has contaminated ocean waters with several heavy metals. One documented example is in the fjord outside the Black Angel zinc mine in Greenland, where the levels of lead in the bottom water are up to 200 micrograms of lead per kilogram of water. These high lead levels are also reflected in seaweed, blue mussels, prawns, and in some fish; in capelin, lead levels are up to 5 micrograms per gram in the bone. However, no one has been able to document any biological effects in the fish. Cadmium levels in the water are also high, up to 2.5 micrograms per kilogram of water; but in contrast to lead, the animals in the fjord have cadmium levels close to background. The cryolite mine in Ivittuut in southern Greenland has also contaminated the nearby water. Lead levels of 18 micrograms per kilogram of water have been measured. At Strathcona Sound in northern Baffin Bay, a lead-zinc mine has released lead, making concentrations in the fjord water one to two orders of magnitude higher than background concentrations in the open ocean. Some of the lead has also been taken up by seaweed and crustaceans. Outside a lead-zinc mine in east Greenland, shorthorn sculpins also have elevated levels of lead, whereas the fish outside the cryolite mine on southern Greenland have not been affected.

The mines at Ivittuut and Strathcona Sound have also contaminated their respective fjords with cadmium, but the levels are much lower than those outside the Black Angel mine. At these sites, the cadmium is not affecting the local sediment, nor are elevated levels found in nearby plants and animals. Metal levels in Arctic Ocean water away from local sources are generally similar to global background levels. Today’s global lead concentrations in oceans are generally more than ten times higher than those in prehistoric times. The levels are consistently higher in surface waters than in deeper layers. One might expect the lead levels in the upper Arctic sediments to mirror this increased long-distance transport, but this does not seem to be the case. Recent seawater analyses from Pechora Bay and Kara and Laptev Seas show very high lead levels, ranging from 0.16 to 0.5 micrograms per kilogram of water. However, these data require confirmation before any conclusions are drawn. Filter-feeders such as mussels take up lead from sediment particles. The concentration increases slightly with increasing shell length, indicating a moderate accumulation as the mussel ages. However, lead levels are low in crustaceans as well as in fish. The highest levels, 0.05 micrograms per gram of liver, have been recorded in Orkdalsfjorden in Norway. Lead does not seem to accumulate in fish-eating birds or in marine mammals. In general, levels in marine mammals are low. Cadmium Levels are High in Marine Biota. Cadmium levels in seawater fall within what could be considered natural background levels. Moreover, there is no indication from sediments that the levels have increased from preindustrial times, nor have temporal trends been detected. An interesting phenomenon relating to cadmium is that its concentration increases farther away from the coast,

COASTAL WATER POLLUTANTS

101

Alpha-HCH in seawater, ng/liter 5 4 3 2 1 0 Chukchi sea Beaufort sea

Bering sea Okoisk sea

North pacific ocean

East china sea

Java sea

Figure 2. Alpha-hexachlorocyclohexane concentration in seawater increases from south to north, illustrating the cold-condensation effect (1).

which is probably connected with the change in salinity of the water. The result is that cadmium levels in both plants and animals are higher in the open ocean than in the inner region of large fjords, even when there are local sources contaminating the water. Cadmium accumulates with age in mussels and crustaceans. In general, the levels in crustaceans are higher than global background levels but show large variations. Mercury levels are high and may be increasing. Several sets of data indicate that mercury levels are higher in the upper layers of Arctic marine sediments than in the layers representing preindustrial inputs see the upper right diagram of Fig. 3. Mercury is enriched even in the marine sediments taken at the North Pole. Natural processes may have caused these profiles, but they could indicate that human activities have increased the environmental mobility of existing stores of mercury.

Radioactive Wastes These can somewhat be divided into high and low level wastes, depending on their activity. Radioactive wastes are characterized by losing their radioactivity with time. Some nuclides lose it quickly; others very slowly. A second consideration is that radioactive elements will enter the biological cycle and therefore the food web. High level radioactive wastes pose a complex problem in their disposal; the low level has been and is being disposed of directly in the sea. Reprocessing plants have added radionuclides to the sea. Spent nuclear fuel is often processed to recover plutonium. Water used in reprocessing contains a mix of different radionuclides, and some of this contaminated water has been released routinely into the sea. In Europe, three reprocessing plants are relevant to the Arctic because of transport of radionuclides by ocean

102

COASTAL WATER POLLUTANTS

0.000 0.0

Figure 3. Mercury concentration at different depths in marine sediment cores (1).

2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22.5 24 25 27.5 29 31 33.5 34 37.5

0.020

0.040

0.060

Hg mg/kg 0.080

0.100

0.120

0.140

0.160

North pole Alesund Eastern hudson bay Central west greenland

Eastern hudson bay Central west greenland

North pole

Alesund

Depth cm

Figure 4 indicates the rates of liquid discharges from 1952 to 1992. Cesium-137 dominates. The peak of the release for most radionuclides was in the mid-tolate 1970s.

TBq/y 6000 5000 4000 3000 2000 1000 0 1950 1955 1960 1965 1970 1975 1980 1985 1990 1995 106Ru

241Pu

90Sr

137Cs

AMAP Figure 4. Discharges of beta-emitters from the Sellafield nuclear reprocessing plant, 1012 bequerels per year.

currents: Sellafield (formerly Windscale) in Cumbria on the northwest coast of England; La Hague near Cherbourg, France; and Dounreay in northeast Scotland. Sellafield has been the most important source of radionuclides to the Arctic marine environment because of the scale of its discharge. The effluent has been released into the sea and carried north by ocean currents. The releases, which started in 1952, are well documented.

Underwater Weapon Tests Have Contaminated Chernaya Bay. Chernaya Bay is a fjord inlet, connected to the Barents Sea, on the southwestern coast of Novaya Zemlya. The former Soviet Union used the bay to conduct underwater tests of nuclear bombs in 1955 and 1957 and in the vicinity of the bay in 1961. As a result of these detonations, the bottom sediments of the bay are contaminated by elevated levels of radioactive plutonium and cesium, as well as other radioactive isotopes. However, the mobility of radionuclides in sediment is low and may at present cause only insignificant exposure of people. Exposure of biota is unknown. Today, the inventory of plutonium in Chernaya Bay is similar to other sites of major plutonium contamination, such as the most contaminated areas of Bylot Sound (where a B-52 bomber crashed) and the Irish Sea in the vicinity of the Sellafield reprocessing plant. Three underground nuclear detonations were carried out by the United States on Amchitka Island in the Aleutian Islands in 1965, 1969, and 1971. These detonations caused radioactive contamination of deep groundwater and rock around the shot cavities. Long-term monitoring activity is planned for this site to 2025. In 1996, aboveground radioactive contamination was detected at the site.

COASTAL WATER POLLUTANTS MSr

Bq/m 40

3

Greenland water Barents sea Kara sea

30

20

10

0 1960

1965

1970

1975

1980

1985

1990

1995

A major direct input of radionuclides into the marine environment has been from European nuclear reprocessing plants, particularly Sellafield on the shore of the Irish Sea. Currents transport the material along the Norwegian coast and into the Arctic Ocean. After 6 to 8 years, some of the contamination leaves the Arctic by way of the East Greenland Current, but much of it stays in the Arctic Basin much longer. Environmental radiocesium levels have been measured since the early 1970s. As can be seen in Fig. 6, the releases of cesium-137 from Sellafield are virtually mirrored in the levels found in the Barents Sea after a transport time of 4 to 5 years. The peak in concentration in the early 1980s is probably the highest level that has ever occurred in that area of the ocean. The Chernobyl accident in 1986 added cesium to the Arctic Ocean and continues to

Levels in seawater (Bq/m3) Releases from sellafield (TBq) 50 6000 East greenland current Barents sea Sellafield

40

4000 30 3000 20 2000 1000

do so via outflow from the Baltic Sea. Figure 6 shows the recent levels of cesium-137 in seawater around the Arctic. Strontium-90 has been measured in surface seawater collected around Greenland and the Barents Sea; see Fig. 5. During the past 35 years, levels in the waters around Greenland have decreased, approximately half was removed or decayed every 13.5 years. This value is probably representative of the Arctic Ocean as a whole. The highest levels of cesium-137 in people were recorded in the mid-1960s; see the Fig. 6. For the following 20 years, the human body burden decreased by a factor of 3 to 7. However, in 1986, the Chernobyl fallout changed the trend in areas directly affected by the accident. Assessment of Receiving Waters

Figure 5. Time trends of activity concentrations of strontium-90 in seawater.

5000

103

10

0 0 1950 1955 1960 1965 1970 1975 1980 1985 1990 1995 Figure 6. Releases of cesium-137 from Sellafield nuclear reprocessing plant (1012 bequerels per cubic meter).

The effects of the ocean waste disposal are the result of a complex relation between two variables, concentration and time. The effect of the oceanic environment on the effluent is of critical significance. After discharge to the ocean, the effluent experiences changes in its physical and microbiological properties, which vary, each point as a function of time. An accurate prediction of pollution conditions or environmental impact depends on knowledge of the oceanographic conditions in the area. These conditions vary in time and space. A significant and representative quantitative judgement requires refined statistical analysis. A statistical description of receiving coastal waters should be based on adequate observation of an entire area for a sufficiently long period of time (at least 1 year). The factors operating should be recorded simultaneously to provide a comprehensive picture of the physical and microbiological properties of the area. The data collected should enable probability distributions of the variables to be derived to select a coherent and suitable set of design parameters. Continuous and periodic records should be taken to cover all typical oceanographic conditions at stations strategically located at different sites and should record the following: (1) Currents (direction and speed) distributed in time/space to permit a comprehensive study of coastal water circulation patterns. (2) A continuous record of tides and winds in the area. (3) The density field obtained from time and space distribution of salinity and temperature enabling the stratification conditions of the coastal waters to be determined. (4) The waves recorded periodically, covering the most critical conditions and enabling a probability distribution of the wave characteristics (height, period, and direction) to be derived for structural design. (5) Periodic dispersion and diffusion experiments using dye tracers. (6) Simultaneously with the oceanographic survey a program on bacterial concentration decrease should be carried out, covering typical oceanographic conditions. The experiments should be made in situ,

104

COASTAL WATER POLLUTANTS

preferably in the existing continuous sewage field. Simultaneously sampling for bacteria and dye tracer direction will enable bacterial die-off rates and dilution to be investigated and the derivation of bacterial concentration decrease from both these effects. During concentration decrease experiments, the water temperature, solar radiation, and other climatic data should be collected. Mainly tides and winds, to a lesser controlled extent by the density field, govern coastal currents near the shore. A comprehensive study of coastal circulation requires a continuous record of winds and tides and simultaneous measurements of vertical profiles of temperature, salinity, and currents. A detailed vector statistical analysis of the current field is an important requirement. The seasonal or climatological associations between oceanographic and meteorological data and in situ bacterial assays are of paramount importance; their interactions determine major design parameters. Statistical analysis of the sewage field is required to predict the seasonal variations in initial dilution and final height of the rise. This analysis should be based on simulations of alternative diffuser discharges at various density stratification conditions. Oceanic Processes To assess the oceanic equivalent dilution factors, the following formulations may be used. Initial Dilution. The initial dilution, RI, and terminal rise height, Ym , were estimated, assuming that the diffuser is a finite line source of buoyancy flux only, by the equation RI = FI · Ym where FI = f (F, θ )U/q; in which f (f , θ ) is an experimental factor, which depends on θ and a type of Froude number F = u3/B. The current and density stratification are the oceanographic variables that directly interface in the initial dilution. We may consider a critical stratification condition in which this interface extends to a constant depth and the dilution layer increases linearly with the bottom slope according to a relation of the type Ym = a + b(x − x0 ), where x is the distance from a horizontal plume perpendicular to the shore. The subsequent dilutions RM, represent a minor part of the overall receiving water reductions. Thus the total physical dilution can be evaluated as RF = RI · RM, RF = FI(1 + FMx)[a + b(x − x0 )] All soluble pollutants experience this composite dilution factor. Small particulate matter will also disperse proportionately. Floatable matter may disperse to a lesser extent, remaining visually detectable and liable to be carried to shore by currents and surface winds. The removal of this material is required, and an initial dilution of 100 is required to provide sufficient emulsification for these materials.

Dilution Equivalent to Bacterial Concentration Decreases. Most experiments on bacterial concentration decrease have been shown to fit Chick’s law very satisfactorily. C(t) = C0 · 10 − t/T90, where C0 = coliform concentration at the origin, and T90 = time required for 90% reduction. Thus the dilution equivalent to the concentration decrease may be computed directly from RB = C0 /C(x) = 10 x/X90 where X90 = u · T90 = distance required for the coliform concentration reduction of 90%.

Dilution Equivalent to Sedimentation. The disappearance rate of coliforms due to sedimentation of solids depends on the degree of removal in the treatment plant. Data collected by experts revealed the fact that T90 values decrease as the effluent TSS increases. Therefore, sedimentation effects are already incorporated in the field experimental results for the concentration decrease. Dilution Equivalent to Treatment. The relation between removal rates and corresponding factor RT is RT = 100/[100 − T (%)]. Dilution Equivalent to Disinfection. The dilution equivalent to disinfection is evaluated by the equation RD = 1/rd, where rd is the bacteria reduction factor. Chlorination of less treated sewage produces organic chlorine compounds, which are toxic and deleterious to the environment. Therefore, chlorination may compromise the already recognized ecologically beneficial effects of enrichment of coastal waters by the supply of nutrients and organic matter from those effluents. Overall Equivalent Dilution. Assuming that RS = 1, the overall equivalent dilution can be evaluated by the equation Rtotal = RT · RD · RI · RM(x) · RB(x) Rtotal = RR · f (F, θ )(bu/q)[a + b(x − x0 )] · [1 + (KL/bo )x] · 10 x/u · T90 For a given set of parameters, Rtotal = CE/CP, Q and the oceanic parameters are u(u, θ ), T90 and KL are the total equivalent approximation, a linear function of the diffuser length and an exponential quadratic function of the outfall length.

The Effect of Current. Separating the current factors including the current speed u from the above equation, a dilution function may be defined as Fu = u · 10 x/uT90 , which represents the effect of current on overall oceanic dilution. The figure given shows curves of Fu as a function of u for the various values of the parameter u90 = x/T90 , that is, the velocity for a 90% concentration decrease. It may be seen that the curves show ‘‘inflections’’ that have been connected by the line. The inflection increases as u90 decreases or as T90 increases. This inflection divides the graph into two domains:

COASTAL WATER POLLUTANTS

105

The initial dilution domain (linear) The concentration decrease dilution factor dilution factor domain (exponential).

Depending on the type of impact on water organisms, communities, and ecosystems, the pollutants can be grouped in the following order of increasing hazard:

For strong currents, the initial dilution is the dominant factor, and for a given T90 , the only way to increase Fu is by initial dilution mixing depth, which can be attained by increasing the outfall length. For light currents, the dilution factor equivalent to die-off dominates. Thus, the current data is collected, as it is useful in the design in

• substances that cause mechanical impacts (suspensions, films, solid wastes) that damage the respiratory organs, digestive system, and receptive ability; • substances that provoke eutrophic effects (e.g., mineral compounds of nitrogen and phosphorus, and organic substances) that cause mass rapid growth of phytoplankton and disturb the balance, structure, and functions of water ecosystems; • substances that have saprogenic properties (sewage with a high content of easily decomposing organic matter) that cause oxygen deficiency followed by mass mortality of water organisms and appearance of specific microflora; • substances causing toxic effects (e.g., heavy metals, chlorinated hydrocarbons, dioxins, and furans) that damage the physiological processes and functions of reproduction, feeding, and respiration; • substances with mutagenic properties (e.g., benzo(a)pyrene and other polycyclic aromatic compounds, biphenyls, radionuclides) that cause carcinogenic, mutagenic, and teratogenic effects.

1. 2. 3. 4.

analysis of coastal circulation hydrodynamics; prediction of initial dilution; prediction of far field dilution and transport; Prediction of waste field impaction probability.

Thus, a comprehensive oceanographic study is necessary to describe the characteristics of the receiving waters. Therefore, an adequate program of oceanographic investigation in situ, including bacterial concentration decrease, is obligatory. The existence of elevated levels of contaminants in zones of high bioproductivity is extremely alarming ecologically. These zones include the water layer up to 100 m from the water surface (photic layer) and boundaries of natural environments (water–atmosphere and water–bottom sediment, as previously mentioned) as well as enclosed seas, estuaries, and coastal and shelf waters. In particular, in shelf and coastal zones, which occurs in 10% of the world ocean surface and less than 3% of its volume, the most intense processes of bioproduction, including self-reproduction of the main living resources of the sea, take place. The main press of anthropogenic impact is also concentrated here.

EFFECTS OF POLLUTANTS ON MARINE HABITAT To estimate the hazard of different pollutants, we should take into account not only their hazardous properties, but other factors, too. These include the volumes of their input into the environment, the ways and scale of their distribution, the patterns of their behavior in water ecosystems, their ability to accumulate in living organisms, the stability of their composition, and other properties, such as the extreme diversity of marine pollution components, the variety of their sources, the scales of distribution, and the degree of hazards. Pollutants can be classified in different ways, depending on their composition, toxicity, persistence, sources, volumes, and so on. To analyze large-scale pollution and its global effects, it is common to distinguish groups of the most widespread pollutants, which include chlorinated hydrocarbons, heavy metals, nutrients, oil hydrocarbons, surface-active substances, and artificial radionuclides. These substances form the so-called background contamination that exists now any place in the hydrosphere.

Some of these pollutants (especially chlorinated hydrocarbons) cause toxic and mutagenic effects. Others (decomposing organic substances) lead to eutrophic and saprogenic effects. Oil and oil products are a group of pollutants that have complex and diverse composition and various impacts on living organisms—from physical and physicochemical damage to carcinogenic effects. Discharge of heated waters can change the structure and function of coastal marine communities. Impacts of fly ash from coal-fired power plants, hot salty water, and residual chlorine are also important. Dumping of fly ash in coastal waters and into the atmosphere has caused severe impacts on spinner dolphins and mangroves in an area of the south coast of India, and has reportedly changed the number of species of plankton. Effects of Marine Oil Spills Oil spills can have a serious economic impact on coastal activities and on those who exploit the resources of the sea. In most cases, such damage is temporary and is caused primarily by the physical properties of oil creating nuisance and hazardous conditions. The impact on marine life is compounded by toxicity and tainting effects resulting from the chemical composition of oil, as well as by the diversity and variability of biological systems and their sensitivity to oil pollution. Biological Effects of Oil. Simply, the effects of oil on marine life are caused by either the physical nature of the oil (physical contamination and smothering) or by its chemical components (toxic effects and accumulation leading to tainting). Marine life may also be affected by cleanup operations or indirectly through physical damage to the habitats in which plants and animals live.

106

COASTAL WATER POLLUTANTS

The main threat posed to living resources by the persistent residues of spilled oils and water-in-oil emulsions (‘‘mousse’’) is physical smothering. The animals and plants most at risk are those that could come into contact with a contaminated sea surface: marine mammals and reptiles; birds that feed by diving or form flocks on the sea; marine life on shorelines; and animals and plants in mariculture facilities. The most toxic components in oil tend to be those lost rapidly through evaporation when oil is spilled. Because of this, lethal concentrations of toxic components leading to large-scale mortality of marine life are relatively rare, localized, and short-lived. Sublethal effects that impair the ability of individual marine organisms to reproduce, grow, feed, or perform other functions can be caused by prolonged exposure to a concentration of oil or oil components far lower than will cause death. Sedentary animals in shallow waters such as oysters, mussels, and clams that routinely filter large volumes of seawater to extract food are especially likely to accumulate oil components. Although these components may not cause any immediate harm, their presence may render such animals unpalatable if they are consumed by humans, due to the presence of an oily taste or smell, which is a temporary problem as the components that cause the taint are lost when normal conditions are restored. The ability of plants and animals to survive contamination by oil varies. The effects of an oil spill on a population or habitat must be viewed in relation to the stresses caused by other pollutants or by any exploitation of the resource. In view of the natural variability of animal and plant populations, it is usually extremely difficult to assess the effects of an oil spill and to determine when a habitat has recovered to its prespill state. In recognition of this problem, detailed prespill studies are sometimes undertaken to define the physical, chemical, and biological characteristics of a habitat and the pattern of natural variability. A more fruitful approach is to identify which specific resources of value might be affected by an oil spill and to restrict the study to meeting defined and realistic aims related to such resources. Impact of Oil on Specific Marine Habitats. Within each habitat, a wide range of environmental conditions prevails, and often there is no clear division between one habitat and another. Plankton is a term applied to floating plants and animals carried passively by water currents in the upper layers of the sea. Their sensitivity to oil pollution has been demonstrated experimentally. In the open sea, the rapid dilution of naturally dispersed oil and its soluble components, as well as the high natural mortality and patchy, irregular distribution of plankton, make significant effects unlikely. In coastal areas, some marine mammals and reptiles, such as turtles, may be particularly vulnerable to adverse effects from oil contamination because of their need to surface to breathe and to leave the water to breed. Adult fish that live in nearshore waters and juveniles in shallow

water nursery grounds may be at greater risk to exposure from dispersed or dissolved oil. The risk of surface oil slicks affecting the sea bed in offshore waters is minimal. However, restrictions on the use of dispersants may be necessary near spawning grounds or in some sheltered, nearshore waters where the dilution capacity is poor. The impact of oil on shorelines may be particularly great where large areas of rocks, sand, and mud are uncovered at low tide. The amenity value of beaches and rocky shores may require the use of rapid and effective cleanup techniques, which may not be compatible with the survival of plants and animals. Marsh vegetation shows greater sensitivity to fresh light crude or light refined products, although weathered oils cause relatively little damage. Oiling of the lower portion of plants and their root systems can be lethal, whereas even a severe coating on leaves may be of little consequence especially if it occurs outside the growing season. In tropical regions, mangrove forests are widely distributed and replace salt marshes on sheltered coasts and in estuaries. Mangrove trees have complex breathing roots above the surface of the organically rich and oxygen-depleted muds in which they live. Oil may block the openings of the air breathing roots of mangroves or interfere with the trees’ salt balance, causing leaves to drop and the trees to die. The root systems can be damaged by fresh oil that enters nearby animal burrows; the effect may persist for some time and inhibit recolonization by mangrove seedlings. Protection of wetlands, by responding to an oil spill at sea, should be a high priority because physical removal of oil from a marsh or from within a mangrove forest is extremely difficult. Living coral grows on the calcified remains of dead coral colonies, which form overhangs, crevices, and other irregularities, inhabited by a rich variety of fish and other animals. If the living coral is destroyed, the reef itself may be subject to wave erosion. The proportion of toxic components, the duration of oil exposure, as well as the degree of other stresses, largely determine the effects of oil on corals and their associated fauna. The waters over most reefs are shallow and turbulent, and few cleanup techniques can be recommended. Birds that congregate in large numbers on the sea or shorelines to breed, feed, or moult are particularly vulnerable to oil pollution. Although oil ingested by birds during preening may be lethal, the most common cause of death is from drowning, starvation, and loss of body heat following damage to plumage by oil. Impact of Oil on Fisheries and Mariculture. An oil spill can directly damage the boats and gear used for catching or cultivating marine species. Floating equipment and fixed traps extending above the sea surface are more likely to become contaminated by floating oil, whereas submerged nets, pots, lines, and bottom trawls are usually well protected, provided they are not lifted through an oily sea surface. Experience from major spills has shown that

COASTAL WATER POLLUTANTS

the possibility of long-term effects on wild fish stocks is remote because the normal overproduction of eggs provides a reservoir to compensate for any localized losses. Cultivated stocks are more at risk from an oil spill: Natural avoidance mechanisms may be prevented in the case of captive species, and oiling of cultivation equipment may provide a source for prolonged input of oil components and contamination of the organisms. The use of dispersants very close to mariculture facilities is ill advised because tainting by the chemical or by the dispersed oil droplets may result. An oil spill can cause loss of market confidence because the public may be unwilling to purchase marine products from the region, irrespective of whether the seafood is actually tainted. Bans on fishing and harvesting marine products may be imposed following a spill to maintain market confidence and to protect fishing gear and catches from contamination. Mercury levels in marine animals, including bivalves and crustaceans, are generally low, whereas mercury seems to accumulate in fish. The highest values in fish are from northern Canada. For seals and whales, concentrations often exceed 0.5 micrograms per gram of muscle, especially in older individuals. Livers from ringed seals in the western Canadian Arctic have very high levels of mercury; up to 205 micrograms per gram of liver have been measured. Levels in livers of bearded seals from the Amundsen Gulf are higher than those of both global background and other Arctic areas, as are mercury levels in toothed whales and polar bears. Some of the highest levels, 280 microgram per gram liver (wet weight), have been recorded in pilot whales from the Faroe Islands. The effects of these mercury levels on the animals are difficult to assess, because some of the mercury may be inactivated by high selenium levels. Moreover, the scientific focus so far has been on tissues relevant for human consumption, and very little information is available on the target organs for mercury, such as the brain. There are no effect studies from the Arctic. However, even for the most exposed animal populations in the western Canadian Arctic and in Greenland, selenium should be abundant enough to protect against mercury poisoning. Mercury is a major concern because the levels in some animals high in the food chain indicate that the environmental load may have increased in recent years. For example, mercury levels in ringed seals from western Canada show that they accumulated mercury about three times faster during the late 1980s and early 1990s than in the early 1970s. Similar increases have been seen in ringed seals from northwest Greenland taken in 1984 and 1994 and in beluga livers from the western Canadian Arctic. Interpreting these findings is difficult because natural variations that may affect the trends are unknown. Moreover, other data, such as those from Atlantic walrus and ringed seal from central-east Greenland, have not indicated any temporal trends. Very little information is available on temporal trends in Arctic marine fish, but measurements from the Baltic Sea from 1980 to

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1993 seem to confirm observations that mercury levels are increasing. The seas are being polluted by organic and inorganic wastes from sewage, from agricultural and industrial wastes, and from runoff containing oil, hydrocarbons, and heavy metals. All of these contribute to sediment runoff and increased turbidity. Siltation of coral regions is also caused by excessive deforestation and land clearing for commercial crops. Construction and land reclamation has caused changes in water circulation and has increased sedimentation. On the coral reefs, there has been extensive overexploitation of resources by heavy fishing pressure, including very destructive methods such as blasting, coral mining, and cyanide poisoning for live fish collection. Industrial Pollution Industrial pollutants that affect coral reefs include nutrients from sewage and organic matter, fertilizer runoff, detergents containing phosphorus, and thermal discharge—the heated water from the cooling systems of power plants and other industries. These all cause nutrient overload, the growth of aquatic plant life, and depletion of dissolved oxygen, or eutrophication, which retards coral growth by decreasing light penetration and changing the dynamics of fish assemblages. Other industrial pollutants include heavy metals and other toxic substances. The coral reefs bordering major cities throughout Southeast Asia have been largely destroyed. Pollution from oil refineries and drilling platforms, it has been shown, kill reef fish and have negative effects on growth rates, recruitment, and feeding of corals. Thermal pollution from hot water discharge from industrial areas is an additional threat to reef species, many of them cannot withstand sudden and drastic increases in temperature. Sedimentation. Land-based human activities often cause sedimentation, a major source of reef degradation. As more people move to coastal cities on the South China Sea, there has been a big increase in construction and land reclamation. Land reclamation and sedimentation have been particularly intensive in Singapore. Land was reclaimed by dumping sand and dirt directly onto coral reef flats and shallow water. These add to the erosion of beaches and sediment runoff that smothers corals and leads to the degradation of a reef. Increased sedimentation also leads to a change in the composition of marine fauna, favoring more resilient species. Sedimentation also comes from soil erosion from unsound agricultural practices, mismanagement of watersheds, exploitation of mangroves, land reclamation and construction, oil drilling, and dumping of terrestrial and marine mine tailings. Overfishing. Overfishing is a force extremely destructive to corals in the South China Sea. Densities of fish are greatly decreased by overfishing. Coral is damaged by destructive fishing techniques and by removal for trade. It is estimated that 10–15% of the total fish yield in the Philippines comes from coral reef fisheries.

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Fishing degrades a reef in several ways. Destructive and illegal fishing methods are common, especially in the Philippines, Indonesia, and Malaysia. These methods include dynamite blasting and cyanide fishing. Overfishing not only depletes fish stocks of target species but also changes the dynamics of the entire reef. Decreases in herbivores can lead to algal blooms that overtop coral growth and can cause mass mortality. Blooms of noxious algae have increased in the past 20 years worldwide and are being blamed on inputs of excess nutrients due to human activities. Some of these noxious algae produce powerful nerve toxins that can cause massive fish kills or even kill a person who unsuspectingly eats shellfish that was harvested from waters tainted with toxic algae. The case of the ‘‘Cell from Hell’’ now blooming in East Coast waters (North Carolina, Virginia, Maryland) is especially noteworthy. Until recently, Pfiesteria was only a curiosity of academic specialists. In the past few years, this organism has been blamed for fish kills unprecedented in their size and has been linked to neurological damage in people who worked or swam in these waters (memory loss, learning difficulties, and decreases in white blood cell content upward of 20% have been recorded in people who were exposed to Pfiesteria). Blooms of Pfiesteria have been linked to nutrient enrichment of coastal waters due to nonpoint pollution from agriculture. Nutrients in waters allow huge population increases of toxic organisms in water that were unknown or rare. The U.S. EPA has pledged to adopt new standards for nutrient inputs to waters. It is hard to imagine an organism more bizarre than Pfiesteria. When no fish prey are present, it goes into a cyst form and settles to the bottom, lying dormant in the sediments. It can also emerge to form an amoeba that feeds on algae in the water column, and even can become a photosynthetic plankton-like organism, except that it ‘‘steals’’ the chloroplasts of algae from its algal prey and uses photosynthesis only to supplement its nutrient supply in the water column. In the presence of certain species of fish, however, it becomes a ‘‘monster’’ predator capable of mass fish kills. As a ‘‘predatory’’ dinoflagellate, it produces different types of toxins that do an incredible array of damage to fish. Some toxins attack internal organs. Another works on the fish immune system. And one toxin actually strips the skin off of the fish. Those who have witnessed the power of Pfiesteria report thousands of fish flopping and thrashing on the water surface, and fish actually beaching themselves, fleeing the water as if on fire.

SUMMARY The introduction by man, directly or indirectly, of substances or energy into the marine environment (including estuaries) results in such deleterious effects as harm to living resources, hazards to human health, and hindrance to marine activities, including fishing, impairment of quality for use of seawater, and reduction of amenities. As the uses of coastal waters and the ocean have increased, pollution of the ocean waters has increased in turn. River pollution has also had an impact on the ocean as the rivers transport material to the

ocean and, as a result, make it the ultimate sink for the world’s waste. The following chart summarizes the sources of wastes and their effects.

Sources and Effects of Marine Pollution. Type Nutrients

Primary Source/Cause

Effect

Runoff approxiFeed algal blooms in mately 50% coastal waters. sewage, 50% from Decomposing algae forestry, farming, depletes water of and other land use. oxygen, killing Also airborne other marine life. nitrogen oxides Can spur algal from power plants, blooms (red tides), cars, etc. releasing toxins that can kill fish and poison people. Sediments Erosion from mining, Cloud water; impede forestry, farming, photosynthesis and other land-use; below surface coastal dredging waters. Clog gills of and mining. fish. Smother and bury coastal ecosystems. Carry toxins and excess nutrients. Pathogens Sewage, livestock. Contaminate coastal swimming areas and seafood, spreading cholera, typhoid, and other diseases. Alien Several thousand per Outcompete native species day transported in species and reduce ballast water; also biological diversity. spread through Introduce new canals linking marine diseases. bodies of water and Associated with fishery increased incidence enhancement of red tides and projects. other algal blooms. Problem in major ports. Persistent Industrial discharge; Poison or cause toxins wastewater disease in coastal (PCBs, discharge from marine life, heavy cities; pesticides especially near metals, from farms, major cities or DDT, etc.) forests, home use industry. etc.; seepage from Contaminate landfills. seafood. Fat-soluble toxins that bioaccumulate in predators can cause disease and reproductive failure.

TRACE ELEMENT POLLUTION

Oil

46% from cars, heavy Low level machinery, contamination can industry, other kill larvae and land-based sources; cause disease in 32% from oil marine life. Oil tanker operations slicks kill marine and other shipping; life, especially in 13% from accidents coastal habitats. at sea; also offshore Tar balls from oil drilling and coagulated oil litter natural seepage. beaches and coastal habitat. Oil pollution is down 60% from 1981. Plastics Fishing nets; cargo Discarded fishing and cruise ships; gear continues to beach litter; wastes catch fish. Other from plastics plastic debris industry and entangles marine landfills. life or is mistaken for food. Plastics litter beaches and coasts and may persist for 200 to 400 years. Radioactive Discarded nuclear Hot spots of substances submarine and radioactivity. Can military waste; enter food chain atmospheric and cause disease fallout; also in marine life. industrial wastes. Concentrate in top predators and shellfish, which are eaten by people. Thermal Cooling water from Kills off corals and power plants and other temperatureindustrial sites. sensitive sedentary species. Displaces other marine life.

READING LIST Brown, L. (1997). Can we raise grain yields fast enough? Worldwatch 10(4): 9–17. Engleman and LeRoy. (1996). Sustaining Water: Population and the Future of Renewable Water Supplies. Population Action International. EPA. (1994). National Water Quality Inventory, 1992: Report to Congress. Office of Wetlands, Oceans and Watersheds, US EPA, Washington, DC. Kane, H. (1997). Eco-farming in Fiji. Worldwatch 10(4): 29–34. Nixon, S. et al. The fate of nitrogen and phosphorus at the landsea margin of the North Atlantic Ocean. Biogeochemistry 35: 141–180. NRC (National Research Council). (1993). Managing Wastewater in Coastal Urban Areas. NRC, Washington, DC. NRC (National Research Council). (1994). Priorities for Coastal Science. NRC, Washington, DC. Hotta, K. and Dutton, I.M. (Eds.). (1995). Coastal Management in the Asia-Pacific Region: Issues and Approaches. Japan

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International Marine Science and Technology Federation, Tokyo. Web Sites http://www.nos.noaa.gov/icri/state.html http://www.cutter.com/osir/ http://www.igc.apc.org/wri/indictrs/reefrisk.htm http://www.enn.com http://www.env.gov.sg Journals and Newspapers Environet The Straits Times Science and Technology-The Hindu

TRACE ELEMENT POLLUTION BOBBY J. PRESLEY Texas A&M University College Station, Texas

INTRODUCTION All three words in the term ‘‘trace element pollution’’ need to be defined. Defining an element, that is a chemical element, is relatively straightforward. It is a substance consisting entirely of atoms having the same number of protons. All such substances are listed by name and symbol on the periodic charts of the elements found in many science textbooks. Definitions for ‘‘trace’’ and ‘‘pollution’’ are not so straightforward. It is sometimes convenient to classify the chemical elements making up a complex substance or matrix into major, minor, and trace, based on their relative amounts (concentrations). Most authors designate those elements present in a matrix at one part per million or less by weight as trace elements. Although trace elements are in low concentration in the environment, they can be either essential or harmful to organisms, depending on the element and the circumstances. A definition of pollution can be inferred from the widely accepted definition of marine pollution given by the United Nations Group of Experts on the Scientific Aspects of Marine Pollution (GESAMP). They say pollution is ‘‘the introduction by man, directly or indirectly, of substances or energy to the marine environment resulting in such deleterious effects as harm to living resources; hazards to human health; hindrance of marine activities including fishing; impairment of the quality for use of seawater; and reduction of amenities.’’ This definition can be applied to all environments, not just to the marine environment. By the GESAMP definition, pollution must be harmful and must be caused by human activity. In some cases, it is easy to show both a human cause and harm to the environment. In other cases, one or both parts of the definition can be hard to prove. For example, it might be possible to show that the concentration of a trace element, such as mercury, is elevated above normal values in, for example, fish. It might be harder to show harm to the fish or to the consumer of the fish and it might be hard to show

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that the enrichment is because of human activity. In such a case, it might be better to say contamination has been documented, but not pollution.

As a first step in determining the significance of a given amount of a trace element in a particular environment, it is essential to know the ranges in concentration expected to occur naturally in various media (soil, water, organisms, etc.). Most trace elements were tied up in igneous rocks on the primitive earth. Table 1 gives average concentrations of selected elements in the uppermost part of the earth’s crust along with data for organisms, soil, sediment, and water. These values can be considered to be ‘‘background’’ concentrations or at least concentrations not greatly influenced by human activity. As igneous rocks ‘‘weather’’ to give soil and sediment some fraction of each element in the rock, it becomes dissolved in rivers, lakes, and the ocean. For most trace elements, however, a large fraction is retained in the solids during weathering. For this reason, the natural (background) trace element concentration in soils and sediments varies, depending on the igneous rocks they came from. Many trace elements behave similarly during weathering, soil formation, erosion, and deposition. For example, fine-grained clay minerals become enriched in most trace elements and quartz sand and carbonate minerals become depleted. It is, therefore, important to have information on the grain size and mineralogy of soils and sediments when evaluating their trace element content. Thus, a given concentration, say 20 ppm Cu, might be background for a clay sediment but be contamination in a sand. Reliable data on trace elements dissolved in rivers, lakes, and seawater is more difficult to produce than is data on rocks, soil, and sediment. Concentrations are generally a thousand or more times lower (Table 1), and this causes sensitivity and matrix interference problems for even the newest analytical instruments. It is also difficult to avoid contaminating samples with trace elements during

ABUNDANCE, DISTRIBUTION, AND BEHAVIOR OF TRACE ELEMENTS Trace elements occur naturally in the environment. Unlike pesticides, plastics, organic solvents, and other manufactured products, the mere presence of trace elements does not imply pollution, contamination, or even human activity. Sufficiently sensitive analytical techniques can detect some amount of all elements in almost any substance. To determine the amount accurately and precisely is, however, a challenge, and to decide what portion of the element is natural and what portion is because of human activity is an even bigger challenge. In order to evaluate the significance of trace element occurrences in the environment, information is needed on: 1. Amounts (concentrations) in various compartments (air, soil, water, organisms, etc.). 2. Sources to the environment (both natural and human). 3. Transport mechanism and pathways between compartments (continents to oceans, water to organisms, etc.). 4. Transfer mechanism within compartments (shallow to deep water, fish gill surface to other organs, etc.). 5. Ultimate fate of the element (burial in sediments, mixing throughout the ocean, etc.). 6. Effect of the element on organisms (both short and long term).

Table 1. Average Concentration of Selected Elements in Various Materials. Values for Sediment, Soil, Crust, Oysters and Fish are µg/g Dry Weight (ppm) Except Ca and Fe in Percent Dry Weight. Values for Water are µg/l (ppb) Material Continental Crust Gulf Coast Seds Gulf Slope Seds San Joaquin Soil Sediment Criteria Average Seawater Miss. River Water Seawater Criteria Fresh Water Criteria Gulf Oysters Tuna Fish

Ref.

As

Ba

Ca

1 2 3 4 5 6 7 8 9 10 11

1.5 6 8.6 17.7 55 1.7

550 – 660

3.00 – 11.3 1.9 –

–

Cr

Cu

35 44 54 130 145 0.2

25 10 27 34.6 390 0.25 1.5 4.8 13 146 2.8

65 69 340 10.3 3.8

100

1100 16 0.57 0.75

Fe 3.5 1.8 2.76 – 0.05 2.2

294 72

Mn

Ni

Pb

V

Hg

Zn

600 330 300 538 – 0.2 1.4

20. 16 38 88 50 0.5 1.5 74 470 1.77 0.5

20. 15 17 18.9 110 0.002 0.008 210 65 0.64 0.5

60 – 100 112 – 1.5 1.0

– 0.050 0.028 1.4 1.3 0.001

71 60 81 106 270 0.4 0.27 90 120 2150 17.4

14.4 0.6

1.8 1.4 0.13 4.1

1. Average metal levels in upper continental crust (95% igneous rock). Taylor and McClennon (1). 2. Median estuarine (inshore) surface sediment metal levels from the U.S. Environmental Protection Agency’s northern Gulf of Mexico (GOM) Environmental Mapping and Assessment Program 1991–1993. 3. Average surface sediment metal levels observed among 43 stations ion the Gulf of Mexico Slope. BJ Presley, unpublished 4. Agricultural soil from the San Joaquin Valley, CA. US NIST Standard Reference Material #2709. 5. ER-M values from Long and Morgan (2), indicating sediment metal levels at which biological effects are often seen. 6. Average seawater values from Bruland, (3). 7. Average Mississippi River dissolved values from Shiller, (4). 8. Maximum Contaminant Concentration for seawater, US EPA, (5). 9. Maximum Contaminant Concentration for fresh water, US EPA, (5). 10. Average concentration in 485 pooled samples of 20 oysters each from the Gulf of Mexico. Presley (6). 11. Fillets from 16 Large Mediterranean Sea Tuna. International Atomic Energy Agency reference material # 350.

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collection, storage, and analysis. For these reasons, nearly all of the dissolved trace element data published before 1970, and much of the recent data, is unreliable. Much of the published data for both fresh and seawater is too high by factors of up to 100. Only data that has been produced by a lab using a well-documented quality assurance program should be accepted. Trace element concentrations in organisms are generally intermediate between those in sediments and those in water. However, concentrations vary widely with the specie of organism and with the specific organ within organisms. Livers, for example are enriched in some elements and kidneys in others and some species of organisms are highly enriched in one or another trace element. Trace element concentrations in organisms can also change with season, life stage, health, food supply, etc. Thus, identification of abnormal trace element concentrations is difficult unless reliable data is available for the same organism over a wide area and/or over some time period. Good general compilations of reliable trace element data for organisms are not as available as are those for soil and sediment. General guidance can be acquired from Furness and Rainbow (7) or similar publications. Many recent journal articles give trace element data for specific organisms from specific locations, but it is not always clear which are background levels. In addition to natural sources, many different human activities can add trace elements to the environment. Mining and metal processing are classic sources of contamination, but other manufacturing, transportation, and waste disposal practices can also be important. In the United States, the EPA’s ‘‘Superfund’’ program has spent many millions of dollars to clean up trace element contamination at dozens of sites around the country. The kinds of practices that led to this gross contamination are very rare today. Environmental regulations and public pressure have caused industry to greatly reduce trace metal releases to the environment. However, as the world population grows, the Earth’s surface is increasingly disturbed by agriculture, petroleum production, forestry, urban development, civil conflicts, and war, all of which make trace elements more available for uptake by humans and other organisms. As discussed above, numerous possible sources exist of trace elements to the environment. It is almost always difficult to determine which possible source is most important for any given element at any given location. Although it is important to identify the source of trace elements, their environmental impact depends not on source but on concentration and behavior. Behavior, including mobility, transport, transfer, and biological uptake, depends strongly on the chemical and physical form of the trace element. In this respect, the size of the trace element specie or the particle with which it is associated is critical, as this will control its transport and settling behavior in air and water. A given trace element will behave differently physically, chemically, and biologically in each of its different forms, and it will partition itself among the various possible forms in response to environmental conditions. It is important to note that many trace elements are ‘‘particle-reactive’’

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and will quickly associate with particles if added to the environment in a dissolved form. Trace element concentrations are almost always much higher in particles than in dissolved forms. Although the behavior of a trace element, including its biological behavior, depends on its form, there is some form of most trace elements that will affect the health of organisms, including humans. At least 20 trace elements have long been known to be essential to health [e.g., (8)]. Diseases because of trace element deficiencies are well known among both humans and other organisms. A number of trace elements such as Cu, Ni, Zn, and Se are essential to life at very low concentrations but toxic at slightly higher concentrations. Good data on trace element concentrations in the environment are needed in order to know whether too little or too much of a giver element is present. The toxic effects of both essential and nonessential trace elements are well known, in the case of As and Pb, human toxicity has been known for more than 2000 years. For other trace elements, toxic effects are less well recognized. In general, however, for all trace elements, an optimal concentration in the environment and in the organism gives optimal function (growth, reproduction, etc.) and higher or lower concentrations result in less than optimal function and possibly death. In order for a trace element in the environment to have an effect on an organism, the element must, of course, be taken up by the organism. For plankton and other aquatic plants, this uptake is directly from solution, but for animals, some, or most, of the uptake might be from food or from ingestion of nonfood particles. In any case, at some point the trace element must be in a soluble form and be transferred across cell membranes and possibly transferred to some vital organ within the organism. The form of the trace element is very important in controlling these transfers and the resulting effects, but both environmental conditions (pH, temperature, etc.) and the type of organism and its condition (age, health, etc.) also play a role. Factors that influence the toxicity of trace elements have been discussed by Bryan (9), Luoma (10), and many others, and the large differences in sensitivity to trace elements exhibited by different organisms are well known [e.g., (11)]. ASSESSING BIOLOGICAL IMPACTS The effect of a trace element on organisms depends on the abundance, distribution, and behavior of the trace element. As discussed above, these are difficult to determine and are subject to complex and incompletely understood processes. The environmental impact of waste disposal or other human activity is, therefore, often controversial. Environmental groups and industry often engage in public fingerpointing and lawsuits over specific activities. Often, more money is spent on lawyers than on attempts to scientifically document impacts. One reason for this is the difficulty in clearly documenting harmful effects in the field, especially at the population or ecosystem level. Laboratory toxicity testing is not easy, but it can usually show dose-response relationships that allow establishment of trace element concentrations above which harmful effects to a given organism are likely to

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result. Such results are, however, usually difficult to apply to the complex conditions in the field, as is discussed below. The simplest laboratory toxicity tests are those that use death of the organism as the only indicator of effect. This crude measurement has been much criticized, but it does establish the rough relative toxicity for various trace elements to various organisms. This test will show, for example, that Cu is much more toxic to most plankton than is As. More subtle effects can also be sought in laboratory cultures of various organisms, for example, changes in metabolism, ability to reproduce, find food, grow, etc. A vast amount of literature exists on methods for detecting sublethal effects of toxins on organisms [e.g., (2,12)]. Different sublethal tests often give different results in rating the relative toxicities of different trace elements, but they have the potential for indicating possible long-term effects on organisms that might not show up in short-term acute tests. Most laboratory toxicity tests use water as free from trace elements, complexing ligands, organic matter, etc. as is possible, so the response of the test organisms can be more clearly related to the trace element added in the test procedure. Consequently, a trace element is almost always less toxic in the environment than it is in the laboratory, because of complexing, adsorption, and other interactions in the environment. Laboratory toxicity tests, even when they try to imitate the environment by using ambient water, multiple trace elements, varying salinities, temperatures, different life stages of organisms, etc., can never truly duplicate natural conditions. It is useful, then, to look for effects of pollutants in the environment, especially at the population level. This is, however, a difficult task, because of the natural temporal and spatial variability in abundance and health of organisms. As a result of the relative expense and time involved in toxicity tests, and their sometimes ambiguous results, many environmental assessment programs seek only to determine concentrations of trace elements in the environment and to look for enrichments caused by human activity. If a trace element enrichment is detected, its significance can then be resolved by toxicity testing or detailed ecological field analysis. In any case, trace element enrichments could be sought in air, water, sediment, or organisms. Water analysis might seem a logical way to detect trace element enrichments in the environment. Furthermore, the significance of trace element concentrations in water can be judged because the US EPA has published values for each element above which harm to organisms is likely (Table 1). However, ambient water, be it river water, groundwater, rainwater, or seawater, is notoriously hard to collect, store, and analyze for trace element content, as was discussed above. Water concentrations can also change over short time scales in some circumstances, which further complicates their use. Soil or sediment can usually be more easily analyzed accurately for trace element content than can water. Soil and sediment also integrate trace element input over some time scale, so they don’t need to be sampled as often as does water. Another advantage of sediment analysis is

that it gives a historical view of pollutant input at sites where sediment is laid down layer by layer, year after year. Dates can be assigned to the different layers by use of radio-isotopes, pollen identification, or other means. Furthermore, sediment layers from prehistoric times give a background value for each trace element that can be compared with values in near-surface layers in order to quantify human-induced enrichments. Recognizing gross sediment contamination is easy. Any sediment that is several-fold higher in a given trace element than the average crustal abundance of that element is contaminated unless some unusual mineralogy exists. However, it is harder to recognize subtle contamination because of difficulties in establishing an exact background concentration for a given location. Values from prehistoric depths in the sediment column are a possible background, as noted above. Another background is sediment well away from any point source of pollutant input. In using either of these methods, care should be taken to compare similar sediment types or to compare element to element concentration ratios rather than absolute concentrations (13). Another problem with using sediment data is that only some unknown fraction of the trace element in the sediment is likely to be available to organisms, which has been much discussed in the literature [e.g., (14)], especially in conjunction with disposal of dredge spoil [e.g., (15)]. Many authors have suggested leaching sediments with dilute acids or other solutions [e.g., (14)] in order to remove only the trace element that could potentially be removed by an organism. Another suggestion that has been much debated is the ability of sulfide in the sediment to limit availability of trace elements to organisms [e.g., (16)]. Thus, although it generally agreed that only a fraction of the total trace element in sediments is available to organisms, no consensus exists on how to measure that fraction. Long and Morgan [(2) and elsewhere] suggest another way to identify sediment that is potentially harmful to organisms because of chemical contamination. They compiled published matching biological health and chemical data from numerous field, laboratory, and modeling studies. The data was then ranked from the lowest to the highest contaminant concentration where any adverse biological effect was reported. From the ranking they derived two guideline concentrations for each contaminant. These two values separate the data into values that (1) rarely, (2) occasionally, or (3) frequently cause adverse biological effects. These derived values have been widely used in monitoring programs. See Table 1 for some of the actual values. If both water and sediment offer analytical and data interpretation challenges, would it not be better to analyze organisms in order to assess trace element contamination? Certainly, advantages to this approach exist. For one thing, there is no question as to whether the element is available to organisms. For another, concentrations are often high enough to make analyses relatively easy, at least for common trace elements such as Cu and Zn. There are, however, problems, for example, deciding what organisms to analyze. It is not practical to analyze every organism at a given location, or even to analyze a

CORAL REEFS AND YOUR COASTAL WATERSHED

representative specie from each major taxonomic group. What, then, should be analyzed? Farrington (17) summarized the rationale for using common mussels (Mytilus sp.), various oyster species (Crassostrea and Ostrea), and other bivalves as ‘‘sentinel’’ organisms in monitoring studies in the marine environment. This approach has resulted in a very large worldwide data set for trace elements in bivalves. In the United States, the National Oceanic and Atmospheric Administration’s ‘‘National Status and Trends Program’’ (NS&T) has been analyzing bivalves from the entire U.S. coastline since 1986 and has produced an especially useful and highquality data set. As a result of the NS&T and similar data, bivalves should be the first choice for organisms to analyze in marine environmental monitoring programs. Many different kinds of plants and animals have been used in nonmarine environmental monitoring studies, everything from plankton and moss to polar bears. It all depends on what is available and the purpose of the monitoring. In general, organisms that have low natural variability in trace element concentration and are geographically widespread and easy to collect should be selected. Data from good, long-term environmental monitoring programs can help answer the question ‘‘are things getting better or worse.’’ Since strict environmental laws took effect in the United States in the 1970s, billions of dollars have been spent on pollution-control devices and cleanup of polluted sites. Have the efforts worked? (O’Connor (18,19) looked for temporal trends in the NS&T data discussed earlier. He found that for 2744 combinations of 14 chemicals and 196 collection sites over a 10-year time period, only 88 increases and 348 decreases in concentration are significant at the 95% confidence level. Chance alone predicts 69 increases and 69 decreases, so by this analysis, it is quite possible that no real increases occurred over that 10-year period and environmental quality along the U.S. coastline may have improved. This finding is consistent with observations of other environmental scientists. At least for chemical contaminants, environmental laws have worked and the U.S. environment is cleaner now than it was in the 1970s. BIBLIOGRAPHY 1. Taylor, S.R. and McClennon, S.M. (1985). The Continental Crust: Its Composition and Evolution. Blackwell Scientific Publications, Oxford, p. 312. 2. Long, E.R. and Morgan, L.G. (1990). The Potential for Biological Effects of Sediment-sorbed Contaminants Tested in the National Status and Trends Program. NOAA Tech. Memo. NOS OMA 52, Rockville, MD, p. 167. 3. Bruland, K.W. (1983). Trace elements in sea-water. In: Chemical Oceanography. Vol. 8, J.P. Riley and R. Chester (Eds.). Academic Press, London, pp. 158–220. 4. Shiller, A.M. (1997). Dissolved trace elements in the Mississippi River: Seasonal Interannual and decadal variability. Geochemica et Cosmochemica Acta 61: 4321–4330. 5. U.S. Environmental Protection Agency. (2002). National Recommended Water Quality Criteria: 2002. EPA-822-R-02047, US EPA Office of Water, Washington, DC. 6. Presley, B.J. (1990). Trace metals in Gulf of Mexico oysters. The Science of the total Environment 97/98: 551–593.

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7. Furness, R.W. and Rainbow, P.S. (1990). Heavy Metals in the Marine Environment. Lewis Publishers, p. 264. 8. Mertz, W. (1981). The essential trace elements. Science, V. 213: 1332–1338. 9. Bryan, G.W. (1976). Factors influenceing toxicity of heavy metals. In: Marine Pollution. R. Johnson (Ed.). Academic Press, London, pp. 185–302. 10. Luoma, S.N. (1983). Bioavailability of trace metals to aquatic organisms-A review. Science of the Total Environment 28: 1–29 11. U.S. Environmental Protection Agency. (1987). Biomonitoring to Achieve Control of Toxic Effluents. EPA/625/8-87/013, US EPA Office of Water, Washington, DC. 12. U.S. Environmental Protection Agency. (1988). Short-Term Methods For Estimating Chronic Toxicity of Effluents and Receiving Waters to Marine and Estuarine Organisms. Environmental Protection Agency, EPA 60014-87/028 Cincinnati, OH. 13. Trefry, J.H. and Presley, B.J. (1976). Heavy Metals in Sediments from San Antonio Bay and the Northwest Gulf of Mexico. Environmental Geology 1: 283–294. 14. Campbell, P.G.C. (1988). Biologically Available Metals in Sediments. National Research Council of Canada, ISSN 03160114, Ottawa, Canada, p. 298. 15. Lake, J., Hoffman, G.L., and Schimmel, S.C. (1985). Bioaccumulation of Contaminants from Black Rock Harbor Dredged Material by Mussels and Polychaetes. U.S. Army, Corp of Engineers, Tech. Report D-85-2, Vicksburg, MS. 16. Allen, H.E., Fu, G., and Deng, B. (1993). Analysis of acidvolatile sulfide (AVS) and simultaneously extracted metals (SEM) for estimation of potential toxicity in aquatic sediments. Environ. Toxicol. and Chem. 12: 1441–1453. 17. Farrington, J.W. (1983). Bivalves as sentinels of coastal chemical pollution; The mussel (oyster) watch. Oceanus 26: 18–29. 18. O’Connor, T.P. (1998). Mussel Watch results from 1986 to 1996. Mar. Poll. Bull. 37: 14–19 19. O’Connor, T.P. (1990). Recent Trends in Environmental Quality: Results of the First Five Years of the NOAA Mussel Watch Project. NOAA/ORCA, Silver Spring, MD.

CORAL REEFS AND YOUR COASTAL WATERSHED U.S. Environmental Protection Agency—Oceans and Coastal Protection Division

Coral reefs are among the world’s richest ecosystems, second only to tropical rain forests in plant and animal diversity. However, they are extremely sensitive environments that have special temperature, salinity, light, oxygen, and nutrient requirements. If environmental conditions fall outside the acceptable range of these requirements, the health and dynamics of a coral reef community can be severely disrupted. That’s why coral communities are sensitive indicators of water quality

This article is a US Government work and, as such, is in the public domain in the United States of America.

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and the ecological health of the coastal watershed. They respond to alterations within the entire coastal watershed, such as changes in freshwater flows and nutrient inputs. Consequently, pollution from the destruction and alteration of surrounding coastal watersheds can directly affect the health and productivity of a coral reef. WHAT ARE CORAL REEFS AND WHAT DO THEY DO? Coral reef ecosystems are unique, biologically diverse systems recognized as valuable economic and environmental resources. Many people think coral reefs are made of plants or rocks, but they are actually made of animals! A coral polyp is a delicate, limestone-secreting animal. The limestone serves as a skeleton that either is embedded within the living tissue of the coral or encloses the animal. A coral reef is made up of colonies of these coral polyps. There are several benefits of coral reefs. • Coral reefs are an important recreational and aesthetic resource for people visiting or living in coastal areas. People use coral reefs for fishing, underwater photography, scuba diving, and snorkeling. • Coral reefs provide protection for harbors and beaches, which are often found behind reefs because the reefs provide natural protection from heavy wave action caused by coastal storms. • Coral reefs are home to a number of species of fish and other marine species, including many that we rely on for food and economic purposes. • Coral reefs also serve as a laboratory for students and scientists to study and learn about complex ecological and biological processes. In addition, the reefs yield many biological treasures that are increasingly being recognized as natural sources of biomedical chemicals. SOME IMPACTS ON CORAL REEFS Coral reef habitats are extremely sensitive to disturbances, such as various forms of pollution and physical contact. Pollution of coastal watersheds poses a threat to

the existence of coral reefs. Impacts can result from activities occurring near the reef itself or from areas within the coastal watershed that drain to the reef. Disturbances and pollution can lead to diseases in coral such as bleaching (when the algae that give corals their color die). Natural occurrences, such as hurricanes, can adversely impact coral reefs through high-energy storm surges and the resulting resuspension of sediment. However, reefs are usually able to recover from natural disturbances. People using the reef can have an adverse impact on reef resources. Portions of a coral reef can be broken by the impact of boat anchors and boat groundings. Divers and snorkelers can harm the reef by simply touching it or by removing the corals. Suntan oil from swimmers and snorkelers can harm or even kill sensitive corals. Dragging hooks, fishing line, and nets across the coral reef, as well as placing and recovering lobster traps on reefs, can be damaging. Overfishing also harms coral reefs by removing important species that eat the algae growing on corals. When these fish species are removed, the algae overgrow the corals, smothering them. Marine debris, trash floating on the ocean or resting on the ocean floor, comes from many sources, including boaters, divers, improper disposal of trash on land, storm water runoff to rivers and streams, ships and other vessels, and offshore oil platforms. Marine debris can harm fish species and other aquatic organisms that use the reef. Trash that lands on the reef can kill corals by continually rubbing against it or smothering it. An excessive amount of nutrients from improperly treated sewage, atmospheric deposition, agricultural and urban runoff, and cleaning products high in phosphates can harm coral reef habitats. In excess levels, nutrients overstimulate the growth of aquatic plants and algae. When nutrient levels increase, the delicate balance that exists between corals and algae is destroyed and the algae can overgrow the corals. When this situation is prolonged, the corals are smothered and die beneath the algal carpet. This, in turn, affects the fish and other aquatic organisms using the area, leading to a decrease in animal and plant diversity and affecting use of the water for fishing and swimming. Some of the leading causes of nearshore coral decline can be related to land development and nearshore

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like industrial discharges, urban and agricultural runoff, mining activities, and runoff from landfills. Some toxic substances bind to sediment and are transported to coastal waters through sedimentation. These toxic substances can cause scarring, death, or reproductive failure in fish, shellfish, and other marine organisms. In addition, they can accumulate in fish tissue, leading to fish consumption advisories. The sensitivity of corals makes them especially vulnerable to the introduction of toxic substances. WHAT IS EPA DOING TO PROTECT CORAL REEFS?

International Coral Reef Initiative

construction that are not environmentally sensitive. Sediment, silt, and other suspended solids wash off plowed fields, construction and logging sites, urban areas, strip-mined land, and eroded stream banks when it rains. Increases in coastal sediment are also caused by construction of seawalls, docks, and marinas; land-clearing; boats running through shallow waters, disturbing and suspending silts with their propellers; and snorkelers and divers kicking up sediment. Sediment can block sunlight that is essential for the survival of some corals, which live in a very close relationship with microscopic plants (algae) that require sunlight to survive. In addition, heavy sedimentation can bury corals, inhibiting their growth or killing them. Pathogens are disease-causing microorganisms such as viruses, bacteria, and parasites. Pathogens are harmful to corals, causing disease and scarring in many species. These microorganisms enter water bodies from sources such as: inadequately treated sewage, storm water drains, septic systems, runoff from livestock pens, and boats that discharge sewage. Coral reefs are vulnerable to the introduction of a wide variety of toxic substances, including metals (such as mercury and lead), toxic organic chemicals (such as PCBs and dioxin), pesticides, and herbicides found in sources

In 1994, EPA, along with the State Department, the National Oceanic and Atmospheric Administration, and the Department of the Interior, formed an international coalition to coordinate information and bring higher visibility to the need for coral reef ecosystem preservation. The coalition became the International Coral Reef Initiative (ICRI), which now includes a membership of more than 90 countries. EPA’s Watershed Approach. EPA has joined with others to promote the Watershed Approach nationally as a means to further restore and maintain the physical, chemical, and biological quality of our nation’s waters, including coral reefs. By addressing issues on a watershed scale, those areas that pose the greatest risk to human and ecological health can be targeted, several pollutants can be addressed at one time, the public can be involved in cleaning up the environment and protecting coral habitats, and integrated solutions for environmental protection can be considered. This is particularly important given the contribution of activities and sources of pollution within the larger watershed to the decline of coral reefs. Through the Watershed Approach, integrated coastal zone management tools and watershed concepts can be applied in the development of comprehensive management and conservation plans. The Watershed Approach aims

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to determine protective approaches for controlling identified stressors to coral reef ecosystems. EPA’s Coastal Watershed Protection Strategy specifically provides technical assistance and support to priority coastal watersheds, such as National Estuary Programs (NEPs) and other coastal waters identified by states. Other EPA Programs. In assisting coastal states with the development of their Coastal Nonpoint Pollution Control Programs, EPA and other federal agencies developed guidance specifying management measures for sources of nonpoint pollution (diffuse runoff of pollutants) in coastal waters. In its program, a state or territory describes how it will implement nonpoint source pollution controls. EPA also works with other federal agencies to protect human health and coral reefs by reducing marine debris. The efforts include the establishment of the National Marine Debris Monitoring Program, which looks at the origins and amounts of marine debris deposited along U.S. coasts. EPA and the Coast Guard work together to regulate the transportation of municipal and commercial waste on vessels and to issue regulations for the manufacture, maintenance, and efficiency of marine sanitation devices (boat toilets), as well as the establishment of ‘‘no discharge zones’’ for vessel sewage. EPA also regulates the discharge of pollutants from facilities into sensitive marine waters. EPA assists states in the development of water quality standards designed to protect human health and aquatic life. This assistance includes the development of criteria for water quality that accurately reflects the most up-to-date scientific knowledge about the effects of pollutants on aquatic life, such as corals, and human health. What Can You Do to Help Protect Coral Reefs? You can do several things to help protect coral reefs and your coastal watershed:

• Be Informed and Involved. Learn about coral reefs and their importance to your coastal watershed. Participate in training or educational programs that focus on reef ecology. Be an informed consumer; ask the store owner or manager from what country the coral was taken and whether that country has a management plan to ensure that the harvest was legal and sustainable over time. Support the creation and maintenance of marine parks and reserves. Become a citizen volunteer. As a volunteer you might be involved in taking water quality measurements, tracking the progress of protection and restoration projects, or reporting special events like fish kills and storm damage. Volunteer for a reef cleanup or a beach cleanup. If you don’t live near a coast, get involved in your local watershed program. Report dumping or other illegal activities. • Take Responsibility for Your Own Backyard. Determine whether additional nutrients or pesticides are needed before you apply them, and look for alternatives to fertilizers and pesticides where the chance of runoff into surface waters might occur. Even if you live far from a coral reef ecosystem, these products might ultimately affect the waters that support coral. Consider selecting plants and grasses with low maintenance requirements. Water your lawn conservatively; the less water you use, the less runoff will eventually find its way into the oceans. • Practice Good Housekeeping. Learn about procedures for disposing of harmful household wastes so they do not end up in sewage treatment plants that can’t treat them or in landfills not designed to receive hazardous materials. Around the house, keep litter, pet waste, leaves, and grass clippings out of street gutters and storm drains to prevent their entrance into streams that might flow to reefs. Use the minimum amount of water needed when you wash your car to prevent waste and runoff. Never dump any household, automotive, or gardening wastes into a storm drain. They might end up on the reef. Take used motor oil, paints, and other hazardous household

SEA LEVEL AND CLIMATE

materials to proper collection sites such as approved service stations or designated landfills. Always follow label directions for the use and disposal of household chemicals. Keep your septic tank in good working order. The improper disposal of wastes and hazardous materials can lead to water quality problems and harm to the sensitive coral reef habitats. • Respect the Reef. Help keep the reef healthy by following local guidelines, recommendations, regulations, and customs. If you dive, don’t touch the coral. Keep your fins, gear, and hands away from the coral since this contact can hurt you and will damage the delicate coral animals. Stay off the bottom because stirred-up sediment can settle on corals and smother them. Avoid entering sensitive habitat areas with your boat or other motorized watercraft. Maintain your boat engine to prevent oil and gas leaks. Keep all waste produced during your excursions in a safe place to be disposed of properly when you’re back on land. If you go boating near a coral reef, don’t anchor your boat on the reef. Use mooring buoy systems if they are available. Maintain and use your marine sanitation devices properly. Conserve energy and keep your auto in good running condition. By conserving energy, harmful air emissions leading to air deposition are minimized.

SEA LEVEL AND CLIMATE RICHARD Z. POORE CHRISTOPHER TRACEY U.S. Geological Survey Reston, Virginia

RICHARD S. WILLIAMS, JR. U.S. Geological Survey Woods Hole, Massachusetts

INTRODUCTION Global sea level and the Earth’s climate are closely linked. The Earth’s climate has warmed about 1 ◦ C (1.8 ◦ F) during the last 100 years. As the climate has warmed following the end of a recent cold period known as the ‘‘Little Ice Age’’ in the 19th century, sea level has been rising about 1 to 2 millimeters per year due to the reduction in volume of ice caps, ice fields, and mountain glaciers in addition to the thermal expansion of ocean water. If present trends continue, including an increase in global temperatures caused by increased greenhouse-gas emissions, many of the world’s mountain glaciers will disappear. For example, at the current rate of melting, all glaciers will be gone from Glacier National Park, Montana, by the middle of the next century (Fig. 1). In Iceland, about 11 percent of the island is covered by glaciers (mostly ice caps). If warming continues, Iceland’s glaciers will decrease by 40 percent by 2100 and virtually disappear by 2200. This article is a US Government work and, as such, is in the public domain in the United States of America.

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Figure 1. Grinnell Glacier in Glacier National Park, Montana; photograph by Carl H. Key, USGS, in 1981. The glacier has been retreating rapidly since the early 1900’s. The arrows point to the former extent of the glacier in 1850, 1937, and 1968. Mountain glaciers are excellent monitors of climate change; the worldwide shrinkage of mountain glaciers is thought to be caused by a combination of a temperature increase from the Little Ice Age, which ended in the latter half of the 19th century, and increased greenhouse-gas emissions.

Most of the current global land ice mass is located in the Antarctic and Greenland ice sheets (Table 1). Complete melting of these ice sheets could lead to a sea-level rise of about 80 meters, whereas melting of all other glaciers could lead to a sea-level rise of only one-half meter. GLACIAL-INTERGLACIAL CYCLES Climate-related sea-level changes of the last century are very minor compared with the large changes in sea level that occur as climate oscillates between the cold and warm intervals that are part of the Earth’s natural cycle of long-term climate change. Table 1. Estimated Potential Maximum Sea-Level Rise from the Total Melting of Present-Day Glaciers. [Modified from Williams and Hall (1993). See also http://pubs.usgs.gov/factsheet/fs50-98/]

Location East Antarctic ice sheet West Antarctic ice sheet Antarctic Peninsula Greeland All other ice caps, ice fields, and valley glaciers Total

Volume (km3 )

Potential Sea-Level Rise (m)

26,039,200 3,262,000 227,100 2,620,000 180,000

64.80 8.06 .46 6.55 .45

32,328,300

80.32

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During cold-climate intervals, known as glacial epochs or ice ages, sea level falls because of a shift in the global hydrologic cycle: water is evaporated from the oceans and stored on the continents as large ice sheets and expanded ice caps, ice fields, and mountain glaciers. Global sea level was about 125 meters below today’s sea level at the last glacial maximum about 20,000 years ago (Fairbanks, 1989). As the climate warmed, sea level rose because the melting North American, Eurasian, South American, Greenland, and Antarctic ice sheets returned their stored water to the world’s oceans. During the warmest intervals, called interglacial epochs, sea level is at its highest. Today we are living in the most recent interglacial, an interval that started about 10,000 years ago and is called the Holocene Epoch by geologists. Sea levels during several previous interglacials were about 3 to as much as 20 meters higher than current sea level. The evidence comes from two different but complementary types of studies. One line of evidence is provided by old shoreline features (Fig. 2). Wave-cut terraces and beach deposits from regions as separate as the Caribbean and the North Slope of Alaska suggest higher sea levels during past interglacial times. A second line of evidence comes from sediments cored from below the existing Greenland and West Antarctic ice sheets. The fossils and chemical signals in the sediment cores indicate that both major ice sheets were greatly reduced from their current size or even completely melted one or more times in the recent geologic past. The precise timing and details of past sea-level history are still being debated, but there is clear evidence for past sea levels significantly higher than current sea level. POTENTIAL SEA-LEVEL CHANGES If Earth’s climate continues to warm, then the volume of present-day ice sheets will decrease. Melting of the current

Figure 3. Red shows areas along the Gulf Coast and East Coast of the United States that would be flooded by a 10-meter rise in sea level. Population figures for 1996 (U.S. Bureau of the Census, unpublished data, 1998) indicate that a 10-meter rise in sea level would flood approximately 25 percent of the Nation’s population.

Greenland ice sheet would result in a sea-level rise of about 6.5 meters; melting of the West Antarctic ice sheet would result in a sea-level rise of about 8 meters (Table 1). The West Antarctic ice sheet is especially vulnerable, because much of it is grounded below sea level. Small changes in global sea level or a rise in ocean temperatures could cause a breakup of the two buttressing ice shelves (Ronne/Filchner and Ross). The resulting surge of the West Antarctic ice sheet would lead to a rapid rise in global sea level. Reduction of the West Antarctic and Greenland ice sheets similar to past reductions would cause sea level to rise 10 or more meters. A sea-level rise of 10 meters would flood about 25 percent of the U.S. population, with the major impact being mostly on the people and infrastructures in the Gulf and East Coast States (Fig. 3). Researchers at the U.S. Geological Survey and elsewhere are investigating the magnitude and timing of sea-level changes during previous interglacial intervals. Better documentation and understanding of these past changes will improve our ability to estimate the potential for future large-scale changes in sea level. READING LIST Fairbanks, R.G. (1989). A 17,000-year glacio-eustatic sea level record; influence of glacial melting rates on the Younger Dryas event and deep-ocean circulation. Nature 342(6250): 637–642. Williams, R.S., and Hall, D.K. (1993). Glaciers, in Chapter on the cryo-sphere, in Gurney, R.J., Foster, J.L., and Parkinson, C.L., eds., Atlas of Earth Observations Related to Global Change. Cambridge University Press, Cambridge, UK, pp. 401–422.

THE PERMANENT SERVICE FOR MEAN SEA LEVEL Figure 2. Wave-cut terraces on San Clemente Island, California. Nearly horizontal surfaces, separated by step-like cliffs, were created during former intervals of high sea level; the highest terrace represents the oldest sea-level high stand. Because San Clemente Island is slowly rising, terraces cut during an interglacial continue to rise with the island during the following glacial interval. When sea level rises during the next interglacial, a new wave-cut terrace is eroded below the previous interglacial terrace. Geologists can calculate the height of the former high sea levels by knowing the tectonic uplift rate of the island. Photograph by Dan Muhs, USGS.

S. JEVREJEVA S. HOLGATE P.L. WOODWORTH Proudman Oceanographic Laboratory Birkenhead, United Kingdom

Mean sea level (MSL) is the average level of the sea, relative to the level of the land on which the measurements

THE PERMANENT SERVICE FOR MEAN SEA LEVEL

are being made, recorded over an extended period such as a month, year, or the lunar nodal period of 18.6 years by an instrument called a tide gauge (or coastal sea level recorder). MSL data are used in a wide range of scientific applications including studies into climate change, ocean circulation variability and geology, as well as in practical applications such as surveying and the establishment of national leveling datums. The Permanent Service for Mean Sea Level (PSMSL) is the global data bank for such MSL information, and it has since 1933 been responsible for the collection, publication, analysis, and interpretation of sea level data from the global network of tide gauges. It is based at the Proudman Oceanographic Laboratory (POL) in Liverpool, U.K. and is a member of the Federation of Astronomical and Geophysical Data Analysis Services (FAGS) established by the International Council for Science (ICSU). The PSMSL is supported by FAGS, the Intergovernmental Oceanographic Commission of the United Nations Educational, Scientific and Cultural Organisation (IOC/UNESCO), and NERC. The database of the PSMSL contains almost 53,000 station-years of monthly and annual mean values of sea level from nearly 2000 tide gauge stations around the world received from almost 200 national authorities (see Fig. 1). On average, approximately 2000 stationyears of data are entered into the database each year. All data are readily available from the PSMSL website: www.pol.ac.uk/psmsl. Data for all stations are included in the PSMSL METRIC (or total) dataset. The METRIC monthly and annual means for any one station-year are necessarily required to be measured to a common datum, although, at this stage, datum continuity between years is not essential. The year-to-year datum checks become essential, however, if the data are subsequently to be

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included in the PSMSL ‘Revised Local Reference (RLR)’ component of the dataset. The RLR dataset contains records for which time series analysis of sea level changes can be performed. Long records from this dataset have been the basis of all analyses of secular changes in global sea level during the last century. The geographical distribution of longer RLR records contains significant geographical bias toward the Northern Hemisphere, a situation that is being rectified by means of international collaboration. Aside from its central role of operation of the global sea level data bank, the PSMSL has a responsibility as a member of FAGS to provide the sea level community with as full a Service as possible with regard to the acquisition, analysis, and interpretation of sea level data. Consequently, the PSMSL provides a range of advice to tide gauge operators and data analysts. It has occupied a central planning and management role in the development of the Global Sea Level Observing System (GLOSS) of the IOC. Through GLOSS and via other routes, the PSMSL provides advice and training to national sea level authorities and individual sea level scientists and technologists. In addition to the provision of training materials (e.g., tide gauge operation manuals), the PSMSL supplies software packages suitable for tidal data analysis and quality control purposes. In addition to the training courses associated with GLOSS, the PSMSL has every few years hosted important study groups and international conferences on sea level science. The study groups have concerned themselves with topics such as the use of global positioning system (GPS) receivers at tide gauge sites to determine the local rates of vertical land movement and have been held under the auspices of the International Association for the Physical Sciences of the Ocean (IAPSO) Commission on Mean Sea Level and Tides (CMSLT), the scientific body to which

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the PSMSL reports formally. The PSMSL hosted a major Symposium in Vienna in 1991 as part of the International Union of Geodesy and Geophysics (IUGG) Congress, an international conference at the Linnean Society in London in 1993 as part of its 60th anniversary celebrations, co-organized ‘‘Tidal Science 96’’ at the Royal Society in London in 1996, and took a major part in ‘‘A Celebration of UK Sea Level Science’’ at the Royal Society in 2004. The proceedings of each of these conferences have since been published. A further major conference is planned in 2008 for the PSMSL 75th anniversary. Probably the most important recent scientific publications with which the PSMSL is associated are those of the First (1990), Second (1995), and Third (2000) Scientific Assessments of Intergovernmental Panel on Climate Change (IPCC). The PSMSL Director has been a lead author for the sea level chapters in each of the IPCC studies. Major conclusions have been that global sea level has indeed risen by approximately 10–20 cm during the past century and may rise by amounts several times larger during the next 100 years. The PSMSL is conscious that developments in technology have expanded the field of sea and land level studies. During the 1990s, satellite radar altimetry and GPS recording become established techniques, whereas space gravity offers the potential for being an effective source of sea-level-related information in the future. Therefore, the PSMSL maintains full participation with altimeter and space gravity working groups in view of the importance of those techniques to sea level research. PSMSL personnel have Principal Investigator status for the TOPEX/Poseidon, Jason, ERS, and Envisat altimeter missions, in addition to the GRACE and GOCE space gravity projects. The major challenge for the future, to which the PSMSL is committed, is to see the established tide gauge and new space-based techniques closely linked within one coherent global sea level monitoring system. READING LIST Woodworth, P.L. and Player, R. (2003). The Permanent Service for Mean Sea Level: an update to the 21st century. J. Coastal Res. 19: 287–295. Woodworth, P.L., Aarup, T., Merrifield, M., Mitchum, G.T., and Le Provost, C. (2003). Measuring progress of the Global Sea Level Observing System. EOS, Trans. Am. Geophys. Union. 84(50): 565.

MARINE AND ESTUARINE MICROALGAL SEDIMENT TOXICITY TESTS IGNATIO MORENO-GARRIDO ´ M. LUBIAN ´ LUIS ´ JULIAN BLASCO Institute of Marine Sciences of Analucia Cadiz, Spain

The term microphytobentos refers to the microscopic algae that live on the submerged (temporary or not)

floor in fresh water, estuarine, or marine environments. Microphytobenthos are mainly composed of mobile, pennate diatoms and cyanophytes. In marine and estuarine environments, those organisms can be found in habitats such as salt marshes, submerged vegetation beds, intertidal (sand or mud) flats, or subtidal sediments where light permits microalgal growth. As the presence of these photosynthetic organisms is not always evident, MacIntyre et al. (1) borrowed the term ‘‘secret garden’’ from the homonymous book published in 1888 and written by Frances Hodgson Burnett (1849–1874), in order to make a literary allusion to microphytobenthos. In fact, habitats where microphytobentos are the only primary producers are recognized as ‘‘unvegetated’’ areas, but the concentration of chlorophyll a in the upper 0.5 cm of the sediments where those organisms live generally exceed the depth-integrated chlorophyll in the entire overlying water column (2). In some cases, chlorophyll from microphytobenthos can be up to six orders of magnitude higher than that for the overlying water (3). Nevertheless, some authors estimate that primary production in the bottom would be lower than in the free plankton in spite of that exceptional data (4). But in some habitats, biomass from benthic microalgae can match or even exceed biomass of bacteria present in the same space (5). Thus, microphytobenthos necessarily play a key role in the benthic trophic webs (6–8). In certain biocenosis, the organisms of the microphytobenthos are the main—sometimes the only—source of carbon for grazers or bacteria (7,9,10). The microphytobenthos are also very important in relation to the stability of coastal and estuarine sediments. Although some cyanobacteria can show hydrophobicity as a mechanism to attach sediment particles or other cells (11), the main strategy of microphytobenthos to keep attached to the sediment is the production of agglutinant molecules [carbohydrates (CH) or exopolysaccharides (EPS)] (12–15). The presence of these molecules has important trophic implications, but their role in the maintenance of the structure of the upper part of the sediment is also important (16–19), because the film of adherent substances produced by microphytobenthos increases the sediment stability. Toxicants such as herbicides can alter microphytobenthic growth and the production of CH and EPS can be diminished. A loss of sediment stability can induce an increase in the turbidity of the water and a higher rate of deposition of fine particles on submerged higher plants or macroalgae, thus reducing even more the primary production of the whole and adjacent systems (20). The importance of sediments in accumulation of xenobiotics in coastal and estuarine environments has been pointed out. Most chemical contaminants entering marine or estuarine environments eventually accumulate in sediments due to different reasons, including higher salinity values that diminish solubility of such substances in water. Sediment can act as a sink for these substances but also as a subsequent source for the same (25). On the other hand, and except in cases of extreme contamination, chemical data by themselves do not predict hazard (21,22). Thus, bioassays are needed to assess the potential toxicity of sediments.

MARINE AND ESTUARINE MICROALGAL SEDIMENT TOXICITY TESTS

In spite of the importance of microphytobenthos, few efforts have been made to develop standard toxicity test on these organisms. Guidelines from the SETAC (23) and recent revisions (24–27) about toxicity testing on benthic organisms offer good information on macro and meiofauna but completely ignore microphytobenthos. The reason for this ‘‘exclusion’’ cannot be found in the lack of importance of benthic microalgae, because of all the reasons expressed above. Probably the most important reasons for the scarcity of work on microphytobenthos ecotoxicology are the difficulties that this biological material cause. First, it is not easy to efficiently remove microphytobenthos from sediments. Size and weight of microalgae on the sediment match part of the sediment particles mixed with them. There are descriptions of techniques that use the migration capacity of the microphytobenthic organisms in order to remove them from sediment. It is well known that microphytobenthos vertically migrate through the sediment as a function of the light and tide conditions. During low tide and light conditions, cells unbury and remain at the surface of the sediment, but during high tide or night conditions, microphytobenthic cells bury themselves again, in order to avoid being removed by the current or waves or grazed during nonphotosynthetic conditions (1,28,29). This vertical movement can be exploited to make cells migrate through a plankton net separated from the sediment by one or more lens tissue papers (8). Other authors improved on this method by covering the plankton net with a few millimeters of silica powder, where

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living mobile cells accumulate (30). Less effective seems to be the technique that uses the adherent capacity of microphytobenthos for removing cells by the disposition of a cover glass on the sediment (19). It is supposed that cells will attach to the cover glass, but it is not easy to ensure the efficiency and repeatability of the technique. Another approach is to take cores (made of Plexiglas, PVC, or other materials) (31) from the upper sediment and resuspend subsamples of the previously sliced sediment in order to directly count (and taxonomically identify) cells by light microscopy (4). This latter method ensures the integration of all species (motile or not). The use of fluorescence microscopy can help in distinguishing photosynthetic cells from debris: with a blue filter and a barrier filter of 530 nm, chlorophyll emits a bright red fluorescence that clearly reveals cells and facilitates their localization and count, something that is difficult if this technique is not used (32,33) (Fig. 1). Other techniques use density differences to separate diatoms from debris by centrifugation in a Percoll gradient (34). Although this technique seems to be good for isolating cells, the percentage of cell recovery is low (near 5% of total population in natural locations). Due to the difficulties of handling cells among the sediment, several works were limited to analyzing photosynthetic pigments or photosynthesis in the upper sediments (35–37) as biomarkers for the microphytobenthic biomass. The analysis of degraded pigment molecules in sediments can be useful as a biomarker for grazing pressure (38).

Figure 1. Cells of the diatom Cylindrotheca closterium among sediment particles, observed under fluorescence microscopy. Using a barrier filter of 530 nm, the two chloroplasts of each cell are bright red, facilitating the counting and location of the algal cells.

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But difficulties in handling microphytobenthos do not end there. The disposition in ‘‘patches’’ of microphytobenthos in field locations is evident, providing a spatial heterogeneous distribution of the cells (39,40). This spatial distribution is sometimes conditioned by the presence of ridges and runnels (41) in the sediment surface, provoked by the natural dynamics of the ecosystem. This must be taken into account when estimations of microphytobenthic organism density are intended to be developed in actual locations (15). Delgado (4) described spatial heterogeneity in the delta of the Ebro River as being insignificant over distances between 10 cm and 10 m. Another consideration is that in emerged (low tide) conditions, there is a process of gradual compaction of sediment due to dewatering, which implies a higher density but a lower content of total pigments, exopolysaccharides, or individuals (42). In spite of all this, in situ or in vitro bioassays involving microphytobenthos should be considered as powerful tools to determine potential toxicity of sediments (24). Some efforts have been made in this direction. Wong et al. (43) described a microalgal toxicity test on sediments from the coast of Hong Kong, but they used a planktonic (and not benthic) species (the chlorophyte Dunaliella salina) that, additionally, belongs to a genus that demonstrates strong resistance to toxicants. Abalde et al. (44) did not find growth inhibition of populations of D. tertiolecta at levels of 8 mg · L−1 Cu, and Moreno-Garrido et al. (45) found that D. salina was the most resistant species to Cu and Cd among the assayed microalgal strains. Each microalgal species shows different sensitivity for toxicants, but there are references showing similar (on the same order of magnitude) sensitivities for very different taxons (46). In this respect, Isochrysis galbana, Cylindrotheca (Nitzschia) closterium, and Nannochloropsis species showed similar responses to water soluble fraction of petroleum. Phytoplankton (free swimming or floating microalgae), periphyton (microalgae growing on solid substrate), and epipsammon (microalgae growing more or less attached to sand) showed comparable toxicity sensitivity responses to paraquat and simazine (47). Other works describe sediment toxicity tests on elutriates or extracts (48). Tolun et al. (49) described experiments where natural sediment toxicity tests on Phaeodactylum tricornutum, based on elutriates and bulk exposure, are compared: in the case of elutriates, the authors found different degrees of growth inhibition, although for direct exposure to sediments all organisms died, supporting the idea that direct exposure to sediment will give more realistic (or more sensitive) responses because part of the toxicants could not be extracted in elutriates or extracts (50). A very interesting approach to sediment toxicity tests on benthic algae was that made by Dahl and Blank (31), in which epipsammic communities were transported to a laboratory where they were kept and used in subsequent measurements of metabolic activities and short-term toxicity tests. Cairns et al. (51) defended the use of microorganisms in toxicity tests, because they can show very high sensitivities to toxicants and thus should be included in regulatory-proposed guidelines. But incorporation in those guidelines of multispecies tests that could be more ‘‘environmentally realistic’’ are very slow, fundamentally

because of methodological questions (regarding replication and reproducibility). On the other hand, predictions based on multispecies tests are no more accurate than those based on monospecific bioassays, which are cheaper and more reproducible than a multispecies bioassay (51). As far as we know, the first attempts to develop a standardized, repeatable protocol for sediment toxicity testing involving a microphytobenthic strain and direct exposure of microalgal cells to sediment have only recently taken place (32,33). In those works, populations of the benthic diatom Cylindrotheca closterium were exposed to sediment spiked with heavy metals or tensides. The test also considers the effect of particle size distribution on growth of the tested microalgal strain, which could mask actual responses of algae to present toxicants in other experiments such as those described by Tolun et al. (49). The test is simple, repeatable, and cheap, and it does not require special facilities other than those found in any laboratory. It is based on the 72-h algal growth inhibition test from OECD (52), adapted to sediments and marine or estuarine habitats. Cylindrotheca closterium, formerly known as Nitzschia closterium, demonstrated to be a good subject in other toxicity bioassays and there is a good pool of previous information about this species (53–58). This species is cosmopolitan for temperate coastal waters, ubiquitous, easy to handle in the laboratory, fast growing, sensitive to toxicants, and presents very low nutrient requirements. Nevertheless, other species have been assayed and compared with C. closterium in order to detect toxicity in natural sediments. EC50 values were calculated for three benthic diatoms exposed to sediment obtained from six different locations from Aveiro Lagoon (Portugal). Those values are shown in Table 1. Locations I, II, and VI were less toxic to microalgae than locations III, IV, and V, since C. closterium is slightly more sensitive than the other two species assayed for the majority of the samples. When a chemical analysis of the samples was performed and a similarity analysis carried out crossing toxicity values with possible substances involved (heavy metals, C, N, PCBs), it was found that some heavy metals (Sn, Zn, Hg, Cu, and Cr) had a great effect (Sn the greatest), as shown by the more than 50% of similarity between samples that showed significant growth inhibition for the benthic diatoms assayed (unpublished data). Other recent work from Adams and Stauber (59) also describes a whole-sediment toxicity test on a benthic microalgal diatom (Entomoneis cf. punctulata). In this case, a flow cytometer is used to detect viability of living algae by the use of a fluorochrome (fluorescein diacetate,

Table 1. EC50 Values for Sediments from Six Locations at the Aveiro Lagoon (Portugal) and Three Benthic Diatoms Sample Sites Algal Species P. tricornutum C. closterium Navicula sp.

I

II

III

IV

V

VI

N.I.a 92 N.I.

N.I. 90 N.I.

27 62 64

51 31 71

62 61 65

N.I N.I. N.I.

a If 100% of the sediment does not exhibit a growth inhibition value of 50%, EC50 is denoted as N.I. (no inhibition).

MARINE AND ESTUARINE MICROALGAL SEDIMENT TOXICITY TESTS

FDA). In this technique, used previously (60) to determine the toxic response in planktonic microalgal cells, FDA absorbed by living or dead cells is only hydrolyzed by nonspecific esterases inside the living cells. Hydrolyzed FDA is converted into fluorescein and can be detected by flow cytometry techniques. Although the FDA technique is less sensitive than growth inhibition measurements (60), flow cytometry opens a wide field of possibilities for microalgal toxicity testing on sediments, because several different fluorochromes can be used to measure quite different cellular parameters used as biomarkers.

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MARINE STOCK ENHANCEMENT TECHNIQUES JANA DAVIS Smithsonian Environmental Research Center Edgewater, Maryland

The United Nations estimates that 28% of all ocean fishery stocks are overexploited or severely depleted, and another 47% is fully exploited (1). Included in this statistic are species that are long-lived and short-lived, open-ocean and coastal, migratory and sedentary. About 75% of fishery

MARINE STOCK ENHANCEMENT TECHNIQUES

stocks are unable to withstand further exploitation, so only 25% of stocks remain to satisfy the increasing human demand, as human population grows, for fishery products. As a result of the pressure on fishery populations and the degradation of many of these resources, a suite of fishery management techniques have been developed. One such technique, stock enhancement, is broadly defined as the (human) activity by which the population of a species is increased. In contrast, more traditional management techniques generally focus on limiting human activities that reduce the population of a species. Traditional measures may focus on reducing fishing effort, for example, by limiting the number of days fishermen can fish, restricting the amount of gear a fishermen can use, regulating the type of gear a fisherman can deploy, or setting restrictions on the types of individuals caught by size or gender. Not only does this latter method serve to reduce catch, it strives to shift the pressure away from more reproductively important individuals toward those with less likely contributions to the next generation over the course of their lifetimes. GOALS OF STOCK ENHANCEMENT: A NONTRADITIONAL MANAGEMENT TECHNIQUE The technique of stock enhancement, while often used in concert with more traditional catch-reducing techniques, does not describe limitations to human activities but rather the active improvement of a fishery stock. Enhancement efforts can encompass many approaches. The population can be bolstered both by increasing carrying capacity (the number of individuals that can be supported by their habitat) or by increasing the number of individuals themselves (2,3). The former method is effective only when the factor limiting population is related to habitat, food, or some other resource. The latter method, a special case of aquaculture, is effective only when the habitat can hold more individuals than it does at present. The key to success in stock enhancement is to identify the factor(s) that limit the population in question and then increase the level of that factor. Incorrect identification of this factor could lead to wasted time, money, and energy. Consider the hypothetical case of a coral reef fish. If the fish population is not habitat-limited, but is recruitment-limited, adding more coral reef habitat would likely serve only to spread out the population spatially. No additional recruits would be available to settle on the new habitat substrate. On the other hand, if the fish is habitatlimited, adding additional individuals will not increase the ultimate population, as there would be no space for them to occupy. Enhancement directed toward habitat-limited populations has included efforts to enhance stocks of both fish, especially salmonids, and invertebrates (4–7). Often these cases are referred to as ‘‘habitat restoration,’’ stressing the action rather than the consequence. Most often when people refer to ‘‘stock enhancement,’’ they are referring to hatchery-raised individuals, usually juveniles, that are released into the wild to bolster recruitment-limited populations directly. These added juveniles are raised outside

125

of the system, for example, in, an aquaculture facility or a field pen. The juveniles may be offspring of wild parents collected from the field (8) or offspring from generations of parents held in captivity (9). The goal of stock enhancement is to raise the individuals beyond the initial phase of high early lifehistory mortality and then release them into the wild. The stage at which they are released is determined through economic optimization, survivorship maximization, or a combination of both. Maintaining individuals, especially heterotrophs, in a hatchery or a pen is expensive. The expense tends to increase as organisms grow. One might calculate the optimum release size or age as a trade-off between the cost of maintaining an individual in the hatchery and the survivorship advantage that the hatchery offers. Often, as individuals grow, the survivorship advantage of the hatchery environment over the natural environment begins to reverse. Many organisms are cannibalistic when held in extremely high densities (10), a trend that increases as organisms grow. Disease transmission also becomes a problem at high densities. Lowering densities requires an expensive solution of creating more tank or pen space. Therefore, at a certain point, mortality rates are actually lower if the organisms are released into the wild than if held in the hatchery. The optimum release point is different for each species. For example, Kemp’s ridley sea turtles are released at the age of 1 year (11). Blue crabs are released at the age of 2 to 3 months (12). Individuals that are never released but harvested in the hatchery are part of a complete aquaculture program such as, for example, farm-raised salmon, catfish, or shrimp (13,14). In populations truly at risk, the goal of stock enhancement is not simply to provide more fish to catch. Ideally, after release, these hatchery-reared individuals contribute to the spawning stock and have a potential exponential impact on overall population generation after generation, depending on the degree of recruitment limitation. In these cases, a purely economic model to predict optimum size at release is not appropriate. Stock enhancement therefore has a potential role beyond fishery application and into the realm of threatened or endangered species protection. STOCK ENHANCEMENT EXAMPLES Stock enhancement techniques have been applied to many types of organisms worldwide. Finfish populations have been the most common recipients of hatchery-raised juveniles, including many salmon from both the Atlantic and Pacific (15,16), Japanese flounder (17), Hawaiian mullet (18), Nassau grouper (19), and Chesapeake Bay striped bass (20), just to name a few. Holothurians have been hatchery-raised for stock enhancement (21). Bivalves, such as quahogs (22) and soft-shell clams (23), and gastropods, such as abalones (24) and queen conch (25), have been the subject of stock enhancement efforts. Finally, crustaceans are beginning to receive more attention as possible stock enhancement targets; programs have been initiated for decades with American and

126

MARINE STOCK ENHANCEMENT TECHNIQUES

European lobsters (26,27), more recent programs for prawn (28) and the Japanese swimming crab (3), and a new exploratory program established for the Chesapeake Bay blue crab (8). Stock enhancement efforts extend worldwide. In Japan, stock enhancement programs are in place for at least 34 finfish and 12 crustaceans (3). Programs relying on various specific techniques exist in South America, North America, Africa, Europe, Asia, Australia/New Zealand, and in the developed world as well as the developing world (29–31). Specific techniques to optimize stock enhancement depend on the life-history traits of the species. Sessile organisms may be enhanced by seeding areas with young individuals (32). Depending on how closely the success of the enhancement effort is to be monitored, more mobile species may be released into isolated areas where they can be followed, similar to the idea of stocking freshwater lakes (8,28). For extremely mobile open-ocean species, juveniles are released into the open ocean (3), and even if tagged, large numbers cannot be followed over time. PROBLEMS OF STOCK ENHANCEMENT Enhancement has been a controversial management method for several reasons (15,29,33). First, hatcheryraised animals may not survive well in the wild. The hatchery environment can offer conditions very different from the natural environment: food is likely to be different, the method of foraging is different, and therefore the hatchery-raised organisms may have mortality-threatening inexperience with natural prey after their introduction into the wild. They may similarly be inexperienced in avoiding predators. Holding tanks may restrict movement, may have unnatural substrates, may have unnatural light regimes, or may have different flow regimes. As a result of all of these differences, stock enhancement efforts, often funded by the taxpayer, may not be successful (34). Many studies have noted differences in hatchery-raised and wild individuals in factors such as behavior, morphology, growth rates, and therefore survivorship (17,35,36). A second concern is that hatchery-raised individuals may be too successful. Hatchery animals may compete with and displace wild animals (2,37). If at any point the population becomes habitat-limited, rather than recruitment-limited (or if the stock enhancement effort pushes the population over the carrying capacity), the survivorship of wild individuals may actually decrease. The ultimate result would then be a decline in the population of wild individuals, even if the overall population has increased. A third concern is that hatchery-raised organisms may carry diseases into the wild, affecting the survivorship of wild individuals. Fourth, increases in stock size due to hatchery successes may provoke a rise in fishing effort and therefore greater pressure on the remaining wild individuals (34,38). Stock enhancement efforts are generally performed in concert with traditional management techniques, such as lowering catch and/or restricting catch to certain segments of the population. However, if the human population increases, pressure

may be put on managers to reduce fishing regulations. Again, the ultimate result would be a decline in the survivorship rates of wild individuals, even if overall population has increased. Finally, and perhaps the reason for the controversy generating the most attention in recent years is genetics, or ‘‘gene dilution.’’ Even if efforts in the hatchery are expended to obtain brood stock from many different parents, genetic variability in the brood stock is unlikely to approach that of the wild parental stock. The offspring produced will be more similar to each other than the offspring of wild stock. Opponents of stock enhancement fear that releasing genetically relatively homogenous hatchery individuals will reduce genetic variability in wild populations as hatchery and wild individuals interbreed. The ultimate result after several to many generations might be a decline in population because lack of genetic diversity can limit a population’s response to environmental change. The controversy over stock enhancement has been fueled by the fact that most enhancement efforts have not been studied quantitatively (37,41,42). Quantitative, hypothesis-driven study did not begin until the late 1980s (43). However, now that the necessity to understand better the outcomes of enhancement programs has been recognized, in many cases the enhancement process has been refined. Methods to select better candidate species have been developed (44). Advances in tagging techniques have allowed better assessment and comparison of hatchery and wild animal survivorship (44). Most importantly, calls have been made for quantitative study of small-scale enhancement efforts before investment in large-scale programs begins (42,43). OPTIMIZING STOCK ENHANCEMENT SUCCESS Stock enhancement is becoming a more often commonly used method for addressing declines in fished species. Many of the problems identified above can be addressed to a comfortable degree before the program is initiated. For example, several programs are considering conditioning individuals to limit the differences between hatcheryraised and wild organisms before release into the wild (12,45,46). In this way, survivorship rates of hatchery organisms are increased, along with the output per enhancement program dollar spent. In addition, carrying capacities of targeted release areas can be determined before programs are initiated to (1) determine whether the case is appropriate for stock enhancement and (2) determine optimum release densities of hatchery-raised organisms. Certain microhabitats within a targeted release region may have higher carrying capacity and greater food or refuge resources than others, and therefore distribution of hatchery-raised organisms can be optimized. Methods of determining whether stock enhancement is successful are more difficult. Determining success requires knowledge about how well hatchery-raised organisms survive, how well wild organisms survive to allow comparison, and the contribution of hatchery organisms to the total population. As programs are developed, a wide

MARINE STOCK ENHANCEMENT TECHNIQUES

range of survivorship and contribution values have been reported. For example, survivorship to fishery size was 3–4% for stocked panaeid shrimp (28), 21% for red drum in Texas (47), and up to 30% for Japanese flounder (3). Even among programs for the same species, values range widely. For example, in some European lobster programs, no hatchery-raised lobsters were recaptured in the fishery, and in others, 10–35% of landed lobsters were of hatchery origin (37). Often the steps to quantify the success of a stock enhancement program take years of scientific study and require laborious study efforts, such as tagging and resampling individuals over time. Such efforts have been deemed mandatory by critics of stock enhancement programs before public monies are used to support these efforts to bolster fishery stocks, one of the world’s most important natural resources. BIBLIOGRAPHY 1. Food and Agriculture Organization of the United Nations. (2002). The State of World Fisheries and Aquaculture. United Nations, Rome, Italy. 2. Castro, K.M., Cobb, J.S., Wahle, R.A., and Catena, J. (2001). Habitat addition and stock enhancement for American lobsters, Homarus americanus. Mar. Freshwater Res. 52: 1253–1261. 3. Masuda, R. and Tsukamoto, K. (1998). Stock enhancement in Japan: Review and perspective. Bull. Mar. Sci. 62: 337–358. 4. Beamesderfer, R.C.P. and Farr, R.A. (1997). Alternatives for the protection and restoration of sturgeons and their habitat. Environ. Biol. Fish. 48: 407–417. 5. Hendry, K., Cragge-Hine, D., O’Grady, M., Sambrook, H., and Stephen, A. (2003). Management of habitat for rehabilitation and enhancement of salmonid stocks. Fish. Res. 62: 171–192. 6. Peterson, C.H., Grabowsi, J.H., and Powers, S.P. (2003). Estimated enhancement of fish production resulting from restoring oyster reef habitat: Quantitative valuation. Mar. Ecol. Prog. Ser. 264: 149–264. 7. Rodwell, L.D., Barbier, E.B., Roberts, C.M., and McClanahan, T.R. (2003). The importance of habitat quality for marine reserve—fishery linkages. Can. J. Fish. Aquat. Sci. 60: 171–181. 8. Davis, J.L.D., Young-Williams, A.C., Hines, A.H., and Zohar, Y. (in press). Assessing the feasibility of stock enhancement in the case of the Chesapeake Bay blue crab. Can. J. Fish. Aquat. Sci. 9. Iguchi, K., Watanabe, K., and Nishida, M. (1999). Reduced mitochondrial DNA variation in hatchery populations of ayu (Plecoglossus altivelis) cultured for multiple generations. Aquaculture 178: 235–243. 10. Kestemont, P. et al. (2003). Size heterogeneity, cannibalism and competition in cultured predatory fish larvae: Biotic and abiotic influences. Aquaculture 227: 333–356. 11. Caillouet, C.W., Fontaine, C.T., Manzella-Tirpak, S.A., and Shaver, D.J. (1995). Survival of head-started Kemp’s ridley sea turtles (Lepidochelys kempii) released into the Gulf of Mexico or adjacent bays. Chelate Conserv. Biol. 1: 285–292. 12. Davis, J.L.D., Eckert-Mills, M.G., Young-Williams, A.C., Hines, A.H., and Zohar, Y. (in press). Morphological conditioning of a hatchery-raised invertebrate, Callinectes sapidus, to improve field survivorship after release. Aquaculture 133: 1–14.

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13. Fleming, I.A. et al. (2000). Lifetime success and interactions of farm salmon invading a native population. Proc. R. Soc. London Ser. B 267: 1517–1523. 14. Wirth, F.F. and Davis, K.J. (2003). Seafood dealers’ shrimppurchasing behavior and preferences with implications for United States shrimp farmers. J. Shellfish Res. 22: 581–588. 15. Washington, P.M. and Koziol, A.M. (1993). Overview of the interactions and environmental impacts of hatchery practices on natural and artificial stocks of salmonids. Fish. Res. 18: 105–122. 16. Thorpe, J.E. (1998). Salmonid life-history evolution as a constraint on marine stock enhancement. Bull. Mar. Sci. 62: 465–475. 17. Kellison, G.T., Eggleston, D.B., and Burke, J.S. (2000). Comparative behaviour and survival of hatchery-reared versus wild summer flounder (Paralichthys dentatus). Can. J. Fish. Aquat. Sci. 57: 1870–1877. 18. Leber, K.M., Blankenship, H.L., Arce, S.M., and Brennan, N.P. (1996). Influence of release season on sizedependent survival of cultured striped mullet, Mugil cephalus, in a Hawaiian estuary. Fish. Bull. 95: 267–279. 19. Roberts, C.M., Quinn, N., Tucker, J.W., and Woodward, P.N. (1995). Introduction of hatchery-reared Nassau grouper to a coral reef environment. N. Am. J. Fish. Manage. 15: 159–164. 20. Secor, D.H. and Houde, E.D. (1998). Use of larval stocking in restoration of Chesapeake Bay striped bass. ICES J. Mar. Sci. 55: 228–239. 21. Battaglene, S.C., Seymour, J.E., and Ramofafia, C. (1999). Survival and growth of cultured juvenile sea cucumbers, Holothuria scabra. Aquaculture 178: 293–322. 22. Walton, W.C. and Walton, W.C. (2001). Problems, predators, and perception: Management of quahog (hardclam), Mercenaria mercenaria, stock enhancement programs in southern New England. J. Shellfish Res. 20: 127–134. 23. Beal, B.F., Kraus, F., and Gayle, M. (2002). Interactive effects of initial size, stocking density, and type of predator deterrent netting on survival and growth of cultured juveniles of the soft-shell clam, Mya arenaria L., in eastern Maine. Aquaculture 208: 81–111. 24. Tegner, M.J. and Butler, R.A. (1985). The survival and mortality of seeded and native red abalones, Haliotis rufescens, on the Palos Verdes Penninsula. Calif. Fish Game 71: 150–163. 25. Stoner, A.W. and Glazer, R.A. (1998). Variation in natural mortality: Implications for queen conch stock enhancement. Bull. Mar. Sci. 62: 427–442. 26. Bannister, R.C.A. and Addison, J.T. (1998). Enhancing lobster stocks: A review of recent European methods, results, and future prospects. Bull. Mar. Sci. 62: 369–387. 27. Beal, B.F., Chapman, S.R., Irvine, C., and Bayer, R.C. (1998). Lobster (Homarus americanus) culture in Maine: A community-based, fishermen-sponsored, public stock enhancement program. Can. Ind. Rep. Fish. Aquat. Sci. 244: 47–54. 28. Davenport, J., Ekaratne, S.U.K., Walgama, S.A., Lee, D., and Hills, J.M. (1999). Successful stock enhancement of a lagoon prawn fishery at Rekawa, Sri Lanka, using cultured postlarvae of penaeid shrimp. Aquaculture 180: 65–78. 29. Bell, J.D. and Gervis, M. (1999). New species for coastal aquaculture in the tropical Pacific: Constraints, prospects and considerations. Aqua Int. 7: 207–223. 30. Cowx, I.G. (1999). An appraisal of stocking strategies in the light of developing country constraints. Fish. Manage. Ecol. 6: 21–34.

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31. Phillips, B.F. and Liddy, G.C. (2003). Recent developments in spiny lobster aquaculture. In: Phillips, B., Megrey, B.A., and Zhou, Y. (Eds). Proc. 3rd World Fish. Congr.: Feeding the world with fish in the next millenium—the balance between production and environment. American Fisheries Society Symposium Vol. 38, pp. 43–57. 32. Barbeau, M.A. et al. (1996). Dynamics of juvenile sea scallop (Placopecten magellameus) and their predators in bottom seeding trials in Lanenburg Bay, Nova Scotia. Can J. Fish. Aquat. Sci. 53: 2494–2512. 33. Li, J. (1999). An appraisal of factors constraining the success of fish stock enhancement programmes. Fish. Manage. Ecol. 6: 161–169. 34. Hilborn, R. (1998). The economic performance of marine stock enhancement projects. Bull. Mar. Sci. 62: 661–674. 35. Mana, R.R. and Kawamura, G. (2002). A comparative study on morphological differences in the olfactory system of red sea bream (Pagrus major) and black sea bream (Acanthopagrus schlegeli) from wild and cultured stocks. Aquaculture 209: 285–306. 36. Davis, J.L.D. et al. (2004). Differences between hatcheryraised and wild blue crabs (Callinectes sapidus): Implications for stock enhancement potential. Trans. Am. Fish. Soc. 37. Bannister, R.C.A., Addison, J.T., and Lovewell, S.R.J. (1994). Growth, movement, recapture rate and survival of hatcheryreared lobsters (Homarus gammarus (Linnaeus, 1758)) released into the wild on the English East Coast. Crustaceana 67: 156–172. 38. Bannister, R.C.A. (2000). Lobster Homarus gammarus stock enhancement in the U.K.: Hatchery-reared juveniles do survive in the wild, but can they contribute significantly to ranching, enhancement, and management of lobster stocks? Can. Spec. Publ. Fish. Aquat. Sci. 130: 23–32. 39. Tringali, M.D. and Bert, T.M. (1998). Risk to genetic effective population size should be an important consideration in fish stock-enhancement programs. Bull. Mar. Sci. 62: 641–659. 40. Utter, F. (1998). Genetic problems of hatchery-reared progeny released into the wild, and how to deal with them. Bull. Mar. Sci. 62: 623–640. 41. Heppell, S.S. and Crowder, L.B. (1998). Prognostic evaluation of enhancement programs using population models and life history analysis. Bull. Mar. Sci. 62: 495–507. 42. Crowe, T.P. et al. (2002). Experimental evaluation of the use of hatchery-reared juveniles to enhance stocks of the topshell Trochus niloticus in Australia, Indonesia and Vanuatu. Aquaculture 206: 175–197. 43. Leber, K.M. (1999). Rational for an experimental approach to stock enhancement. In: Stock Enhancement and Sea ¨ Ranching. B.R. Howell, E. Moksness, and T. Svasand (Eds.). Blackwell Science, Oxford, pp. 63–75. 44. Blackenship, H.L. and Leber, K.M. (1995). A responsible approach to marine stock enhancement. Am. Fish. Soc. Symp. 15: 167–175. 45. Olla, B.L. and Davis, M.W. (1989). The role of learning and stress in predator avoidance of hatchery-reared coho salmon (Oncorhynchus kisutch) juveniles. Aquaculture 76: 209–214. 46. Berejikian, B.A. (1995). The effects of hatchery and wild ancestry and experience on the relative ability of steelhead trout fry (Oncorhynchus mykiss) to avoid a benthic predator. Can. J. Fish. Aquat. Sci. 52: 2476–2482. 47. McEachron, L.W., Colura, R.L., Bumgaurdner, B.W., and Ward, R. (1998). Survival of stocked red drum in Texas. Bull. Mar. Sci. 62: 359–368.

PHYSICAL AND CHEMICAL VARIABILITY OF TIDAL STREAMS PAOLO MAGNI IMC—International Marine Centre Torregrande-Oristano, Italy

SHIGERU MONTANI Hokkaido University Hakodate, Japan

This article focuses on the variability of major hydrologic parameters of tidal streams that occur on different temporal and spatial scales within an individual estuary. Rapid changes in water temperature and salinity take place on a timescale of 31 to 2.0 psu, coincident with the lower low tide, and rapidly increased up to >31 psu again before the following higher high tide (Fig. 1a). At the subsequent higher low tide, the decrease in salinity was less marked, and minimum values were about 25 psu. An overall more restricted fluctuation of salinity occurred in the 1996

PHYSICAL AND CHEMICAL VARIABILITY OF TIDAL STREAMS

(a)

2 Depth, m

(b)

May 30−31, 1995 Spring tide

1 A.M.

Higher high tide

1

2

129

May 28−29, 1996 Neap tide

1 A.M.

1

Temperature, °C

Lower low tide 0

0

22

24 23

21

22

20

21

19

20 19

18

18

Salinity, psu

17

17

16

16

30

30

20

20

10

Dissolved oxygen, mg/L

0

10

(100%)

0 7.0 mg/L (88%)

17.4 mg/L (224%) 15

6 12 9

4

(100%) 6 2

(20%)

3 1.4 mg/L

0 10:00

20:00 6:00 Time, h

1.3 mg/L

0 16:00

10:00

survey, consistent with the reduced amplitude of the tide (Fig. 1b) and a rainfall regime more limited in spring 1996 than in spring 1995 (1). These results show that saline intrusion is a strong function of tidal state (e.g., low vs. high tide) and amplitude (e.g., spring vs. neap tide). They also demonstrate that tidal streams may experience, within a few hours, strong changes in salinity that cause very different habitat conditions at the same location in an estuary. These may vary from oligohaline (0.51–5 psu)

20:00 6:00 Time, h

(20%) 16:00

Figure 1. Daily fluctuations of water depth, temperature, salinity and dissolved oxygen (DO) concentration of a tidal stream in an estuarine intertidal zone (Seto Inland Sea, Japan). Measurements were made using a CTD cast placed ca. 10 cm from the bottom sediment at the end of May of 2 consecutive years (a: 1995; b: 1996). Relevant sensors were logged every 15 min. Notes: horizontal dashed lines in temperature and DO boxes individuate major differences within the same range of values (17–22 ◦ C and 20–100% of air saturation, respectively) recorded in the 2 years. In DO boxes, absolute minimum and maximum values are also indicated. The vertical dashed line is arbitrarily depicted at 1 A.M. to highlight the temporal differences in tidal state and amplitude between the two surveys.

to euhaline (30.1–40 psu) conditions, in agreement with salinity ranges given by the Venice Conference (5). The daily fluctuation in water temperature varied according to the tidal cycle and showed a temporal pattern opposite to that of salinity during both surveys (Fig. 1a, b). The range of temperature was relatively larger in May 1996 (17.7–24.5 ◦ C) than in 1995 (16.9 ◦ C–22.3 ◦ C). In both cases, it was apparent that warmer waters were brought into the estuary by the freshwater runoff during

PHYSICAL AND CHEMICAL VARIABILITY OF TIDAL STREAMS

were conducted (see above section and Fig. 1). Between January 1995 and April 1996, the water temperature varied from 3.6 ◦ C ± 0.3 (December) to 29.0 ◦ C ± 1.8 (July) (Fig. 2). Statistically significant differences were demonstrated for the specific geographical area (latitude 34◦ 21 N, longitude 43◦ 21 E) between a warm period (May–October, water temperature 23.5 ± 4.2 ◦ C) and a cold period (November–April, water temperature 10.0 ± 4.4 ◦ C) (1). The seasonal variability in low-tide salinity (Fig. 2) was consistent with the results obtained from the shortterm measurements of May 1995 and May 1996 during a complete tidal cycle (Fig. 1). In particular, the salinity recorded during the seasonal survey was lowest on May 16, 1995 at 4.6 ± 1.5 psu (Fig. 2). Accordingly, the short-term measurements of May 1995 showed a salinity decrease during low tide to 25 psu between the end of September 1995 and April 1996 and reached the highest values in November 1995 at 31.4 ± 1.1 psu (Fig. 2). Consistently, also during the short-term measurements of May 1996, the decrease in salinity during low tide was more limited than that found in May 1995 (Fig. 1a, b). The DO concentration also varied greatly from 5.4 ± 1.7 mg/L (September 1995) to 15.1 ± 1.1 mg/L (April 1996)

Salinity, psu

40

30

20

10

Dissolved oxygen, mg/L

0 15 12 9 6 3 0 20 Jan 17 Feb 17 Mar 15 Apr 16 May 30 May 14 Jun 12 Jul 10 Aug 7 Sep 29 Sep 30 Oct 27 Nov 26 Dec 24 Jan 22 Feb 21 Mar 17 Apr

ebb flow. These results are consistent with the period and site of measurements; physical processes of heat transfer in spring–summer (and water cooling in winter) are most effective in the upper and shallower riverine zone of the estuary (1). Dissolved oxygen (DO) concentration was subjected to strong daily fluctuations, partly as a function of the tidal state (and water depth). The two surveys also showed major differences (Figs. 1a and b). In the 1995 survey, DO concentration was mostly within normoxic values, ranging from 1.4 mg/L to 7.0 mg/L. Differently in the 1996 survey, DO concentration showed a larger variation, in much higher values up to 17.4 mg/L. Elevated DO concentrations are related to ecosystem processes of primary production (6). It is known that shallow lagoons and coastal areas dominated by seagrass or macroalgae are subjected to oversaturation of DO (i.e., >100% of air saturation), especially during the warm period (7,8). Viaroli et al. (7) reported DO oversaturation up to 150% in the near-bottom water of a coastal lagoon dominated by the macroalga Ulva. This was followed by the outbreak of a dystrophic crisis, complete anoxia through the water column at some stations of the lagoon. The development of large amounts of macroalgae (Ulva sp.) also tends to occur in the estuarine sand flat of this study coincident with increasing temperature and solar radiation during the spring. Extended beds of macroalgal biomass were present during the field measurements of the 1996 survey (personal observations). Accordingly, during the daytime measurements of May 28, DO concentration rapidly rose to oversaturation; a major increase of >200% of air saturation occurred between 16:00 and 16:30 (Fig. 1b), indicating a period of major oxygen production by macroalgae. Similar to the extremely high DO values found in this study, Piriou & M´enesguen (9) reported that the in vitro growth of Ulva under light saturation and nutrient enrichment raised the DO concentration to 22 mg/L, 4 hours after the experiment started. By contrast, during nighttime, a progressive decrease in DO concentration occurred, down to hypoxic values of 20. If the water height is small, the cnoidal wave profile becomes the sinusoidal one. The cnoidal wave is periodic with the profile given by   
x t ,κ (5) − η(x, t) = hCn2 2K(κ) L T




2π d tan h L




 =

gL 2π




tan h

2π d L

 (2)

The solitary wave is a progressive wave consisting of a single crest and is not oscillatory as the other examined. The wave form is (4) 

3H 2 (x − Ct) (6) η(x, t) = H sec h 4d3 where C (wave celerity) is defined by  
H C = gd 1 + 2d

(7)

According to the dispersion relation of the linear wave theory, the celerity of water waves in shallow water (Eq. 3) is smaller than solitary wave phase velocity because of the inclusion of terms that depend on H/d. The solitary wave describes enough well waves approaching shallow water, even if wave period or wave

SHALLOW WATER WAVES

length are not associated with the theory. When the wave pass from deep to shallow water their crests peak up and are separated by flat troughs appearing like a series of solitary waves. Solitary theory appears reasonable even if the periodicity is neglected, because in shallow water the period is not particularly significant, but rather the water depth becomes important. As the solitary wave advances into shallow water, its height increases, the crests becomes greater and sharper, the trough becomes longer and flatter: in this condition, a wave is well represented by solitary waves. The solitary wave is not an oscillatory wave, as those obtained with other theories, but a translation one. Water particles, as a wave passes, are subjected to a translation in wave direction, whereas in the oscillatory waves it moves forward and backward, remaining after a period in its original position. Wave profiles, according to the theories illustrated above, are presented in Fig. 1. WAVE REFRACTION AND SHOALING Waves passing from deep to shallow water are subjected to refraction in which the direction of their travel changes in such a way that approaching the coast the crests tends to become more parallel to the depth contours (Fig. 2). To

determine the variation of the wave direction, Snell’s law can be applied sin θ2 sin θ1 = = cons C1 C2

Linear theory wave S.W.L.

Cnoidal wave

S.W.L.

Solitary wave

S.W.L. Figure 1. Wave profiles.

a1

Celerity CA

a2

b1

Greater depth

Contour line

B

A Shallower depth

B

Wave ray

Contour line

Cel e CB rity

a1

Greater depth A

(8)

where q1 and q2 are the angles between adjacent wave crests and the respective bottom contours, whereas C1 and C2 are the wave celerity at the two depths. With a regular bottom (straight and parallel contours), the relation can be applied directly between the angle at any depth and the deep water angle approach. With an irregular bottom, wave refraction may cause a spreading or a convergence of the wave energy. This effect can be easily illustrated taking in account the wave rays (Fig. 2), defined as the lines drawn normal to the wave crests and directed in the wave advance. If the wave rays spread, the wave crests become longer and the energy flux, assumed constant between two rays, must be extended over a greater length. The opposite (energy concentration) happens if the waves rays converge. Actually, the calculations of the rays are made by software based on models such as a mild slope, parabolic, or Boussinesq wave model. Another effect of the change in the wave length in shallow water is that the wave height increases. This effect is the consequence of the energy conservation in

Shallower depth

a2

136

Shoreline Figure 2. Wave refraction and wave rays.

b2

Shoreline

SHALLOW WATER WAVES

concert with the decrease of the celerity approaching the shallow water. This phenomenon is referred to as shoaling. The effects of shoaling and refraction in water can be expressed by the following formula: H = H0 Ks Kr

(9)

where H0 is wave height in deep water, Ks is the shoaling coefficient    CG0 1  Ks = = (10)  2π d CG 2n tan h L   with n =

1 1 + 2π d 2 L

tion coefficient

 1   and Kr is the refrac
2π d  senh L

b0 Kr = (11) b

where b0 and b are the distance between two adjacent rays, respectively, in deep water and at a generic depth. For straight and parallel contours lines, the refraction coefficient becomes

cos θ0 (12) Kr = cos θ WAVE BREAKING Waves shoaling causes the increasing of wave height until its physical limit because of steepness of wave H/L. When this limit is reached, the wave breaks and dissipates energy. Battjes (5) has shown that the breaking wave characteristics can be correlated to a parameter, called surf similarity x, which is defined as tan β ξ=

H0 L0

— spilling (x < 0.5), in which each wave gradually peaks until the wave becomes unstable and cascades down as white water, bubbles, and foam; — plunging (0.5 < x < 3.3), in which the shoreward face of the wave becomes vertical, curls over, and

Spilling

Plunging

plunges forward and downward as an intact mass of water; — surging (3.3 < x < 5), in which the base of the wave, while it is peaking up, finds the shore, and then the crest collapses and disappears. When the surf similarity x is >5, reflection happens and no breaking occurs. According to Galvin (6), a fourth breaker, called collapsing, intermediate between plunging and surging types, exists. It is difficult to identify which type of breaker can verify, because it depends on the individual heights and interactions of the waves. Plunging and surging breakers can be seen on a beach during the same storm. However, spilling breakers are typical of very low sloping beaches with waves of high steepness values; plunging waves occur in steeper beaches and waves of intermediate steepness; surging, instead, is associated with high gradient beaches with waves of low steepness. The breaking happens when the water particle velocity is greater than wave celerity. According to solitary wave theory, this condition is described by (7)


Hb db

 = 0.78

(14)

max

where the subscript b denotes the breaking. Laboratory tests pointed out that Equation (14) is verified more for oscillatory waves than solitary wave. However, it is considered fundamental to express the relation between the relative depth and the breaking condition. Other parameters playing a role in wave breaking exist, such as beach slope and bottom roughness. An empirical relation considers the beach slope m (8):


(13)

where tan b is the beach slope and H0 and L0 are the wave height and length in deep water, respectively. Three common type of breakers are recognized (Fig. 3):

137

Hb db

 = 0.75 + 25m − 112m2 + 3870m3

(15)

max

in which as the slope increases, the breaking happens more and more nearshore. Another expression was developed by Goda (9)  
L0 15π db (1 + 15m4/3 ) 1 − exp − db L0 max (16) where L0 is the deep water wave lenght. For irregular waves (waves with different height and period), Kamphuis (10) proposed two criteria based on


Hb db



= 0.17

Collapsing

Surging

Figure 3. Breaking types.

138

WATER WAVES

extensive model tests




Hsb = 0.095e4m Lbp tan h

2π db Lbp

(17)

hurricane may generate huge waves. This storm usually causes disasters. Tidal and wind waves contain a lot of energy. Electric power can be generated by them.

(18)

BASIC CHARACTERISTICS OF A WATER WAVE



Hsb = 0.56e3.5m db

where Hsb is the significant wave breaking, Lbp the breaking wave length, and db is the breaking depth. BIBLIOGRAPHY 1. Airy, G.B. (1845). Tides and waves. Encyclopedia Metropolitana. 192: 241–396. 2. Korteweg, D.J. and de Vries, G. (1895). On the change of form of long waves advancing in a rectangular canal and on a type of long stationary wave. Phil. Mag. 5: 442–443. 3. Wiegel. (1960). 4. Russell, J.S. (1884). Report on waves. 14th Meeting of the British Association for the Advancement of Science. pp. 311–390. 5. Battjes, J.A. (1974). Surf similarity. Proceedings of the 14th Coastal Engineering Conference, ASCE. pp. 466–497. 6. Galvin, C.J. (1968). Breaker type classification on three laboratory beaches. J. Geophys. Res. 73: 3651–3659. 7. Munk, W.H. (1949). The solitary wave theory and its application to surfs problems. Ann. New York Acad. Sci. 51: 376–424. 8. SPM. (1984). 9. Goda, Y. (1970). A synthesis of breaker indices. Trans. Japan Soc. Civil Engrs. 2: 227–230. 10. Kamphuis, J.W. (1991). Wave transformation. Co. Eng. 15: 173–184.

FURTHER READING CERC U.S. Corps of Eng. (1984). Shore Protection Manual. Vicksburg, VA.

WATER WAVES

The water wave was studied mathematically in the nineteenth century as a form of oscillatory wave. In 1847, Stokes (1) published his famous paper entitled ‘‘On the theory of oscillatory waves.’’ Stokes’ wave theory has been widely used till now. The elements of an oscillatory wave are wave height, wave period, and wavelength. The wave height H is the distance between the wave crest and trough. The wavelength L is the distance between successive crests, and the wave period T the time difference between successive crests. In deep water, the water depth H is larger than the half of the wavelength L. After Stokes, L is a function of the wave period T in the following equation: gT 2 (1) L= 2π The wave celerity C is then C=

National Taiwan University Taipei, Taiwan

INTRODUCTION As one stands at the coast, an endlessly moving succession of irregular humps and hollows can be seen reaching to the shore. This is the water wave generated by wind. There are different kinds of water waves, which are driven by different forces. Beside wind waves, the tide and the tsunami are other well-known water waves. Different Water waves are distinguished by their wavelength, which is defined as the length between successive humps. The wavelength of a wind wave is shorter and is easily recognized. A water wave is an important physical phenomenon in an ocean, sea, or lake. It influences beach morphology, maritime structures, and human activities very much. Because the wind varies tremendously, a typhoon or

(2)

in which g is the earth’s gravitational acceleration. Obviously, the celerity C is proportional to the wave period T. The water depth is less than L/2, so then the wave particle movement touches the bottom. Tides and tsunamis usually are shallow water waves or long waves. Equations 1 and 2 become the following: 
2π h gT 2 (3) tanh L= 2π L
 2π h gT C= tanh (4) 2π L A wave contains kinetic and potential energy. The average energy E per unit sea surface area is the following: E=

NAI KUANG LIANG

gT L = T 2π

1 ρgH 2 8

(5)

The wave energy propagates in a group velocity, Cg, which is given by the following equation (2):
 4π h/L C 1+ (6) Cg = 2 sinh 4π h/L The wave energy flux P is as follows: P = E · Cg

(7)

A real ocean wave is not oscillatory, as Stokes’ theory described, but irregular in a stochastic process. Pierson (3,4) introduced the technique of communication engineering to the ocean wave and proposed the random wave theory. The ocean wave would be a superposition of sinusoidal wave components of different directions, amplitudes, frequencies, and angular phases, in which the phase is a random variable of equal probability density between—π and π . Then the ocean wave can be represented by a power spectrum. However, for

WOODS HOLE: THE EARLY YEARS

convenience, engineers usually use the significant wave to represent the ocean wind wave. For a group of N wave heights measured at a point, waves are ordered from the largest to the smallest and assigned a number from 1 to N to them. The significant wave height H1/3 is defined as the average of the first (highest) N/3 wave heights (5). The order of the wave period is accompanied by its wave height as a pair. The significant wave period T1/3 is defined as the average of the first N/3 wave periods. WIND WAVE FORECASTING During the Second World War, two famous American oceanographers, Dr. H.U. Sverdrup and Dr. W.H. Munk, were assigned by the U.S. Navy to develop a wind wave forecasting scheme for the Normandy landing operations. The work was originally completed in 1943 and classified and published in 1947(6). Later, the scheme was extensively patched and amended by Bretschneider(7). Therefore, the scheme has been named the SMB method. In the SMB method, the wind speed, wind duration, and fetch are the main parameters, where fetch is defined as the wind blowing distance in the water area. Modern ocean wave modeling was initiated in 1956 and extensively developed and revised (8). So far it is still developing. The basic concept is the evolution of energy spectrum F governed by the energy balance equation: df = Sin + Snl + Sds dt

(8)

in which Sin is the energy input flux from wind to wave spectrum components, Snl the energy flux exchange due to nonlinear wave–wave interaction, and Sds the energy flux output due to dissipation. The present operating ocean models are WOM, NWW3, etc. TYPHOON WAVE The typhoon or hurricane is an atmospheric eddy that originates in tropical or subtropical ocean regions. The typhoon wind speed is usually very high so that the typhoon wave plays a significant role in the design of coastal and offshore structures. Because the wind velocity changes rapidly, the generation process within typhoons is complicated. Parametric typhoon or hurricane wave prediction models were developed (9,10), which pointed out that the maximum wave exists at the right side of the typhoon center as one faces the forward direction. The radius of a typhoon is about 200–400 km. When a typhoon is still far away from a location, the swell may arrive because the swell energy propagates usually faster than the typhoon moves. The typhoon can be regarded as a point source of wave generation. If Person A, who does not move, throws balls to Person B at a fixed time interval and the speed of the ball relative to the ground is constant, then Person B receives balls at the same time interval. If Person A moves toward Person B, Person B receives balls at a shorter time interval. If the ball is like energy, Person A throws more energy flux to Person B, as Person A moves toward Person B. This is the same as the well-known Doppler effect. As a whistling

139

train moves toward an observer, the sound heard by the observer is higher in frequency than the actual whistle at the source. The observer receives more sound energy flux. As shown in Equation 7, wave energy flux P is wave energy E × group velocity Cg, which is proportional to the wave period. The wave energy is proportional to H 2 (Equation 5). Because the typhoon wave moving speed has little effect on the swell period, then the swell height is enhanced, as the typhoon approaches a station (11). There were cases where swell heights were larger than those inside the typhoon because the moving velocity was close to the group velocity of the swell. BIBLIOGRAPHY 1. Stokes, G.G. (1847). On the theory of oscillatory waves. Trans. Cambridge Philos. Soc. 8: 441. 2. Kinsman, B. (1965). Wind Waves: The Generation and Propagation on the Ocean Surface. Prentice-Hall, Englewood Cliffs, NJ. 3. Pierson, W.J., Jr. (1952). A Unified Mathematical Theory for the Analysis, Propagation and Refraction of Storm Generated Ocean Surface Waves, Parts I and II. New York University, College of Engineering, Research Division, Department of Meteorology and Oceanography prepared for the Beach Erosion Board, Department of the Army, and Office of Naval Research, Department of the Navy. 4. Pierson, W.J., Jr. (1955). Wind-generated gravity waves. In: Advances in Geophysics. Vol. 2. Academic Press, New York, pp. 93–178. 5. Dean, R.G. and Dalrymple, R.A. (1984). Water Wave Mechanics for Engineers and Scientists. Prentice-Hall, Englewood Cliffs, NJ. 6. Sverdrup, H.U. and Munk, W.H. (1947). Wind, Sea and Swell: Theory of Relations for Forecasting. U. S. Navy Hydrographic Office Pub. No. 601. 7. Bretschneider, C.L. (1952). The generation and decay of wind waves in deep water. Trans. Am. Geophys. University 33(3): 381–389. 8. Komen, G.J., Cavaleri, L., Donelan, M., Hasselmann, K., Hasselmann, S., and Janssen, P.A.E.M. (1994). Dynamics and Modeling of Ocean Waves. Cambridge University Press. 9. Bretschneider, C.L. and Tamaye, E.E. (1976). Hurricane wind and wave forecasting techniques. Proc. 15th Int. Coastal Eng. Conf., Honolulu, HI, pp. 202–237. 10. Young, I.R. (1988). Parametric hurricane wave prediction model. J. Waterway Port Coastal Ocean Eng. 114(5): 637–652. 11. Liang, N.K. (2003). The typhoon swell Doppler effect. Ocean Eng. 30: 1107–1115.

WOODS HOLE: THE EARLY YEARS Northeast Fisheries Science Center—NOAA

The beginning of Woods Hole dates back to the early 17th century. Five years before the settlement of Jamestown, This article is a US Government work and, as such, is in the public domain in the United States of America.

140

WOODS HOLE: THE EARLY YEARS

Virginia, and 18 years before the Pilgrims landed at Provincetown and Plymouth, Bartholomew Gosnold coasted along Cape Cod and Marthas Vineyard, and about May 31, 1602, he is believed to have landed at what is now known as Woods Hole. The Town of Falmouth, of which Woods Hole is presently a part, was first settled in 1659–61 when several persons were granted permission to purchase land. The date of the settlement of Woods Hole took place 17 years later. The town (Falmouth) was incorporated on June 4, 1686, and called Succonessett, the name which later, probably in 1694, was changed to Falmouth. On July 23, 1677, the land around Little Harbor of Woods Hole was divided among the 13 settlers in ‘‘lots of 60 acres upland to a share’’ and an ‘‘Indian deed’’ confirming the land title was signed by Job Notantico on July 15, 1679. Fishing, hunting, and sheep breeding were the principal occupations of the early settlers and their descendents. Later on a grist mill was built and salt was made by solar evaporation of sea water in pans built along the banks of Little Harbor. These quiet, rural conditions, devoid of adventure, persisted until about 1815, when Woods Hole became an important whaling station from which ships operated on the high seas. The whaling industry in the United States became a very profitable business, and Woods Hole was a part of it. In 1854, the total receipts for the American whaling fleet amounted to $10.8 million, the largest part of this amount resulted from whaling carried out by Massachus tts captains. Woods Hole participated in these activities and prospered. It is known that between 1815 and 1860, not less than nine whaling ships were making port at the Bar Neck wharf, which was located where the U.S. Navy building of the Woods Hole Oceanographic Institution now stands. The place was busy processing oil and whalebone and outfitting ships. A bake house for making sea biscuits for long voyages stood next to the present ‘‘Old Stone Building’’ built in 1829 as a candle factory. This conspicuous old landmark on Water Street of Woods Hole, identified by an appropriate bronze plaque, now serves as a warehouse for the Marine Biological Laboratory for storing preserved zoological specimens. About 1860, whaling became less profitable and Woods Hole entered into the second phase of its economic life which was dominated by the establishment and operation of a new commercial venture known as the Pacific Guano Works. During the years from 1863 to 1889, when the Pacific Guano Works was in operation the life of Woods Hole centered around the plant which was built at Long Neck near the entrance to what is known now as Penzance Point. Many large sailing vessels carrying sulphur from Italy, nitrate of soda from Chile, potash from Germany, and many schooners under the American flag loaded with guano and phosphorus from the Pacific Coast of South America were anchored in Great Harbor waiting for their turn to unload their cargoes. The number of laborers regularly employed by the Guano Company varied from 150 to 200 men, mostly Irishmen brought in under contract. Several local fishermen found additional employment as pilots for guano ships. The company maintained a store where various goods such as leather,

lead pipe, tin, coal, wood, and other items were bought and sold. The store acted also as a labor housing agency. Through efforts of the business manager of the Guano Company, the Old Colony Railroad was persuaded to extend its branch from Monument Beach to Woods Hoie. The establishment of well-organized and reliable transportation to Boston was an important factor in the future life of the community. The Pacific Guano Works was established by the shipping merchants of Boston who were seeking cargo for the return voyage of their ships. The guano deposits of one of the Pacific islands seemed to furnish this opportunity. As soon as the joint stock company was organized in 1859 with the capital of $1 million, arrangements were made almost immediately by which the newly formed concern came into possession and control of Howland Island. This island is located in the middle of the Pacific Ocean at longitude 177 deg. W., a short distance north of the Equator, about 1, 500 miles true south from Midway Island of the Hawaiian archipelago. At the same time appropriate plant and docking facilities were built at Woods Hole and 33 large sailing ships became available for hauling guano. Unlike the well-known guano islands off the coast of Peru, Howland Island is located in the zone of abundant rainfall. Consequently, the guano deposits of the island were leached of organic components and consisted of highly concentrated phosphate of lime. Fertilizer produced by the company was made by restoring the lost organic matter of the phosphate rock

1887 Map of Woods Holl (Fisheries is bottom left)

WOODS HOLE: THE EARLY YEARS

by adding the right proportion of organic constituents which were obtained from menhaden, pogy, and other industrial fish which abound in Cape Cod waters. The rock was pulverized and purified by washing; fish brought in by local fishermen were first pressed to extract oil, and the residue digested with sulphuric acid, washed, and dried. Acid was produced locally from sulphur imported from Sicily, and the digestion of fish flesh was carried out in large lead-lined vats. The plant was well equipped with machinery needed for the process and even had a chemical laboratory where chemists made the necessary analyses. Various sheds for storage and drying, barracks for laborers, and a business office completed the facilities. When the deposits of phosphate rock on Howland Island were exhausted, the company acquired title to the Greater and Lesser Swan Islands from the U.S. Government. These islands are located in the Caribbean Sea at latitude 17 deg. N. and longitude 83 deg. W. off the coast of Honduras. The islands are only 400 miles from Key West, Florida, and 500 miles from New Orleans. They contained good-quality phosphate rock and being much closer to Woods Hole greatly reduced the voyage time and cost of delivery. Further expansion of the company consisted in the acquisition of Chisolm’s Island near the coast of South Carolina, construction of a plant for cracking and washing phosphate rock on the Ball River side of the island, and establishment of a processing plant in Charleston, S.C. From the initial production (in 1865) of 7, 540 sacks of fertilizer weighing 200 pounds each, the output reached 11, 420 tons in 1871 and continued to grow until the combined annual production in 1879 of the works at Woods Hole and Charleston reached from 40, 000 to 45, 000 tons of guano fertilizer. Spencer Baird, Secretary of the Smithsonian Instution and first commissioner of the U.S. Commission of Fish and Fisheries arrived in Woods Hole in 1871. Baird was greatly impressed by the idea of utilizing menhaden and other fishes for the production of guano fertilizer and considered it a worthwhile project. In a letter dated October 18, 1875, to John M. Glidden, treasurer of the Pacific Guano Works Company, Baird urged him ‘‘to make a display of your wares at the centennial (in Philadelphia),

141

as this is one of the most important interests in the United States.’’ He writes further that ‘‘there is no species (of fish) worked up elsewhere comparable to the movement with the menhaden, or pogy, as to numbers and the percentage of oil. The combination, too, of the pogy scrap with the South Carolina phosphates and the guanos of the West Indies and of the PacificA are also quite novel, and as being especially an American industry, are eminently worthy of full appreciation.’’ While the scientists, agriculturalists, and stockholders of the company thought very highly of the guano works, the existence of a malodorous plant was not appreciated by the residents of Woods Hole who suffered from a strongly offensive odor whenever the wind was from the west. Woods Hole might have continued to grow as one of the factory towns of Massachusetts but, fortunately for the progress of science and good fortune of its residents (except those who invested their savings in the shares of Pacific Guano Works), the company began to decline and became bankrupt in 1889. Cessation of business and heavy monetary losses brought financial disaster to many residents of Woods Hole. The gloom prevailing in the village after the closing of the guano works began to dissipate, however, with the development of Woods Hole as a place of scientific research and with the increasing tourist trade. The factory buildings were torn down, the chimney which dominated the Woods Hole landscape was dynamited, and over 100,000 pounds of lead lining the acid chambers were salvaged. Large cement vats and the remnants of the old wharf remained; in the following years the latter became a favored place for summer biologists to collect interesting marine animals and plants. The years from 1871 to the death of Baird in 1887 were the formative period of the new era of Woods Hole as a scientific center. In historical documents and in old books the present name Woods Hole is spelled in a different way. The old name ‘‘Woods Holl’’ is considered by some historians of Cape Cod to be a relic of the times prior to the 17th century when the Norsemen visited the coast. The ‘‘Holl’’, supposed to be the Norse word for ‘‘hill’’, is found in the old records. The early settlers gave the name ‘‘Hole’’ to inlets or to passages between the islands, such as ‘‘Robinson’s Hole’’ between Naushon and Pasque Islands, or ‘‘Quick’s Hole’’ between Pasque and Nashawena Islands, and Woods’ Hole between the mainland and Nonamesset Island. In 1877 the Postmaster General ordered the restoration of the original spelling ‘‘Wood’s Holl’’, which remained in force until 1896 when the United States Post Office changed it back to Woods

142

WOODS HOLE: THE EARLY YEARS

Hole and eliminated the apostrophe in Wood’s. The change was regretted by the old timers and by C. O. Whitman who had given the specific name ‘‘hollensis’’ to some local animals he described. At the time of his arrival at Woods Hole in 1871, Baird was well known to the scientific circles of this country and abroad as a naturalist, student of classification and distribution of mammals and birds, and as a tireless collector of zoological specimens. He maintained voluminous correspondence with the scientists in the United States and Europe, and was Permanent Secretary of the recently organized American Association for the Advancement of Science. To the general public he was known as a contributor to a science column in the New York Herald and author of many popular magazine articles. His newly acquired responsibilities as Commissioner of Fisheries greatly added to his primary duties as Assistant Secretary of the Smithsonian Institution which was primarily responsible for the establishment of the National Museum in Washington. As a scientist, Baird belonged to the time of Louis Agassiz, Th. H. Huxley, and Charles Darwin. Like Agassiz he attended medical college but never completed his studies, although the degree of M. D. honoris causa was later conferred upon him by the Philadelphia Medical College. In the words of Charles F. Holder, ‘‘he was a typical American of the heroic type. A man of many parts, virtues, and intellectual graces, and of all the zoologists science has given the world.. . . he was most prolific in works of practical value to man and humanity.’’ Commissioner Baird attended many Congressional hearings and conferences with state officials and fishermen at which the probable causes of the decline of fisheries were discussed and various corrective measures suggested. From the lengthy and frequently heated discussions and evidence presented by the fishermen and other persons familiar with the fisheries problems, he became convinced that an alarmingly rapid decrease in the catches of fish had continued for the last 15 or 20 years. Such a decline was particularly noticeable in the case of scup, tautog, and sea bass in the waters of Vineyard Sound. It was logical, therefore, that the new Commissioner of Fisheries would select for his initial activities the New England

coastal area where the fishing industry was of greatest importance as a politico-economical factor. Woods Hole, however, was not a significant fishing center. In the ‘‘Fisheries and Fishlng Industry of the United States’’ prepared and edited by Goode (1884–87) for the 1880 Census, the fishing activity at Woods Hole is described in the following words: ‘‘Of the male inhabitants only seven are regularly engaged in fishing, the remainder being employed in the guano factory, in farming and other minor pursuits.. . . There is one ship carpenter in Wood’s Holl, but he finds employment in his legitimate business only at long intervals. Of sailmakers, riggers, caulkers, and other artisans there are none. Four men are employed by Mr. Spindel, during the height of the fishing season, in icing and boxing fish. The boat fishery is carried on by seven men from April until September, inclusive. Only three species of fish are usually taken, namely, scup, tautog, and sea bass. The total catch of each fisherman is about 15 barrels, or about 2400 pounds. In addition about 6,720 lobsters are annually taken.’’ Before selecting a location for permanent headquarters for the work on fishery management and conservation, Baird undertook extensive explorations of the fishing grounds off the entire New England Coast. Section 2 of the Joint Resolution Number 8 of Congress gave the Commissioner full authority to carry out the necessary research. In part it reads as follows ‘‘and further resolved, That it shall be the duty of the said Commissioner to prosecute investigations and inquiries on the subject, with the view of ascertaining whether any and what diminution in the number of the food-fishes of the coast and the lakes of the United States has taken place; and, if so, to what causes the same is due; and also, whether any and what protective, prohibitory, or precautionary measures should be adopted in the premises; and to report upon the same

WOODS HOLE: THE EARLY YEARS

to Congress.’’ Section 4 of the same Resolution contains an important clause which authorizes the Commissioner of Fisheries ‘‘to take or cause to be taken, at all times, in the waters of the seacoast of the United States, where the tide ebbs and flows, and also in the waters of the lakes, such fish or specimens thereof as many in his judgement, from time to time, be needful or proper for the conduct of his duties as aforesaid, any law, custom, or useage of any State to the contrary notwithstanding.’’ The significant words ‘‘where the tide ebbs and flows’’ were interpreted by Baird in a very broad scientific sense which extended the authority for his investigations to the offshore areas of the open ocean. Pounds and weirs were most frequently accused by the public as destructive methods of fishing responsible for the decline in the abundance of food fishes along the coast. Although Baird gave very serious consideration to the possible destructiveness of fixed nets, traps, pounds, pots, fish weirs, and other stationary apparatus, he was fully aware of the complexity of the factors which may cause the decline in fish populations. He discusses this difficult problem in a paper entitled ‘‘Report on the condition of the sea fisheries of the south coast of New England’’ and published as the first section of the voluminous First Report of the Commissioner of Fish and Fisheries for 1871. Of the causes which may have contributed to the decrease of summer shore fisheries of the south side of Massachusetts and Rhode Island, a fact which he considered as well established by the testimonies of competent persons, he lists the following: (1) decrease or disappearance of the food of commercial fishes; (2) migration of fishes to other localities; (3) epidemic diseases and ‘‘peculiar atmospheric agencies, such as heat, cold, etc.’’; (4) destruction by other fishes; (5) man’s activities resulting in the pollution of water, in overfishing, and the use of improper apparatus. The biologist of today will recognize in this statement Baird’s broad philosophical approach to the major problem of fishery biology. The outlined program combined oceanographical and meteorological investigations with the studies of biology, ecology, parasitology, and population dynamics of various fish species. Baird’s program of research is as comprehensive and valid today as it was 90 years ago. No time was lost in initiating this program. Woods Hole was selected as the base of the sea coast operations during the first summer and Vinal N. Edwards became the first permanent federal employee of the fisheries service. In spite of the insignificance of local fisheries, this locality offered a number of advantages which were recognized by Baird. Communication with Boston, New York, and Washington was good and promised to be better with the expected opening of the railroad branch in 1872. Being centrally located in relation to principal fishing grounds of New England and having good dock facilities and water of sufficient depth for sea going vessels, Woods Hole was a suitable base for visiting the offshore grounds. Furthermore, it was believed that the alleged decrease in food fishes was most clearly manifested in the region around Vineyard Sound. The small yacht Mazeppa of the New Bedford Custom House and the revenue-cutter

143

Moccasin attached to the custom-house at Newport, R.I., were placed at the disposal of Baird; and the LightHouse Board granted permission to occupy some vacant buildings and the wharf at the buoy-station on the west bank of Little Harbor. The Secretary of the Navy came to Baird’s assistance by placing at his command a small steam launch which belonged to the Boston Navy Yard and by giving many condemned powder tanks which could be used for the preservation of specimens. Nets, dredges, tanks, and other gear were provided by the Smithsonian Institution. Cooperation of the various governmental agencies was authorized by Congress which in Section 3 of the Resolution specified that ‘‘the heads of the Executive Departments be, and they are hereby directed to cause to be rendered all necessary and practicable aid to the said Commissioner in the prosecution of the investigations and inquiries aforesaid.’’ This provision of the law was of great value. It is apparent, however, that the success in obtaining cooperation authorized by law depended a great deal on the personal characteristics of Baird, his great ability of getting along with people, and his remarkable power of persuasion, These qualifications played the major role in his success in organizing the Commission’s work and also in obtaining the cooperation of scientists as well as that of fishermen and businessmen. The investigation during the first summer consisted primarily in collecting large numbers of fishes and studying their spawning, rate of growth, distribution, and food. In the course of this work nearly all the fish pounds and traps, some 30 in number, in the vicinity of Woods Hole, were visited and their location recorded. There was no difficulty in obtaining the owners’ permission to examine these installations and to collect the needed

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THE INTERNATIONAL MARITIME ORGANIZATION–LONDON CONVENTION ANNUAL OCEAN DUMPING REPORTS

specimens. Altogether 106 species of fish were secured, photographed, and preserved for the National Museum. Of this number 20 or more species had not previously been known from Massachusetts waters. Information gained in this manner was supplemented by the testimonies of various fishermen who presented their ideas either for or against the use of traps and pounds. Among them was Isaiah Spindel, who at the request of Baird, prepared a description of a pound net used at Woods Hole and explained its operation. In the following years Spindel became an influential member of the group of local citizens who supported Baird’s plan of establishing a permanent marine station at Woods Hole. The ship Moccasin under the command of J. G. Baker was engaged in taking samples of plankton animals, in determining the extent of beds of mussels, starfish, and other bottom invertebrates, and in making temperature observations. One of the principal collaborators in the studies conducted at Woods Hole in 1871 was A. E. Verrill of Yale University, a professor whom Baird appointed as his assistant and placed in charge of the investigations of marine invertebrates. Dredging for bottom animals during the first summer was carried out on a relatively small scale from a chartered sailing yacht Mollie and a smaller vessel used in the immediate vicinity of Woods Hole. Extensive collections were made by wading on tidal flats exposed at low water. Zoological work attracted considerable interest among the biologists of this country. Many of them stopped at Woods Hole for greater or lesser periods and were encouraged by Baird to use the facilities of the Fish Commission. The group included such well known men as L. Agassiz, A. Hyatt, W. G. Farlow, Theodore Gill, Gruyure Jeffries of England, and many others. The first year’s work extended until the early part of October. Before returning to Washington, Baird commissioned Vinal N. Edwards of Woods Hole to continue the investigation as far as possible. By the end of the first year a general plan of study of the natural histories of the fishes and the effect of fishing on fish populations was prepared with the assistance of the well-known ichthyologist, Theodore N. Gill. His old ‘‘Catalogue of the fishes of the Eastern Coast of North America from Greenland to Georgia’’, was revised and the next text including the recently collected data concerning the Massachusetts fishes, appeared in the First Report of the U.S. Commissioner of Fish and Fisheries. The plan of investigation suggested by Gill was adopted by Baird as a guide for the work of his associates for the purpose of ‘‘securing greater precision in the inquiries.’’ The plan is composed of 15 sections, such as Geographical distribution, Abundiance, Reproduction, etc., with detailed subdivisions under each one. A questionnaire containing 88 different items was included in order to facilitate the inquiries conducted among the fishermen. The scope of the highly comprehensive program is complete enough to be useful today; marine biologists of today would probably only rephrase it, using modern terminology. During the first year of operations conducted at Woods Hole, Baird and his associates laid down the foundation of the new branch of science which we now call fishery biology or fishery science.

AN ANALYSIS OF THE INTERNATIONAL MARITIME ORGANIZATION–LONDON CONVENTION ANNUAL OCEAN DUMPING REPORTS CHRISTINE DICKENSON IVER W. DUEDALL Florida Institute of Technology Melbourne, Florida

Trends are analyzed for types and quantities of permitted wastes, primarily dredged material, sewage sludge, and industrial waste, to be dumped at sea by member countries to the London Convention (LC) from 1992 and 1995 through 1998. In 1972, the Inter-Governmental Conference on the Dumping of Wastes at Sea led to the creation of the London Dumping Convention (now called the London Convention) to help regulate the dumping of wastes at sea. The act of dumping, as defined by the LC, is international disposal at sea of any material and in any form, from vessels, aircraft, platforms, or other artificial structures. The first consultative meeting of the LC contracting parties was held in 1976 by the InterGovernmental Maritime Consultative Organization [now called the International Maritime Organization (IMO)]. During this meeting, the procedure for the reporting of permits issued, on an annual basis, for dumping at sea by contracting parties, was determined. The IMOLC annual reports on permitted wastes list the number of permits issued by member countries, the types and quantities of wastes permitted for dumping at sea, and the location and designation of dump sites. Now with nearly 25 years of dumping records available, we are able to see trends in permitted dumping activity. In 1976, the first year of permitted dumping records, the combined amount of permitted wastes was nearly 150 million tons. In the last four years (1995 through 1998), where data are available, the total amount of wastes permitted to be dumped by the LC contracting parties as between 300 and 350 million tons. Currently, a majority of these wastes are being disposed of in the East Asian Seas and the North Sea and the largest quantity of waste being dumped is dredged material. Although the disposal of wastes at sea is considered to be a major issue, it is only responsible for about 10% of the total anthropogenic contaminants entering the ocean. Unfortunately, the longterm impacts of this dumping in the ocean are still largely unknown. INTRODUCTION ‘‘Historically, most coastal countries used the sea for waste disposal. It was generally the most economic way to manage waste, since land usually had, and still has, a high price tag while the sea has no private owner in the normal sense. In addition, dilution processes served the illusion that dumping at sea does not cause any permanent damage. So why risk contaminating land or drinking water with wastes if the sea was close by?’’ (1). ‘‘Accurate

THE INTERNATIONAL MARITIME ORGANIZATION–LONDON CONVENTION ANNUAL OCEAN DUMPING REPORTS

worldwide records on the amounts of wastes disposed at sea prior to 1976 are virtually impossible to obtain’’ (1). However, as a result of the international activities leading to the creation of the London Dumping Convention (LC) in 1972, information is now available on the number of permits issued by many countries for disposal at sea, their dumpsite locations, and the kinds and quantities of wastes that have been dumped (1). According to the LC, the most common form of ocean dumping today is disposal from ships or barges (1,2). Wastes are loaded on these vessels and then taken to the dumpsites. Dumpsites are chosen based on the kind of waste and the ocean’s properties at each site (2). Liquid wastes are generally disposed of in more dispersive environments, where mixing will rapidly dilute the dumped material. Solid wastes, on the other hand, are usually disposed of in less dispersive near-shore sites to keep the solids confined. Here we specifically analyze trends in types and quantities of permitted wastes to be dumped at sea by member countries to the LC from 1992 and 1995 through 1998. The types of wastes that have been dumped in the ocean include industrial waste, sewage sludge, dredged material, incineration at sea, and radioactive wastes. Industrial waste may include both liquid and solid wastes and it may contain such items as acid-iron waste, fishprocessing liquids, metal refinery wastes, and gas pipeline flushing wastes (2). An overall reduction in the dumping of these types of wastes has been achieved over the years by switching to alternative disposal methods, reusing wastes, and creating cleaner production technologies. The dumping of industrial wastes at sea has been prohibited since 1996 (3). ‘‘Sewage sludge is an anaerobic waste product from treatment of municipal wastewater. The sludge is in aqueous form containing about 3% suspended particles by weight’’ (2). Alternatives for the disposal of sewage sludge at sea include incineration, deposit on land, and agriculture use (3). ‘‘Dredged materials range from clean sands to heavily contaminated fine grained materials’’ (2). ‘‘Physical properties of dredged materials, including grain size, bulk density, water content, and geotechnical characteristics, are especially variable due to the kind or type of sediment being dredged, which is itself dependent on geological and watershed characteristics, as well as to the operational procedures used in dredging and disposal’’ (2). Incineration at sea is defined as the burning of liquid chlorinated hydrocarbons as well as other halogenated compounds in which all ash is dumped into the sea (3). This type of dumping was phased out early in 1991. The dumping of radioactive wastes, however, might be the most harmful practice of all. Although the dumping of high level radioactive wastes has never been allowed under the London Convention, it has still occurred in some cases, and even though a moratorium was placed on the dumping of low level radioactive wastes in 1983, this type of dumping still occurred (3,4). It was not until 1994 that this act became legally binding. Finally, the other waste category includes such wastes as inert geological materials, decommissioned vessels, scrap metals, and fish wastes (3).

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METHODS Annual reports on permitted dumping at sea by contracting parties were obtained from the IMO-LC (5). There is approximately a three-year delay before the most recent year’s records are put into report format and published. Older records that only existed in hard copy form had to be entered into Excel spreadsheets to create a new database. Early on, it was decided that all wastes would be reported in tons to make comparisons from year to year easier. Many of the older reports had wastes recorded in cubic meters. These records were converted into tons by multiplying the cubic meters by the density of the waste (1.3 g/cm3 for dredged material and other varying factors as indicated by each permit). Newer records, however, were received in Excel format and, for the most part, they were already converted into tons, making them very easy to add to the database. The primary categories of wastes of importance to this study include dredged material, sewage sludge, industrial waste, and other matter. From the 1992 and 1995 through 1998 reports (5), it was desired to find out which waste was being dumped in the largest quantity, as well as which countries were dumping the most wastes and which water bodies were receiving the most wastes. The more recent data is compared to earlier reports from 1976 to 1985. Comparisons of the different sources of pollutants in the oceans and comparisons between land dumping and sea dumping are also made. Then the past, as well as the future, of ocean dumping can be assessed. RESULTS Tables in the IMO-LC annual reports (5) on permitted wastes (IMO, 1992, 1995–1998) list the number of permits issued to member countries, the types and quantities of wastes permitted for disposal at sea, and the location and designation of dumping sites. It is important to know that these reports are reflective of what has been permitted to be dumped and not of what has actually been dumped. The accuracy of these records, therefore, is somewhat questionable (1). It is also important to remember that not all of the contracting parties report their activities every year and some reports from contracting parties may be incomplete (2). For example, in 1995, thirty-eight of the seventy-five contracting parties registered did not report their activities. There is also no way of knowing how much noncontracting parties are dumping (2). In 1992, 473 million tons were reported as permitted for disposal at sea. In the last four years (1995 through 1998) where data are available, the total amount of wastes permitted to be dumped by the LC contracting parties was 351 million tons in 1995, 312 million tons in 1996, 309 million tons in 1997, and 348 million tons in 1998. Figure 1 illustrates these data. As seen in Fig. 2, a majority of these wastes are being disposed of in the East Asian Seas (Fig. 3) and the North Sea (Fig. 4). The ‘‘other areas’’ category is represented mostly by the United States and its disposal in the Gulf of Mexico (Fig. 5).

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THE INTERNATIONAL MARITIME ORGANIZATION–LONDON CONVENTION ANNUAL OCEAN DUMPING REPORTS Total tons permitted for disposal at sea per year

500,000,000 400,000,000 300,000,000 200,000,000 100,000,000 0

Year 1992

1995

1996

1997

1998

Figure 1. London Convention permitted dumping from 1992 and 1995 through 1998.

The countries responsible for dumping the most wastes, in descending order, are China, the United States, Hong Kong, the United Kingdom, Germany, and Belgium (Table 1). As shown in Fig. 6, the largest quantity of waste being dumped is dredged material. At 327 million tons in 1998, it is the largest quantity by far, with sewage sludge coming in as the next highest quantity with only 15 million tons. DISCUSSION In 1976, the first year of permitted dumping records, the combined amount of permitted wastes was nearly 150 million tons (1). By 1985, this amount had grown to approximately 300 million tons (1). This apparent doubling, however, partly corresponds to an increase in the number of contracting parties to the LC. It is also important to remember that not all of the contracting parties report their activities every year and some reports from contracting parties may be incomplete. Therefore, it is not possible to provide a highly accurate interpretation of these data and the reader should be cautioned that any analysis must be considered approximate.

Figure 3. China and Hong Kong dumping in the East Asian Seas. (From Ref. 6).

Looking back to the data in Fig. 1, waste disposal in the ocean seems to continue to increase into the early 1990s with 473 million tons in 1992. However, by 1995 through 1998 this amount had leveled off to between 300 and 350 million tons. This decrease could be representative of the many changes in policy occurring within the LC in the 1990s to intentionally decrease the amount of waste disposed of in the ocean (3). It is also known that a major dredging project took place in Hong Kong in the early 1990s in order to expand the airport there (3). This may explain why the amount of waste dumped in 1992 was so high. The East Asian Seas section of Fig. 2 supports this theory, with data from 1992 being much higher than in the following years. Yearly fluctuations like this occur due to the variation in maintenance dredging and new works associated with shipping activities (3). According to the IMO-LC web page, ‘‘in early 1991, incineration at sea operations came to a halt, ahead of

Permitted quantities (in tons) per location 250,000,000 200,000,000 North, central, & south america

East asia & australasia

North east atlantic & adjacent waters 150,000,000 100,000,000 50,000,000 0 Atlantic ocean

North English sea channel

Irish sea

Other areas

1992

South west pacific 1995

East asian seas 1996

Indian ocean

1997

Other NW NE areas Atlantic Pacific ocean ocean 1998

Figure 2. Locations of dumping sites with their respective amounts of wastes dumped.

Arctic seas

Other areas

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Table 1. Highest Combined Tons for 1992 and 1995–1998

Figure 4. The United Kingdom, Germany, and Belgium dumping in the North Sea. (From Ref. 6).

the agreed global deadline of December 31, 1992. . .. In 1990, contracting parties to the LC agreed to phase-out sea disposal of industrial waste effective by January 1, 1996. . .. In 1993, contracting parties started a detailed

Rank

Country

Tons

1 2 3 4 5 6

China United States Hong Kong United Kingdom Germany Belgium

295,482,060 293,272,190 285,437,056 246,759,318 132,149,000 127,507,436

review of the LC, leading to the adoption of a few crucial amendments. . .. These amendments consolidated in a legally binding manner the prohibition to dump all radioactive wastes or other radioactive matter and of industrial wastes, the latter as per January 1, 1996, as well as the prohibition of incineration at sea of industrial wastes and of sewage sludge’’ (3). According to reports by contracting parties, no permits for the dumping of industrial waste have been issued since 1996. Before this phase-out, Japan and the Republic of Korea were responsible for most of the industrial waste being dumped (3). The amount of sewage sludge being dumped at sea decreased in the early 1990s, reflecting the phase-out of this practice by several countries, Ireland and the United Kingdom being the most recent. Currently, only Japan, the Philippines, and the Republic of Korea dispose of sewage sludge at sea (3).

Figure 5. The United States dumping in the Gulf of Mexico and the northwest Atlantic. (From Ref. 6).

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Waste category comparisons

500,000,000 450,000,000 400,000,000 350,000,000 300,000,000 250,000,000 200,000,000 150,000,000 100,000,000 50,000,000 0 1992

1995

1996

Figure 6. Dredged material, by far, makes up the largest quantity of waste being dumped at sea.

Dredged material makes up the majority of what is being dumped at sea (see Fig. 6). Unlike other wastes, the amount of dredged material dumped each year tends to fluctuate greatly due to variable maintenance dredging and new works associated with shipping activities (3). For example, there were huge dredging efforts going on in Hong Kong with the extension of the airport there in 1992. One hundred forty-four million tons of dredged material were permitted to be disposed of at sea that year by Hong Kong. Arguably, this could be responsible for the 1994 anomaly of the combined total of 473 million tons, meaning that the average amount of wastes dumped annually actually leveled out to between 300 and 350 million tons much earlier than the later 1990s. The interactions of the wastes dumped at sea with seawater and the toxicity of these wastes to organisms, including humans, are of great importance to the scientific community. Although the long-term effects of disposal at sea are unknown, many studies have looked at some of the short-term effects, particularly in relation to radioactive wastes that have been dumped. Studies conducted in the Arctic by the IAEA and the U.S. Congress Office of Technology Assessment found that (1) releases from dumped objects were small and confined to the immediate vicinity of the dumped objects, (2) projected future doses to the general public were small, (3) doses to marine organisms were insignificant in the context of populations, and (4) remediation on radiological grounds was not warranted (7). A study by Hill et al. (8) assesses the possible effects of 40 years of dredging, and how it might have affected benthos in the Irish Sea. Zooplankton and micronekton communities, off eastern Tasmania, as studied by Bradford et al. (9), were found to be affected by the presence of a warm-core eddy instead of jarosite wastes dumped within the vicinity. As de La Fayette (7) said, it is difficult to draw conclusions from data that has been collected because more research, monitoring, and prevention projects are still needed for us to understand all of the factors that are involved in knowing how, when, and where ocean stored wastes might affect us and the environment.

Dredged material Sewage sludge Other matter Industrial waste 1997

1998

Year

Although such concern is felt about disposal at sea, it is not the largest contributor of pollutants to the oceans. As seen in Fig. 7, the IMO-LC puts it in fourth place, making up only 10% of the total pollutants in the oceans. Runoff and land-based discharges are the largest source, making up 44%. Even maritime transportation, at an estimated 12%, pollutes the oceans more than dumping does. In an attempt to compare ocean dumping to disposal in landfills, we look at the United States. According to Zero Waste America (10), in 2001, 409 million tons of municipal waste were generated in the United States. Of that, 278 million tons were disposed of in landfills. Table 2 illustrates dumping on land and in the ocean by the United States during 1992 and 1995 through 1998. Those five years of ocean dumping, totaling 293 million tons, is only 15 million tons more than what was disposed of in landfills in 2001 alone. In conclusion, we note that much more waste is being disposed of in landfills than in the ocean. Therefore, shouldn’t we be more concerned about the land? CONCLUSION Although the environmental impacts of dumping at sea are still largely unknown, it is comforting that the LC is moving in a positive direction toward more regulations for the better protection of our oceans. With industrial

Dumping in relation to other sources of pollutants in the oceans 10% 1% Run-off and land-based discharges

12% 44%

Land-based discharges through the atmosphere Maritime transportation Dumping Offshore productions

33%

Figure 7. At 10%, dumping ranks fourth compared to the other sources of pollutants in the ocean.

MARINE SOURCES OF HALOCARBONS Table 2. Ocean Dumping versus Land Dumping in the United States Tons in Millions Year

Land

Ocean

1992 1995 1996 1997 1998

241 249 238 236 238

67 58 46 53 69

Data taken from Reference 10.

waste already phased out and sewage sludge on its way to becoming completely phased out, that leaves dredged material and ‘‘other matter’’ as the future of ocean dumping. Other matter mostly makes up inert geological material from mining and excavations; bulky wastes such as steel equipment, scrap metal, and concrete; fish wastes; obsolete ammunition and explosives; discontinued oil platforms; spoiled food; and other random wastes. This category may also eventually be phased out, but dredged material will most likely continue to be dumped since this form of waste, in the normal sense, either came ‘‘from the ocean floor’’ or somewhere close to it being that twothirds of dredged material is connected with maintenance operations to keep harbors, rivers, and other waterways from being blocked up. Unfortunately, according to the IMO-LC, about 10% of dredged material is moderately to heavily contaminated from a variety of sources including shipping, industrial and municipal discharges, and land runoff. Whether we find somewhere to dispose of dredged material on land, or we continue to dispose of it in the sea, it will continue to affect our environment. Acknowledgments We are grateful to Rene Coenen, Office of the London Convention, International Maritime Organization, for kindly providing the LC dumping records. We are also thankful to Ruth Caulk and Melissa Sheffer for assisting in the preparation of the ocean dumping spreadsheets.

BIBLIOGRAPHY 1. Duedall, I.W. (1990). A brief history of ocean disposal. Oceanus 33(2): 29. 2. Connell, D.W. and Hawker, D.W. (1992). Pollution in Tropical Aquatic Systems. CRC Press, Boca Raton, FL. 3. London Convention. (2003). Available: http://www.londonconvention.org. 4. Ahnert, A. and Borowski, C. (2000). Environmental risk assessment of anthropogenic activity in the deep sea. J. Aquat. Ecosys. Stress Recov. 7: 299–315. 5. IMO. (1972). Convention on the Prevention of Marine Pollution by Dumping of Wastes and Other Matter, London Convention 1972. Final Reports on Permits Issued in 1992, 1995, 1996, 1997, and 1998. 6. Karls, F. (1999). World History: The Human Experience—The Modern Era. Glencoe/McGraw-Hill, New York. 7. de La Fayette, L. (1998). The London Convention 1972: preparing for the future. Int. J. Mar. Coastal Law 13(4): 515–536.

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8. Hill, A.S. et al. (1999). Changes in Irish Sea benthos: possible effects of 40 years of dredging. Estuarine Coastal Shelf Sci. 48: 739–750. 9. Bradford, X. et al. (1999). 10. Zero Waste America. (2001). Available: http://www.zerowasteamerica.org/Statistics.htm.

MARINE SOURCES OF HALOCARBONS ROBERT M. MOORE Dalhousie University Halifax, Nova Scotia, Canada

Halogenated gases are important in the atmosphere by virtue of their ability to carry chlorine, bromine, and iodine to high altitude, where they can act as efficient catalysts of ozone destruction. The effect on ozone of chlorine release in the stratosphere by certain manufactured chlorofluorocarbons is well known, but naturally occurring halogen carriers account for a part of the background cycle of ozone breakdown. The oceans constitute an enormous reservoir of dissolved halogens in the form of halide ions, but the transport of this material to the upper atmosphere is inefficient because sea salt particles, introduced to the atmosphere as sea spray from breaking waves and bubblebursting, have short lifetimes, being readily washed out by rain. For this reason the relatively minute concentrations of dissolved halogenated gases, such as methyl chloride occurring in surface seawater at concentrations around 10−10 M, have the potential to drive significant transport of chlorine into the upper atmosphere. Since the role of anthropogenic chlorine and bromine gases in stratospheric ozone depletion was recognized, there has been renewed interest in the ocean as a source of halogenated trace gases. The emission of iodine compounds from the ocean has more recently been shown to be potentially important in aerosol production as well as in affecting ozone concentrations over the ocean (1). The fluxes of trace gases such as those between the ocean and atmosphere is most commonly calculated from concentration measurements in the surface ocean and the overlying atmosphere together with empirical relationships between gas exchange coefficients and wind velocity (2). Supersaturation of a gas at the sea surface with respect to the atmosphere will support an outward flux, the magnitude of which is strongly dependent on wind speed, and to a lesser extent on temperature. Gas fluxes can be highly variable spatially and seasonally, and, in general, estimated global fluxes will have a substantial degree of uncertainty due to the sparseness of the concentration measurements combined with a large uncertainty in the exchange velocity. Ideally, gas fluxes would be calculable based on a firm understanding of the processes governing both the production and loss of the gas in the upper ocean and the processes controlling gas exchange. This remains a distant goal, and current studies are directed at developing an understanding of these processes. Some of the first measurements made of halogenated methanes in seawater suggested that the ocean could be

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MARINE SOURCES OF HALOCARBONS

the predominant source of methyl chloride accounting for most of the 4 million tonnes that must be added annually to the atmosphere to support the observed concentrations (3). More recently, it has been demonstrated that, in the case of methyl chloride, the ocean is just one of many sources (including biomass burning, woodrot fungi, and vegetation) and can probably account for only about 15% of the total flux to the atmosphere (4). In contrast, the ocean is still recognized to be the most important source of atmospheric methyl iodide. Impetus for the study of sources of methyl bromide came with the discovery of the efficiency with which bromine could catalyze stratospheric ozone loss, working either directly or in concert with chlorine. This led to concern about the possible role of methyl bromide, manufactured for use as a fumigant, in observed ozone loss. The existence of natural sources and sinks for methyl bromide made it more difficult to quantify the deleterious effect of the anthropogenic component. Extensive studies of the distribution of methyl bromide in the oceans and of its flux across the air–sea interface in different regions demonstrated the existence of both production and loss processes within the ocean (5). The balance between these determines whether a particular region will emit CH3 Br to the atmosphere or take it up. A major oceanic sink results from the reaction between CH3 Br and chloride ion, the rate being strongly dependent on temperature. Other losses include chemical hydrolysis and also biological uptake. Bromine is carried to the atmosphere by many other gases produced in the ocean, notably bromoform and dibromomethane (6). Early studies pointed to an association between certain halogenated compounds in seawater, such as methyl iodide, and coastal beds of macrophytes. The more general phenomenon of organohalogen production by seaweeds has long been known, and a wide range of halogenated organic compounds of varying molecular weight and complexity have been identified (7). The focus here is on halogenated compounds that play a role in atmospheric chemistry, so these are typically volatile compounds. Two such compounds that have been shown to have major sources in macrophytes are di- and tribromomethane (CH2 Br2 and CHBr3 ). The latter has been found to occur within certain seaweeds at concentrations up to the percent level. It is apparent that haloperoxidase enzymes are involved in the biosynthesis of organohalogens by macrophytes. In spite of the variety and abundance of halogenated methanes in seaweeds and the flux of halogen to the atmosphere that can thus be supported in the coastal zone, the overall importance of these algae on an oceanwide scale is small on account of their limited distribution (8), although, in the case of bromoform and dibromomethane, Carpenter and Liss, (9) estimated that macrophytes are a major source. For this reason there has been interest in the possible production of the same compounds by marine microalgae, which are ubiquitous in the sunlit waters of the world’s oceans. An approach that has been adopted with some success to study this question has been the use of laboratory cultures of specific algae. This has demonstrated the capacity of a range of phytoplankton of

different algal classes to produce compounds such as CH3 I, CH2 Br2 , CHBr3 , and CH2 I2 (10,11), although species vary in the spectrum of compounds that they produce as well as in the production rates. There has been some progress in elucidating the production mechanism of these compounds with the identification of haloperoxidase enzymes in a few phytoplankton species (11). However, much uncertainty remains regarding the oceanic source of methyl chloride, bromide, and iodide. For although there are published studies that point to the capacity of some phytoplankton to produce methyl halides (12), the measured rates when normalized to biomass concentrations have typically been found to be quite inadequate to account for the observed fluxes from the ocean into the atmosphere (10). This may be due in part to the limitations of laboratory studies of phytoplankton cultures: relatively few species can be grown successfully in the laboratory and the conditions are very dissimilar to those in the ocean itself. Alternative possible explanations include the involvement of microbial processes (13), zooplankton, and photochemistry, or a combination of sources. One laboratory study has provided some evidence for the production of methyl iodide in seawater irradiated with simulated sunlight (14), but it has yet to be demonstrated that this process is significant in the ocean. Open ocean studies of halocarbon distributions show that there is no simple correlation with phytoplankton biomass, quantified as chlorophyll a concentration. This means that the production of a particular compound cannot be attributed uniformly to all species of phytoplankton. More success has been obtained in studies that measured a series of photosynthetic pigments and evidence is forthcoming from one such study that CH3 Br has a source in Prymnesiophytes (15). Measurements of dissolved methyl halides in the ocean typically show relatively high concentrations in the surface mixed layer, frequently with a maximum directly beneath, declining to levels at or near detection limits in deep waters (16). Such profiles are broadly consistent with a source at or near the surface and net consumption at greater depths. The maximum may be explained as occurring in a zone of production beneath the surface mixed layer that is poorly ventilated, so with reduced emission to the atmosphere. However, certain halocarbons have quite different distributions with depth. The chlorofluorocarbons such as CFC11 and CFC12 may have higher concentrations below the surface and at intermediate depths, with levels diminishing to low values in some of the deepest waters of the ocean. These distributions are well understood to be the result of transport of these anthropogenic gases from the atmosphere either by direct exchange at the surface or via downward mixing and advection of cold, dense, ventilated waters to intermediate and abyssal depths. There is now evidence that some halocarbons that have relatively short atmospheric lifetimes (e.g., dichloromethane and tri- and tetrachloroethylene) are also introduced to the ocean by the same processes, and that these gases have much longer lifetimes in the ocean than in the atmosphere. This apparently explains the presence of dichloromethane in the near bottom waters of the Labrador Sea (17).

FOOD CHAIN/FOODWEB/FOOD CYCLE

Chloroform is likely to undergo the same processes but may have an additional in situ source. Loss processes occurring in the ocean can be inferred for most organohalogens by their lower concentrations at depth than at the surface. They have been studied most thoroughly for methyl bromide and, for this compound, are known to include bacterial removal and chemical loss through hydrolysis and reaction with chloride ion (18). Other loss processes affecting a range of halogenated compounds apparently exist in waters that are depleted in dissolved oxygen (19). It should be noted a loss process for one halocarbon might represent a source for another: thus, reaction of both CH3 Br and CH3 I with chloride ion yields CH3 Cl. The photocatalyzed loss of CH2 I2 is a source of CH2 ICl. Continued measurement of halocarbons both in the ocean and atmosphere will improve our knowledge of the magnitude, geographic distribution, and seasonality of their fluxes, but a greater challenge is likely to be identifying more accurately their sources. That knowledge is essential if we hope to be able to predict how the fluxes of these atmospherically reactive gases may change in the future. BIBLIOGRAPHY 1. O’Dowd, C.D. et al. (2002). Marine aerosol formation from biogenic iodine emissions. Nature 417: 632–636. 2. Wanninkhof, R. (1992). Relationship between wind speed and gas exchange over the ocean. J. Geophys. Res. 97: 7373–7382. 3. Harper, D.B. (2000). The global chloromethane cycle: biosynthesis, biodegradation and metabolic role. Nat. Prod. Rep. 17: 337–348. 4. Moore, R.M. (2000). The solubility of a suite of low molecular weight organochlorine compounds in seawater and implications for estimating the marine source of methyl chloride to the atmosphere. Chemosphere: Global Change Sci. 2: 95–99. 5. Lobert, J.M. et al. (1995). Science 267: 1002–1005. 6. Quack, B. and Wallace, D.W. (2003). Air–sea flux of bromoform: controls, rates, and implications. Global Biogeochem. Cycles 17(1): 1023–1029. 7. Gribble, G.W. (2003). The diversity of naturally produced organohalogens. In: The Handbook of Environmental Chemistry. Vol. 3, Part P, G.W. Gribble (Ed.). Springer, New York, pp. 1–15. 8. Baker, J.M. et al. (2001). Emissions of CH3 Br, organochlorines and organoiodines from temperate macroalgae. Chemosphere Global Change Sci. 3: 93–106. 9. Carpenter, L.J. and Liss, P.S. (2000). On temperate sources of bromoform and other reactive organic bromine gases. J. Geophys. Res. 105: 20,539–20,547. 10. Manley, S.L. and de la Cuesta, J. (1997). Methyl iodide production from marine phytoplankton cultures. Limnol. Oceanogr. 42: 142–147. 11. Moore, R.M., Webb, M., Tokarczyk, R., and Wever, R. (1996). Bromoperoxidase and iodoperoxidase enzymes and production of halogenated methanes in marine diatom cultures. J. Geophys. Res. 101: 20,899–20,908. 12. Scarratt, M.G. and Moore, R.M. (1996). Production of methyl chloride and methyl bromide in laboratory cultures of marine phytoplankton. Mar. Chem. 54: 263–272.

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13. Amachi, S., Kamagata, Y., Kanagawa, T., and Muramatsu, Y. (2001). Bacteria mediate methylation of iodine in marine and terrestrial environments. Appl. Environ. Microbiol. 67: 2718–1722. 14. Moore, R.M. and Zafiriou, O.C. (1994). Photochemical production of methyl iodide in seawater. J. Geophys. Res. 99: 16,415–16,420. 15. Baker, J.M. et al. (1999). Biological production of methyl bromide in the coastal waters of the North Sea and open ocean of the northeast Atlantic. Mar. Chem. 64: 267–285. 16. Moore, R.M. and Groszko, W. (1999). Methyl iodide distribution in the ocean and fluxes to the atmosphere. J. Geophys. Res. 104: 11,163–11,171. 17. Moore, R.M. (2004). Dichloromethane in N. Atlantic Ocean waters. J. Geophys. Res. 109: 9004. 18. King, D.B. and Saltzman, E.S. (1997). Removal of methyl bromide in coastal seawater: chemical and biological rates. J. Geophys. Res. 102: 18,715–18,721. ¨ (1996). Reduction ¨ 19. Tanhua, T., Fogelqvist, E., and Basturk, O. of volatile halocarbons in anoxic seawater, results from a study in the Black Sea. Mar. Chem. 54: 159–170.

FOOD CHAIN/FOODWEB/FOOD CYCLE ANDREW JUHL Lamont–Doherty Earth Observatory of Columbia University Palisades, New York

A representation of the feeding relationships of the organisms, or groups of organisms, within an ecological community will be shown here. By showing which organisms feed on which other organisms, the pathways of energy flow through the ecosystem can be followed. ‘‘Food chain’’ is an older term and is currently less used than ‘‘foodweb’’ or ‘‘food cycle.’’ The newer terms were coined to acknowledge the complexity of feeding relationships within most ecosystems. Examples of marine planktonic and intertidal foodwebs will be used below to illustrate specific points, but the concepts are applicable to any ecosystem. The simplest food chain is organized in a strict hierarchy with primary producers (organisms that generate organic matter by fixing inorganic carbon, usually through photosynthesis) eaten by herbivores (primary consumers), the herbivores eaten by carnivores (secondary consumers), on to tertiary consumers, quaternary consumers, and soon. Each level within the food chain is termed a trophic level. Figure 1 shows an example of a very simple food chain for a planktonic ecosystem where phytoplankton are the primary producers, eaten by zooplankton herbivores. Zooplankton, in turn, are eaten by fish. Each group of organisms is represented by a shape, and the feeding connection and direction of energy flow are shown with an arrow. Simple food chains have heuristic value and are the basis of many quantitative ecosystem models. However, they can be criticized for oversimplification and for missing major groups of organisms. One step toward a more realistic picture of an ecosystem is to disaggregate the trophic levels. Ideally, each level

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Fish

Herring adults Herring larvae

Zooplankton Ciliates

Calanus copepods

Small phytoplankton

Large phytoplankton

Phytoplankton Figure 1. A simple food chain for a planktonic ecosystem.

could be broken down into its constituent species. One problem to consider is whether rare species should be depicted. A food chain diagram could include any feeding relationship that occurs within a system, whether common or unusual. This could lead to so many boxes and connecting arrows that any heuristic value would be jeopardized. In addition, the required information to fully disaggregate species groups may not be available. For example, microorganisms can be difficult to distinguish to the species level. Practical considerations often lead researchers to lump microorganisms into size classes and functional groups (e.g., primary producers >2 µm, 2–20 µm, 99.8% of the mass of the total dissolved salts in the ocean, are known as the major ions and are defined as dissolved species having a concentration in seawater of >1 mg kg−1 (1 ppm) (Table 1). Na+ and Cl− alone account for 86% of all dissolved chemical species in seawater. The source of Cl− , and most anions in seawater, is outgassing of the Earth’s interior, or volcanic emissions. The majority of the major cations are derived from the effects of continental weathering and are delivered to the ocean via rivers. The gross composition of seawater is the result of the partitioning of elements between continental rock and seawater over geologic time. The major ions represent the more soluble elements, which have preferentially partitioned into seawater. The major ions are found to occur in nearly constant ratios to each other throughout most of the world’s oceans. That is, although the salinity or the total amount of salt dissolved in seawater varies from location to location in the ocean, the ratios of the major ions to each other remains nearly constant, which is true from ocean to ocean as well as from surface to deep waters (Table 2). Table 1. Major Ions in Seawater

Chloride, Cl− Sodium, Na+ Sulfate, SO4 2− Magnesium, Mg2+ Calcium, Ca2+ Potassium, K+ Bicarbonate, HCO3 2− Bromide, Br− Boron, B (as B(OH)3 Strontium, Sr2+ Fluoride, F−

Table 2. Ion to Chloride Ratios in Various Oceans and Selected Seas for Na+ and K+ Ocean or Depth Interval Atlantic Pacific Indian Red Sea Mediterranean 0–100 m 700–1500 m >1500 m MEAN

Na+ /Cl−

K+ /Cl−

0.5552 0.5555 0.5554 0.5563 0.5557 0.5554 0.5557 0.5555 0.5555+ /−0.0007

0.0206 0.0206 0.0207 0.0206 0.0206 0.0206 0.0206 0.0206 0.0206+ /−0.0002

Also shown are ion to chloride ratios for selected depth intervals in the world ocean. Mean values are for the world ocean over all depth intervals.

MARIE DE ANGELIS

Major Ion

159

g kg−1 Seawater at Salinity of 35.000 ppt

% by wt

g ion/g Cl−

19.353 10.781 2.712 1.284 0.4119 0.399 0.126 0.0673 0.0257 0.00794 0.00130

55.30 30.77 7.75 3.69 1.18 1.14 0.41 0.19 0.013 0.023 0.0037

1 0.5561 0.1400 0.0668 0.0213 0.0206 0.0075 0.0035 0.00024 0.00041 0.00006

Concentrations are presented in terms of g of ion per kg of seawater, % by weight in seawater, and the weight ratio g ion per g Cl− .

This observation is known as ‘‘The Rule of Constant Proportions’’ or conservative behavior. As Cl− is the single most abundant ionic species in seawater, the Rule of Constant Proportion is usually expressed as the ratio of major ions to chloride ion (e.g., Na+ /Cl− ). Exceptions to this rule exist for calcium (Ca2+ ), strontium (Sr2+ ), and bicarbonate (HCO3 2− ) as a small fraction of the total concentration of these species participate in biological reactions resulting in slight variations of ion/Cl− ratios between surface and deep water. The residence time (τ ) of a chemical species in the ocean can be defined as the average time an individual atom for a given element remains in seawater before being permanently removed. Assuming that the major source of dissolved salts to the ocean is from riverine input, residence time of a given ion can be calculated as the total mass (in g) of ion in the ocean (Mocean ) divided by the mass of ion delivered by rivers annually (Fluxin ) (in g yr−1 ). τ (years) =

Mocean Fluxin

Flux in can be estimated from the average concentration of the ion in the world’s rivers (in g L−1 ) multiplied by the total volume of river water entering the ocean annually (in L yr−1 ). Mocean can be estimated from mean seawater concentrations (in g L−1 ) multiplied by the volume of the ocean (in L). The residence times for the major ions calculated by this method are extremely large, on the order of millions to hundreds of millions of years (Table 3). The long residence times of major ions reflects the relatively low reactivity of these chemical species. Except Table 3. Mean Seawater and River Water Concentrations and Residence Times for Several Major Ions Species Cl− Na+ Mg2+ SO4 2− K+ Ca2+ Br−

Mean SW (mg L−1 )

Mean RW (mg L−1 )

19, 350 10, 760 1, 294 2, 712 399 412 67

5.75 5.25 3.35 8.25 1.3 13.4 0.02

SW RW 3, 365 2, 090 386 338 307 31 3, 350

τ (106 years) 123 75 14 12 11 1.1 123

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for Ca2+ , Sr2+ , and HCO3 2− , no biological processes exist that accelerate the removal of these species to the oceanic sediment. The main removal mechanisms for major ions are kinetically slow reactions such as precipitation or incorporation into clay minerals in the sediment or oceanic crust. Thus, major ions are conservative or remain in constant proportion to each other because the major ions have very long residence times relative to either the residence time of water in the ocean (40,000 years) or the oceanic circulation time (1000 years). Even if the salt content (salinity) of seawater changes from location to location in the ocean, the ratios of the major ions do not change because they react on much longer time scales than the removal or addition of water by precipitation or evaporation. As the residence time of major ions greatly exceeds the average oceanic circulation time of approximately 1000 years, no change is seen in major ion concentrations from ocean to ocean or between surface water and deep water The major ion composition of river water and seawater are different (Table 3). The composition of present-day river water reflects the weathering of continental rock, whose composition reflects the long-term partitioning of chemical species between the continents and seawater. The composition of seawater with respect to its major ion composition is not believed to have changed over the past several million years despite the constant input of river water of different composition. The oceans are believed to be at steady-state equilibrium with respect to the major ions. That is, the composition of seawater remains constant because the flux of major ions into the ocean equals the flux of major ions out of the ocean. Although the major source of major ions is river input, hydrothermal circulation of seawater through oceanic basalt, particularly at relatively low temperatures, may be a significant source of some major cations (Na+ , K+ ). To achieve steady-state equilibrium, the sum of the sinks or fluxes out of major ions must balance the input from rivers. Several processes, most occurring at extremely slow rates over large areas of the oceanic seafloor, have been identified as sinks for major ions. These processes include: (a) Cation exchange in which clay minerals within the oceanic sediment as well as clays delivered to the oceans by rivers exchange cations in seawater to form new clay minerals relatively enriched in Na, K, and Mg at the expense of Ca. (b) Trapping and eventual burial of seawater within interstitial water of marine sediments. This process removes the more concentrated ions in seawater (Na, Cl). (c) Evaporite formation during some periods of Earth’s history, in which large deposits of minerals derived from seawater have formed when seawater was trapped and evaporated from shallow, closed basins. Although the areal extent of such basins is limited at the present time, formation of evaporites is an important removal mechanism for some major ions, including Na, Cl, and SO4 2− , over geologic time. (d) Sea spray transported to land can result in net removal of Na and Cl. (e) Reverse weathering involving reactions between ions in seawater and cation-poor aluminosilicates derived from continental weathering. This process results in the formation of new cation-rich clay minerals and CO2 , resulting in the net

removal of Na, K, and Mg from seawater. (f) Hydrothermal circulation involving the reaction of major ions between seawater and oceanic basalt is a net sink for Mg2+ and SO4 2− . READING LIST Mackenzie, F.T. and Garrels, R.M. (1966). Chemical balance between rivers and oceans. Amer. J. Sci. 264: 507–525. Holland, H.D. (1984). The Chemical Evolution of the Atmosphere and Oceans. Princeton University Press, Princeton, NJ, p. 584. Pilson, M.E.Q. (1998). An Introduction to the Chemistry of the Sea. Prentice-Hall, Englewood Cliffs, NJ, p. 431.

TSUNAMI Tsunami (also called Seismic Sea Wave, and popularly, Tidal Wave), an ocean wave produced by a submarine earthquake, landslide, or volcanic eruption. These waves may reach enormous dimensions and have sufficient energy to travel across entire oceans. From the area of the disturbance, the waves will travel outward in all directions, much like the ripples caused by throwing a rock into a pond. The time between wave crests may be from 5 to 90 minutes, and the wave speed in the open ocean will average 450 miles per hour. Tsunamis reaching heights of more than 100 feet have been recorded. As the waves approach the shallow coastal waters, they appear normal and the speed decreases. Then, as the tsunami nears the coastline, it may grow to great height and smash into the shore, causing much destruction. 1. Tsunamis are caused by an underwater disturbance—usually an undersea earthquake. Landslides, volcanic eruptions, and even meteorites can also generate a tsunami. 2. Tsunamis can originate hundreds or even thousands of miles away from coastal areas. Local geography may intensify the effect of a tsunami. Areas at greatest risk are less than 50 feet above sea level and within one mile of the shoreline. 3. People who are near the seashore during a strong earthquake should listen to a radio for a tsunami warning and be ready to evacuate at once to higher ground. 4. Rapid changes in the water level are an indication of an approaching tsunami. 5. Tsunamis arrive as a series of successive ‘‘crests’’ (high water levels) and ‘‘troughs’’ (low water levels). These successive crests and troughs can occur anywhere from 5 to 90 minutes apart. They usually occur 10 to 45 minutes apart. The Tsunami Warning System, a cooperative international organization and operated by the United States Weather Service, has been in operation since the 1940s. The headquarters of the center is located in Hawaii. An associated Alaska Regional Tsunami Warning System is located in Alaska. Tsunami prediction essentially

TSUNAMI

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PACIFIC

NORTH

SOUTH

PACIFIC

commences with earthquake monitoring and prediction information. Inputs from these systems are linked with information from a series of tide monitoring installations. Locations of tide stations and of seismograph stations in the Tsunami Warning System are shown in Fig. 1. When inputs indicate conditions are favorable for a tsunami, a watch is issued for the probable affected area. Warnings are issued when readings from various tidal stations appear to match the seismographic information. Because of the complexity of the factors involved and a large degree of uncertainty nearly always present, there is a tendency to issue watches and warnings as a safety precaution even though a tidal wave of significance may not develop. Unfortunately, after awhile, persons in likely areas to be affected grow callous to watches and tend to ignore them. A tsunami that hit Hilo, Hawaii in 1960 killed 60 residents even though they had been warned of the coming event an ample 6 hours in advance of the strike. This, however, was not the reason for the most devastating Tsunami in history which occurred on December 26, 2004 in the Indian Ocean as a result of a 9.0 earthquake off the shore of Banda Aceh, Indonesia. The resulting Tsunami quickly hit the Indonesian coastline, but hours later struck Sri Lanka, India, The Maldives, and later yet, Keyna and Somalia on the East coast of Africa. More than 150,000 lives were lost because no network existed to communicate the likely result of the 9.0 quake in the Indian Ocean. No such event had occurred in this area of the

Figure 1. Network of tide and seismograph stations that are part of the Tsunami Warning System, headquartered in Hawaii. (National Oceanic and Atmospheric Administration.).

world in over 120 years since Kakatoa erupted in 1883. Twenty three earthquake monitoring stations picked up the seismic shocks in Indonesia itself, and the U.S. Geological Survey’s worldwide monitoring system, with 120 stations, pinpointed the quake immediately. However, on that fateful Sunday, the few warning telephone calls that were made went unanswered. Thus, lack of communication, more than lack of technology, caused the extreme loss of life. The material destruction, however, touching so many countries, was a result of the movement of the Earth’s crust beneath the Indian Ocean extending along a fracture believed to have been 600 miles in length, ultimately creating a tidal wave nearly 1000 miles long. Such horror can never again touch the Indian Ocean nations as they continue to build a Tsunami Watch System containing all the elements of the early warning capability that has existed in the Pacific Ocean since the middle of the 20th century. In the late 1970s, scientists suggested an improved method for making tsunami predictions. For a number of years, specialists have suggested that better analysis and interpretation of seismic waves produced by earthquakes may improve the prediction of tsunamis. Seismic waves range from very short-period waves that result from the sharp snap of rocks under high stress to very long-period waves, due mainly to the slower movements of large sections of the ocean floor. Many researchers believe that tsunamis result mainly from the vertical movement of these large blocks, leading to a tentative conclusion that the strength of seismic waves of very long period may be

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TSUNAMI

the best criterion for an earthquake’s ability to generate a tsunami. Part of the problem is that most seismographs installed in the system are not very sensitive to very long-period waves, and thus a given earthquake cannot be analyzed effectively in terms of its potential for producing a tsunami. Equipment has been refined so that, today, shorter period waves are used to locate an earthquake and 20-second waves are used to calculate the magnitude. However, some scientists feel that the true magnitude of some earthquakes can only be determined by measuring the characteristics of longer waves, such as 100-second waves. Brune and Kanamori (University of California at San Diego) have observed that the Chilean earthquake of 1960 had a magnitude of 8.3 when calculated on the basis of 20-second waves, but its magnitude was 9.5, or more than 10 times larger in wave amplitude and more than 60 times larger in energy released, when calculated by Kanamori’s method, which attempts to include the energy release represented by very long-period seismic waves. Other scientists are coming to the viewpoint that many warnings could be omitted if predictions were based on longer waves. Two of the first long-period seismographs incorporated in the Tsunami Warning System were installed in Hawaii and on the Russian island of Yuzhno-Sakhalinsk, which is northeast of Vladivostok. Later, an installation was made at the Alaska Warning Center in Palmer. Whether the latest reasoning proves successful must await a number of years of experience with the earthquakes of the future and the resulting tsunamis. CONTINUING TSUNAMI RESEARCH During the 2000s, research pertaining to the fundamentals of tsunamis and the development of mathematical models of the phenomenon continues. Considerable attention is being directed to specific regions, including the west coasts of Mexico and Chile, the southwestern shelf of Kamchatka (Russia) and, in the United States, the generation of tsunamis in the Alaskan bight and in the Cascadia Subduction Zone off the west coasts of Washington (Puget Sound) and Oregon. Research also is being directed toward the development of simple and more economic warning systems, particularly in the interest of the developing countries, the coasts of which border on the Pacific Ocean. The Tsunami Warning System previously described requires millions of dollars for equipment, maintenance, and operation, well beyond the means of some countries. Also, some scientists believe that more localized equipment installations could possibly serve local shore communities better while costing less. These observations, however, do not challenge the need and validity for the larger tsunami network. The National Oceanic and Atmospheric Administration (NOAA) has developed a system costing in the range of $20,000 that can be installed and operated by non experts. The system has been undergoing trials at Valparaiso, Chile, a city that has been struck by nearly 20 tsunamis within the past two centuries. A sensor (accelerometer) is installed in bedrock under the city and can measure tectonic activity in excess of 7.0 on the

Richter scale. These measurements are interlocked with level sensors. Researchers B.F. Atwater and A.L. Moore (University of Washington), in their attempts to model earthquake and tsunami activity in the area over the last thousand years, have reported what they believe to have been a large earthquake on the Seattle fault some time between 1000 and 1100 years ago. The researchers report, ‘‘Water surged from Puget Sound, overrunning tidal marshes and mantling them with centimeters of sand. One overrun site is 10 km northwest of downtown Seattle; another is on Whidbey Island, some 30 km farther north. Neither site has been widely mantled with sand at any other time in the past 2000 years. Deposition of the sand coincided—to the year or less—with abrupt, probably tectonic subsidence at the Seattle site and with landsliding into nearby Lake Washington. These findings show that a tsunami was generated in Puget Sound, and they tend to confirm that a large shallow earthquake occurred in the Seattle area about 1000 years ago.’’ Simulations of tsunamis from great earthquakes on the Cascadia subduction zone have been carried out by M. Ng, P.H. Leblond, and T.S. Murty (University of British Columbia). A numerical model has been used to simulate and assess the hazards of a tsunami generated by a hypothetical earthquake of magnitude 8.5 associated with rupture of the northern sections of the subduction zone. The model indicates that wave amplitudes on the outer coast are closely related to the magnitude of seabottom displacement (5 meters). The researchers observe, ‘‘Some amplification, up to a factor of 3, may occur in some coastal embayments. Wave amplitudes in the protected waters of Puget Sound and the Straits of Georgia are predicted to be only about one-fifth of those estimated on the outer coast.’’ READING LIST Atwater, B.F. and Moore, A.L (December 4, 1991). A tsunami about 1000 years ago in Puget Sound, Washington. Science: 1614. Bernard, E.N. (Ed.). (July 1989). Tsunami Hazard: A Practical Guide for Tsunami Hazard Reduction, (Papers from Symposium at Novosibirsk, Russia). International Union of Geodesy and Geophysics, New York. Bryant, E. (2001). Tsunami: The Underrated Hazard. Cambridge University Press, New York. Collins, E. (February 1988). Wave Watch. Sci. Amer: 28. Kubota, I. (Spring 1987). Japan’s Weather Service and the Sea. Oceanus: 71. Lander, J.F., Lockridge, P.A., and Kozuch, M.J. (1997). Tsunamis Affecting the West Coast of the United States, 1806–1992. DIANE Publishing Company, Collingdale, PA. Lander, J.F. (1997). Tsunamis Affecting Alaska, 1737–1996. DIANE Publishing Company, Collingdale, PA. Ng, M.K-F., Leblond, P.H., and Murty, T.S. (November 30, 1990). Simulation of Tsunamis from Great Earthquakes on the Cascade Subduction Zone. Science: 1248. Prager, E.J. (1999). Furious Earth: The Science and Nature of Earthquakes, Volcanoes, and Tsunamis. McGraw-Hill Professional Book Group, New York.

TSUNAMI Soloviev, S.L., Kim, K.S., Solovieva, O.N. et al. (2000). Tsunamis in the Mediterranean Sea, 2000 B.C. –2000 A.D. Kluwer Academic Publishers, Norwell, MA. Staff. (1997). Long-Wave Runup Models. World Scientific Publishing Company, Inc., River Edge, NJ. Tsuchiya, Y. and Shuto, N. (1995). Tsunami: Progress in Prediction, Disaster Prevention and Warning. Kluwer Academic Publishers, Norwell, MA.

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WEB REFERENCES International Tsunami Information Center (ITIC): http://www. shoa.cl/oceano/itic/frontpage.html. National Tsunami Hazard Mitigation Program: http://www.pmel. noaa.gov/tsunami/. Tsunami Links: http://www.pmel.noaa.gov/tsunami-hazard/links. html.

METEOROLOGY BALLOONING AND METEOROLOGY IN THE TWENTIETH CENTURY LINDA VOSS U.S. Centennial of Flight Commission

Balloons are ideal for gathering meteorological information and have been used for that purpose throughout their history. Meteorological measurements of wind and air pressure have gone hand in hand with the earliest balloon launches and continue today. Balloons can climb through the denser air close to the Earth to the thinner air in the upper atmosphere and collect data about wind, the different layers of the atmosphere, and weather conditions as they travel. The first meteorological balloon sondes, or ‘‘registering balloons,’’ were flown in France in 1892. These balloons were relatively large, several thousand cubic feet, and carried instruments to record barometric pressure (barometers), temperature (thermometers), and humidity (hygrometers) data from the upper atmosphere. They were open at the base of the balloon and were inflated with a lifting gas, which could be hydrogen, helium, ammonia, or methane. The lifting gas in the balloon exited through the

A balloon equipped for meteorological observations. A German balloon ascent in the late 1800s. 17 Balloon Equipped for Meteorological Observations.

A zero-pressure balloon being inflated at Alice Springs, Australia.

opening as the balloon expanded during its ascent and the air became thinner and the pressure dropped. At the end of the day, as the lifting gas cooled and took up less space, the balloon descended very slowly. The meteorologists had to wait until the balloon descended all the way to Earth to

Two men performing balloon tests for the U.S. Weather Bureau.

This article is a US Government work and, as such, is in the public domain in the United States of America. 164

BALLOONING AND METEOROLOGY IN THE TWENTIETH CENTURY

Weather balloons are used daily to carry meteorological instruments to 20 miles (30 kilometers) and above into the atmosphere to measure temperature, pressure, humidity, and winds. The balloons are made of rubber and weigh up to 2.2 pounds (one kilogram). More than 200,000 weather soundings are made with such balloons worldwide each year.

Preparing to launch America’’s first ‘‘ballon-sonde.’’ Since this first launch on September 15, 1904, in St. Louis, Missouri, literally millions of weather balloons have been launched by the National Weather Service and its predecessor organization. From: The Principles of Aerography, by Alexander McAdle, 1917.

retrieve their instruments, which often had drifted up to 700 miles (1,126 kilometers) from their launch point. The German meteorologist Assmann solved the problem of drifting balloons and retrieval of instruments in 1892 by introducing closed rubber balloons that burst when they reached a high altitude, dropping the instruments to Earth by parachute much closer to the launch

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site. These balloons also had fairly constant rates of ascent and descent for more accurate temperature readings. Assmann also invented a psychrometer, a type of hygrometer used to measure humidity in the air that laboratories generally use. In the 1930s, meteorologists were able to get continuous atmospheric data from balloons when the radiosonde was developed. A radiosonde is a small, radio transmitter that broadcasts or radios measurements from a group of instruments. Balloons, usually unmanned, carry the transmitter and instruments into the upper atmosphere. The radiosonde transmits data to Earth while measuring humidity, temperature, and pressure conditions. Today, three types of balloons are commonly used for meteorologic research. Assmann’s rubber, or neoprene, balloon is used for measuring vertical columns in the atmosphere, called vertical soundings. The balloon, inflated with a gas that causes the balloon to rise, stretches as it climbs into thin air, usually to around 90,000 feet (27,400 meters). Data is taken as the balloon rises. When the balloon has expanded from three to six times its original length (its volume will have increased 30 to 200 times its original amount), it bursts. The instruments float to Earth under a small parachute. The neoprene balloon can either carry radiosondes that transmit meteorological information or be tracked as a pilot balloon, a small balloon sent aloft to show wind speed and direction. Around the world, balloons equipped with radiosondes make thousands of soundings of the winds, temperature, pressure, and humidity in the upper atmosphere each day. But these balloons are launched and tracked from land, which limits what the radiosondes can measure to less than one-third of the Earth’s surface. Zero-pressure plastic (usually polyethylene) balloons were first launched in 1958. They carry scientific instruments to a predetermined atmospheric density level. Zero-pressure balloons are the best for extremely high altitudes because the balloons can be lighter and stress on them can be distributed over the surface of the balloon. About the same time, the Air Force Cambridge Research Laboratories (AFCRL) started working on super-pressure balloons, which were made from mylar. The development of mylar plastic films and advances in electronic miniaturization made constant-altitude balloons possible. Mylar is a plastic that can withstand great internal pressure. The mylar super-pressure balloon does not expand as it rises, and it is sealed to prevent the release of gas as the balloon rises. By the time the balloon reaches the altitude where its density equals that of the atmosphere, the gas has become pressurized because the heat of the sun increases the internal gas pressure. However, because mylar can withstand great internal pressure, the volume of the balloon remains the same. By carefully calculating the weight of the balloon and whatever it is carrying, the altitude at which the balloon will achieve equilibrium and float can be calculated. As long as the pressure inside the balloon remains the same, it will remain at that altitude. These balloons could be launched to remain aloft at specified altitudes for weeks or months at a time.

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Moreover, satellites could be used to track and request data from many balloons in the atmosphere to obtain a simultaneous picture of atmospheric conditions all over the globe. Another advantage of super-pressure balloons is that, since they transmit their data to satellites, they can gather data from over oceans as well as land, which is a limitation of balloons equipped with radiosondes. The AFCRL program resulted in the Global Horizontal Sounding Technique (GHOST) balloon system. With GHOST, meteorologists at last achieved their goal of semipermanent platforms floating high in the atmosphere. Eighty-eight GHOST balloons were launched starting in March of 1966. The GHOST balloons and their French counterpart, EOLE, (the name Clement Ader used for one of his aircraft—named after the Greek god of the wind) used strong, plastic super-pressure balloons to trace air circulation patterns by drifting with the wind at constant density altitudes. Many super-pressure balloons were aloft at a time, grouped at constant density levels. Each balloon had a sensing device and transmitting system for gathering information on its position and weather data and transmitted atmospheric and weather data to weather satellites. They first transmitted their data to the NASA Nimbus-4 meteorological satellite in 1970. In 1966, a GHOST balloon circled the Earth in 10 days at 42,000 feet (12,801 meters). By 1973, NASA had orbited scientific instrument packages aboard sealed balloons at altitudes up to 78,000 feet (23,774 meters). Other GHOST balloons remained aloft for up to a year. The program lasted for 10 years. The ultimate of the super-pressure balloons was the balloon satellite Echo I. Launched into space in 1960, the balloon inflated to a sealed volume by residual air, benzoic acid, and a chemical called anthraquinone. Constant-altitude, super-pressure balloons continue to fly over the oceans and land surfaces. These balloons have been relied on for decades to provide extensive knowledge of global meteorology and improve worldwide weather forecasting.

concise.asp?ti=761576250&sid=3#s3. http://www.encarta.msn .com. ‘‘Measuring the Weather,’’ USAToday.com. http://www.usatoday .com/weather/wmeasur0.htm. ‘‘Meteorology,’’ Microsoft Encarta Online Encyclopedia 2000. 1997–2000 Microsoft Corporation. http://encarta.msn.com. Warner, Lucy, ed. ‘‘Forecasts: Observing and Modeling the Global Atmosphere,’’ UCAR at 25. University Center for Atmospheric Research. Boulder, Colorado. Oct. 17, 2000. http://www.ucar.edu/communications/ucar25/forecasts.html. ‘‘Weather Instruments to Make.’’ http://asd-www.larc.nasa.gov/ SCOOL/psychrometer.html To find out how you can make your own psychrometer, link to The CERES S’COOL Project at http://asdwww.larc.nasa.gov/SCOOL. The link has lots of information about making weather observations. Making a Psychrometer is in the Table of Contents or go to http://explorer.scrtec.org/explorer/explorer-db/html/783750680447DED81.html. To learn more about weather instruments and even set up your own weather station to report to the U.S. National Weather Service, go to http://www.usatoday.com/weather/wmeasur0.htm.

Educational Organization

Standard Designation (where Content of applicable) Standard

National Science Education Standards

Content Standard D

International Technology Education Association International Technology Education Association

Standard 9

Crouch, Tom D. The Eagle Aloft: Two Centuries of the Balloon in America. Washington, DC.: Smithsonian Institution Press, 1983. Tannenbaum, Beulah and Harold E. Making and Using Your Own Weather Station. New York. Venture Books, 1989. Vaeth, J. Gordon. ‘‘When the Race for Space Began.’’ U.S. Naval Institute Proceedings. August 1961.

International Technology Education Association

Standard 7

ONLINE REFERENCES

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Standard 10

READING LIST

Cowens, John. ‘‘Building a Psychrometer,’’ Greenwood Elementary, La Grande, OR 97850. September 28, 1993. http://explorer.scrtec.org/explorer/explorer-db/html/783750680447DED81.html ‘‘Hygrometer,’’ Microsoft Encarta Online Encyclopedia 2000. 1997–2000 Microsoft Corporation. http://encarta.msn.com. Lally, Vincent E. ‘‘Balloon: Modern Scientific Ballooning,’’ Microsoft Encarta Online Encyclopedia 2000. 1997–2000 Microsoft Corporation. http://www.encarta.msn.com//find/

Students should develop an understanding of energy in the earth system. Students will develop an understanding of engineering design. Students will develop an understanding of the role of trouble shooting, research and development, innovation, and experimentation in problem solving. Students will develop an understanding of the influence of technology on history.

TODD RASMUSSEN The University of Georgia Athens, Georgia

Water levels in wells are often observed to fluctuate as the air pressure changes. Blaise Pascal described this effect in 1663 (1) and was the first to attribute the water level changes in wells to changes in atmospheric pressure. He

BAROMETRIC EFFICIENCY

CONFINED AQUIFERS For confined aquifers, Jacob (4) showed that increasing the load on the ground surface increases the load on the aquifer. This additional weight is either carried by the mineral grains or by the water within the aquifer pores. If the entire weight is borne by the mineral grains, and these grains do not deform with the increase in load, then the total head within the aquifer remains unchanged, and the barometric efficiency is 100%. The barometric efficiency is smaller if some of the weight is carried by the fluid. A very low barometric efficiency in confined aquifers occurs when the fluid within

Barometric effciency, BE = 1

Total head Water level Barometric pressure Barometric effciency, BE = 0

5

5

4

4

3

3 Head

where consistent units (e.g., feet of water, mm of Hg, hPa) are used for both water levels and barometric pressure (2,3). To understand why this relationship exists, we can conceive of an aquifer that is entirely isolated from the atmosphere. Such an aquifer maintains a constant total head, H = Ho , within the aquifer and is entirely unaffected by atmospheric pressure changes. Water level elevations, W, in the aquifer are measured in an open borehole. The total head within the well is the sum of the water level elevation plus the barometric pressure exerted on the water surface, H = W + B. If the total head within the well is equal to the total head within the aquifer, Ho = W + B, and the total head within the aquifer is constant, then the water level varies inversely with barometric pressure, W = Ho − B. The change in water level is just the negative of the change in barometric pressure,
W =
(Ho − B) = −
B, so that the barometric efficiency is 100 percent, BE = −
W/
B = −(−
B)/
B = 1. A second example assumes that the total head in the hypothetical aquifer increases with barometric pressure, H = Ho + B, so that the head in the aquifer goes up and down over time. In this case, the water levels in the well remain unchanged, and the barometric efficiency of the aquifer is zero. These two extremes are shown in Fig. 1. Note that the barometric pressure is identical in both cases. In the left figure, the barometric efficiency is 100%, so that the water level varies inversely with barometric pressure. The total head is the sum of the barometric pressure and water levels, so it remains unchanged. In the right figure, the barometric efficiency is zero, so that the water level is unaffected by barometric pressure changes and the total head varies directly with barometric pressure. Most aquifers lie between these two extremes, however, and the actual response depends, in large part, on whether the aquifer is confined or unconfined. Both of these cases are described below.

Total head Water level Barometric pressure

Head

noted that water levels declined as the barometric pressure increased, and vice versa. The barometric efficiency, BE, is used to relate changes in water levels,
W, to changes in barometric (air) pressure,
B:
W (1) BE = −
B

2 1

2 1

0

0

−1

−1

0

10

20 Time

30

167

0

10

20

30

Time

Figure 1. Effect of barometric pressure on water levels in wells. Both figures show barometric pressure (lower, dashed line), total head (solid line), and water levels (dotted line). Left figure shows effects when the barometric efficiency is 100%, BE = 1, and right figure shows effects when the barometric efficiency is zero, BE = 0.

the aquifer bears most of the weight. Examples of such aquifers include poorly consolidated sedimentary aquifers or horizontal fractures that extend great distances. Figure 2 shows the extreme conditions. In the left figure, the entire increase in load caused by an increase in barometric pressure is carried by the mineral grains. In this case, the pore fluids are not affected by the increase in load, the total head remains unchanged, and water levels drop in an amount equal to the increase in barometric pressure. In the right figure, the mineral grains do not carry the load and the fluid carries the increased load, causing an increase in total head. In confined aquifers, the barometric response is virtually instantaneous. A change in barometric pressure should cause an immediate change in water levels in wells. Water levels in large-diameter wells may not respond immediately, however, because of the time required for water levels to adjust to the new level. Instead of a rapid

B

Aquitard

Mineral grains ∆W = − ∆B α=1

B

Aquitard

Pore fluids ∆H = ∆B α=0

Figure 2. Effect of barometric pressure on total head in confined aquifers. The mineral grains carry the load in the left figure, whereas the pore fluids carry the load in the right figure. Most confined aquifers, however, are intermediate between these two extremes.

BAROMETRIC EFFICIENCY

response to pressure changes, a slow response may be found in these wells. This phenomena is called borehole storage and is a function of the diameter of the borehole, the length of the screened interval relative to the aquifer thickness, the aquifer hydraulic properties (e.g., transmissivity and storativity), and the rate at which the barometric pressure is changing. The effects of borehole storage can be eliminated by placing a packer in the well below the water surface, but above the screened interval, which eliminates the need for the water level to change in response to changes in barometric pressure. Some change in volume may result from the slight compressibility of water, but this effect is very small. A gauge-type pressure sensor (i.e., internally vented to the atmosphere) is then placed below the packer to measure the water pressure. Several techniques are available for estimating the confined aquifer barometric efficiency, including linear regression and Clark’s Method (5). Clark’s method is an unbiased technique for estimating the barometric efficiency that performs well when a water level trend is present in the data (6). Regression provides a better estimate, but only when the trend function can be accurately specified (7).

UNCONFINED AQUIFERS Barometric pressure changes commonly do not affect water levels in wells located in shallow, unconfined aquifers, which is because the air pressure moves rapidly through the unsaturated zone and causes an immediate increase in total head within the aquifer. In deeper unconfined aquifers, however, the typical response to barometric pressure changes is to see an immediate inverse response (i.e., an increase in barometric pressure causes an equivalent and opposite water level response), followed by a slow decay back to the original water level (5,8,9). To understand this complex response, we focus on how total head responds to barometric pressure changes. Initially, no increase in total head within the unconfined aquifer occurs because the water held within pores is not confined by an overlying confining unit. The total head within the aquifer can rise over time, however, as air diffuses downward through the unsaturated zone. The total head in the unconfined aquifer increases once the air pressure change reaches the water table. The time required for the total head to respond to barometric pressure changes is a function of the depth of the water table and the air diffusivity within the unsaturated zone. The air diffusivity is higher in coarse, dry soils and is lower in wet soils or in soils with a low air permeability. As the water level response in observation wells is the total head minus the barometric pressure, the fact that the total head does not change initially means that a rapid water-level response occurs to barometric pressure changes. In fact, the barometric efficiency is one in wells with a deep unsaturated zone—or where the air permeability of the unsaturated zone is low—followed

Unsaturated zone barometric response 1.5 1 0.5 Head

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0 −0.5 −1 −1.5

Barometric pressure Total head Water level

0

5

10

15 Time

20

25

30

Figure 3. Effect of barometric pressure on total head and water levels in confined aquifers. The barometric pressure in this example takes a step from zero to one at t = 10. The total head slowly rises as the barometric pressure makes its way through the unsaturated zone. The water levels initially fall to maintain the constant total head, but then return to normal once the barometric pressure signal reaches the water table.

by a gradual decrease as air diffuses downward to the water table. This concept is illustrated graphically in Fig. 3. Note that the barometric pressure is simplified to a simple step change from zero to one. The total head does not initially respond, but slowly rises over time as the barometric pressure moves through the unsaturated zone and eventually reaches the water table. The observed water level in a well placed in this aquifer responds immediately (as long as borehole storage can be neglected), but eventually decays back to its original level The barometric efficiency of unconfined aquifers is not single valued like the confined aquifer barometric efficiency. Instead, the response includes a delay that can be estimated using regression deconvolution (5,9). The resulting response can be used to estimate the unsaturated air diffusivity of the unsaturated zone. BIBLIOGRAPHY 1. Pascal, B. (1973). The Physical Treatises of Pascal. Octagon Books, New York. 2. Bear, J. (1972). Dynamics of Fluids in Porous Media. Dover Publications, New York. 3. Walton, W.C. (1970). Groundwater Resource Evaluation. McGraw-Hill, New York. 4. Jacob, C.E. (1940). On the flow of water in an elastic artesian aquifer. Amer. Geophys. Union Trans. 21: 574–586. 5. Rasmussen, T.C. and Crawford, L.A. (1997). Identifying and removing barometric pressure effects in confined and unconfined aquifers. Ground Water. 35(3): 502–511. 6. Clark, W.E. (1967). Computing the barometric efficiency of a well. J. Hydraul. Divi. Amer. Soc. Civil Engin. 93(HY4): 93–98. 7. Davis, D.R. and Rasmussen, T.C. (1993). A comparison of linear regression with Clark’s method for estimating barometric

CERES: UNDERSTANDING THE EARTH’S CLOUDS AND CLIMATE efficiency of confined aquifers. Water Resources Res. 29(6): 1849–1854. 8. Weeks, E.P. (1979). Barometric fluctuations in wells tapping deep unconfined aquifers. Water Resources Res. 15(5): 1167–1176. 9. Spane, F.A., Jr. (2002). Considering barometric pressure in groundwater flow investigations. Water Resources Res. 38(6): 1078.

READING LIST Spane, F.A., Jr. and Mercer, R.B. (1985). HEADCO: A Program for Converting Observed Water Levels and Pressure Measurements to Formation Pressure and Standard Hydraulic Head. Pacific Northwest National Laboratory, Rockwell Hanford Operations, Richland, WA. Report PNL-10835.

CERES: UNDERSTANDING THE EARTH’S CLOUDS AND CLIMATE NASA—Langley Research Center

The Clouds and the Earth’s Radiant Energy System (CERES) instrument is one of several launched aboard the Earth Observing System’s (EOS) Aqua spacecraft in 2002. Scientists use observations from the CERES instrument to study the energy exchanged between the Sun; the Earth’s atmosphere, surface and clouds; and outer space. The CERES Aqua instruments are the fourth and fifth CERES instruments in orbit. NASA launched the first CERES instrument aboard the Tropical Rainfall Measuring Mission satellite or TRMM in November 1997. Results of the TRMM mission show that the first CERES provided better measurement capabilities than any previous satellite instrument of its kind. Two other CERES instruments are currently orbiting the Earth on the EOS Terra spacecraft, launched in late 1999. Early CERES Terra results give new insights into the effects of clouds on climate and how the climate system changes from decade to decade. Two CERES instruments on each of the Terra and Aqua spacecraft will provide global coverage of energy radiated and reflected from the Earth. Scientists use measurements from both satellites’ orbits to improve observations of the daily cycle of radiated energy. NASA Langley Research Center manages the CERES mission. Langley’s highly successful Earth Radiation Budget Experiment (ERBE) provided the foundation for the design of the CERES instrument. ERBE used three satellites to provide global energy measurements from 1984 through the 1990s. The TRW Space & Electronics Group in Redondo Beach, Calif., built all six CERES instruments.

This article is a US Government work and, as such, is in the public domain in the United States of America.

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WHAT CERES WILL MEASURE CERES measures the energy at the top of the atmosphere, as well as estimate energy levels in the atmosphere and at the Earth’s surface. Using information from very high resolution cloud imaging instruments on the same spacecraft, CERES will determines cloud properties, including altitude, thickness, and the size of the cloud particles. All of these measurements are critical for advancing the understanding of the Earth’s total climate system and the accuracy of climate prediction models. BALANCING THE EARTH’S ENERGY BUDGET The balance between Earth’s incoming and outgoing energy controls daily weather and climate (long-term weather patterns). Sunlight or solar energy is the planet’s only incoming energy source. Heat emitted from the sunlight reflected by the Earth’s surface, atmosphere and clouds make up the planet’s outgoing energy. Scientists have been working for decades to understand this critical energy balance, called the Earth’s ‘‘energy budget.’’ The energy received from the Sun is at short wavelengths, while the energy emitted by the surface of the Earth, the atmosphere and clouds is at long wavelengths. Greenhouse gases in the atmosphere absorb the long wavelength energy or heat emitted by the Earth. Increases in the amounts of greenhouse gases produced by both natural processes or human activities can lead to a warming of the Earth’s surface. Such changes may, in turn, alter the planet’s daily weather and climate. Clouds and small particles in the atmosphere called aerosols also reflect some sunlight back into space. Major sources of aerosols include windblown dust, emissions from the burning of fossil fuels, such as gasoline, and the burning of forests and agricultural fields. CLOUD EFFECTS One of the most intriguing questions facing climate modelers today is how clouds affect the Earth’s climate

CERES detects low (blue and white) to high (yellow) amounts of emitted heat

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CHINOOK

Earth’s radiation budget is the balance between incoming and outgoing energy

and vice versa. The U.S. Global Change Research Program classifies understanding the role of clouds and the Earth’s energy budget as one of its highest scientific priorities. Understanding cloud effects requires a detailed knowledge of how clouds absorb and reflect sunlight, as well as how they absorb and re-emit outgoing heat emitted by the planet. For example, low, thick clouds primarily reflect incoming solar energy back to space causing cooling. Thin, high clouds, however, primarily trap outgoing heat and produce warming. To date, satellite studies have found that clouds have an overall cooling effect on the Earth. Analyses of satellite data also indicate that clouds which form over water are very different from clouds which form over land. These differences affect the way clouds reflect sunlight back into space and how much heat emitted from the Earth the clouds absorb and re-emit. For example, over the equator in the eastern Pacific Ocean ˜ events, there is a significant decrease during El Nino in the amount of energy emitted by the Earth due to ˜ events occur when portions increased cloudiness. El Nino of the eastern Pacific Ocean become considerably warmer than normal, causing an increase in cloudiness over the region. These changes can affect weather patterns around the world. WATER VAPOR EFFECTS Water vapor in the atmosphere also impacts our daily weather and climate, though scientists are only beginning to understand how this complex mechanism works. Water vapor acts like a greenhouse gas and absorbs outgoing heat to warm the Earth. Because water vapor also condenses to make clouds, additional water vapor in the atmosphere also may increase the amount of clouds. FUTURE MISSIONS One additional CERES instrument is available to fill the gap between Aqua and the next generation of highly accurate Earth radiation budget measurements. These

observations are expected to be made on the National Polar-orbiting Operational Environmental Satellite System (NPOESS) starting around 2010. To continue the 22-year record of global energy measurements, the next CERES mission should launch in 2007. EDUCATIONAL OUTREACH As a CERES instrument passes overhead, students worldwide are observing clouds and then sending their observations to NASA Langley’s Atmospheric Sciences Data Center (ASDC). At the ASDC, scientists store data for further analysis by the CERES science team. The student observations are part of a global educational outreach program called the Students’ Cloud Observations On-Line (S’COOL) project. Since the project began five years ago, S’COOL has reached over 1,000 schools in all 50 states and 57 other countries on five continents. COMMERCIAL APPLICATIONS CERES supports commercial applications by providing data about weather and sunlight at the Earth’s surface for the renewable energy industry via an innovative Web site (http://eosweb.larc.nasa.gov/sse/). The Surface Meteorology and Solar Energy Project maintains the site. In the first three years of operation, the number of registered users of the Web site, including major energy companies, financial institutions and federal agencies, has grown to over 2,000 from nearly 100 countries. With 35,000 hits per month since January 2001, SSE is the most accessed Web site at the ASDC.

CHINOOK ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

Wind is defined as the movement of air. Although we commonly define wind as a gentle breeze or a harsh gust of

GLOBAL CLIMATE CHANGE

cold air, there is a wind phenomenon that can increase the temperature instead of lowering it. It is called a Chinook wind. Chinooks are most commonly associated with the Rocky Mountain range in North American but can also be found in the Swiss Alps and the Andes. They can increase temperatures high enough to melt the snow in their path as they travel down the mountainside. The Chinook wind falls under the classification of katabatic wind, which means that it moves downhill. The name ‘‘Chinook’’ was taken from the Chinook Indians that lived along the Rocky Mountains until the early 1800s when the tribe became extinct due to disease. Although the tribe died off, their legends and tales live on. One such legend is that of the Chinook wind, which in its literal translation means ‘‘snow-eater.’’ It is possible for a Chinook to take place anytime during the year, but its effects are much more dramatic during the winter months. Chinook winds cause dramatic increases in the temperature on the eastern side of the Rockies and can send temperatures into the fifties and sixties. A temperature change of this magnitude can take anywhere from an hour to a day. The heat produced is a reaction formed from the change of gas to liquid in the atmosphere. These warm gusts of air then cause the evaporation of any snow on the ground, hence the name snow-eater. There is no definite length of time that a Chinook will last. On average, it can last hours or days. Chinooks are the end result of the warm moist air of the Pacific Ocean moving up and down the Rocky Mountains. A westerly wind collects the warm moist air in the Pacific Ocean and carries it to the coast where it meets the western side of the Rockies. As the air makes its way up the mountain, the air becomes cooler, and the precipitation is squeezed out of it. Through this process warm, dry air is produced as it makes its way down the eastern side of the mountain. This process is repeated over and over again on each mountain that is in the way of the wind and each time produces warmer, dryer air. The air that makes its way down the last mountain is extremely dry and warm. This wind is called the Chinook wind. These warm gusts of air can reach speeds up to 100 mph. The results from Chinook winds are both positive and negative. Chinook winds cause evaporation of the snow covering the ground, so the length of the grazing season is extended. A longer grazing season decreases the need to stock up on feed for animals. A negative result is the decreased amount of precipitation due to the quick evaporation of snow. Less precipitation causes hardships in planting. READING LIST Alberta Agriculture, Food, and Rural Development. (2002). Weather in Alberta. Available: http://www.agric.gov.ab.ca/ agdex/000/7100001b.html. (March 11). The Alberta Traveller. (2002). Chinook Winds. Available: http://www.traveller.babelfish.com/weather guides/Chinook. html. (March 11). Black Hills Weather. (2002). The Snow Eating Wind. Available: http://www.blackhillsweather.com/chinook.html. (March 11). Ho, I. (2002). Snow-Eater. The Catalyst. Available: http://www. brown.edu/Students/Catalyst/archive/spring99/ch.html. (March 11).

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Under the Chinook Arch. http://www.nucleus.com/∼cowboy/ Misc/Chinooks.html. (11 March 2002).

GLOBAL CLIMATE CHANGE Geophysical Fluid Dynamics Laboratory—NOAA

In 1967, two GFDL scientists, Syukuro Manabe and Richard Wetherald, published what is now regarded as the first credible calculation of the effect of increased carbon dioxide on the climate. They calculated that a doubling of atmospheric carbon dioxide would warm the earth’s surface by about 2 ◦ C. This result laid the foundation for what has become an international, multi-disciplinary research effort on global warming. Manabe, in collaboration with oceanographer Kirk Bryan and other scientists at GFDL, has continued to lead the international effort to develop the coupled oceanatmosphere climate models that are crucial to understanding and predicting the impact of greenhouse gases.

Three-dimensional view of projected surface air temperature and ocean warming ( ◦ C) due to greenhouse gases as calculated by a low-resolution GFDL coupled ocean-atmosphere climate model. The top panel shows the surface air temperature change over North and South America and surrounding regions. The three-dimensional box illustrates the depth to which a 1 ◦ C and 0.2 ◦ C warming has penetrated in the model’s Pacific Ocean. The gray surface depicts the model’s ocean floor. Note the deep mixing of the ocean warming signal in the southern hemisphere ocean near Antarctica. The temperature changes are projections of the warming due to greenhouse gases by the latter half of the twenty-first century in the absence of further increases in sulfate aerosol forcing. Results shown are based on years 61–80 of a transient CO2 increase experiment (+1% per year compounded). [Source: adapted from Syukuro Manabe and Ronald Stouffer, Nature, 15 July 1993.]

This article is a US Government work and, as such, is in the public domain in the United States of America.

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OBSERVATIONS OF CLIMATE AND GLOBAL CHANGE FROM REAL-TIME MEASUREMENTS

is enough to cause the ocean’s global thermohaline circulation to almost disappear in the model. The global thermohaline circulation is important because it is responsible for a large portion of the heat transport from the tropics to higher latitudes in the present climate. In addition, sea level continues rising steadily for centuries after the CO2 increase is halted. From this perspective, global warming can no longer be viewed as just as a problem of our own lifetimes, but as a legacy—with uncertain consequences—now being passed forward to many future generations. EVALUATING CLIMATE MODELS GFDL scientists, including Tony Broccoli and Tom Delworth, are searching for innovative ways to evaluate climate models and to distinguish between human-induced climate change and natural climate variability. Measurements of the current climate, historical observations, and glimpses of earth’s climate during the ice ages and other past climates all provide opportunities to test climate models. Through research on climate models and observations, scientists at GFDL will continue to evaluate and refine the climate models that will be needed to help answer critical policy-relevant questions about global warming and its consequences.

OBSERVATIONS OF CLIMATE AND GLOBAL CHANGE FROM REAL-TIME MEASUREMENTS DAVID R. EASTERLING THOMAS R. KARL

Impact of increasing CO2 on the earth’s climate as simulated in a GFDL coupled ocean-atmosphere climate model. Shown are timeseries of: a) prescribed CO2 concentration on a logarithmic scale in comparison to present levels; b) global mean surface air temperature ( ◦ C); c) global mean increase of sea level (cm) due to thermal expansion; and d) intensity of the North Atlantic Ocean’s meridional overturning circulation (10**6 m**3/sec). The labels ‘‘Control’’, ‘‘2xCO2 ’’, and ‘‘4xCO2 ’’ refer to separate experiments where CO2 either remains constant (Control), or increases at 1% per year (compounded) to double (2xCO2 ) or quadruple (4xCO2 ) the current concentration. Note that the sea level rise estimates do not include the effect of melted continental ice sheets. With this effect included, the total rise could be larger by a substantial factor. [Source: Syukuro Manabe and Ronald Stouffer, Nature, 15 July 1993.]

A PROBLEM FOR CENTURIES TO COME? In a recent paper, published 26 years after Manabe’s pioneering one-dimensional CO2 sensitivity study, he and Ron Stouffer of GFDL used a three-dimensional coupled ocean-atmosphere model to examine possible CO2 -induced climate changes over several centuries. Earlier studies had focused on shorter time horizons. In their scenario, CO2 quadruples over a period of 140 years, then no longer increases. This perturbation

(from Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003)

INTRODUCTION Is the planet getting warmer? Is the hydrologic cycle changing? Is the atmospheric/oceanic circulation changing? Is the weather and climate becoming more extreme or variable? Is the radiative forcing of the climate changing? These are the fundamental questions that must be answered to determine if climate change is occurring. However, providing answers is difficult due to an inadequate or nonexistent worldwide climate observing system. Each of these apparently simple questions are quite complex because of the multivariate aspects of each question and because the spatial and temporally sampling required to address adequately each question must be considered on a global scale. A brief review of our ability to answer these questions reveals many successes, but points to some glaring inadequacies that must be addressed in any attempt to understand, predict, or assess issues related to climate and global change.

OBSERVATIONS OF CLIMATE AND GLOBAL CHANGE FROM REAL-TIME MEASUREMENTS

IS THE PLANET GETTING WARMER? There is no doubt that measurements show that nearsurface air temperatures are increasing. Best estimates suggest that the warming is around 0.6 ◦ C (+−0.2 ◦ C) since the late nineteenth century (1). Furthermore, it appears that the decade of the 1990s was the warmest decade since the 1860s, and possibly for the last 1000 years. Although there remain questions regarding the adequacy of this estimate, confidence in the robustness of this warming trend is increasing (1). Some of the problems that must be accounted for include changes in the method of measuring land and marine surface air temperatures from ships, buoys, land surface stations as well as changes in instrumentation, instrument exposures and sampling times, and urbanization effects. However, recent work evaluating the effectiveness of corrections of sea surface temperatures for time-dependent biases, and further evaluation of urban warming effects on the global temperature record have increased confidence in these results. Furthermore, by consideration of other temperature-sensitive variables, e.g., snow cover, glaciers, sea level and even some proxy non-realtime measurements such as ground temperatures from boreholes, increases our confidence in the conclusion that the planet has indeed warmed. However, one problem that must be addressed is that the measurements we rely upon to calculate global changes of temperature have never been collected for that purpose, but rather to aid in navigation, agriculture, commerce, and in recent decades for weather forecasting. For this reason there remain uncertainties about important details of the past temperature increase and our capabilities for future monitoring of the climate. The IPCC (1) has summarized latest known changes in the temperature record, which are summarized in Fig. 1. Global-scale measurements of layer averaged atmospheric temperatures and sea surface temperatures from instruments aboard satellites have greatly aided our ability to monitor global temperature change (2–4), but the situation is far from satisfactory (Hurrell and Trenberth, 1996). Changes in satellite temporal sampling (e.g., orbital drift), changes in atmospheric composition (e.g., volcanic emissions), and technical difficulties related to overcoming surface emissivity variability are some of the problems that must be accounted for, and reduce the confidence that can be placed on these measurements (5). Nonetheless, the space-based measurements have shown, with high confidence, that stratospheric temperatures have decreased over the past two decades. Although perhaps not as much as suggested by the measurements from weather balloons, since it is now known that the data from these balloons high in the atmosphere have an inadvertent temporal bias due to improvements in shielding from direct and reflected solar radiation (6).

IS THE HYDROLOGIC CYCLE CHANGING? The source term for the hydrologic water balance, precipitation, has been measured for over two centuries in

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some locations, but even today it is acknowledged that in many parts of the world we still cannot reliably measure true precipitation (7). For example, annual biases of more than 50% due to rain gauge undercatch are not uncommon in cold climates (8), and, even for more moderate climates, precipitation is believed to be underestimated by 10 to 15% (9). Progressive improvements in instrumentation, such as the introduction of wind shields on rain gauges, have also introduced time-varying biases (8). Satellitederived measurements of precipitation have provided the only large-scale ocean coverage of precipitation. Although they are comprehensive estimates of largescale spatial precipitation variability over the oceans where few measurements exist, problems inherent in developing precipitation estimates hinder our ability to have much confidence in global-scale decadal changes. For example, even the landmark work of Spencer (10) in estimating worldwide ocean precipitation using the microwave sounding unit aboard the National Oceanic and Atmospheric Administration (NOAA) polar orbiting satellites has several limitations. The observations are limited to ocean coverage and hindered by the requirement of an unfrozen ocean. They do not adequately measure solid precipitation, have low spatial resolution, and are affected by the diurnal sampling inadequacies associated with polar orbiters, e.g., limited overflight capability. Blended satellite/in situ estimates also show promise (11); however, there are still limitations, including a lack of long-term measurements necessary for climate change studies. Information about past changes in land surface precipitation, similar to temperature, has been compared with other hydrologic data, such as changes in stream flow, to ascertain the robustness of the documented changes of precipitation. Figure 1 summarizes some of the more important changes of precipitation, such as the increase in the mid to high latitude precipitation and the decrease in subtropical precipitation. Evidence also suggests that much of the increase of precipitation in mid to high latitudes arises from increased autumn and early winter precipitation in much of North America and Europe. Figure 2 depicts the spatial aspects of this change, reflecting rather large-scale coherent patterns of change during the twentieth century. Other changes related to the hydrologic cycle are summarized in Fig. 1. The confidence is low for many of the changes, and it is particularly disconcerting relative to the role of clouds and water vapor in climate feedback effects.∗ Observations of cloud amount long have been made by surface-based human observations and more recently by satellite. In the United States, however, human observers have been replaced by automated measurements, and neither surface-based or spaced-based data sets have proven to be entirely satisfactory for detecting changes in clouds. Polar orbiting satellites have an enormous difficulty to overcome related to sampling aliasing and satellite drift (12). For human observers changes in observer schedules, observing biases, and incomplete sampling have created major problems in data ∗

An enhancement or diminution of global temperature increases or decreases due to other causes.

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OBSERVATIONS OF CLIMATE AND GLOBAL CHANGE FROM REAL-TIME MEASUREMENTS Surface temperature indicators Ocean

Land

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* 1990s warmest decade and 1998 warmest year since instrument records began (1861) * 1990s warmest decade of the millennium and 1998 warmest year for at least the N. Hemisphere * N.Hemisphere warming for 20th Century greatest of past 10 centuries Since the retreat of the last glacial maximum (18,000 years ago): * Local changes > 3°C/10yr * Global increases ~ 2°C/1000yr ** N. Hemisphere snow cover extent: since 1987, ** Marine air temperature: 10% below 1973-86 mean 0.4 to 0.7°C increase since late 19th Century

*** Widespread retreat of mountain glaciers during 20th century

* Lake and river ice retreat since the late 19th century (nearly 2-weeks decrease in ice ** Sea surface temperature: duration) 0.4 to 0.8˚C increase *** Land air temperatures: 0.4 to 0.8˚C increase since the late 19th since late 19th century century. *** Reduction in freeze-free season over much of the mid-to-high-latitude region ** Land nighttime air temperature increases at twice the rate as daytime temperatures since 1950

Figure 1. Schematic of observed variations of selected indictors regarding (a) temperature and (b) the hydrologic cycle (based on Ref. 1).

interpretations, now compounded by a change to new automated measurements at many stations. Nonetheless, there is still some confidence (but low) that global cloud amounts have tended to increase. On a regional basis this is supported by reductions in evaporation as measured by pan evaporimeters over the past several decades in Russia and the United States, and a worldwide reduction in the land surface diurnal temperature range. Moreover, an increase in water vapor has been documented over much of North America and in the tropics (1). Changes in water vapor are very important for understanding climate change since water vapor is the most important greenhouse gas in the atmosphere. The measurement of changes in atmospheric water vapor is hampered by both data processing and instrumental difficulties for both weather balloon and satellite retrievals. The latter also suffers from discontinuities among successive satellites and errors introduced by changes in orbits and calibrations. Upper tropospheric water vapor is particularly important for climate feedbacks, but, as yet, little can be said about how it has varied over the course of the past few decades. IS THE ATMOSPHERIC/OCEANIC CIRCULATION CHANGING? Surprisingly, there is a considerable lack of reliable information about changes in atmospheric circulation,

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*** Virtually certain (probability > 99%) ** Very likely (probability > 90% but < 9%) * Likely (probability > 66% but < 90%) ? Uncertain (probability > 33% but < 66%)

even though it is of daily concern to much of the world since it relates to day-to-day weather changes. Analyses of circulation are performed every day on a routine basis, but the analysis schemes have changed over time, making them of limited use for monitoring climate change. Moreover, even the recent reanalysis efforts by the world’s major numerical weather prediction centers, whereby the analysis scheme is fixed over the historical record, contains time-varying biases because of the introduction of data with time-dependent biases and a changing mix of data (e.g., introducing satellite data) over the course of the reanalysis (13). Even less information is available on measured changes and variations in ocean circulation. A few major atmospheric circulation features have been reasonably well measured because they can be represented by rather simple indices. This includes the ˜ El Nino–Southern Oscillation (ENSO) index, the North Atlantic Oscillation (NAO) index, and the Pacific–North American (PNA) circulation pattern index. There are interesting decadal and multidecadal variation, but it is too early to detect any long-term trends. Evidence exists that ENSO has varied in period, recurrence interval, and strength of impact. A rather abrupt change in ENSO and other aspects of atmospheric circulation seems to have occurred around 1976–1977. More frequent ENSOs with rare excursions into its other extreme ˜ (La Nina) became much more prevalent. Anomalous circulation regimes associated with ENSO and largeamplitude PNA patterns persisted in the North Pacific

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Surface hydrological and storm-related indicators Land

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** No widespread changes in tropical storm frequency/intensity during the 20th century

** 5 to 10% increase in N. Hemisphere mid-to-high latitude precipitation since 1900, with much of it due to heavy and extreme precipitation events * Widespread significant increases in surface water vapor in the N.H., 1975–1995 ***Virtually

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Figure 2. Precipitation trends over land 1900–1999. Trend is expressed in percent per century (relative to the mean precipitation from 1961–1990) and magnitude of trend is represented by area of circle with green reflecting increases and brown decreases of precipitation.

from the late 1970s into the late 1980s, affecting temperature anomalies. Moreover, the NAO has been persistent in its association with strong westerlies into the European continent from the late 1980s until very recently when it abruptly shifted. As a result, temperature anomalies and storminess in Europe have abruptly changed over the past 2 years compared to the past 7 or 8 years. Increases in the strength of the Southern Hemisphere circumpolar vortex during the 1980s have been documented (14,15) using station sea level pressure data. This increase was associated with a delayed breakdown in the stratospheric polar vortex and ozone deficit in the Antarctic spring. A near-global sea level pressure data set has been used to identify changes in circulation patterns in the Indian Ocean. Allan et al. (16) and Salinger et al. (17) find that circulation patterns in the periods 1870–1900 and 1950–1990 were more meridional than those in the 1900–1950 period, indicating intensified circulation around anticyclones. These changes may be related to changes in the amplitude of longwave troughs to the south and west of Australia and the Tasman Sea/New Zealand area and a subsequent decrease in precipitation in Southwest Australia (18,19).

IS THE WEATHER AND CLIMATE BECOMING MORE EXTREME OR VARIABLE? Climate and weather extremes are of great interest. Due to inadequate monitoring as well as prohibitively expensive access to weather and climate data held by the world’s national weather and environmental agencies, only limited reliable information is available about large-scale changes in extreme weather or climate variability. The time-dependent biases that affect climate means are even more difficult to effectively eliminate from the extremes of the distributions of various weather and climate elements. There are a few areas, however, where regional and global changes in weather and climate extremes have been reasonably well documented (20). Interannual temperature variability has not changed significantly over the past century. On shorter time scales and higher frequencies, e.g., days to a week, temperature variability may have decreased across much of the Northern Hemisphere (8). Related to the decrease in high-frequency temperature variability there has been a tendency for fewer low-temperature extremes, but widespread changes in extreme high temperatures have not been noted. Trends in intense rainfall have been examined for a variety of countries. Some evidence suggests an increase in

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intense rainfalls (United States, tropical Australia, Japan, and Mexico), but analyses are far from complete and subject to many discontinuities in the record. The strongest increases in extreme precipitation are documented in the United States and Australia (21). Intense tropical cyclone activity may have decreased in the North Atlantic, the one basin with reasonably consistent tropical cyclone data over the twentieth century, but even here data prior to World War II is difficult to assess regarding tropical cyclone strength. Elsewhere, tropical cyclone data do not reveal any longterm trends, or if they do they are most likely a result of inconsistent analyses. Changes in meteorological assimilation schemes have complicated the interpretations of changes in extratropical cyclone frequency. In some regions, such as the North Atlantic, a clear trend in activity has been noted, as also in significant wave heights in the northern half of the North Atlantic. In contrast, decreases in storm frequency and wave heights have been noted in the south half of the North Atlantic over the past few decades. These changes are also reflected in the prolonged positive excursions of the NAO since the 1970s.

IS THE RADIATIVE FORCING OF THE PLANET CHANGING? Understanding requires a time history of forcing global change. The atmospheric concentration of CO2 , an important greenhouse gas because of its long atmospheric residence time and relatively high atmospheric concentration, has increased substantively over the past few decades. This is quite certain as revealed by precise measurements made at the South Pole and at Mauna Loa Observatory since the late 1950s, and from a number of stations distributed globally that began operating in subsequent decades. Since atmospheric carbon dioxide is a long-lived atmospheric constituent and it is well mixed in the atmosphere, a moderate number of well-placed stations operating for the primary purpose of monitoring seasonal to decadal changes provides a very robust estimate of global changes in carbon dioxide. To understand the causes of the increase of atmospheric carbon dioxide, the carbon cycle and the anthropogenic carbon budget must be balanced. Balancing the carbon budget requires estimates of the sources of carbon from anthropogenic emissions from fossil fuel and cement production, as well as the net emission from changes in land use (e.g., deforestation). These estimates are derived from a combination of modeling, sample measurements, and high-resolution satellite imagery. It also requires measurements for the storage in the atmosphere, the ocean uptake, and uptake by forest regrowth, the CO2 and nitrogen fertilization effect on vegetation, as well as any operating climate feedback effects (e.g., the increase in vegetation due to increased temperatures). Many of these factors are still uncertain because of a paucity of ecosystem measurements over a sustained period of time. Anthropogenic emissions from the burning of fossil fuel and cement production are the primary cause of the atmospheric increase.

Several other radiatively important anthropogenic atmospheric trace constituents have been measured for the past few decades. These measurements have confirmed significant increases in atmospheric concentrations of CH4 , N2 O, and the halocarbons including the stratospheric ozone destructive agent of the chloroflourocarbons and the bromocarbons. Because of their long lifetimes, a few wellplaced high-quality in situ stations have provided a good estimate of global change. Stratospheric ozone depletion has been monitored both by satellite and ozonesondes. Both observing systems have been crucial in ascertaining changes of stratospheric ozone that was originally of interest, not because of its role as a radiative forcing agent, but its ability to absorb ultraviolet (UV) radiation prior to reaching Earth’s surface. The combination of the surfaceand space-based observing systems has enabled much more precise measurements than either system could provide by itself. Over the past few years much of the ozonesonde data and satellite data has been improved using information about past calibration methods, in part because of differences in trends between the two observing systems. Figure 3 depicts the IPCC (9) best estimate of the radiative forcing associated with various atmospheric constituents. Unfortunately, measurements of most of the forcings other than those already discussed have low or very low confidence, not only because of our uncertainty about their role in the physical climate system, but because we have not adequately monitored their change. For example, estimates of changes in sulfate aerosol concentrations are derived from model estimates of source emissions, not actual atmospheric concentrations. The problem is complicated because of the spatially varying concentrations of sulfate due to its short atmospheric lifetime. Another example is measurements of solar irradiance, which have been taken by balloons and rockets for several decades, but continuous measurements of top-of-the-atmosphere solar irradiance did not begin until the late 1970s with the Nimbus

Trend (%/century) in annual precipitation 1900–1999 85N 55N 30N 10N 10S 30S 55S 85S −50% −40% −30% −20% −10% 0%

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Figure 3. Estimates of globally and annually averaged radiative forcing (in W/m−2 ) for a number of agents due to changes in concentrations of greenhouse gases and aerosols and natural changes in solar output from 1750 to the present day. Error bars are depicted for all forcings (from Ref. 1).

OBSERVATIONS OF CLIMATE AND GLOBAL CHANGE FROM REAL-TIME MEASUREMENTS

7 and the Solar Maximum Mission satellites. There are significant absolute differences in total irradiance between satellites, emphasizing the critical need for overlaps between satellites and absolute calibration of the irradiance measurements to determine decadal changes. Spectrally resolved measurements will be a key element in our ability to model the effects of solar variability, but at the present time no long-term commitment has been made to take these measurements. Another important forcing that is estimated through measured, modeled, and estimated changes in optical depth relates to the aerosols injected high into the atmosphere by major volcanic eruptions. The aerosols from these volcanoes are sporadic and usually persist in the atmosphere for at most a few years. Improved measurements of aerosol size distribution and composition will help better understand this agent of climate change. WHAT CAN WE DO TO IMPROVE OUR ABILITY TO DETECT CLIMATE AND GLOBAL CHANGE? Even after extensive reworking of past data, in many instances we are incapable of resolving important aspects concerning climate and global change. Virtually every monitoring system and data set requires better data quality, continuity, and fewer time-varying biases if we expect to conclusively answer questions about how the planet has changed, because of the need to rely on observations that were never intended to be used to monitor the physical characteristics of the planet of the course of decades. Long-term monitoring, capable of resolving decade-to-century-scale changes, requires different strategies of operation. In situ measurements are currently in a state of decay, decline, or rapid poorly documented change due to the introduction of automated measurements without adequate precaution to understand the difference between old and new observing systems. Satellite-based systems alone will not and cannot provide all the

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necessary measurements. Much wiser implementation and monitoring practices must be adopted for both spacebased and surface-based observing systems in order to adequately understand global change. The establishment of the Global Climate Observing System (GCOS) is a high priority (22), and continued encouragement by the World Meteorological Organization (WMO) of a full implementation of this system in all countries is critical. Furthermore, in the context of the GCOS, a number of steps can be taken to improve our ability to monitor climate and global change. These include: 1. Prior to implementing changes to existing environmental monitoring systems or introducing new observing systems, standard practice should include an assessment of the impact of these changes on our ability to monitor environmental variations and changes. 2. Overlapping measurements in time and space of both the old and new observing systems should be standard practice for critical environmental variables. 3. Calibration, validation, and knowledge of instrument, station, and/or platform history are essential for data interpretation and use. Changes in instrument sampling time, local environmental conditions, and any other factors pertinent to the interpretation of the observations and measurements should be recorded as a mandatory part of the observing routine and be archived with the original data. The algorithms used to process observations must be well documented and available to the scientific community. Documentation of changes and improvements in the algorithms should be carried along with the data throughout the data archiving process. 4. The capability must be established to routinely assess the quality and homogeneity of the historical

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Figure 4. Global, annual-mean radiative forcings (Wm−2 ) due to a number of agents for the period from pre-industrial (1750) to the present. The height of the vertical bar denotes the central or ‘‘best’’ estimate, no bar indicates that it is not possible to provide a ‘‘best’’ estimate. The vertical line indicates an estimate of the uncertainty range and the level of scientific understanding is a subjective judgement about the reliability of the forcing estimate based on such factors as assumptions, degree of knowledge of the physical/chemical mechanisms, etc. (From Ref. 1).

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5.

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database for monitoring environmental variations and change, including long-term high-resolution data capable of resolving important extreme environmental events. Environmental assessments that require knowledge of environmental variations and change should be well integrated into a global observing system strategy. Observations with a long uninterrupted record should be maintained. Every effort should be made to protect the data sets that document long-term homogeneous observations. Long term may be a century or more. A list of prioritized sites or observations based on their contribution to long-term environmental monitoring should be developed for each element. Data-poor regions, variables, regions sensitive to change, and key measurements with inadequate temporal resolution should be given the highest priority in the design and implementation of new environmental observing systems. Network designers, operators, and instrument engineers must be provided environmental monitoring requirements at the outset of network design. This is particularly important because most observing systems have been designed for purposes other than long-term monitoring. Instruments must have adequate accuracy with biases small enough to resolve environmental variations and changes of primary interest. Much of the development of new observation capabilities and much of the evidence supporting the value of these observations stem from researchoriented needs or programs. Stable, long-term commitments to these observations, and a clear transition plan from research to operations, are two requirements in the development of adequate environmental monitoring capabilities. Data management systems that facilitate access, use, and interpretation are essential. Freedom of access, low cost, mechanisms that facilitate use (directories, catalogs, browse capabilities, availability of metadata on station histories, algorithm accessibility and documentation, on-line accessibility to data, etc.), and quality control should guide data management. International cooperation is critical for successful management of data used to monitor long-term environmental change and variability.

BIBLIOGRAPHY 1. IPCC (2001). Climate Change, 2001: The Scientific Basis. Contribution of Working Group 1 to the Third Assessment Report of the Intergovernmental Panel on Climate Change. J.T. Houghton, Y. Ding, D.J. Griggs, M. Noguer, P.J. van der Linden, X. Dai, K. Maskell, and C.A. Johnson (Eds.). Cambridge University Press, New York. 2. Spencer, R.W. and Christy, J.R. (1992). Precision and radiosonde validation of satellite gridpoint temperature anomalies, Part I: MSU channel 2. J. Climate 5: 847–857.

3. Spencer, R.W. and Christy, J.R. (1992). Precision and radiosonde validation of satellite gridpoint temperature anomalies, Part II: A tropospheric retrieval and trends during 1979–90. J. Climate 5: 858–866. 4. Reynolds, R.W. (1988). A real-time global sea surface temperature analysis. J. Climate 1: 75–86. 5. National Research Council (NRC) (2000). Reconciling Observations of Global Temperature Change, Report of the Panel on Reconciling Temperature Observations. National Academy Press, Washington, DC. 6. Luers, J.K. and Eskridge, R.E. (1995). Temperature corrections for the VIZ and Vaisala radiosondes. Appl. Meteor. 34: 1241–1253. 7. Sevruk, B. (1982). Methods of correcting for systematic error in point precipitation measurements for operational use. Hydrology Rep. 21. World Meteorological Organization, Geneva, 589. 8. Karl, T.R., Knight, R.W., and Plummer, N. (1995). Trends in high-frequency climate variability in the twentieth century. Nature 377: 217–220. 9. IPCC (1992). Climate Change, 1992, Supplementary Report, WMO/UNEP. J.T. Houghton, B.A. Callander, and S.K. Varney (Eds.). Cambridge University Press, New York, pp. 62–64. 10. Spencer, R.W. (1993). Global oceanic precipitation from the MSU during 1979–1992 and comparison to other climatologies. J. Climate 6: 1301–1326. 11. Huffman, G.J., Adler, R.F., Arkin, P., Chang, A., Ferraro, R., Gruber, A., Janowiak, J., McNab, A., Rudolf, B., and Schneider, U. (1997). The global precipitation climatology project (GPCP) combined precipitation dataset. Bull. Am. Meteor. Soc. 78: 5–20. 12. Rossow, W.B. and Cairns, B. (1995). Monitoring changes in clouds. Climatic Change 31: 175–217. 13. Trenberth, K.E. and Guillemot, C.J. (1997). Evaluation of the atmospheric moisture and hydrologic cycle in the NCEP reanalysis. Clim. Dyn. 14: 213–231. 14. van Loon, H., Kidson, J.W., and Mullan, A.B. (1993). Decadal variation of the annual cycle in the Australian dataset. J. Climate 6: 1227–1231. 15. Hurrell, J.W. and van Loon, H. (1994). A modulation of the atmospheric annual cycle in the Southern Hemisphere. Tellus 46A: 325–338. 16. Allan, R.J., Lindesay, J.A., and Reason, C.J.C. (1995). Multidecadal variability in the climate system over the Indian Ocean region during the austral summer. J. Climate 8: 1853–1873. 17. Salinger, M.J., Allan, R., Bindoff, N., Hannah, J., Lavery, B., Leleu, L., Lin, Z., Lindesay, J., MacVeigh, J.P., Nicholls, N., Plummer, N., and Torok, S. (1995). Observed variability and change in climate and sea level in Oceania. In: Greenhouse: Coping with Climate Change. W.J. Bouma, G.I. Pearman, and M.R. Manning (Eds.). CSIRO, Melbourne, Australia, 100–126. 18. Nicholls, N. and Lavery, B. (1992). Australian rainfall trends during the twentieth century. Int. J. Climatology 12: 153–163. 19. Allan, R.J. and Haylock, M.R. (1993). Circulation features associated with the winter rainfall decrease in southwestern Australia. J. Climate 6: 1356–1367. 20. Easterling, D.R., Meehl, G., Parmesan, C., Changnon, S., Karl, T., and Mearns, L. (2000). Climate extremes: observations, modeling, and impacts. Science 289: 2068–2074.

OVERVIEW: THE CLIMATE SYSTEM 21. Easterling, D.R., Evans, J.L., Groisman, P.Ya., Karl, T.R., Kunkel, K.E., and Ambenje, P. (2000). Observed variability and trends in extreme climate events: A brief review. Bull. Am. Meteor. Soc. 81: 417–426. 22. Spence, T. and Townsend, J. (1995). The Global Climate Observing System (GCOS). Climatic Change 31: 131–134.

READING LIST Elliott, W.P. (1995). On detecting long-term changes in atmospheric moisture. Climatic Change 31: 219–237. Groisman, P.Y., and Legates, D.R. (1995). Documenting and detecting long-term precipitation trends: Where we are and what should be done. Climatic Change 31: 471–492. Hurrell, J.W. and Trenberth, K.E. (1996). Satellite versus surface estimates of air temperature since 1979. J. Climate 9: 2222–2232.

OVERVIEW: THE CLIMATE SYSTEM ROBERT E. DICKINSON (from The Handbook of Weather, Climate, and Water: Dynamics, Climate, Physical Meteorology, Weather Systems, and Measurements, Wiley 2003)

The climate system consists of the atmosphere, cryosphere, oceans, and land interacting through physical, chemical, and biological processes. Key ingredients are the hydrological and energy exchanges between subsystems through radiative, convective, and fluid dynamical mechanisms. Climate involves changes on seasonal, year-to-year, and decadal or longer periods in contrast to day-to-day weather changes. However, extreme events and other statistical measures are as, or more, important than simple averages. Climate is seen to impact human activities most directly through the occurrence of extremes. The frequency of particular threshold extremes, as, for example, the number of days with maximum temperatures above 100 ◦ F, can change substantially with shifts in climate averages.

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increases with altitude in the stratosphere in response to increasing heating per the unit mass by ozone absorption of ultraviolet radiation. The variation of temperature structure with latitude is indicated in Fig. 2. The troposphere is deepest in the tropics because most thunderstorms occur there. Because of this depth and stirring by thunderstorms, the coldest part of the atmosphere is the tropical tropopause. In the lower troposphere temperatures generally decrease from the equator to pole, but warmest temperatures shift toward the summer hemisphere, especially in July. Longitudinally averaged winds are shown in Fig. 3. Because of the geostrophic balance between wind and pressures, winds increase with altitude where temperature decreases with latitude. Conversely, above about 8 km, where temperatures decrease toward the tropical tropopause, the zonal winds decrease with altitude. The core of maximum winds is referred to as the jet stream. The jet stream undergoes large wavelike oscillations in longitude and so is usually stronger at a given latitude than in its longitudinal average. These waves are especially noticeable in the winter hemisphere as illustrated in Fig. 4. GLOBAL AVERAGE ENERGY BALANCE Solar radiation of about 342 W/m−2 entering Earth’s atmosphere is absorbed and scattered by molecules. The major gaseous absorbers of solar radiation are water vapor in the troposphere and ozone in the stratosphere. Clouds and aerosols likewise scatter and absorb. Clouds are the dominant scatterer and so substantially enhance the overall planetary reflected radiation, whose ratio to incident solar radiation, about 0.31, is referred to as albedo. Thermal infrared radiation, referred to as longwave, is controlled by clouds, water vapor, and other greenhouse gases. Figure 5 (4) illustrates a recent estimate of the various terms contributing to the global 120 Thermosphere

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The atmosphere is described by winds, pressures, temperatures, and the distribution of various substances in gaseous, liquid, and solid forms. Water is the most important of these substances. Also important are the various other radiatively active (‘‘greenhouse’’) gases, including carbon dioxide and liquid or solid aerosol particulates. Most of the mass of the atmosphere is in the troposphere, which is comprised of the layers from the surface to about 12 km (8 km in high latitudes to 16 km at the equator) where the temperature decreases with altitude. The top of the troposphere is called the tropopause. Overlying this is the stratosphere, where temperatures increase with altitude to about 50 km or so (Fig. 1). The tropospheric temperature decreases with altitude are maintained by vertical mixing driven by moist and dry convection. The temperature

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energy balance. The latent heat from global average precipitation of about 1.0 m per year is the dominant nonradiative heating term in the atmosphere. Because of the seasonally varying geometry of Earth relative to the sun, and the differences in cloudiness and surface albedos, there are substantial variations in the distribution of absorbed solar radiation at the surface and in the atmosphere, as likewise in the transfer of latent heat from the surface to the atmosphere. This heterogeneous distribution of atmospheric heating drives atmospheric wind systems, either directly or through the creation of available potential energy, which is utilized to maintain random occurrences of various kinds of instabilities, such as thunder-storms and wintertime

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cyclonic storm systems. These dynamical systems hence act to redistribute energy within the atmosphere and so determine the distributions of temperature and water vapor. Likewise, the balances at the surface between fluxes of radiative, latent, and thermal energies determine surface temperatures and soil moistures. The properties of the near-surface air we live in are determined by a combination of surface and atmospheric properties, according to processes of the atmospheric boundary layer. Thus climatic anomalies in surface air may occur either because of some shift in atmospheric circulation patterns or through some modification of surface properties such as those accompanying deforestation or the development of an urban area.

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THE ATMOSPHERIC BOUNDARY LAYER The term boundary layer is applied in fluid dynamics to layers of fluid or gas, usually relatively thin, determining the transition between some boundary and the rest of the fluid. The atmospheric boundary layer is the extent of atmosphere that is mixed by convective and mechanical stirring originating at Earth’s surface. Such stirring is commonly experienced by airplane travelers as the bumps that occur during takeoff or landing, especially in the afternoon, or as waves at higher levels in flying over mountainous regions. The daytime continental boundary layer, extending up to several kilometers in height, is most developed and vigorously mixed, being the extent to which the daytime heating of the surface drives convective overturning of the atmosphere. The land cools radiatively at night, strongly stabilizing the atmosphere against convection, but a residual boundary layer extends up to about 100 m stirred by the flow of air over the underlying rough surface. This diurnal variation of fluxes over the ocean is much weaker and the boundary layer is of intermediate height. The temperature of the atmosphere, when stirred by dry mixing, decreases at a rate of 9.8 K/km. Above the boundary layer, temperatures decrease less rapidly with height, so that the atmosphere is stable to dry convection. A layer of clouds commonly forms at the top of the daytime and oceanic boundary layers and contributes to the convection creating the boundary layer through its radiative cooling (convection results from either heating at the bottom of a fluid or cooling at its top). Also, at times the clouds forming near the top of the boundary layer can be unstable to moist convection, and

Figure 4. Mean 500-mb contours in January, Northern Hemisphere. Heights shown in tens of meters (3).

so convect upward through a deep column such as in a thunderstorm. ATMOSPHERIC HYDROLOGICAL CYCLE The storage, transport, and phase changes of water at the surface and in the atmosphere are referred to as the hydrological cycle. As already alluded to, the hydrological cycle is closely linked to and driven by various energy exchange processes at the surface and in the atmosphere. On the scale of continents, water is moved from oceans to land by atmospheric winds, to be carried back to the oceans by streams and rivers as elements of the land hydrological cycle. Most of the water in the atmosphere is in its vapor phase, but water that is near saturation vapor pressure (relative humidity of 100%) converts to droplets or ice crystals depending on temperature and details of cloud physics. These droplets and crystals fall out of the atmosphere as precipitation. The water lost is replenished by evaporation of water at the surface and by vertical and horizontal transport within the atmosphere. Consequently, much of the troposphere has humidities not much below saturation. Saturation vapor pressure increases rapidly with temperature (about 10% per kelvin of change). Hence, as illustrated in Fig. 6, the climatological concentrations of water vapor vary from several percent or more when going from near-surface air to a few parts per million near the tropical tropopause. Water vapor concentrations in the stratosphere are close to that of the tropical tropopause, probably because much of the air in the lower stratosphere has been pumped through the tropical tropopause by moist convection.

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OVERVIEW: THE CLIMATE SYSTEM Reflected solar radiation 107 W m-2

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CLIMATE OF THE STRATOSPHERE The dominant radiative processes in the stratosphere are the heating by absorption of solar ultra violet (UV) radiation and cooling by thermal infrared emission from carbon dioxide and, to a lesser extent, ozone molecules. The stratospheric absorption of UV largely determines how much harmful UV reaches the surface. Ozone in the upper troposphere and lower stratosphere additionally adds heat by absorption of thermal emission from the warmer surface and lower layers. The stratosphere, furthermore, enhances the greenhouse warming of CO2 in the troposphere through substantial downward thermal emissions to the troposphere.

How changes of ozone change stratospheric and tropospheric radiative heating depends on the amounts of overlying ozone and, for thermal effect, on pressure and radiative upwelling depending on underlying temperatures. Besides radiative processes, stratospheric climate is characterized by its temperature and wind patterns and by the chemical composition of its trace gases. At midstratosphere, temperature increases from winter pole to summer pole with an accompanying eastward jet stream in the winter hemisphere extending upward from the tropospheric jet steam. This wind configuration allows planetary wave disturbances to propagate into the stratosphere, contributing significant temporal and longitudinal

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variabilities. Conversely, the westward jet, found in the summer stratosphere attenuates wave disturbances from below, and so is largely zonally symmetric, changing only with the seasonal heating patterns.

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4. Kiehl, J.T. and Trenberth, K.E. (1997). Earth’s annual global mean energy budget. J. Clim. 78: 197–208. 5. Peixoto, J.P. and Oort, A.H. (1992). Observed atmospheric branch of the hydrological cycle, Chapter 12.3. Physics of Climate. American Institute of Physics, New York.

THE CRYOSPHERE The term cryosphere refers to the components of the climate system dominated by water in its frozen phase, that is, in high latitudes and extratropical winter conditions. Elements include snow, its distribution and depths, sea ice, its distribution and properties, highlatitude ice caps, and temperate glaciers. The largest volume of frozen water is stored in ice caps, and glaciers. This storage acts to remove water from the oceans. How it changes with climate change is, hence, of interest for determining changing sea levels. Ice is highly reflective of sunlight, especially in crystal form. The loss of solar heating because of this high albedo acts to substantially reduce high-latitude temperatures especially in spring and early summer where near-maximum solar radiation sees white snow-covered surfaces. This high albedo can be substantially masked by cloud cover and, over land, tall vegetation such as conifer forests. THE OCEAN Oceans are a major factor in determining surface temperatures and fluxes of water into the atmosphere. They store, release, and transport thermal energy, in particular, warming the atmosphere under wintertime and high-latitude conditions, and cooling it under summer and tropical conditions. How the oceans carry out these services depends on processes coupling them to the atmosphere. Atmospheric winds push the oceans into wind-driven circulation systems. Net surface heating or cooling, evaporation, and precipitation determine oceanic densities through controlling temperature and salinity, hence oceanic buoyancy. This net distribution of buoyancy forcing drives ‘‘thermohaline’’ overturning of the ocean, which acts to transport heat. Climate of the surface layers of the ocean includes the depth to which waters are stirred by waves and net heating or cooling. Heating acts to generate shallow warm stable layers, while cooling deepens the surface mixed layers. Under some conditions, convective overturning of cold and/or high-salinity water can penetrate to near the ocean bottom. BIBLIOGRAPHY 1. Hartmann, D.L. (1994). Atmospheric temperature, Chapter 1.2. In: Global Physical Climatology. Academic Press, San Diego. 2. Grotjahn, R. (1993). Zonal average observations, Chapter 3. In: Global Atmospheric Circulations: Observations and Theories. Oxford University Press, New York. 3. Holton, J.R. (1992). The observed structure of extratropical circulations, Chapter 6.1. In: An Introduction to Dynamic Meteorology. Academic, San Diego.

CLIMATE AND SOCIETY MICHAEL H. GLANTZ (from The Handbook of Weather, Climate, and Water: Atmospheric Chemistry, Hydrology, and Societal Impacts, Wiley 2003)

At the turn of the twentieth century, scholars who wrote about the interplay between climate and society did so based on their perceptions of climate as a boundary constraint for the development prospects of a society. Perceptions of climate were used as an excuse to dominate societies in Africa, Asia, and Latin America. As a result, climate–society studies soon became viewed as a colonial ploy to control populations in developing areas in the tropics. Perhaps the most cited book in this regard was written by Ellsworth Huntington, Climate and Civilization, published in 1915 (1). In his view, inhabitants of the tropics were destined to accept lower levels of economic and social development because their climate setting was not conducive to lively (i.e., productive) human activity or an aggressive work ethic. According to Huntington, tropical climate was the main culprit causing people in the tropics to be less productive than people in temperate regions. Huntington argued that the temperate climate has an energizing effect on humans. With the growing belief that such an argument was racist in intent, Huntington’s work was challenged, and discussion of the various ways in which climate might influence human behavior was stifled for decades, notwithstanding a few notable exceptions. One such exception is entitled Climate and the Energy of Nations (2) in which Markham referred to the ‘‘air-conditioning revolution,’’ a revolution based on the development and spread of a new technology into the tropical areas. Markham asserted that technology brings islands of temperate-zone climate into the tropics, thereby generating a more aggressive work ethic. Following the end of World War II and the onset of the Cold War between Soviet-style communism and Western capitalism and democracy, attention of governments turned to Cold War conflicts, avoidance of nuclear war, searches for allies, and decolonization. The major Cold War nations were in a competition to show that their approach to economic development was the only way for the newly independent countries to follow. A main stated objective was their intent to assist these countries to become food secure based on the nation’s resources. Consideration of climate was making its way back into the discussions of economic development in developing countries. Once again interest was raised with regard to climatic constraints to economic development in tropical countries. In the 1950s and 1960s, attention focused on decolonization and political development of the newly

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independent states (e.g., 3). In the mid-1970s, a World Bank report about the economic prospects for developing countries—The Tropics and Economic Development: A Provocative Inquiry into the Poverty of Nations—hinted at the economic, social, and political problems caused by climate variability from one year to the next. Its author (4) noted that recurrent droughts in northeast Brazil are a chronic constraint on the region’s economic development prospects. His reference to interannual climate variability was brief and unelaborated. However, climate as a boundary constraint was starting to give way to climate as something that societies might be able to forecast and cope with, at least in its extremes. In the 1970s, attention focused on how the vagaries of weather exposed hundreds of millions of people to hunger and, depending on the socioeconomic situation in a particular country, to famine as well (e.g., 5,6). Thus, there was a growing number of examples of the notion that climate was not really a boundary constraint to the level of development that a people or culture could expect to attain. This notion began to give way to the belief that variability in climate, from one year to the next or one decade to the next, could be coped with so as to soften the impacts of climate variability and weather extremes on agriculture and livestock and, more generally, on the productivity of the land’s surface (e.g., 7,8). Recall that the 1970s was a disruptive decade with respect to climate: 5 years of drought in the West African Sahel (5); failure of the Soviet harvest and subsequent large-scale, low-cost grain purchases by the Soviet Union in the early 1970s (9); the global food crisis (10); talk of a possible return to an ice age (e.g., 11,12); the Ethiopian famine (13); drought-related coups in subSaharan Africa; drought in the wheat-producing Canadian prairie provinces (7); the first drop in global fish catches since the end of World War II (10), and so forth. A devastating 5-year drought from 1968 to 1973 in the West African Sahel and its associated death and environmental destruction in the region drew attention to the impacts on household and village responses to prolonged, multiyear droughts. Widespread droughts around the globe in 1972–1973, famines in West Africa and Ethiopia, ˜ event, along blamed for the most part on an El Nino with the drop in fish landings, prompted the U.N. Secretary General to convene a series of UN-sponsored world conferences on food (1974), population (1974), human settlements (1976), water (1977), desertification (1977), climate (1979) and technology (1979). Thus, toward the middle of the 1970s, at least five new major climate-related scientific issues emerged: the effect of chlorofluorocarbons (CFCs) on the ozone layer in the stratosphere, talk of an impending Ice Age suddenly shifted to talk of a human-induced global warming, acid ˜ Each of these issues rain, desertification, and El Nino. raised interest in climate–society interactions to higher levels among researchers in different disciplines, government agencies, economic sectors, the media, and the public. Societies around the globe responded (and continue to respond) in different ways to each of these climate-related issues. For example, desertification is an environmental issue that is of great concern to African countries.

North Americans, however, refused to accept the view that desertification could occur in the U.S. West as a result of mismanagement of the land’s surface, while noting that desertification was the plight of poor developing countries in Africa. The term desertification first appeared in a report on the destruction of dry forests in central Africa by a French forester (14). Since then, the concept of desertification has been expanded to include such land degradation processes as soil erosion, wind deflation, soil salinization, water logging, livestock overgrazing, and soil trampling. While many of these processes were exposed during the prolonged drought in the West African Sahel and then labeled as desertification, it is not difficult to show that similar processes also take place in the U.S. West. The acid rain issue was addressed in the United States with the implementation by the U.S. Congress of a decade-long national assessment called NAPAP (National Acid Precipitation Assessment Program). Stratospheric ozone depletion was addressed globally in the 1980s with the development of international legal instruments culminating in the Montreal Protocol of 1987 and, later, amendments to it (15). It was in the early 1970s, 1972–1973 to be exact, ˜ event (defined briefly as an invasion of that an El Nino warm water from the Western Pacific into the central and eastern equatorial Pacific Ocean) attracted global attention. An event in 1982–1983, the biggest in a century until that time, captured the full attention of the scientific community and various governments as a natural phenomenon that spawned hazards around the globe. Such hazards included, but were not limited to, droughts, floods, frosts, fires and food shortages, famine, and disease. Ever since the mid-1970s, research funding ˜ of El Nino–related research has been growing along with international interest in the phenomenon and its societal ˜ and environmental impacts. The extraordinary El Nino ˜ and its cold event of 1997–1998 helped to make El Nino ˜ household words throughout much counterpart, La Nina, of the world. Only at the end of the twentieth century ˜ events become of serious interest to the did La Nina ˜ research and forecasting communities (16). This El Nino belated interest is even more surprising given the scientific observation that tropical storms and hurricanes in the Atlantic Basin and in the Gulf of Mexico tend to increase ˜ events and drop in number in number during La Nina ˜ events. during El Nino Global warming is an environmental issue that arose out of discussions and governmental and scientific concerns about the possibility of a global cooling. It was first suggested in 1896 by Swedish chemist Arrhenius (17,18) that the burning of coal by human societies would add enough extra carbon dioxide into the air to eventually heat up Earth’s atmosphere by a few degrees Celsius. This issue was revisited in the 1930s by Callendar (19), who thought that a human-induced global warming of the atmosphere could stave off the imminent recurrence of an ice age. The issue was again revisited in the 1950s when global warming was looked at in neutral terms, as an experiment that societies were performing on the chemistry of the atmosphere, for which the outcome is unknown (20).

CLIMATE AND SOCIETY

It was not until the mid-1970s that human-induced global warming began to be viewed as an adverse event for future generations of human societies and ecosystems that might not be able to adapt to the rate of warming expected to occur. The cause of the warming was attributed to the increasing amounts of greenhouse gases (CO2 , CFCs, CH4 , NOx , collectively referred to as GHGs) being emitted into the atmosphere as a result of human activities. Carbon dioxide is a product of the burning of fossil fuels, and its amount in the atmosphere has been rising since the onset of the Industrial Revolution in the late 1700s. Tropical deforestation also contributes carbon dioxide to the atmosphere. Tropical forests have served as sinks for carbon dioxide, pulling it out of the air and storing it. When trees are felled, decompose or burned, the stored carbon is emitted into the air. Chlorofluorocarbons (CFCs), a greenhouse gas as well as a stratospheric ‘‘ozone eater,’’ are man-made chemicals first discovered in the 1920s for use as a refrigerant. Methane resulting from livestock rearing (e.g., cattle, pigs) and from the increasing number of landfills is another greenhouse gas. Nitrous oxides are used by farmers in fertilizers and have been widely applied to agricultural lands around the globe in increasing amounts since the end of World War II. Of these major greenhouse gases, carbon dioxide is seen at the main culprit in the global warming debate. Current scientific research suggests that the level of climate change that might be expected (at current rates of greenhouse gas emissions) is on the order of 1.5 to 4.5 ◦ C by the end of the twenty-first century (21–23). Concerned with the prospects of a changing global climate, many nations have come together to call for a technical assessment of the state of the science through the Intergovernmental Panel on Climate Change (IPCC). The degree of warming, however, is dependent on numerous factors: the rate at which GHGs continue to be emitted into the atmosphere, the shift by societies to alternative energy sources, the rate of tropical deforestation, the residence time of GHGs in the atmosphere (several of these gases will remain in the atmosphere for decades to centuries), the development of methods to sequester carbon (i.e., taking it from the atmosphere and binding it in some way in Earth’s land surface, vegetation, or oceans), and so forth. Some degree of global warming is inevitable, given the residence time of the GHGs already emitted into the atmosphere. This means that societies around the globe, from local to national, must attempt to ascertain how a warmer global climate regime might affect regional and local climates. Will there be more extreme climate events (such as droughts, floods, frosts, fires) or fewer? These societies must also seriously consider nationally, as well as collectively in cooperation with other countries, the most effective way(s) to cope with the potentially adverse impacts of some degree of human-induced global warming. Coping mechanisms for climate change likely to occur decades in the future can be divided into three categories: preventive, mitigative and adaptive measures. Preventive measures are designed to prevent the increased buildup of GHGs in human-induced global warming, acid rain,

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˜ Each of these issues raised desertification, and El Nino. interest in climate–society interactions to higher levels among researchers in different disciplines, government agencies, economic sectors, the media, and the public. Societies around the globe responded (and continue to respond) in different ways to each of these climate-related issues. For example, desertification is an environmental issue that is of great concern to African countries. Climate-related surprise is not a black-and-white condition. People are hardly ever either totally surprised or never surprised. There are shades of surprise with regard to human responses to the same climate-related event. They can be hardly surprised, mildly surprised, somewhat surprised, very surprised, extremely surprised or totally surprised (NB: each of these examples was taken from the scientific literature). Myers (24, p. 358) introduced the interesting notion of ‘‘semisurprised.’’ Thus, surprise may best be described in ‘‘fuzzy’’ terms with the degree of surprise dependent on several intervening variables such as personal experience, core beliefs, expectations, or knowledge about a phenomenon or about a geographic location. One could argue that there are knowable as well as unknowable surprises (25). Knowable refers to the fact that some climate surprises can be anticipated (24). For example, certain parts of the globe are drought prone. It is known that drought will likely recur. What is not known is exactly when it will take place, how long it will last, how intense it will be, or where its most devastating impacts ˜ is in this category. While we are likely to occur. El Nino have now come to expect these events to recur, we do not know when that will happen or what it will be like. The uncertainty then cascades down the ‘‘impacts chain,’’ and ˜ the as we speculate about likely impacts of an El Nino, degree of uncertainty will increase. ˜ Even with Take, for example, the 1997–1998 El Nino. the best monitoring and observing system in the world focused on minute changes in various aspects of the tropical Pacific Ocean, forecasters and modelers were unable to predict the onset of one of the biggest El ˜ events in the past 100 years. Nor were they able Nino to predict the course of development of that event. They were better than in earlier times, however, at predicting some of its impacts on societies in certain parts of the globe, especially those where the influences of changes in the sea surface temperatures in the tropical Pacific are known to be strong. Societies (and their scientists) are on a learning curve with regard to the various ways that climate variability and climate change might affect climate-related human activities. They must avoid becoming complacent as a result of a belief that they fully understand atmospheric processes or their impacts. They must accept that there will be climate surprises in the future, even if the global climate does not change. They must learn from past experiences on how best to cope with the vagaries of climate (26). Many countries now realize that climate-related problems do not stop at international boundaries. There are many transboundary issues that demand regional (if not international) cooperation, given that countries share river basins, inland seas, airsheds, the global

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atmosphere as well as the onslaught and impacts of extreme meteorological events such as droughts, floods, and tropical storms. While climate-related anomalies cannot be prevented, societal preparation for, and response to, their adverse impacts can be improved through better knowledge of the direct and indirect ways in which atmospheric processes interact with human activities and ecological processes. The enhancement of such knowledge will lead to better forecasts as well as better computer modeling of the interactions among land, sea, and air. A society forewarned of climate-related hazards is forearmed to cope with those hazards more effectively. BIBLIOGRAPHY 1. Huntington, E. Civilization and Climate. Yale University Press, New Haven, Connecticut, 1915, rev, ed. 1924, reprinted by University Press of the Pacific, 2001. 2. Markham, S.F. (1944). Climate and the Energy of Nations. Oxford University Press, London. 3. Pye, L. (1966). Aspects of Political Development. Little, Brown and Co., Boston, Massachusetts. 4. Kamarck, A.M. (1976). The Tropics and Economic Development: A Provocative Inquiry into the Poverty of Nations. The Johns Hopkins University Press, Baltimore, Maryland. 5. Glantz, M.H. (Ed.). (1976). The Politics of Natural Disaster: The Case of the Sahel Drought. Praeger Press, New York. 6. Sen, A. (1981). Poverty and Famines: An Essay on Entitlement and Deprivation. Oxford University Press, Oxford, UK. 7. Glantz, M.H. (Ed.). (1977). Desertification: Environmental Degradation in and around Arid Lands. Westview Press, Boulder, Colorado. 8. Hare, K. (1977). Connections between climate and desertification. Environ. Conserv. 4(2): 82. 9. Trager. (1975). 10. Brown, L., and Eckholm, E.P. (1974). By Bread Alone. Praeger Press, New York. 11. Ponte, L. (1976). The Cooling. Prentice-Hall, Inc., Englewood Cliffs, New Jersey. 12. Weather Conspiracy: The Coming of the New Ice Age, Ballentine Books, New York, compiled by The Impact Team, 1977. 13. Wolde Mariam, M. (1984). Rural Vulnerability to Famine in Ethiopia, 195777, Vikas Publishing House, New Delhi. 14. Aubreville, A. (1949). Climate, Forests and Desertification in Tropical Africa, Soci´ete d’Editions G´eographiques, Maritimes et Coloniales. 15. Benedick, R.E. (1998). Ozone Diplomacy: New Directions in Safeguarding the Planet. Harvard University Press, Cambridge, Massachusetts, Enlarged Edition. 16. Glantz, M.H. (Ed.). (2002). La Nina ˜ and Its Impacts: Facts and Speculation. United Nations University Press, Tokyo, Japan. 17. Arrhenius, S. (1896). On the influence of carbonic acid in the air upon the temperature of the ground. Philos. Mag. 41: 237–276. 18. Arrhenius, S. (1908). Worlds in the Making. Harper & Brothers, New York. 19. Callendar, G.S. (1938). The artificial production of carbon dioxide and its influence on temperatures. Q. J. Roy. Meteor. Soc. 64: 223–237.

20. Revelle, R.R., and Suess, H.E. (1957). Carbon dioxide exchange between atmosphere and ocean and the question of an increase of atmospheric CO2 during the past decades. Tellus 9: 18–27. 21. IPCC (1990). Climate Change: The IPCC Scientific Assessment. Cambridge University Press, Cambridge, UK. 22. IPCC (1996). Climate Change 1995: The Science of Climate Change, Contribution of Working Group I to Second Assessment Report, Cambridge University Press, Cambridge, UK. 23. IPCC (Intergovernmental Panel on Climate Change). (2001). Climate Change 2001: Impacts, Adaptation, and Vulnerability, Contribution of Working Group II to Third Assessment Report, Cambridge University Press, Cambridge, UK. 24. Myers, N. (1995). Environmental unknowns. Science 269: 358–360. 25. Streets, D.G., and Glantz, M.H. (2000). Exploring the concept of climate surprise. Global Environ. Chang. 10: 97–107. 26. Glantz, M.H. (Ed.). (1988). Societal Responses to Regional Climatic Change: Forecasting by Analogy. Westview Special Study, Boulder, Colorado.

READING LIST Smith, J.B., Bhatti, N., Menzhulin, G., Benioff, R., Bodyko, M.I., Campos, M., Jallow, B., and Rijsberman, F. (Eds.). (1996). Adapting to Climate Change: An International Perspective. Springer-Verlag, New York.

WHAT IS CLIMATOLOGY? National Drought Mitigation Center

‘‘CLIMATE IS WHAT YOU EXPECT. WEATHER IS WHAT YOU GET.’’ Weather is the condition of the atmosphere over a brief period of time. For example, we speak of today’s weather or the weather this week. Climate represents the composite of day-to-day weather over a longer period of time. People in Minneapolis–St. Paul expect a white Christmas, and people in New Orleans expect very warm, humid summers. And a traveler going to Orlando, Florida, in March will not pack the same kind of clothing as a traveler going to Vail, Colorado, in March. These examples show how climate influences our daily lives. Additionally: • Our houses are designed based on the climate where we live. • Farmers make plans based on the length of the growing season from the last killing freeze in the spring to the first freeze in the fall. • Utility companies base power supplies on what they expect to be the maximum need for heating in the winter and the maximum need for cooling in the summer. This article is a US Government work and, as such, is in the public domain in the United States of America.

CLOUD SEEDING

A climatologist attempts to discover and explain the impacts of climate so that society can plan its activities, design its buildings and infrastructure, and anticipate the effects of adverse conditions. Although climate is not weather, it is defined by the same terms, such as temperature, precipitation, wind, and solar radiation. Climate is usually defined by what is expected or ‘‘normal’’, which climatologists traditionally interpret as the 30-year mean. By itself, ‘‘normal’’ can be misleading unless we also understand the concept of variability. For example, many people consider sunny, idyllic days normal in southern California. History and climatology tell us that this is not the full story. Although sunny weather is frequently associated with southern California, severe floods have had a significant impact there, including major floods in 1862 and 1868, shortly after California became a state. When you also factor in severe droughts, most recently those of 1987–94, a more correct statement would be that precipitation in southern California is highly variable, and that rain is most likely between October and April. The misconception that weather is usually normal becomes a serious problem when you consider that weather, in one form or another, is the source of water for irrigation, drinking, power supply, industry, wildlife habitat, and other uses. To ensure that our water supply, livelihoods, and lives are secure, it is essential that planners anticipate variation in weather, and that they recognize that drought and flood are both inevitable parts of the normal range of weather. HOW DOES CLIMATOLOGY HELP US PREPARE FOR DROUGHT? Climatology provides benchmarks, such as the drought of record. The drought of record is the drought remembered as having the greatest impact on a region. Most of us are not consciously aware of how much the climate fluctuates from one decade or century to the next. One way for reservoir managers, municipal water suppliers, and other planners to check reality is to compare expectations of water supply against a region’s drought of record. But caution is necessary here: although the weather conditions could recur, the impacts would likely be very different. For most of the country, the drought of record was 30 to 60 years ago, and population concentrations and water use patterns have shifted substantially since then. Planners need to consider and watch out for a variety of problems and misconceptions. Specifically, climatology answers crucial questions such as: • • • •

How often does drought occur in this region? How severe have the droughts been? How widespread have the droughts been? How long have the droughts lasted?

Examining water supplies and understanding the impact of past droughts help planners anticipate the effects of drought:

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• What would happen if the drought of record occurred here now? • Who are the major water users in the community, state, or region? • Where does our water supply come from and how would the supplies be affected by a drought of record? • What hydrological, agricultural, and socioeconomic impacts have been associated with the various droughts? • How can we prepare for the next drought of record?

CLOUD SEEDING ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

Cloud seeding is the deliberate treatment of certain clouds or cloud systems to affect the precipitation process within those clouds. The treating of clouds can alter weather conditions, thus cloud seeding is also called weather modification. The technology that led to cloud seeding has been developed only during the past 60 years. Yet as this technology becomes more efficient, increased worldwide application of cloud seeding for practical use is likely to continue into the future. THE CLOUD SEEDING PROCESS To understand how cloud seeding works, one must understand some facts about the weather. All air in the atmosphere contains moisture. Warm air rises from the earth’s surface and begins to cool. As the air cools, the moisture condenses into tiny droplets that make up clouds. A cloud is almost 100% air. The tiny droplets composing clouds are not heavy enough to fall to the ground until they merge with millions of other droplets at temperatures below 32 ◦ F and interact with dust, salt, or sand particles. One type of cloud seeding, known as cold cloud seeding, introduces silver iodide and other agents to enhance ice crystallization in clouds colder than 32 ◦ F. This is often called as the static seeding effect. Once the droplets within the clouds freeze, the resulting ice crystals grow at the expense of the water droplets surrounding them, a process called sublimation. Other crystals grow through contact with neighboring droplets; this is known as riming. Through these two processes, the tiny crystals form snowflakes. If these snowflakes are heavy enough, they fall from the clouds and, depending on the temperatures below, will come to the earth as snow, raindrops, or a mixture. The other type of cloud seeding, warm cloud seeding, produces rainfall from clouds that are above 32 ◦ F. This process involves introducing additional condensation nuclei (salt particles), which cause additional water droplets to condense within the clouds. If the collision of these particles makes the droplets heavy enough, precipitation can fall from the clouds. In each case, it will usually take the clouds 20 to 30 minutes to produce rain, making it crucial to monitor cloud movement.

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CONDENSATION

USES OF WEATHER MODIFICATION The primary goal of most weather modification projects is to increase levels of precipitation or to suppress fog or hail. Water agencies, local municipalities, farmers, ranchers, hydroelectric power facility operators, ski resort owners, and others sponsor cloud seeding activities. To date, cloud seeding has successfully stimulated precipitation in more than 50 countries. More than half of the United States now has some type of regulation concerning cloud seeding programs. More recently, cloud seeding has been used to suppress certain damaging effects of weather. Airports have employed programs to disperse fog levels and increase flight visibility. And in areas damaged by hail, programs have been undertaken to decrease hailstorms. As technology continues to enhance the weather modification process, it can be assumed that more entities will sponsor cloud seeding programs. Cloud seeding technology is highly portable and very flexible to changing weather conditions. For those who use it, cost/benefit ratios are typically very favorable.

still contain only about 3% of the moisture, leaving 97% available. Furthermore, some analyses of precipitation data downwind from seeding projects have indicated small increases in precipitation. To date, no scientific studies have shown that some areas receive precipitation at the expense of their neighbors. FUTURE OF CLOUD SEEDING Despite all of the advances made during the past 60 years, weather modification continues to receive relatively little support. Many states and the federal government acknowledge and regulate cloud seeding, but few provide significant funding for weather modification studies or projects. Most scientists insist that more research must be done and more data must be gathered before weatherchanging programs garner public trust on a larger scale. However, most researchers are optimistic that new studies and technological advances will continue to advance the science of weather modification. READING LIST

EFFECTIVENESS OF CLOUD SEEDING From the earliest experiment, which produced just a few droplets of rain, cloud seeding has progressed to make some significant and valuable impacts on weather. Various studies have documented the effects of different programs. For the most part, these programs are overwhelmingly successful. To augment precipitation, well-designed and well-conducted projects yield an average winter precipitation of 5 to 20% more in continental regions and 5 to 30% more in coastal areas; they have yielded as much as 100% more in warm weather. Increases depend on a variety of factors, including spatial coverage of suitable cloud systems and the frequency of different systems. ENVIRONMENTAL HEALTH CONCERNS OF WEATHER MODIFICATION To date, no significant environmental problems have been attributed to cloud seeding programs, though government agencies, private firms, and research institutions have conducted many studies. Researchers believe that any negative effects are minimal because relatively little seeding material is used compared to the large surface area that is targeted. For example, the most common seeding material, silver iodide, will yield a concentration of less than 0.1 microgram per liter in rainwater or snow. The U.S. Public Health Service states that the acceptable amount of silver iodide in water is 50 micrograms per liter. One other common question about cloud seeding is whether the stimulation of rainfall in one area results in decreased rainfall in other areas. It does not. Clouds are inefficient in the way they gather and distribute moisture. They never gather or release all the moisture that is available; rather, clouds gather only about 1% of the moisture in the atmosphere. Even if a seeding program tripled the efficiency of cloud formation, the cloud would

American Society of Civil Engineers. (1995). Guidelines for Cloud Seeding to Augment Precipitation. ASCE Manuals and Reports on Engineering Practice No. 81. Stauffer, N.E. and Williams, K. (2000). Utah Cloud Seeding Program, Increased Runoff/Cost Analysis. Technical Report, Utah Department of Natural Resources, Division of Water Resources. Weather Modification Advisory Board. (1978). The Management of Weather Resources. Report to the Secretary of Commerce, 2 volumes. Klien, D.A. (1978). Environmental Impacts of Artificial Ice Nucleating Agents. Dowden, Hutchinson & Ross, Inc., Stroudsburg, PA. Brown, K.J., Elliot, R.D., and Edelstein, M.W. (1978). Transactions of Total-Area Effects of Weather Modification. Report to the National Science Foundation on a workshop held August 8–12, 1977, Fort Collins, CO. Cloud Seeding (Weather Modification) Frequently Asked Questions—http://www.xmission.com/∼nawc/wmfaq.html. The Science of Cloud Seeding—http://twri.tamu.edu/twripubs/ WtrResrc/v20n2/text-1.htm. Cloud Seeding Circus.com—www.cloudseedingcircus.com. Al Weather Modification Page—www.atmos-inc.com/weamod. html.

CONDENSATION ALDO CONTI Frascati (RM), Italy

Condensation is the physical process by which a vapor changes to its liquid state. Condensation happens when the temperature of the vapor decreases below the so-called dew point. In physics, this sort of process is called a phase change because the matter involved changes its state (i.e., from gaseous to liquid, in this case). Water condensation

COSMIC WATER

is evident in our atmosphere; it produces fog, mist, dew, clouds, and rain, depending on the conditions. It can be seen on a cold sheet of glass, where condensed water forms a maze of droplets. Condensation depends mainly on temperature, and it is a process that happens at the molecular level. When the temperature is high, the molecules in a vapor have plenty of energy and collide at high speed, which means that the molecules bounce immediately and do not stay together long enough to establish a bond. In this situation, the vapor remains. But when the temperature, and therefore the speed, decreases, then the molecules can stick to each other. The result is a droplet of liquid. Condensation is ruled not only by temperature, but by pressure as well, which is the reason why one speaks about dew point and not dew temperature. As a general rule, the dew point temperature increases with pressure. Condensation, with evaporation, is very important in the water cycle of the earth. By condensation, water falls back as rain, hail, or snow and becomes available again for human use.

COSMIC WATER D.L. MARRIN Hanalei, Hawaii

Once believed by science to be the substance that distinguished the earth from the rest of the universe, it is now understood that water is ubiquitous in the cosmos—not only as ice and vapor, but also as liquid. Sophisticated scientific instruments can detect cosmic water on the basis of the light, or other electromagnetic, waves that it absorbs and emits (1). Unlike the legendary ‘‘waters of chaos’’ that gave rise to the material world, water’s component hydrogen and oxygen atoms owe their existence to the Big Bang and to the stars, respectively. HYDROGEN AND OXYGEN Hydrogen is both the simplest and the most abundant atom in the universe. It represents about 75% of the atomic mass in the cosmos. The word hydrogen literally means ‘‘water-forming’’ according to the Greek language from which it is derived. Hydrogen atoms are generally traced back to the so-called Big Bang, when a tremendous amount of energy was released and subsequently expanded into what we call our universe. As the newborn universe began to cool, subatomic and atomic particles (e.g., quarks, protons, electrons) were initially created and later drawn together by a number of fundamental forces to form atoms. Possessing one proton and one electron, hydrogen is believed to have been the first atom created. As more hydrogen atoms were created in the early universe, they coalesced into dense gas clouds that contained much of the conventional matter. Oxygen is the third most abundant atom in the cosmos—behind hydrogen and helium. Because helium is a very inert (nonreactive) atom, water is sometimes

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described as an interaction between the two most abundant ‘‘reactive’’ atoms in the cosmos. Unlike hydrogen, the origins of the oxygen atom are rooted in dying stars rather than in the Big Bang. As stars near the end of their stellar life, they begin to cool and to switch from hydrogen to helium as a source for nuclear fusion. The cooling stars then enter a phase where they become increasingly dense as intense gravitational energy compresses them into an extremely unstable state that may explode during the final stages of compression (2). This explosion or supernova releases the outer layer of the star, which contains many common heavier atoms (e.g., oxygen, carbon), into space as interstellar clouds. Dust grains comprising these clouds are composed of both silicate (oxygen-rich) and carbonaceous (carbon-rich) minerals that are available to react with hydrogen (1).

INTERSTELLAR SPACE In the interstellar realms of galaxies, water exists predominantly as ice—adhering to the tiny particles that comprise the ubiquitous dust and gas clouds. Water is the primary molecular ice attached to these particles, although methane, carbon monoxide, and water–ammonia mixtures may also be present depending on physical conditions in the gas clouds (3). Water ice in the 10 K temperatures and vacuum conditions of interstellar space is what physical chemists refer to as amorphous ice or glassy water; it is relatively unstructured compared to the highly crystalline ices that are formed at higher temperatures (e.g., those characteristic of the earth’s surface and atmosphere). Amorphous ice is so unstructured that it can flow, not unlike a viscous version of liquid water (4). Some astrophysicists posit that the simple organic molecules responsible for biological life may have been created in this strange ice (4). As interstellar temperatures rise above 150 K (as often occurs near stars), amorphous ice irreversibly transits to more familiar crystalline ice. The various phases and molecular structures of water as a function of temperature are shown in Table 1. Although water’s component atoms are plentiful in interstellar dust and gas clouds, creation of molecular water requires either converting O and OH species directly to water ice on the surfaces of dust grains or producing water vapor via heat energy—usually in the form of stellar radiation (5). The latter process requires that water vapor adhere to dust grains, where the newly formed water molecule is protected from the same ionizing radiation that created it. Scientists currently believe that stars facilitate the creation of water vapor and also that water vapor assists in the birthing of stars. Stars are being born and are dying on an ongoing basis, such that star birthing regions (e.g., the Orion Cloud Complex of the Milky Way Galaxy) generate up to 20% of a galaxy’s luminosity as gas and dust clouds are gravitationally compressed into newborn stars. Recent data indicate that these cloud complexes contain an extremely high concentration of water vapor, which has been estimated of the order of 1 part in 2000 (6). The superabundance of water in stellar nurseries (about 20 times greater than that in similar interstellar clouds)

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COSMIC WATER Table 1. Representative Temperature Ranges and Cosmic Locations for the Three Phases of Water

Location

Temperature, Ka

Phase

Stellar surfaces

4000 to 50,000

None

Stellar/planetoid surfaces

100

Solid (crystalline)

Interstellar space

10 to 150

Solid (amorphous)

Comments Water molecules do not exist. Only hydrogen and oxygen atoms or plasma. Water exists at the surface of cool stars and in cooler regions of some hot stars. Few places in the cosmos possess the requisite temperatures and pressures. Crystalline ices possess both a cubic and a more common hexagonal structure. The depths of interstellar space are cold enough to produce glassy water.

a These temperatures, which are measured on the Kelvin scale, are only approximate and vary depending on environmental conditions (e.g., pressure, rate of cooling or heating).

may permit the gas and dust to cool sufficiently so that condensation can proceed and stars are eventually formed. As hot winds (in the form of shock waves) are sent out during stellar birthing, the cloud must be cooled—initially by molecular hydrogen and subsequently by water and other simple gases (7). Water vapor is created during the interstellar cloud shocks as oxygen reacts explosively with hydrogen, causing the water vapor concentration to increase substantially during star birthing. Scientists have theorized one of two eventual fates for water created in star birthing. One is that the intense heat of the fledgling star rapidly dissociates water into its component atoms. The other is that the water is deposited on dust grains that later form the star’s planetoids. The origin of earthly water is usually attributed to either this second process or the impact of large comets, which are believed to have been more prevalent during the earth’s early history. STARS Two of the brightest supergiants in our galaxy, Betelgeuse and Antares, have water in their photospheres, which constitutes the visible portion of a star (8). A star’s photosphere is where its gases transit from opaque to transparent, permitting us to see the stars that are located closest to the earth. This stellar water is actually present within the star itself, not simply as a component of the surrounding dust and gas cloud from which the star was birthed. Aging supergiant stars release massive amounts of water as they die; however, the exact source and role of this water are not yet known. In addition to cool stars, water has been discovered in the photosphere of at least one hot or main sequence star—the Sun. Water cannot exist on the surface of the Sun, where temperatures of 6000 K dissociate the water molecule into its component hydrogen and oxygen atoms. Water can exist in the dark centers of sunspots, where temperatures are less than 3500 K (9). Sunspots are relatively calm solar regions where strong magnetic fields filter out the energy emanating from the intense interior, rendering them both the coolest and darkest regions of the Sun. Water is a major player in determining a star’s radiative opacity, which describes the extent to which light escapes from stars into interstellar space (10).

In this role, water impedes the outward flow of radiation from stars by absorbing energy within certain wavelengths and, thus, renders the star more opaque than it would otherwise appear.

COMETS AND METEORS Comets are one of the few interstellar objects that are commonly associated with water, predominantly as ice. Comets are composed primarily of water ice that incorporates many of the other simple molecules in interstellar dust and gas clouds (e.g., carbon monoxide, methane, ammonia). Comets are most easily recognized by their unmistakable tails, which can extend millions to hundreds of millions of kilometers behind the icy body of the comet. The tail consists of dust and ionized particles (mostly water ice) that are always transported away from the Sun by the solar wind, which is an ionized stream of particles consisting predominantly of protons and electrons. The ionization of water ices is the primary mechanism influencing the properties of a comet’s tail, including the steam jets that release tons of water vapor per second from the comet. It is now believed that these steam jets result from solar-induced changes in ice’s molecular geometry (e.g., a transition from the amorphous to crystalline structure). Large comets are generally accepted as a source of planetary water, but controversy surrounds the hypothesis that many small comets hitting the planet’s upper atmosphere also contribute significantly to the volume of water on the earth. The first liquid water in the solar system, it was projected, made its appearance on meteors just 20 million years after our Sun and its debris emerged from the interstellar dust and gas cloud (11). Although liquid water is rarely present on the surface of meteors today, the chemical interaction of water with primitive rocks produced carbonate minerals, suggesting that the chemical processes of water evaporation and condensation were among the earliest in the solar system. Recently, a small meteorite found in southwestern North America contained actual liquid water within its salt crystals, which were believed to have been created from the original interstellar cloud that gave rise to the solar system.

THE WATER CYCLE

PLANETOIDS Most planet-sized bodies in our solar system (and probably in others) are now known or suspected to contain water in some form. A number of recent missions have revealed a Martian landscape that almost certainly indicates the large-scale flow of liquid water. The surface features of Mars (e.g., flood plains, river beds, mud deposits) suggest the recent presence of liquid water and also the mineralogy of Martian rocks could have resulted only from aqueous processes. Moreover, it has been hypothesized that Mars may have also once possessed surface oceans. The Jovian moon, Europa, is another of the solar system planetoids that probably contains liquid water located tens of kilometers beneath its icy surface. The liquid water underlying Europa’s surface ice is believed to be an ocean containing saltwater that may be similar in composition to the seawater of the earth’s oceans. Unlike the earth, the heat required to maintain water in a liquid phase on Europa is believed to originate from an internal source such as volcanic activity rather than from the heat of the Sun. Data collected from the Infrared Space Observatory indicate the presence of water in the upper atmospheres of our solar system’s gas giant planets and on one of Saturn’s moons (1). The source of water in these planet’s atmospheres is attributed to comets or to water-containing interplanetary dust. Based on recently developed techniques for measuring a suite of stellar characteristics (e.g., orbital velocity, position, brightness), the search for planets has been extended beyond our solar system to other star systems in the galaxy (12). Various planets have been identified orbiting stars in the constellations of Leo, Pegasus, Virgo, and Ursa Major that probably possess surface temperatures ranging from slightly less than 100 ◦ C down to −100 ◦ C. Planets or moons characterized by this temperature range could possess water in solid, gaseous, and liquid phases. BIBLIOGRAPHY 1. Salama, A. and Kessler, M. (2000). ISO and Cosmic Water. European Space Agency Bulletin 104, pp. 31–38. 2. Lewis, J. (1995). Physics and Chemistry of the Solar System. Academic, San Diego, pp. 30–31. 3. Tielens, A.G.G.M., Hagen, W., and Greenberg, J.M. (1983). Interstellar ice. J. Phys. Chem. 87: 4220–4229. 4. Blake, D. and Jenniskens, P. (2001). The ice of life. Sci. Am. August: 44–51. 5. O’Neill, P.T. and Williams, D.A. (1999). Interstellar water and interstellar ice. Astrophys. Space Sci. 266: 539–548. 6. Cowen, R. (1998). Unveiling the hidden universe. Sci. News 153: 328–330. 7. Harwit, M., Neufield, D.A., Melnick, G.J., and Kaufman, M.J. (1998). Thermal water vapor emission from shocked regions in Orion. Astrophys. J. Lett. 497: 105–108. 8. Jennings, D.E. and Sada, P.V. (1998). Water in Betelgeuse and Antares. Science 279: 844–847. 9. Oka, T. (1997). Water on the sun: Molecules everywhere. Science 277: 328–329. 10. Tsuji, T. (1986). Molecules in stars. Annu. Rev. Astron. Astrophys. 24: 89–125.

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11. Endress, M., Zinner, E., and Bischoff, A. (1996). Early aqueous activity on primitive meteorite parent bodies. Nature 379: 701–703. 12. Boss, A. (1996). Extrasolar planets. Phys. Today September: 32–38.

THE WATER CYCLE NASA—Goddard Space Flight Center

INTRODUCTION As seen from space, one of the most unique features of our home planet is the water, in both liquid and frozen forms, that covers approximately 75% of the Earth’s surface. Believed to have initially arrived on the surface through the emissions of ancient volcanoes, geologic evidence suggests that large amounts of water have likely flowed on Earth for the past 3.8 billion years, most of its existence. As a vital substance that sets the Earth apart from the rest of the planets in our solar system, water is a necessary ingredient for the development and nourishment of life. HYDROLOGIC HISTORY The notion that water is continually circulating from the ocean to the atmosphere to the land and back again to the ocean has interested scholars through most of recorded history. In Book 21 of the lliad, Homer (ca. 810 B.C.) wrote of ‘‘the deep-flowing Oceanus, from which flow all rivers and every sea and all springs and deep wells.’’ Thales (ca. 640 B.C.–ca. 546 B.C.) and Plato (ca. 427 B.C.–347 B.C.) also alluded to the water cycle when they wrote that all waters returned by various routes to the sea. But it wasn’t until many centuries later that scientific measurements confirmed the existence of a water (or hydrologic) cycle. Seventeenth century French physicists Pierre Perrault (1608–1680) and Edmond Mariotte (1620–1684) separately made crude precipitation measurements in the Seine River basin in France and correlated these measurements with the discharge of the river to demonstrate that quantities of rainfall and snow were adequate to support the river’s flow. WATER, WATER, EVERYWHERE Water is everywhere on Earth and is the only known substance that can naturally exist as a gas, liquid, and solid within the relatively small range of air temperatures and pressures found at the Earth’s surface. In all, the Earth’s water content is about 1.39 billion cubic kilometers (331 million cubic miles) and the vast bulk of it, about 96.5%, is in the global oceans. Approximately 1.7% is

This article is a US Government work and, as such, is in the public domain in the United States of America.

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THE WATER CYCLE

Table 1. One Estimate of Global Water Distribution. Estimates of Groundwater are Particularly Difficult and Vary Widely Amongst Sources, with the Value in this Table Being Near the High End of the Range. Using the Values in This Table, Groundwater Constitutes Approximately 30% of Fresh Water, Whereas Ice (Including Ice Caps, Glaciers, Permanent Snow, Ground Ice, and Permafrost) Constitute Approximately 70% of Fresh Water. With Other Estimates, Groundwater is Sometimes Listed as 22% and Ice as 78% of Fresh Water Volume (1000 km3 ) Oceans, Seas, & Bays Ice caps, Glaciers, & Permanent Snow Groundwater Fresh Saline Soil Moisture Ground Ice & Permafrost Lakes Fresh Saline Atmosphere Swamp Water Rivers Biological Water Total

Percent of Percent of Total Water Fresh Water

1,338,000 24,064

96.5 1.74

— 68.7

23,400 (10,530) (12,870) 16.5 300

1.7 (0.76) (0.94) 0.001 0.022

— 30.1 — 0.05 0.86

176.4 (91.0) (85.4) 12.9 11.47 2.12 1.12 1,385,984

0.013 (0.007) (0.006) 0.001 0.0008 0.0002 0.0001 100.0

— 0.26 — 0.04 0.03 0.006 0.003 100.0

Source: Gleick, P.H., 1996: Water resources. In Encyclopedia of Climate and Weather, ed. by S. H. Schneider, Oxford University Press, New York, vol. 2, pp. 817–823.

Figure 1. In the hydrologic cycle, individual water molecules travel between the oceans, water vapor in the atmosphere, water and ice on the land, and underground water.

of their leaves. Together, evaporation, sublimation, and transpiration, plus volcanic emissions, account for all the water vapor in the atmosphere. While evaporation from the oceans is the primary vehicle for driving the surface-toatmosphere portion of the hydrologic cycle, transpiration is also significant. For example, a cornfield 1 acre in size can transpire as much as 4000 gallons of water every day. After the water enters the lower atmosphere, rising air currents carry it upward, often high into the atmosphere,

stored in the polar icecaps, glaciers, and permanent snow, and another 1.7% is stored in groundwater, lakes, rivers, streams, and soil. Finally, a thousandth of 1% exists as water vapor in the Earth’s atmosphere (Table 1). A MULTI-PHASED JOURNEY The hydrologic cycle describes the pilgrimage of water as water molecules make their way from the Earth’s surface to the atmosphere, and back again. This gigantic system, powered by energy from the sun, is a continuous exchange of moisture between the oceans, the atmosphere, and the land (Fig. 1). Studies have revealed that the oceans, seas, and other bodies of water (lakes, rivers, streams) provide nearly 90% of the moisture in our atmosphere. Liquid water leaves these sources as a result of evaporation, the process by which water changes from a liquid to a gas. In addition, a very small portion of water vapor enters the atmosphere through sublimation, the process by which water changes from a solid (ice or snow) to a gas. (The gradual shrinking of snow banks, even though the temperature remains below the freezing point, results from sublimation.) The remaining 10% of the moisture found in the atmosphere is released by plants through transpiration (Fig. 2). Plants take in water through their root systems to deliver nutrients to their leaves, then release it through small pores, called stomates, found on the undersides

Figure 2. Plants return water to the atmosphere through transpiration. In this process, water evaporates from pores in the plant’s leaves, after being drawn, along with nutrients, from the root system through the plant.

THE WATER CYCLE

where the air cools and loses its capacity to support water vapor. As a result, the excess water vapor condenses (i.e., changes from a gas to a liquid) to form cloud droplets, which can eventually grow and produce precipitation (including rain, snow, sleet, freezing rain, and hail), the primary mechanism for transporting water from the atmosphere back to the Earth’s surface. When precipitation falls over the land surface, it follows various routes. Some of it evaporates, returning to the atmosphere, and some seeps into the ground (as soil moisture or groundwater). Groundwater is found in two layers of the soil, the ‘‘zone of aeration,’’ where gaps in the soil are filled with both air and water, and, further down, the ‘‘zone of saturation,’’ where the gaps are completely filled with water. The boundary between the two zones is known as the water table, which rises or falls as the amount of groundwater increases or decreases (Fig. 3). The rest of the water runs off into rivers and streams, and almost all of this water eventually flows into the oceans or other bodies of water, where the cycle begins anew (or, more accurately, continues). At different stages of the cycle, some of the water is intercepted by humans or other life forms. Even though the amount of water in the atmosphere is only 12,900 cubic kilometers (a minute fraction of Earth’s total water supply that, if completely rained out, would cover the Earth’s surface to a depth of only 2.5 centimeters), some 495,000 cubic kilometers of water are cycled through the atmosphere every year, enough to uniformly cover the Earth’s surface to a depth of 97 centimeters. Because water continually evaporates, condenses, and precipitates, with evaporation on a global basis approximately equaling global precipitation, the total amount of water vapor in the atmosphere remains approximately the same over time. However, over the continents, precipitation routinely exceeds evaporation, and conversely, over the oceans, evaporation exceeds precipitation. In the case of the oceans, the routine excess of evaporation over precipitation would eventually leave

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the oceans empty if they were not being replenished by additional means. Not only are they being replenished, largely through runoff from the land areas, but, over the past 100 years, they have been over-replenished, with sea level around the globe rising by a small amount. Sea level rises both because of warming of the oceans, causing water expansion and thereby a volume increase, and because of a greater mass of water entering the ocean than the amount leaving it through evaporation or other means. A primary cause for increased mass of water entering the ocean is the calving or melting of land ice (ice sheets and glaciers). Throughout the hydrologic cycle, there are an endless number of paths that a water molecule might follow. Water at the bottom of Lake Superior may eventually fall as rain in Massachusetts. Runoff from the Massachusetts rain may drain into the Atlantic Ocean and circulate northeastward toward Iceland, destined to become part of a floe of sea ice, or, after evaporation to the atmosphere and precipitation as snow, part of a glacier. Water molecules can take an immense variety of routes and branching trails that lead them again and again through the three phases of ice, liquid water, and water vapor. For instance, the water molecules that once fell 100 years ago as rain on your great grandparents’ farmhouse in lowa might now be falling as snow on your driveway in California. THE WATER CYCLE AND CLIMATE CHANGE Amongst the highest priorities in Earth science and environmental policy issues confronting society are the potential changes in the Earth’s water cycle due to climate change. The science community now generally agrees that the Earth’s climate will undergo changes in response to natural variability, including solar variability, and to increasing concentrations of greenhouse gases and aerosols. Furthermore, agreement is widespread that these changes may profoundly affect atmospheric water vapor concentrations, clouds, and precipitation patterns. For example, a warmer climate, directly leading to increased evaporation, may well accelerate the hydrologic cycle, resulting in an increase in the amount of moisture circulating through the atmosphere. Many uncertainties remain, however, as illustrated by the inconsistent results given by current climate models regarding the future distribution of precipitation. THE AQUA MISSION AND THE WATER CYCLE

Figure 3. The water table is the top of the zone of saturation and intersects the land surface at lakes and streams. Above the water table lies the zone of aeration and soil moisture belt, which supplies much of the water needed by plants.

As mentioned earlier, the hydrologic cycle involves evaporation, transpiration, condensation, precipitation, and runoff. NASA’s Aqua satellite monitors many aspects of the role of water in the Earth’s systems, and will do so at spatial and temporal scales appropriate to foster a more detailed understanding of each of the processes that contribute to the hydrologic cycle. These data and the analyses of them nurture the development and refinement of hydrologic process models and a corresponding improvement in regional and global climate models, with a direct anticipated benefit of more-accurate weather and climate forecasts.

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CYCLONES

Aqua’s contributions to monitoring water in the Earth’s environment involves all six of Aqua’s instruments: the Atmospheric Infrared Sounder (AIRS), the Advanced Microwave Sounding Unit (AMSU), the Humidity Sounder for Brazil (HSB), the Advanced Microwave Scanning Radiometer-Earth Observing System (AMSR-E), the Moderate Resolution Imaging Spectroradiometer (MODIS), and Clouds and the Earth’s Radiant Energy System (CERES). The AIRS/AMSU/HSB combination provides more-accurate space-based measurements of atmospheric temperature and water vapor than have ever been obtained before, with the highest vertical resolution to date as well. Since water vapor is the Earth’s primary greenhouse gas and contributes significantly to uncertainties in projections of future global warming, it is critical to understand how it varies in the Earth system. The water in clouds is examined with MODIS, CERES, and AIRS data; and global precipitation is monitored with AMSR-E. The cloud data includes the height and areal coverages of clouds, the liquid water content, and the sizes of cloud droplets and ice particles, the latter sizes being important to the understanding of the optical properties of clouds and their contribution to the Earth’s albedo (reflectivity). HSB and AMSR-E, both making measurements at microwave wavelengths, have the ability to see through clouds and detect the rainfall under them, furthering the understanding of how water is cycled through the atmosphere. Frozen water in the oceans, in the form of sea ice, is examined with both AMSR-E and MODIS data, the former allowing routine monitoring of sea ice at a coarse resolution and the latter providing greater spatial resolution but only under cloud-free conditions. Sea ice can insulate the underlying liquid water against heat loss to the often frigid overlying polar atmosphere and also reflects sunlight that would otherwise be available to warm the ocean. AMSR-E measurements allow the routine derivation of sea ice concentrations in both polar regions, through taking advantage of the marked contrast in microwave emissions of sea ice and liquid water. This continues with improved resolution and accuracy, a 22year satellite record of changes in the extent of polar ice. MODIS, with its finer resolution, permits the identification of individual ice floes, when unobscured by clouds. AMSR-E and MODIS also provide monitoring of snow coverage over land, another key indicator of climate change. Here too, the AMSR-E allows routine monitoring of the snow, irrespective of cloud cover, but at a coarse spatial resolution, while MODIS obtains data with much greater spatial detail under cloud-free conditions. As for liquid water on land, AMSR-E provides an indication of soil moisture, which is crucial for the maintenance of land vegetation, including agricultural crops. AMSR-E’s monitoring of soil moisture globally should permit, for example, the early identification of signs of drought episodes. THE AQUA SPACECRAFT Aqua is a major mission of the Earth Observing System (EOS), an international program centered in NASA’s

Earth Science Enterprise to study the Earth in detail from the unique vantage point of space. Focused on key measurements identified by a consensus of U.S. and international scientists, EOS is further enabling studies of the complex interactions amongst the Earth’s land, ocean, air, ice and biological systems. The Aqua spacecraft circles the Earth in an orbit that ascends across the equator each day at 1:30 p.m. local time and passes very close to the poles, complementing the 10:30 a.m. measurements being made by Terra, the first of the EOS spacecraft, launched in December 1999. The instrument complement on Aqua is designed to provide information on a great many processes and components of the Earth system, including cloud formation, precipitation, water vapor, air temperature, cloud radiative properties, sea surface temperature, surface wind speeds, sea ice concentration and temperature, snow cover, soil moisture, and land and ocean vegetation. The individual swaths of measurements will be compiled into global images, with global coverage of many variables being obtained as frequently as every two days or, with the help of numerical models, combined every 6 or 12 hours into comprehensive representations of the Earth’s atmospheric circulation and surface properties. In combination with measurements from other polar orbiting satellites, Aqua measurements also provide accurate monthly-mean climate assessments that can be compared with and assimilated into computer model simulations of the Earth’s climate. The Earth Observing System has three major components: the EOS spacecraft, an advanced ground-based computer network for processing, storing, and distributing the collected data (the EOS Data and Information System); and teams of scientists and applications specialists who study the data and help users in universities, industry, and the public apply it to issues ranging from weather forecasting and climate prediction to agriculture and urban planning.

CYCLONES ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

Cyclones are hazardous weather conditions characterized by extreme gusts of wind moving in a circular pattern, low pressure, and intense rain. They form over tropical or subtropical waters; however, some can reach land, where they have in the past inflicted great amounts of damage on buildings and communities. Based on wind strength, cyclones are given various names. Tropical depressions have maximum surface winds of less than 39 mph. Those cyclones whose maximum winds are between 39 and 74 mph are known as tropical storms. Upon attaining 74-mph surface winds, a cyclone is known as one of a number of names that are all regional equivalents of the same type of storm. The names of cyclones are typhoon (in the NW Pacific Ocean), hurricane (in the North Atlantic, NE Pacific, and South Pacific Ocean), tropical cyclone (in the SW Indian

CYCLONES

Ocean), severe cyclonic storm (in the North Indian Ocean), and severe tropical cyclone (in the SW Pacific and SE Indian Oceans). Cyclones always build over tropical seas. There are seven basins in which the majority of tropical cyclones form: the Atlantic Basin, the North Indian Basin, the Southwest Indian Basin, the Southeast IndianAustralian Basin, the Australian-Southwest Pacific Basin, the Northwest Pacific Basin, and the Northeast Pacific Basin. Heat gives cyclones their energy, and thus the water over which a cyclone forms must be at least as warm as 80◦ F (i.e., tropical). Other conditions necessary for a cyclone include a fast cooling atmosphere, moisture in the middle troposphere area, a distance of more than 300 miles from the equator to be influenced by the Coriolis force, and minimal vertical wind shear. This vertical wind shear is the result of differences between winds in the lower and upper portions of the atmosphere. However, the main contributor to the formation of a cyclone is a disturbance in the form of a thunderstorm or group of showers. When all of these factors come together, conditions are right for a tropical cyclone, although the presence of each of these factors does not guarantee the formation of a cyclone. Cyclones are spontaneous; a minute variation in one variable can be the difference between a hurricane and a thunderstorm. The circular area in the center of a cyclone, known as the ‘‘eye,’’ has conditions quite different from those in the region surrounding it. Calmness and a light breeze characterize the eye. Temperatures are generally higher than those in the surrounding area, and the sky is usually very clear. Since the 1970s, cyclones have been regularly named by various people and agencies, depending on where the cyclones originated. They are named to ensure ease of description among newscasters, weatherpeople, and the general public. Names are often taken from rotating lists, one name for each letter of the alphabet. When an exceptional cyclone occurs, its name is taken out of use (‘‘retired’’) to avoid confusion. Strong cyclones often cause extensive damage to whatever they encounter. Damage can range from crop destruction to total devastation of building structures, depending on the severity of the cyclone. They become most dangerous as they hit land and spawn tornadoes. These are formed when tropical cyclones begin to lose their power. The diameter of a tropical cyclone is measured in kilometers; the diameter of a tornado is measured in meters. In 1980, one of the most destructive cyclonespawned tornadoes in the United States caused nearly $100,000,000 worth of damage to the Austin, Texas, area. Along with destruction, cyclones (and the subsequent tornadoes) cause death. Large objects lifted from the path of the extreme wind are tossed about like lethal weapons. In 1964, twenty-two people were killed by a tornado that hit the Los Angeles area in the United States. The most damaging cyclone in history was Hurricane Andrew, which caused over $26 billion in damage to the Southeastern United States. The most deadly cyclone ever may have been the Bangladesh Cyclone, which killed at least 300,000 people in 1970. One of the ways in which a hurricane will cause damage is through the storm surge. This is the name

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given to the phenomenon whereby sea level in the area of a cyclone rises due to cyclonic winds. Once a cyclone reaches 74 mph maximum winds and is considered a hurricane, it’s intensity is rated by the Saffir–Simpson Scale, used since the 1970s by the National Oceanic and Atmospheric Administration (NOAA). Hurricanes are referred to as category one, category two, category three, category four, or category five. Category one hurricanes have winds between 74 and 95 mph. They cause negligible damage to buildings and other structures, although they can damage mobile homes and road signs. They carry a storm surge of about 5 feet. Category two hurricanes have maximum winds of 96–110 mph. They can cause minor damage to buildings and sizable damage to mobile homes and trees. Their storm surge is usually around 7 feet. Category three hurricanes have winds from 111–130 mph. They cause noticeable damage to buildings and trees; buildings near the shoreline are often destroyed by flooding. Evacuation of certain low-lying areas can be necessary for category three hurricanes, and their storm surge ranges from 9–12 feet. Category four hurricanes have winds between 131 and 155 mph. These hurricanes can cause massive damage to smaller structures. They destroy trees and mobile homes utterly. Evacuation of large areas can be necessary, as the storm surge can reach up to 18 feet. Category five hurricanes are the most powerful; they have maximum sustained winds of more than 155 mph. Residential as well as industrial structures are often destroyed. Damage is catastrophic, and major evacuations usually take place. The storm surge exceeds 18 feet above normal. Cyclones can range from relatively simple tropical storms to devastating hurricanes whose winds swirl at furious speeds. Their effects can be disastrous and longlasting. READING LIST Bureau of Meteorology. (2002). Surviving Cyclones. Available: http://www.bom.gov.au/info/cyclone/. (March 13). Department of Public Safety. (2002). ‘‘General Hurricane Information.’’ Available: http://www.escambia-emergency.com/geninfo. asp. (March 13). Landsea, Christopher W. FAQ: Hurricanes, Typhoons, and Tropical Cyclones. Available: http://www.aoml.noaa.gov/hrd/tcfaq/ tcfaqA.html. (March 13). National Weather Service Houston/Galveston. (2002). Tropical Cyclone FAQ. Available: http://www.srh.noaa.gov/hgx/index3/ tropical3/geninfo3.htm. (March 13). Hurricane Research Division: FAQ Available: http://www.aoml. noaa.gov/hrd/tcfaq/F1.html. Hurricane Research Division: FAQ Available: http://www.aoml. noaa.gov/hrd/tcfaq/E7.html. Hurricane Research Division: FAQ Available: http://www.aoml. noaa.gov/hrd/tcfaq/B1.html. Hurricane Research Division: FAQ Available: http://www.aoml. noaa.gov/hrd/tcfaq/A10.html. Hurricane Research Division: FAQ Available: http://www.aoml. noaa.gov/hrd/tcfaq/A1.html. Wonders of Weather. Available: http://school.discovery.com/lessonplans/programs/wondersofweather/vocab.html.

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WATER CYCLE

The Saffir–Simpson Hurricane Scale. Available: http://www.nhc. noaa.gov/aboutsshs.shtml.

WATER CYCLE ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

The water cycle—a series of steps through which virtually all water on earth is constantly cycling. The names of these steps vary, but they generally follow this order: storage, evaporation, transpiration, condensation, precipitation, infiltration, and runoff. Due to the nature of the water cycle, there is really no ‘‘beginning’’ or ‘‘end’’, but for simplicity, the cycle can be said to begin at the storage stage. Storage is the storing of water on, in, and above the earth. There are about 340 million cubic miles of water on the earth. This water is stored in many places: glaciers, seas, rivers, lakes, polar ice caps, and all living things on the planet. The largest percentage of water, approximately 97.25%, exists as salt water in the ocean. Water is also stored in the earth’s atmosphere and as groundwater. Groundwater in the top few miles of the earth’s crust is easily obtainable, whereas groundwater further down is chemically attached to rocks and usually cannot be accessed. Evaporation is an essential way in which water is transferred from the earth to the atmosphere. During evaporation, water molecules change from a liquid state directly into a gaseous state, known as water vapor. Changes in temperature and air pressure on the earth cause millions of gallons of water to evaporate into the atmosphere each day. The majority of the water that is transferred into the atmosphere is primarily in a liquid state before the change; however, a small percentage of new water vapor comes from ice. The process of water changing state from a solid directly to a gas is called sublimation. Sublimation occurs on glaciers and polar ice sheets. When the sun strikes these massive ice structures, the temperature and pressure allow no room for a liquid, only a solid and a gas. Transpiration is another way that water is transferred into the atmosphere. Although similar in process to evaporation, transpiration differs because the medium from which the water comes is not earth, it is plants. Transpiration is the process through which water absorbed from the ground by plant roots evaporates into the atmosphere. After the plants absorb the water from the soil, the water moves up through the veins of the plant, eventually reaching the leaves (or the plant’s equivalent). These leaves have pores, or stomata, which allow water to evaporate. Large forests can release massive amounts of water into the atmosphere through transpiration. Once the water has been transferred into the atmosphere by either evaporation or transpiration, the molecules will eventually condense. This process is called condensation. Condensation is the reciprocal process of evaporation. During condensation, water changes state from a gas to a liquid. Once again, a change in

either temperature or pressure initiates the change. Condensation is most visible in the formation of dew, fog, and clouds. Dew usually appears in the morning and is water that has condensed on any solid object during the rapid temperature change that occurs from day to night. If the water vapor is dense enough around objects when the temperature change happens, dew is formed. Fog is really a low level cloud formation. When the conditions are right, fog is formed on or near the ground. The only difference between fog and cloud formations is their locations with reference to the earth. Clouds form when water vapor in the atmosphere encounters a critical temperature or air pressure, causing the water vapor to change state into a liquid. Water molecules then ‘‘stick together’’ causing physical cloud structures to emerge. When the molecules in the cloud that are stuck together reach a certain critical mass, they release the moisture causing precipitation to begin. Precipitation is the process by which water is transferred from the medium of the atmosphere to the medium of the earth. Precipitation literally comes in all shapes and sizes, including rain, snow, and ice. The differentiating factor that determines the form water takes when precipitating is air temperature. The lower the temperature, the more likely it is that the precipitation will be either snow or ice (the closer the temperature is to the freezing point of water, 32 ◦ F or 0 ◦ C the more likely). Droplets of rain form around small particles of dust or dirt to form a cohesive raindrop as a vehicle for the trip to the earth. Regardless of the form it takes on its descent, water will take one of two paths once it arrives on the earth’s crust. It will either seep into the ground or flow into larger bodies of water such as streams, seas, oceans, and rivers. The former process is infiltration, and the latter is runoff. Infiltration restores groundwater to the water table that has been lost from wells. During the infiltration process, the water is purified by cascading amongst rocks and minerals that draw impurities out of the water. Water that does not infiltrate the earth’s crust runs off into streams, rivers, seas, and oceans. Runoff replenishes bodies of water that have lost water due to evaporation. Runoff can occur either above the land or below it. Large amounts of water can run across the land immediately after precipitation. Smaller amounts of water run through the ground and along the water table until they reach a larger body of water. So thusly does the water complete its cycle throughout the different media of the earth. This process is essential for the survival of life on the earth because it allows water to permeate every accessible region of the crust and atmosphere, thereby allowing life to flourish wherever water exists. READING LIST Hydrologic Cycle. (2002). Available: http://www.und.nodak.edu/ instruct/eng/fkarner/pages/cycle.htm. (April 1). The Water Cycle: Scientific Concepts. (1995–1998). Available: http://mbgnet.mobot.org/fresh/cycle/concepts.htm. (March 28, 2002).

DEGREE DAY METHOD

approaches based on the time of the year or the number of days (3). Because of the close coupling between the timescale of plants and temperature, Ritchie and NeSmith (6) proposed that the most appropriate term to describe plant development would be ‘‘thermal time’’ or degree days (DD), whose units are ◦ C-day. The degree day method is based on the effects of temperature on developmental rates, rather than on the duration of a phase. Figure 1b shows the linear relationship between temperature and rate of development, the reciprocal of the duration of the phase in days. The base temperature (Tb ), temperature at which the rate of development is zero, is obtained by extrapolating this linear relationship to the intersection with the x-axis:

Gedzelman, S.D. (2001). The Water Cycle. In: Microsoft Encarta Encyclopedia Deluxe 2001.

DEGREE DAY METHOD CARLOS D. MESSINA University of Florida Gainesville, Florida

An adequate description and prediction of plant development is critical for understanding plant growth and its responses to the environment. Only then we can assess the impact of a changing environment on plant productivity and survival, understand plant adaptation, design adequate agricultural production systems, and optimize natural resources management. Of particular importance for water resource management are the effects of plant ontogeny on plant water use and productivity throughout the regulation of leaf area. This determines the patterns of plant water demand and the partitioning between transpiration and evaporation components. Therefore, plant development dictates the selection of cultivars that best fit the water availability patterns under dryland conditions, and the amount, frequency and timing for irrigation otherwise. Plant development is responsive to environmental cues such as photoperiod, water availability, and temperature. Figure 1a shows an example of the close association between ambient temperature and plant development for boll growth duration in cotton. Rameur first suggested in 1735 that the duration of any developmental stage was related to temperature and that this duration can be characterized by a thermometric constant, which can be predicted by the sum of daily air temperatures (2). Since then, several methods based on the concept of normalizing time by temperature to predict developmental rates were developed (2,3) and applied in life sciences. Although we can formulate some hypothesis to explain the relationship between plant development and temperature based on the effects of temperature on circadian oscillations (4), lipid composition and membranes fluidity, and cell division (5), the scientific basis of these empirical methods remains elusive. It was shown, however, that degree day based methods significantly improved the description and prediction of phenological events relative to other

100 90

y = 0.1503x 2 − 10.7x + 226.15 R 2 = 0.9952

80 70 60 50

20

25

30

Tb = −

y0 α

where y0 is the intercept and α is the slope of the linear regression equation (Fig. 1b). Then, the rate of development (RD) can be calculated as RD(t) = α (Ta − Tb ) where Ta is the mean temperature. Integrating the rate of development in time,

RD dt = α (Ta − Tb ) dt and considering a daily time step (
t = 1) for integration, we can estimate the cumulative development (CD) at time n (tn ), provided that t0 = 0, as CD = α



(Ta − Tb )

n

For constant temperatures, given that development is complete when CD equals one, the degree days above a certain base temperature for a given phase are DD = (Ta − Tb ) n or DD =

1 α

For the example of cotton shown in Fig. 1, boll growth requires 1000 ◦ C-days above a base temperature

0.03 0.025 0.02 0.015 a 0.01 0.005

40 30 15

(b) Rate boll growth duration, 1/days

Boll growth duration, days

(a)

197

0 0 35 Mean temperature, ° C

y = 0.001x − 0.0065 R 2 = 0.9922

Tb 10

20

30

Figure 1. Responses of boll growth duration in cotton (Gossypium hirsutum, L.) development to air temperature. (a) Boll growth duration, (b) rate of boll growth duration (1/d). Tb denotes base temperature (see below), and α is the slope of the regression line. Data is from (1).

198

DEGREE DAY METHOD

of 6.5 ◦ C to reach maturity. Although this formulation for calculating DD is useful for estimating the requirements of the plant to complete a given developmental phase, in natural environments, temperature fluctuates and makes this procedure inadequate. The following canonical form to calculate DD is frequently used: DD =

  Tmax + Tmin  n

2

 − Tb

where Tmax and Tmin are maximum and minimum daily temperatures, n is the number of days of temperature observations, and [(Tmax + Tmin )/2] = Tb when [(Tmax + Tmin )/2] < Tb . An alternative method, frequently used in calculating DD for corn, compares Tmax and Tmin with Tb individually before calculating their average. Significant discrepancies in calculated DD between these procedures may arise under some circumstances (3). These procedures for calculating DD are adequate, provided that (1) the daily temperature does not fall below Tb and does not exceed an upper threshold for a significant part of the day, (2) the temperature of the growing plant tissue is the same as the air temperature, and (3) the response of the rate of development to temperature is linear over the range of temperature that the crop experiences (5). When these assumptions are not satisfied, alternative approaches are required. For example, McMaster et al. (5) and Vinocur and Ritichie (7) used soil or apex temperatures to calculate DD because the growing points were close to the soil rather than at the height where the temperature is normally measured. In simulation models such as CERES (8), hourly temperatures are approximated from Tmax and Tmin , and if the temperature is below Tb , DD is set to zero for that part of the day. For a thorough discussion of the limitations of the degree day method, see Ritchie and NeSmith (5). The degree day method can be found in the literature in a different mathematical form under the concept of physiological day. This model is generally used to simulate development in legumes and is particularly adequate for incorporating the effects of photoperiod on development (9). The rate of development is calculated as the product of two functions: RD(t) = f (P) × f (T) One function accounts for the effects of photoperiod (P) and the other for the effects of temperature (T) on development. The model predicts relative development; the maximum rate of development is standardized to 1.0. At optimum temperature (Topt ) and photoperiod (P < Pmin in short day species and P > Pmax in long day species), the rate of progress in calendar days equals the rate of progress in physiological days. When conditions deviate from the optimum, the rate of progress per day decreases and becomes a fraction of a physiological day. Then, cumulative development is measured in photothermal time units. The effects of the photoperiod on the rate of development are

nonlinear: f (P) = 0,

P − Pmin f (P) = 1 − , Pmax − Pmin f (P) = 1,

if P > Pmax if Pmin < P ≤ Pmax if P ≤ Pmin

In short day species, the relative rate of development is maximum below a threshold Pmin , above which the development decreases to zero at Pmax . Both Pmin and Pmax are characteristic of the species and cultivar (Table 1). A similar function is used to describe the effects of temperature: f (T) = 0, T − Tb f (T) = , Topt1 − Tb f (T) = 1, T − Topt2 f (T) = 1 − , Tupper − Topt2 f (T) = 0,

if T ≤ Tb if Tb < T < Topt1 if Topt2 ≥ T ≥ Topt1 if Tupper > T > Topt2 if T ≥ Tupper

where Topt1 and Topt2 define a plateau whose rate of development is maximum. For temperatures above Topt2 and below Topt1 , the rate of development decreases linearly to zero at Tupper and Tb , respectively. Table 1 shows characteristic values for selected cereals and legumes. In this chapter, a derivation of the degree day method is shown along with a mathematical variant that incorporates the effects of the photoperiod. The literature about models that uses degree days is vast, it is not

Table 1. Cardinal Temperatures and Photoperiods for Selected Models of Cereals and Legumesa Temperature, ◦ C Tb Topt1 Topt2 Tupper

Photoperiod, h Pmin

Pmax

Legumes Soybean Vegetative 7 Early reproductive 6 Late reproductive −48 Bean Vegetative 4 Early reproductive 5 Late reproductive 0 Peanutb Vegetative 11 Early reproductive 11 Late reproductive 5

28 26 26

35 30 34

45 45 45

11.7–14.6 15.5–21.0

27 22 18

35 35 35

45 45 45

12.2–13.2

— —

28 28 26

30 28 26

55 55 55

—

—

Cereals Wheat & Barley Millet Rice Maize Sorghum a

0 10 9 8 8

15 36 — 34 34

— — — — —

— — — — —

20 — 12 125–150c 11.7–12.8 35–189c 12.5 0.3–0.8d 12.5–13.6 30–90c

Adapted from Reference 10. insensitive to photoperiod. c DD ( ◦ C-day) per hour increase in photoperiod. d Units are in days per hour increase in photoperiod (11). b

DESERTIFICATION

the intent to review it here. However, the rationale and concept presented here covers most of the models currently used. Summerfield et al. (12) discuss additive photothermal models and Jones et al. (13) provide further details on multiplicative ones. Ritchie and NeSmith (6) and McMaster and Wilhelm (3) discuss some limitations of the degree day method. As emphasized in these previous papers, to obtain reliable and accurate predictions or descriptions of plant development, the parameters and the method used for their estimation must be consistent. BIBLIOGRAPHY 1. Reddy, K.R., Davidonis, G.H., Johnson, A.S., and Vinyard, B.T. (1999). Temperature regime and carbon dioxide enrichment alter cotton roll development and fiber properties. Agron. J. 91: 851–858. 2. Wang, J.Y. (1960). A critique of the heat unit approach to plant-response studies. Ecology 41: 785–790. 3. McMaster, G.S. and Wilhelm, W.W. (1997). Growing degreedays: one equation, two interpretations. Agric. Forest Meteorol. 87: 291–300. 4. Michael, T.P., Salome’, P.A., and McClung, C. (2003). Two arabidopsis circadian oscillators can be distinguished by differential temperature sensitivity. Proc. Nat. Acad. Sci. USA 100: 6878–6883. 5. McMaster, G.S. et al. (2003). Spring wheat leaf appearance and temperature: extending the paradigm? Ann. Bot. 91: 697–705. 6. Ritchie, J.T. and NeSmith, D.S. (1991). Temperature and crop development. In: Modeling Plant and Soil Systems. R.J. Hanks and J.T. Ritchie (Eds.). Agronomy 31. American Society of Agronomy, Madison, WI, pp. 5–29. 7. Vinocur, M.G. and Ritchie, J.T. (2001). Maize leaf development biases caused by air-apex temperature differences. Agron. J. 93: 767–772. 8. Ritchie, J.T., Singh, U., Godwin, D.C., and Bowen, W.T. (1998). Cereal growth, development and yield. In: Understanding Options for Agricultural Production. G.Y. Tsuji, G. Hoogenboom, and P.K. Thornton (Eds.). Kluwer Academic, Dordrecht, the Netherlands, pp. 79–98. 9. Boote, K.J., Jones, J.W., and Hoogenboom, G. (1998). Simulation of crop growth: CROPGRO model. In: Agricultural Systems Modeling and Simulation. R.M. Peart and R.B. Curry (Eds.). Marcel Dekker, New York, pp. 651–693. 10. Jones, J.W., Tsuji, G.Y., Hoogenboom, G., Hunt, L.A., Thornton, P.K., Wilkens, P.W., Imamura, D.T., Bowen, W.T., and Singh, U. (1998). Decision support system for agrotechnology transfer. In: Understanding Options for Agricultural Production. G.Y. Tsuji, G. Hoogenboom, and P.K. Thornton (Eds.). Kluwer Academic, Dordrecht, the Netherlands, pp. 157–177. 11. Kiniry, J.R. (1991). Maize phasic development. In: Modeling Plant and Soil Systems. R.J. Hanks and J.T. Ritchie (Eds.). Agronomy 31. American Society of Agronomy, Madison, WI, pp. 55–69. 12. Summerfield, R.J. et al. (1993). Towards the reliable prediction of time to flowering in 6 annual crops .2. soybean (Glycine max). Exp. Agric. 29: 253–289. 13. Jones, J.W., Boote, K.J., Jagtap, S.S., and Mishoe, J.W. (1991). Soybean development. In: Modeling Plant and Soil Systems. R.J. Hanks and J.T. Ritchie (Eds.). Agronomy 31. American Society of Agronomy, Madison, WI, pp. 71–90.

199

DESERTIFICATION ALDO CONTI Frascati (RM), Italy

Desertification is a degradation of the top layer of the soil that reduces its ability to support plant life and to produce food. As a result, the soil becomes dusty and dry, and it is easily carried away by erosion, which affects wild and domestic animals, wild plants and crops and, finally, humans. Desertification can be the result of human activities or of natural climate changes. In the latter case, the process is very slow and can take several thousand years to produce its effects. A common misconception about desertification is that it spreads from a desert core. The truth is that land degradation can occur where land abuse has become excessive. If it is not stopped in time, desertification spreads from that spot. Eventually, many of these spots merge and form a large homogeneous area. Another misconception is that desertification is the result of droughts. In fact, well-managed land can recover from even a long period of drought with very little adverse effect as soon as rains return. But it is true that droughts can increase the pace of desertification already taking place. Desertification is a term that has been in use since at least 1949, when Aubreville, a perceptive and wellinformed botanist and ecologist, published a book on Climate, Forets, et Desertification de l’Afrique Tropicale. Aubreville observed desertification in tropical Africa and understood immediately that the culprit was not the Sahara desert gaining land. He noticed instead that the main reasons behind desertification were tree cutting, indiscriminate use of fire to clear the land, and cultivation. Many processes can lead to desertification. Logging, for instance, makes the soil unstable on mountain slopes. Eventually, all the soil runs down as a dangerous landslide or mud river and leaves behind exposed rocks, unable to support any life. What is left behind are barren mountains, particularly evident in China. In Europe, one of the leading causes of desertification is overgrazing. In the past, wild animals used to move following the rainfall, always grazing the richest areas. In modern times, the use of fences prevented these movements, and the result was heavy overgrazing that left the soil exposed to erosion. In many areas, desertification is the result of agriculture. A typical example of this is salinization of soil that happens normally when the soil is overirrigated. The water that is not used by plants evaporates, leaving behind salts that concentrate in the soil. Eventually, the concentration of salts in the soil becomes so high that plants cannot survive, again exposing the land to erosion. In same cases, desertification is the more direct result of urbanization, mining, and recreational activities. In any case, the adverse effects are still the same. Nowadays, desertification is a serious problem that affects, according to some estimates, up to 30% of dry lands. Worldwide, desertification is making approximately 12 million hectares useless for cultivation every year. But land degradation and desertification are by no means new problems, despite the attention focused on them in

200

DEW

recent years. During the first conquest of Africa, it was normal to clear patches of land with fire and then use them to grow crops. After three or four seasons, the land was depleted of nutrients and unable to support any plant life. Moreover, there is some historical evidence that serious and extensive land deterioration was already occurring several centuries before. It started in arid regions, and it had three epicenters: the Mediterranean Sea and other places where destructive changes in soil and plant cover had occurred, but were small in extent or not well known. Luckily, desertification in many areas has been stopped, but very little effort has been made to restore the land to its original productivity. Today, desertification can be defeated using techniques already known, if financial resources are available and the political will to act is present. For instance, only in the last few decades, satellite images have allowed a better understanding and monitoring of the problem on a large scale. There are several possible remedies available, even at the local level. The first is to avoid cutting trees or, at least, to replace them with new ones. Plants, a major soil stabilizer, can alone stop erosion. Moreover, the use of available local water and ways to control the salinity of the soil can be very effective. On this topic, genetic engineering is trying to help. Scientists are working on the development of crops that can survive higher salinity, both as a way to use the land and produce food and as a way to save lost soils. Curiously, one of the remedies until now used against desertification is to pollute the soil. In Iran, oil is sprayed over semiarid land with crops. The oil covers the seedlings, retains the moisture, and prevents them from being blown away by the wind until they grow large enough. As stated at the beginning, desertification is sometimes a slow natural process. As an example, it should be enough to say that 20,000 years ago, the Sahara Desert was a lush forest. This is proved by the fossils of the animals that used to live in this forest. Moreover, pictures taken using radar from the Space Shuttle allow us to identify numerous dry riverbeds under a few meters of sand. This is already happening again due to the global warming on the earth. Many deserts are expanding, even though it is not possible to find a human reason for that. Desertification seems to threaten, in particular, all countries in the Mediterranean region. The coasts of North Africa are already disappearing into the sand, but the same might start happening in Europe. The heat wave of 2003 caused lots of problems. In Italy, many crops were heavily affected, and in the whole Mediterranean region, but in particular in Spain and Portugal, large areas of forest were destroyed by fires.

DEW ALDO CONTI Frascati (RM), Italy

Dew forms when water condenses on objects on the surface near the ground and forms a thin layer or many

droplets. Dew forms normally during the night when the air temperature decreases and approaches the dew point. Objects on the surface cool down, too, by radiative cooling, facilitating the condensation of small droplets of water. Dew formation is helped by the high humidity of the bottom layer of air, close to the surface. This layer can supply the needed water and prevent the evaporation of the dew already deposited. Strong winds can inhibit the formation of dew. Turbulence mixes a larger layer of air and creates a more homogeneous distribution of humidity and heat, thus preventing the formation of the right conditions near the ground. Dew forms more easily on surfaces that cool efficiently, like metals, which is the reason that cars are often covered with dew in the morning. But dew is often seen on grass or plants because their transpiration creates a thin layer of very humid air. Dew is very important in the ecology of many deserts, especially those along the western coasts of Africa and South America. In some of them, water from condensation of dew caused by night cooling often exceeds that of rainfall. But dew can be an interesting source of water for human consumption, too. For instance, data recorded in the Negev desert of Israel, a country where water is really scarce, have shown that dew falls for 200 days each year. Moreover, dew is plentiful on many little islands that, surrounded by water, have high humidity but where no water dwells. Some islands of the Mediterranean Sea suffer from a chronic water shortage, which is the reason that one of the first dew-collecting plants has been built in Ajaccio (Corsica Island in France). It is also interesting to describe an old project studied in India and unfortunately never actually realized. The idea was to pump water at 4.5 ◦ C from the sea at a depth of 500 meters. The pumping scheme called for the use of four 1.2-meter pipes and wind-powered pumps. A heat exchanger of 130,000 square meters could then condense, every day, more than 600 cubic meters of dew. But many animals and plants have already learned how to survive on dew. A place particularly interesting for this kind of adaptation is the Namib desert in Namibia (Africa). Here, many insects in the early morning sit on top of sand dunes trying to catch some dew (or fog) on their wings and legs. In particular, the beetle Stenocara has even modified the surface of its hard shell surface to trap mist more efficiently. This beetle shell is covered with bumps whose peaks are smooth like glass and attract water. The troughs around the bumps are covered with a wax that repels water. The water is therefore collected by the peaks, and when a droplet is big enough to touch the water repelling valley it rolls down to the animal’s mouth. Some researchers think that the easy trick of the animal can be used in dew collecting devices. A similar surface can be prepared in many ways, even with an ink-jet printer that sprays hydrophilic ink onto an acetate sheet. Dewcollecting devices have the problem of making the water run off the collecting surface. Another Namibian creature well adapted to survive on dew is the plant Welwitschia mirabilis. Despite its look, it is a close relative of pine trees. It has a short trunk split in two and from each side grows a single leaf that can be several meters long. Welwitschias are among the oldest

DEW DESERTS

plants on the earth, and some of them are 1500 years old. The trick that allows this plant to survive in one of the driest deserts of the planet is a two-part root system. Welwitschias have a long root that goes deep in the soil trying to reach the water table and extends for several meters away from the plant. These roots collect all the dew that forms on the top layer of soil. Moreover, all the water that condenses on the leaves forms big drops that eventually run down, so that the plant can water its own roots.

DEW DESERTS GIORA J. KIDRON The Hebrew University of Jerusalem Jerusalem, Israel

The term dew deserts denotes deserts that receive dew precipitation throughout the year. Dew is defined as condensation of atmospheric vapor and does not include vapor condensation that stems from the wet ground, i.e., distillation. We term dew desert as a desert that receives at least 10 mm of dew in >0.1 mm of daily precipitation. As a result of dew precipitation, primary production in dew deserts is much higher, affecting the entire food chain. Like the Atacama Desert in South America, the Namib Desert in South Africa, and the Sonora Desert in Baja California (North America), dew only affects those portions of the desert close to a sea or ocean, i.e., to a large body of water that can serve as an adequate vapor source. As most deserts occupy mainly the interior of the continents, the term applies only to a small part of near sea deserts. At present, information regarding the extent of dew precipitation within deserts is scarce. Most of the available information pertains to the heart of the Negev Desert where annual precipitation is 90–100 mm (1). There, a total of 33 mm was annually measured (2). Nevertheless, based on rain isohyets and the proximity to a sea or ocean (3), one may hypothesize that parts of the Sahara and the Arabian Peninsula are in fact dew deserts. Certain areas within the fog deserts may also be dew deserts (4–6). Although daily dew amounts may usually range between 0.1 and 0.3 mm and the total annual sum may be much smaller than that of rain precipitation, the occurrence of a steady and constant source may be of great importance in arid and semiarid zones (7–9). As the threshold for organism activity was found to be 0.03 mm of dew (10), amounts of 0.1–0.3 mm are still sufficient to allow growth activity. Although small, the dew amounts received may allow the establishment and distribution of many species, otherwise not capable of inhabiting the site. Consequently, higher biomass and higher species diversity may characterize the dew deserts. Biomass increase per millimeter of dew may be much higher than that calculated for rain (7) and that extrapolated, by adding the addition of water from dew to that of rain. The dew may thus be an important source of moisture for the primary food chain in arid and semiarid zones (11).

201

Many groups of micro-organisms were hypothesized to use dew. Among them, cyanobacteria, hypolithic algae and cyanobacteria that occupy the underside of stones (12,13), and lichens (14,15). Among the lichens are the endolithic lichens, embedded within calcite or dolomite crystals at the upper 1–5 mm of the rock and stone surface (16); epilithic and epedaphic lichens that dwell on the rock and soil surface, respectively (17); foliose (18); and fruticose lichens (18,19). In addition, bugs, ants, and beetles were reported to use dew directly by drinking it (20), whereas isopods and beetles were reported to use it indirectly by consuming wet food (20). When wet, snails in the Negev Desert were reported to successfully graze on endolithic lichens. By doing so, the snails disintegrate the limestone, a weathering phenomenon that can take place only upon rock moistening (21). Although cyanobacteria, hypolithic algae, and cyanobacteria, epedaphic, endolithic, epilithic, foliose, and fruticose lichens were all hypothesized to use dew, positive evidence for dew use under field conditions was obtained only from endolithic, epilithic, foliose, and fruticose lichens. These lichens were shown to photosynthesize for 2–4 h in the morning following dew in the Negev Desert (Fig. 1) (10,18,22). One should note that whereas lithobiontic lichens (inhabiting rocks and stones) were shown to use dew (22), dew use by micro-organisms inhabiting the soil is still controversial. Whereas Lange et al. (22) and Veste et al. (23) showed the use of dew by epedaphic lichens inhabiting soil and sand, respectively, and indirect evidence, expressed by the development of sexual and vegetative reproduction organs, were also monitored in mosses in the Negev Desert (24), the use of dew by endedaphic cyanobacteria inhabiting sand is not certain. According to Jacobs et al. (25), dew may moisten the surface. However, other reports indicate that dew moistening of the soil is rare (24,26). According ¨ to Bunnenberg and Kuhn (26), an amount of 0.13 mm of dew at 9 cm above ground amounted to 0.03 mm only at the surface. Similarly, Kidron et al. (24) showed that average dew precipitation of 0.1 mm measured on glass plates at 0.7 cm above ground amounted to only 0.034 mm at the surface. As cyanobacteria need liquid water for growth (27), and the necessary threshold for net photosynthesis was 0.1 mm (28), only rarely sufficient dew will moisten the surface to reach >0.1 mm, and thus the contribution of dew to the growth and development of endedaphic cyanobacteria may be negligible. The uncertainty regarding the use of dew by certain micro-organisms and at certain habitats is mainly linked to two factors: (a) Does dew condense at all habitats? (b) Is the amount supplied sufficient for use and for net carbon gain? As for the second question, rapid dew evaporation during the morning may result in a net loss of carbon, as was shown for the lichen Ramalina maciformis following mornings of low dew precipitation (10). Whereas the efficiency to which different microorganisms may use dew may be species-dependent, no controversy exists as to the perquisite conditions necessary, i.e., the capability of the dew to condense at their habitat. This ability depends, of course, on the existence of sufficient moisture in the air. It also depends

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mg CO2 g−1 h1

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1.00

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0.75

Balance Assim.-yield − resp.-loss = net yield

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0 −0.25

60 Water content

% H2O

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0000

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Figure 1. Apparent CO2 exchange rate (above) and the thalli water content (below) of Ramalina maciformis during 24 h in September 1967. Net carbon yield resulted in 0.146 mg C per gram of dry weight a day (18).

on substrate temperatures that will dictate whether the dew point temperature, necessary for vapor condensation, will be reached. Once reached, dew condensation will take place. As dew was found to condense at a more or less constant rate (29), the time during which the dew point is reached may dictate, to a large extent, the overall amount of dew that will be condensed. Furthermore, as phototrophic microorganisms necessitate light hours for photosynthesis, the length of time during which dew is available may dictate, to a large extent, the overall net gain of organic carbon owing to dew use. Thus, variables that may affect dew condensation and dew duration should be considered once dew availability in different habitats is examined. Factors such as topographical elevation, height above ground, aspect, location along the slope, and angle may all affect dew condensation. Recent dew measurements carried in the Negev Desert aimed to examine the role of the above-mentioned factors. Thus, in order to obtain continuous dew amounts and duration, a simple and inexpensive method was adapted. The Cloth-Plate Method (CPM) consists of 10 × 10 × 0.2 glass plates glued at their bottom to 10 × 10 × 0.5 cm plywood plates, thus creating an identical substratum (30). Velvet-like cloth (6 × 6 × 0.15 cm) are attached each afternoon to the center of the glass plate and collected throughout the following morning into glass jars that are immediately sealed and then weighed in a nearby lab, oven dried (in 70 ◦ C) and then weighed again, and their moisture content is calculated. By placing plates next to each other within a certain habitat, the CPM facilitates inexpensive large-scale continuous measurements (Fig. 2). The readings are also not affected by wind, as might be the case with some other devices (31).

Figure 2. Dew measurements by the CPM in the Negev Highlands.

The CPM was used to assess the possible role of altitude and distance from the sea. When simultaneous measurements at three locations at 250, 550, and 1000 m above sea level being 37, 55, and 98 km from the Mediterranean Sea were carried out in the Negev Desert, positive correlations between dew and fog amounts and altitude were found (Fig. 3). Dew precipitation, as well as fog precipitation, increased with altitude, with the most elevated location receiving more than twice the amount obtained at the topographically lowest location, which was the case, although the most elevated location was also the farthest away from the Mediterranean Sea, i.e., from a vapor source. The data thus indicate that, within 100 km from the sea, topographical height may compensate for

DEW DESERTS 0.45 0.4

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r = 0.93

0.05 0 5.5

6

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9.5 10 10.5 11

Figure 4. A typical condensation and evaporation pattern during the morning hours (30).

0.4 0.35 0.3 Dew amounts (mm)

the increase in distance from the vapor source. In this regard, one should note that elevation plays a major role in controlling dew and fog precipitation also in the Atacama Desert (4). The CPM facilitates careful analysis of dew precipitation. When continuous dew measurements were carried out, a typical condensation and evaporation pattern was obtained during the morning hours (Fig. 4). Dew condensation was found to continue also after dawn and sunrise, explained by radiation-induced air turbulence (30). As for the relationship between dew amounts and height above ground, dew values showed high variability (33,34). When dew was measured on a hilltop in the Negev Highlands at 0.7, 10, 20, 30, 40, and 50 cm above ground, maximal values were recorded at 10 cm above ground (Fig. 5). The findings reflected two effects: the warming effect of the soil as a result of a nocturnal heat flux that rises from the deeper horizons of the soil, responsible for the decrease in dew values near the ground (31,34,35), and the air turbulence at height, resulting from higher wind velocities, responsible for lower dew values farther away from the ground (35,36). High variability in dew amounts and duration was also found in the Negev Desert when dew was measured at 0.7 and 40 cm above ground along limestone slopes (approximately 50 m long) of four aspects (north, south, east, and west) within a second-order drainage basin (Fig. 6). Whereas at 0.7 cm above ground, the hilltops and the bottom parts of the northern and western aspects received the highest amounts, the wadi beds received approximately half these values (Fig. 6a). The lowest values were received at the south-facing midslope (located at the lee side of the prevailing north north-west winds), being approximately a quarter of the maximal values recorded. At 40 cm above ground, a reverse trend was obtained with the south-facing midslope and the wadi beds receiving the highest values (Fig. 7). And thus, whereas the 0.7 cm above-ground results were not in accordance with the classical model that predicts high dew values

203

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Figure 3. The relationships between dew and fog precipitation with altitude in the Negev Desert (32).

at the wadi beds (because of nocturnal katabatic wind) and higher amounts at the lee side of the wind, i.e., at the south-facing slope (because of undisturbed inversion), the 40 cm above ground measurements corresponded to the classical model. The discrepancy was explained by the overwhelming impact of the rock surface temperatures on the dew values at 0.7 cm above ground, with southfacing rock surfaces being 3–8 degrees higher than the north-facing rock surfaces throughout the night. As for the higher amounts of dew received at 0.7 cm above ground at the hilltops in comparison with the wadi

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Figure 6. Dew amounts (a) and duration (b) within 18 stations located at 4 aspects of a second-order drainage basin in the Negev Desert. Top = hilltop, Up = Upper slope, Mid = Midslope, Bot = Bottom slope (37).

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beds, it was hypothesized that the afternoon winds act as a cooling agent, facilitating a much earlier drop in surface temperatures (and thus reaching much faster the necessary dew point temperature) at the wind-exposed hilltops (37). Aspect and slope location also dictate dew duration, seen as being of even greater ecological importance to organisms than the maximal amount (38–40). Generally, a positive correlation between dew amounts and duration was found (34,37), and thus, by affecting dew amounts, aspect and slope position also affect dew duration (Fig. 6b). Nevertheless, dew duration was also affected by slope

0.3

2

N/S E/W Wadi

Figure 7. Dew measurements at 0.7 and 40 cm above ground at three habitats within a second-order drainage basin in the Negev Desert. N = northern exposure, S = southern exposure, Wadi N/S = wadi between the northern and southern exposure, Wadi E/W = wadi between the eastern and western exposures (37).

location and aspect beyond these relations. Although condensation was found to take place in all habitats also after sunrise (for usually 0.5–1.0 h after sunrise), condensation at the sun-sheltered habitats of the bottom north- and west-facing slopes continued for up to 1.5 h following sunrise (41). As a result of the higher maximal values and the limited desiccation effect of the sun in these habitats, these habitats were characterized by longer dew duration. The substrate angle is another factor found to affect dew amounts and duration. When cloths were attached to 50 × 50 × 10 cm wooden boxes, having sides of different angles (30, 45, 60, 75, and 90◦ ) placed at different aspects (facing north, south, west, and east) on a hilltop, a decrease in dew amounts with an increase in angle, from 30◦ to 90◦ , was found and dew amounts were positively correlated with cos(θ ) (Fig. 8). Thus, dew amounts obtained at an angle of 90◦ was approximately a quarter of the values obtained at a horizontal surface and at an angle of 30◦ , both of which received similar values. This difference was explained by the slower rate of nocturnal cooling that takes place with an increase in angle in accordance with the lower proportion of sky seen by the substrate (43). No preferential condensation in accordance with aspect was found. Nevertheless, dew duration was aspect-dependent with a decrease in duration following the order west>north>south>east (Fig. 9). Thus, daylight dew duration was approximately double at the westfacing sun-sheltered angle than at the sun-exposed eastfacing aspect. The variability in dew amounts and duration may have important consequences for micro-organisms. For

DEW DESERTS (a)

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1

Figure 9. The relationships between dew duration with (a) angle (in degrees) and (b) cos(θ) in the Negev Desert (42).

Cos (∅)

most micro-organisms, the ability to use dew is primarily a function of the amount condensed and the duration during which dew is available, both of which are habitatdependent. Thus, when cloths were attached to different substrates, high variability in dew amounts was received on loose cobbles, partially embedded cobbles, and rock surfaces (Fig. 10). Loose cobbles that rapidly cool at night were found to receive approximately twice the amounts of embedded cobbles and more than 4 times the amounts condensed on a nearby rock (44). Whereas the cobbles were primarily inhabited by endolithic lichens, the rock surfaces were mainly inhabited by epilithic lichens. Similarly, lichen growth on loose cobbles was controlled by angle-induced dew. Thus, the top of loose cobbles, inhabited by lichens, received approximately twice the amount condensed on the uninhabited side of the cobble (45). Consequently, dew contribution to the ecosystem biomass may be highly important. For instance, in research conducted by Kappen et al. (22,46) in the Negev Desert, the biomass of lichens (with Ramalina maciformis predominating) in the northern aspect was over 200 g m−2 , three to five times as much than at the other aspects. Although some of the differences can certainly be attributed also to use of rain-induced moisture, the fact that the main water source for Ramalina maciformis and other endolithic and epilithic lichens in the Negev Desert was dew pointed to the importance of dew on lichen biomass. The difference in lichen cover of

the sun-sheltered and the sun-exposed slopes points also to the role of dew in lichen distribution (47). As lichens may serve as food for snails, the whole food chain may be affected by dew precipitation. Dew precipitation may also affect vascular plants. The effect may not always be positive, as dew may hasten plant fungal infection (7,48). Dew may preferentially accumulate on leaves because of their lower temperature (0.5–2 ◦ C lower in comparison to the ambient air) (35,36) and thus, wet the leaves for several hours during the morning (34).

B

0.3

B

0.25 Dew amounts (mm)

Figure 8. The relationships between dew amounts with (a) angle (in degrees) and (b) cos(θ) as measured in the Negev Desert (42).

J H

B

Loose Cobbles Embedded Cobbles Rock

B

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1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 Days of measurements

Figure 10. Simultaneous dew amounts as obtained on loose and embedded cobbles inhabited by endolithic lichens, and rock surface inhabited by epilithic lichens (44).

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However, no conclusive evidence of direct dew use by vascular plants is yet available, and the research conducted on this topic is controversial (7,8,31,49,50). Dew, however, may have an indirect effect, facilitating stomata opening (51), longer hours of photosynthesis (39), reducing transpiration (7), and affecting flowering (7). Dew may facilitate seedling survival (52), recovery from wilting (34), plant growth (34), and yield (34,53). By affecting micro-organisms and plant growth, and by facilitating insect survival (20), dew also indirectly affects the nutrient cycling (54) and soil-forming processes (55), which in turn has an important impact on the ecosystem and may thus call for a specific awareness of the dew deserts. BIBLIOGRAPHY 1. Rosenan, N. and Gilad, M. (1985). Meteorological data. In: Atlas of Israel. Carta, Jerusalem. 2. Evenari, M. (1981). Ecology of the Negev Desert, a critical review of our knowledge. In: Developments in Arid Zone Ecology and Environmental Quality. H. Shuval (Ed.). Balaban ISS, Philadelphia, PA, pp. 1–33. 3. Reitan, C.H. and Green, C.R. (1968). Appraisal of research on weather and climate of desert environments. In: Deserts of the World, An Appraisal of Research Into Their Physical and Biological Environments. W.G. McGinnies, B.J. Goldman, and P. Paylore (Eds.). The University of Arizona Press, Tucson, AZ, pp. 19–92. 4. Redon, J. and Lange, O.L. (1983). Epiphytic lichens in the region of a Chilean ‘‘fog oasis’’ (Fray Jorge). 1. Distributional patterns and habitat conditions. Flora. 174: 213–243. 5. Lange, O.L., Meyer, A., Ullmann, I., and Zellner, H. (1991). Microclimate conditions, water content and photosynthesis of lichens in the coastal fog zone of the Namib Desert: measurements in the fall. Flora. 185: 233–266. 6. Lange, O.L., Meyer, A., Zellner, H., and Heber, U. (1994). Photosynthesis and water relations of lichen soil crusts: field measurements in the coastal fog zone of the Namib Desert. Functional Ecol. 8: 253–264. 7. Wallin, G.R. (1967). Agrometeorological aspects of dew. Agric. Meteorol. 4: 85–102. 8. Evenari, M., Shanan, L., and Tadmor, N. (1971). The Negev, the Challenge of a Desert. Harvard Univ. Press, Cambridge, MA, p. 345. 9. Noy-Meir, I. (1973). Desert ecosystems: environment and producers. Ann. Rev. Ecol. Syst. 4: 25–51. 10. Kappen, L. etal. (1979). Ecophysiological investigations on lichens of the Negev Desert, IV: Annual course of the photosynthetic production of Ramalina maciformis (Del.) Bory. Flora. 168: 85–108. 11. Shachak, M. and Steinberger, Y. (1980). An algae-desert snail food chain: energy flow and soil turnover. Oecologia. 46: 402–411. 12. Berner, T. (1974). The Ecophysiology of the Hypolithic Algae in the Negev Highlands of Israel. Ph.D. Thesis, The Hebrew University of Jerusalem. 13. Berner, T. and Evenari, E. (1978). The influence of temperature and light penetration on the abundance of the hypolithic algae in the Negev Desert of Israel. Oecologia. 33: 255–260. 14. Friedmann, E.I., Lipkin, Y., and Ocampo-Paus, R. (1967). Desert algae of the Negev. Phycologia. 6: 185–195. 15. Friedmann, E.I. and Galun, M. (1974). Desert algae, lichens and fungi. In: Desert Biology II. G.W. Brown (Ed.). Academic Press, New York, pp. 165–212.

16. Fry, E.I. (1922). Some types of endolithic lichens. Ann. Bot. 35: 541–562. 17. Golubic, S., Freiedmann, I., and Schneider, J. (1981). The lithobiontic ecological niche, with special reference to microorganisms. J. Sedimentary Petrol. 51: 475–478. 18. Lange, O.L., Schulze, E-D., and Koch, W. (1970). Ecophysiological investigations on lichens of the Negev Desert, II: CO2 gas exchange and water conservation of Ramalina maciformis (Del.) Bory in its natural habitat during the summer dry period (Technical translation 1655. National Research Council of Canada). Flora. 159: 38–62. 19. Lange, O.L. (1969). Ecophysiological investigations on lichens of the Negev Desert. I. CO2 gas exchange of Ramalina maciformis (Del.) Bory under controlled conditions in the laboratory (Technical translation 1654. National Research Council of Canada). Flora. 158: 324–359. 20. Broza, M. (1979). Dew, fog and hygroscopic food as a source of water for desert arthropods. J. Arid. Environ. 2: 43–49. 21. Shachak, M., Jones, C.G., and Granot, Y. (1987). Herbivory in rocks and the weathering of a desert. Science 236: 1098–1099. 22. Lange, O.L., Schulze, E-D., and Koch, W. (1970). Ecophysiological investigations on lichens of the Negev Desert, III: CO2 gas exchange and water metabolism of crustose and foliose lichens in their natural habitat during the summer dry period. Flora. 159: 525–538. 23. Veste, M., Littmann, T., and Friedrich, H. (2001). Microclimate boundary conditions for activity of soil lichen crusts in sand dunes of the north-western Negev Desert, Israel. Flora. 196: 465–474. 24. Kidron, G.J., Hernstadt, I., and Barzilay, E. (2002). The role of dew as a moisture source for sand microbiotic crusts in the Negev Desert, Israel. J. Arid. Environ. 52: 517–533. 25. Jacobs, A.F.G., Heusinkveld, B.G., and Berkowicz, S.M. (2000). Dew measurements along a longitudinal sand dune transect, Negev Desert, Israel. Int. J. Biometeorol. 43: 184–190. ¨ 26. Bunnenberg, C. and Kuhn, W. (1977). Application of the βabsorption method to measure dew on soil and plant surfaces. Int. J. Appl. Radiation Isotopes 28: 751–754. 27. Lange, O.L., Kilian, E., and Ziegler, H. (1986). Water vapor uptake and photosynthesis of lichens: performance differences in species with green and blue-green algae as phycobionts. Oecologia 71: 104–110. 28. Lange, O.L. et al. (1992). Taxonomic composition and photosynthetic characteristics of the ‘‘biological soil crusts’’ covering sand dunes in the Western Negev Desert. Func. Ecol. 6: 519–527. 29. Zangvil, A. (1996). Six years of dew observation in the Negev Desert, Israel. J. Arid. Environ. 32: 361–372. 30. Kidron, G.J. (1998). A simple weighing method for dew and fog measurements. Weather 53: 428–433. 31. Noffsinger, T.L. (Ed.). (1965). Survey of techniques for measuring dew. In: Humidity and Moisture, Measurement and Control in Science and Industry. A. Wexler (Ed.). Reinhold Publishing Corporation, New York, pp. 523–531. 32. Kidron, G.J. (1999). Altitude dependent dew and fog in the Negev desert, Israel. Agric. Forest Meteorol. 96: 1–8. 33. Lloyd, M.G. (1961). The contribution of dew to the summer water budget of Northern Idaho. Bull. Am. Meteorol. Soc. 42: 572–580. 34. Duvdevani, S. (1964). Dew in Israel and its effect on plants. Soil Sci. 98: 14–21. 35. Long, I.F. (1958). Some observations on dew. Meteorol. Mag. 87: 161–168.

DEW POINT 36. Monteith, J.L. (1957). Dew. Quart. J. Royal Meteorol. Soc. 83: 322–341.

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37. Kidron, G.J., Yair, A., and Danin, A. (2000). Dew variability within a small arid drainage basin in the Negev highlands, Israel. Quart. J. Royal Meteorol. Soc. 126: 63–80. 38. Lange, O.L., Geiger, I.L., and Schulze, E-D. (1977). Ecophysiological investigations on lichens in the Negev Desert. V. A model to simulate net photosynthesis and respiration of Ramalina macifomis. Oecologia. 28: 247–259. 39. Kappen, L. et al. (1980). Ecophysiological investigations on lichens of the Negev Desert, VII: The influence of the habitat exposure on dew imbibition and photosynthetic productivity. Flora. 169: 216–229. 40. Lange, O.L. and Tenhunen, J.D. (1982). Water relations and photosynthesis of desert lichens. J. Hattori. Bot. Lab. 53: 309–313. 41. Kidron, G.J. (2000). Analysis of dew precipitation in three habitats within a small arid drainage basin, Negev Highlands, Israel. Atmos. Res. 55: 257–270. 42. Kidron, G.J. (in press). Angle and aspect dependent dew precipitation in the Negev Desert, Israel. J. Hydrol. 43. Oke, T.R. (1978). Boundary Layer Climates. John Wiley & Sons, New York, p. 372. 44. Kidron, G.J. (2000). Dew moisture regime of endolithic and epilithic lichens inhabiting calcareous cobbles and rock outcrops, Negev Desert, Israel. Flora. 195: 145–153. 45. Kidron, G.J. (2002). Causes of two patterns of lichen colonization on cobbles in the Negev Desert, Israel. Lichenologist. 34: 71–80. 46. Kappen, L., Lange, O.L., Schulze, E-D., Evenari, M., and Buschbom, U. (1975). Primary production of lower plants (lichens) in the desert and its physiological basis. In: Photosynthesis and Productivity in Different Environments. J.P. Cooper (Ed.). Cambridge University Press, Cambridge, UK, pp. 133–143. 47. Danin, A. and Garty, J. (1983). Distribution of cyanobacteria and lichens on hillsides of the Negev Highlands and their impact on biogenic weathering. Z. Geomorph. 27: 423–444. 48. Duvdevani, S., Reichert, I., and Palti, J. (1946). The development of downy and powdery mildew of Cucumbers as related to dew and other environmental factors. Palestine J. Bot. (Rehovot Series). 2: 127–151. 49. Stone, E.C. (1957). Dew as an ecological factor. I. A review of the literature. Ecology 38: 407–413. 50. Waisel, Y. (1958). Dew absorption by plants of arid zones. Bull. Res. Counc. Israel 6D: 180–186. 51. Schulze, E-D. et al. (1972). Stomatal responses to changes in humidity in plants growing in the desert. Planta 108: 259–270. 52. Stone, E.C. (1957). Dew as an ecological factor. II. The effect of artificial dew on the survival of Pinus ponderosa and associated species. Ecology 38: 414–422. 53. Rotem, J. and Reichert, J. (1964). Dew - a principal moisture factor enabling early blight epidemics in a semi-arid region of Israel. Plant Disease Reptr. 48: 211–215. 54. Jones, C.G. and Shachak, M. (1990). Fertilization of the desert soil by rock-eating snails. Nature 345: 839–841. 55. Shachak, M., Jones, C.G., and Brand, S. (1995). The role of animals in an arid ecosystem: Snails and isopods as controllers of soil formation, erosion and desalinization. Adv. GeoEcol. 28: 37–50.

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Frascati (RM), Italy

The dew point is the temperature to which air must be cooled at constant pressure until it reaches saturation and the water, or any other vapor, begins condensing. The result is the formation of fog. When the dew point falls below freezing, it is called the frost point. In this case, water forms ice crystals directly. The dew point is also a good indicator of the amount of water contained in the air. The more humid the air, the higher the dew temperature. The dew point is often reached in the evening, when the air cools down, which is why fog is more common in the evening, or why one normally finds that dew or frost are common in the morning but disappear as soon as the temperature increases. Dew point and relative humidity can be measured with an instrument called a wet- and dry-bulb psychrometer. The dry bulb is a normal thermometer that measures the actual air temperature. The wet bulb is another thermometer, but has a bulb wrapped in a piece of cloth dampened by a string that dips into a bottle of distilled water. Unless the air is so humid that it is very close to saturation, the wet bulb thermometer measures a lower temperature, because it is cooled by the evaporation of water. The difference between the dry-bulb and wet-bulb temperatures is called the wet-bulb depression. It is then possible to calculate the dew point as a function of the wet-bulb depression and of the dry-bulb temperature. To calculate the dew point starting from the psychrometer readings, the first step is to calculate the saturation vapor pressure in millibars, corresponding to the dry- and wet-bulb temperature: Es = 6.11 × 10.0[7.5T/(237.7 + T)] Eswb = 6.11 × 10.0[7.5Twb /(237.7 + Twb )] where T and Twb are the readings of the dry- and wet-bulb temperatures. Now we are ready to calculate the actual mixing ratio of the air: W = [(T − Twb )(Cp ) − Lv (Eswb /P)]/[−(T − Twb )(Cpv ) − Lv ] where

Cp = specific heat of dry air at constant pressure (J/g)∼1.005 J/g Cpv = specific heat of water vapor at constant pressure (J/g)∼4.186 J/g Lv = Latent heat of vaporization (J/g)∼2500 J/g T = air temperature in ◦ C Twb = wet bulb temperature in ◦ C Eswb = saturation vapor pressure at the wet-bulb temperature (mb) P = atmospheric pressure at the surface ∼1013 mb at sea level

We can now use the following formula to obtain the saturation mixing ratio for the air: Ws = Es /P

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RH = W/Ws Now we can use the relative humidity to calculate the actual vapor pressure (E) of the air as follows: E = RH × Es The dew point temperature is then Td = [−430.22 + 237.7 × ln(E)]/[− ln(E) + 19.08] Let us now assume a psychrometer that gives the following readings at a pressure of 1013 mbar: ◦

T = 30 C ◦

Twb = 20 C Then, Eswb = 23.34 mbar Es = 42.31 mbar The mixing ratio becomes W = 0.019 Ws = 0.042 The relative humidity is then, RH = 0.45 (or 45%) The actual vapor pressure is then, E = 18.95 mbar and the dew point is ◦

Td = 16.7 C Instead of going through all these formulas, it is possible to use tables that give the dew point temperature as a function of the readings for different pressures.

DROUGHTS ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

A drought is a period of time during which weather is in one way or another excessively dry. A drought’s severity is determined by the degree of moisture deficiency in the affected area, as measured by various methods. Droughts are a standard, recurring, and normal feature of longterm climate. There are four generally accepted categories of drought: meteorological, socioeconomic, hydrologic, and agricultural.

Meteorological droughts involve a lack of precipitation during a given time period. Their specific definition depends mostly on the region in question; a lack of precipitation for some areas would be excessive precipitation in others. Some regions receive rain yearround, and some receive virtually none. Likewise, some regions receive precipitation consistently throughout the year, and others have seasonal precipitation patterns. A departure from any of these climatological characteristics as abnormally low precipitation can cause a meteorological drought. Meteorological droughts are usually measured in terms of their duration and in contrast to established precipitation averages for specified time periods for a specified area. This type of drought generally precedes the other three types. Agricultural droughts need not involve a severe lack of precipitation. They are marked by damage to crops and plants due to insufficient water supplies. Because water from rain and snow is more likely to run into creeks and rivers instead of filtering into the soil due to human intervention, agricultural droughts can sometimes occur in spite of noticeable rainfall. Depending on a plant’s species and stage of growth, it needs varying amounts of water from the soil; when the majority of crops cannot obtain the amount of water they need, an area is said to be undergoing an agricultural drought. This type of drought is most damaging to farmers and those in poor countries, and it generally comes after the onset of a meteorological drought. A hydrologic drought involves water shortages in reservoirs, lakes, and streams due to lack of precipitation. These droughts are measured by stream flow and water levels in local reservoirs. Local climate is not the only factor that causes a hydrologic drought; altered land usage, dams, and degraded soil quality can all have effects as well. Reduced precipitation in one area can cause a hydrologic drought in another because various areas are hydrologically connected by their rivers, lakes, and other bodies of water. The primary cause of this type of drought is generally a meteorological drought; hydrologic droughts usually come some time after agricultural droughts. The final category of drought is socioeconomic. This type of drought, like the other three, involves a lack of precipitation. It occurs when reduced precipitation causes a shortage of water as related to the demands of the local populace. The socioeconomic drought is the only type of drought that has a major effect on the general population, and it is generally the last of the four types to occur. Aside from lack of rain, the three most important variables to look at during a drought are air temperature, humidity, and circulation patterns in the atmosphere. Atmospheric circulation patterns that produce little or no precipitation are often associated with droughts. Climate abnormalities are another component of a drought. Precipitation levels that would be considered a drought in one area are normal in another that has a different climate. To understand the cause of a drought, scientists study the circulation patterns of the atmosphere across global distances, using computer-generated atmospheric simulator models.

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Droughts also have direct effects on communities that experience this phenomenon. Both water quality and cost are negatively affected by droughts. Changes in the taste and odor of drinking water can also occur. Those who draw their water from wells may have to drill more deeply to reach the lowered level of the water table. During a drought, water conservation is important. Restrictions on watering gardens, washing cars, and using sprinkler systems are applied when a drought emergency or warning is in effect. There are generally three stages of drought preparedness: drought watches, drought warnings, and drought emergencies. During a drought watch, no restrictions are enforced, but citizens are encouraged to curb water usage because the possibility of drought has been established for an area. During a drought warning, restrictions are still not enforced, but citizens are instructed to reduce water usage further in response to impending drought conditions. The final stage is a drought emergency. This is the most serious stage, often involving mandatory water usage guidelines to ensure that sufficient water is available for critical needs. At this stage, officials try to avoid local shortages by evenly distributing the water available. Droughts can have a great impact in many ways. There are three main types of impact they can have: economic, social, and environmental. A lack of water in the soil for crops can cause farmers great financial difficulty. The resulting decrease in crop production has a ripple effect on the economy, causing short-term and long-term problems. One of the biggest short-term problems is unemployment. Certain industries can also be affected by long-term problems. For instance, loss of tax revenue can hinder tourism. Brush and trees can become very dry, and an outbreak of destructive fires can result. Drought can harm the logging industry, fisheries, and hydroelectric power generation, all economic effects. It can also cause insect problems in agriculture, as well as erosion and disease. Environmental effects include reduced biodiversity, degraded air quality, and perhaps most importantly, devastating forest fires. Social impacts include induced emigration, rampant famine, and, indirectly, greater poverty. These are merely a few of the multitude of impacts, direct and indirect, that drought can have.

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Understanding and Defining Drought. Available: http://www.drought.unl.edu/whatis/concept.htm. Drought Information Center. Available: http://www.dep.state.pa. us/dep/subject/hotopics/drought/drtterm.htm.

DROUGHT INDICES MICHAEL J. HAYES Climate Impacts Specialist—National Drought Mitigation Center

INTRODUCTION Drought indices assimilate thousands of bits of data on rainfall, snowpack, streamflow, and other water supply indicators into a comprehensible big picture. A drought index value is typically a single number, far more useful than raw data for decision making. There are several indices that measure how much precipitation for a given period of time has deviated from historically established norms. Although none of the major indices is inherently superior to the rest in all circumstances, some indices are better suited than others for certain uses. For example, the Palmer Drought Severity Index has been widely used by the U.S. Department of Agriculture to determine when to grant emergency drought assistance, but the Palmer is better when working with large areas of uniform topography. Western states, with mountainous terrain and the resulting complex regional microclimates, find it useful to supplement Palmer values with other indices such as the Surface Water Supply Index, which takes snowpack and other unique conditions into account. The National Drought Mitigation Center is using a newer index, the Standardized Precipitation Index, to monitor moisture supply conditions. Distinguishing traits of this index are that it identifies emerging droughts months sooner than the Palmer Index and that it is computed on various time scales. Most water supply planners find it useful to consult one or more indices before making a decision. What follows is an introduction to each of the major drought indices in use in the United States and in Australia.

READING LIST PERCENT OF NORMAL All About Droughts. (2002). Available: http://205.156.54.206/om/ drought.htm. (March 13). Hanson, Ronald L. Evaporation and Droughts, U.S. Geological Survey. (2002). Available: http://geochange.er.usgs.gov/sw/ changes/natural/er/. (March 13). Impacts of Drought. (2002). Available: http://enso.unl.edu/ndmc/ enigma/impacts.htm. (March 13). McNab, A. and Karl, T. Climate and Droughts, National Oceanic and Atmospheric Administration. (2002). Available: http://geochange.er.usgs.gov/sw/changes/natural/drought/. (March 13). Stormfax Guide to Droughts, Stormfax. Available: http://www.stormfax.com/drought.htm. Understanding Your Risk: Impacts of Drought. Available: http://www.drought.unl.edu/risk/impacts.htm. All About Droughts. Available: http://www.nws.noaa.gov/om/drought.htm.

Overview: The percent of normal is a simple calculation well suited to the needs of TV weathercasters and general audiences. Pros: Quite effective for comparing a single region or season. Cons: Easily misunderstood, as normal is a mathematical construct that does not necessarily correspond with what we expect the weather to be. The percent of normal precipitation is one of the simplest measurements of rainfall for a location. Analyses using This article is a US Government work and, as such, is in the public domain in the United States of America.

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the percent of normal are very effective when used for a single region or a single season. Percent of normal is also easily misunderstood and gives different indications of conditions, depending on the location and season. It is calculated by dividing actual precipitation by normal precipitation—typically considered to be a 30-year mean—and multiplying by 100%. This can be calculated for a variety of time scales. Usually these time scales range from a single month to a group of months representing a particular season, to an annual or water year. Normal precipitation for a specific location is considered to be 100%. One of the disadvantages of using the percent of normal precipitation is that the mean, or average, precipitation is often not the same as the median precipitation, which is the value exceeded by 50% of the precipitation occurrences in a long-term climate record. The reason for this is that precipitation on monthly or seasonal scales does not have a normal distribution. Use of the percent of normal comparison implies a normal distribution where the mean and median are considered to be the same. An example of the confusion this could create can be illustrated by the long-term precipitation record in Melbourne, Australia, for the month of January. The median January precipitation is 36.0 mm (1.4 in.), meaning that in half the years less than 36.0 mm is recorded, and in half the years more than 36.0 mm is recorded. However, a monthly January total of 36.0 mm would be only 75% of normal when compared to the mean, which is often considered to be quite dry. Because of the variety in the precipitation records over time and location, there is no way to determine the frequency of the departures from normal or compare different locations. This makes it difficult to link a value of a departure with a specific impact occurring as a result of the departure, inhibiting attempts to mitigate the risks of drought based on the departures from normal and form a plan of response (1). STANDARDIZED PRECIPITATION INDEX (SPI) Overview: The SPI is an index based on the probability of precipitation for any time scale. Who uses it: Many drought planners appreciate the SPI’s versatility. Pros: The SPI can be computed for different time scales, can provide early warning of drought and help assess drought severity, and is less complex than the Palmer. Cons: Values based on preliminary data may change. Developed by: T.B. McKee, N.J. Doesken, and J. Kleist, Colorado State University, 1993. SPI Values 2.0+ 1.5 to 1.99 1.0 to 1.49 −.99 to .99 −1.0 to −1.49 −1.5 to −1.99 −2 and less

extremely wet very wet moderately wet near normal moderately dry severely dry extremely dry

Monthly maps: http://www.drought.unl.edu/monitor/ spi.htm; http://www.wrcc.dri.edu/spi/spi.html. The understanding that a deficit of precipitation has different impacts on groundwater, reservoir storage, soil moisture, snowpack, and streamflow led McKee, Doesken, and Kleist to develop the Standardized Precipitation Index (SPI) in 1993. The SPI was designed to quantify the precipitation deficit for multiple time scales. These time scales reflect the impact of drought on the availability of the different water resources. Soil moisture conditions respond to precipitation anomalies on a relatively short scale. Groundwater, streamflow, and reservoir storage reflect the longer-term precipitation anomalies. For these reasons, McKee et al. (2) originally calculated the SPI for 3-, 6-, 12-, 24-, and 48-month time scales. The SPI calculation for any location is based on the long-term precipitation record for a desired period. This long-term record is fitted to a probability distribution, which is then transformed into a normal distribution so that the mean SPI for the location and desired period is zero (3). Positive SPI values indicate greater than median precipitation, and negative values indicate less than median precipitation. Because the SPI is normalized, wetter and drier climates can be represented in the same way, and wet periods can also be monitored using the SPI. McKee et al. (2) used the classification system shown in the SPI values table to define drought intensities resulting from the SPI. McKee et al. (2) also defined the criteria for a drought event for any of the time scales. A drought event occurs any time the SPI is continuously negative and reaches an intensity of −1.0 or less. The event ends when the SPI becomes positive. Each drought event, therefore, has a duration defined by its beginning and end, and an intensity for each month that the event continues. The positive sum of the SPI for all the months within a drought event can be termed the drought’s ‘‘magnitude’’. Based on an analysis of stations across Colorado, McKee determined that the SPI is in mild drought 24% of the time, in moderate drought 9.2% of the time, in severe drought 4.4% of the time, and in extreme drought 2.3% of the time (2). Because the SPI is standardized, these percentages are expected from a normal distribution of the SPI. The 2.3% of SPI values within the ‘‘Extreme Drought’’ category is a percentage that is typically expected for an ‘‘extreme’’ event (Wilhite 1995). In contrast, the Palmer Index reaches its ‘‘extreme’’ category more than 10% of the time across portions of the central Great Plains. This standardization allows the SPI to determine the rarity of a current drought, as well as the probability of the precipitation necessary to end the current drought (2). The SPI has been used operationally to monitor conditions across Colorado since 1994 (4). Monthly maps of the SPI for Colorado can be found on the Colorado State University website (http://ulysses.atmos.colostate.edu/SPI.html). It is also being monitored at the climate division level for the contiguous United States by the National Drought

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Mitigation Center and the Western Regional Climate Center (WRCC). You can download the SPI program and sample files here. PALMER DROUGHT SEVERITY INDEX (THE PALMER; PDSI) Overview: The Palmer is a soil moisture algorithm calibrated for relatively homogeneous regions. Who uses it: Many U.S. government agencies and states rely on the Palmer to trigger drought relief programs. Pros: The first comprehensive drought index developed in the United States. Cons: Palmer values may lag emerging droughts by several months; less well suited for mountainous land or areas of frequent climatic extremes; complex—has an unspecified, built-in time scale that can be misleading. Developed by: W.C. Palmer, 1965. Weekly maps: http://www.cpc.ncep.noaa.gov/products/ analysis monitoring/regional monitoring/palmer .gif. Palmer Classifications 4.0 or more 3.0 to 3.99 2.0 to 2.99 1.0 to 1.99 0.5 to 0.99 0.49 to −0.49 −0.5 to −0.99 −1.0 to −1.99 −2.0 to −2.99 −3.0 to −3.99 −4.0 or less

extremely wet very wet moderately wet slightly wet incipient wet spell near normal incipient dry spell mild drought moderate drought severe drought extreme drought

In 1965, W.C. Palmer developed an index to measure the departure of the moisture supply (5). Palmer based his index on the supply-and-demand concept of the water balance equation, taking into account more than just the precipitation deficit at specific locations. The objective of the Palmer Drought Severity Index (PDSI), as this index is now called, was to provide measurements of moisture conditions that were standardized so that comparisons using the index could be made between locations and between months (5). The PDSI is a meteorological drought index, and it responds to weather conditions that have been abnormally dry or abnormally wet. When conditions change from dry to normal or wet, for example, the drought measured by the PDSI ends without taking into account streamflow, lake and reservoir levels, and other longer-term hydrologic impacts (6). The PDSI is calculated based on precipitation and temperature data, as well as the local Available Water Content (AWC) of the soil. From the inputs, all the basic terms of the water balance equation can be determined, including evapotranspiration, soil recharge, runoff, and moisture loss from the surface layer. Human impacts on the water balance, such as irrigation, are not considered.

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Complete descriptions of the equations can be found in the original study by Palmer (5) and in the more recent analysis by Alley (7). Palmer developed the PDSI to include the duration of a drought (or wet spell). His motivation was as follows: an abnormally wet month in the middle of a long-term drought should not have a major impact on the index, or a series of months with near-normal precipitation following a serious drought does not mean that the drought is over. Therefore, Palmer developed criteria for determining when a drought or a wet spell begins and ends, which adjust the PDSI accordingly. Palmer (5) described this effort and gave examples, and it is also described in detail by Alley (7). In near-real time, Palmer’s index is no longer a meteorological index but becomes a hydrological index referred to as the Palmer Hydrological Drought Index (PHDI) because it is based on moisture inflow (precipitation), outflow, and storage, and does not take into account the long-term trend (6). In 1989, a modified method to compute the PDSI was begun operationally (8). This modified PDSI differs from the PDSI during transition periods between dry and wet spells. Because of the similarities between these Palmer indices, the terms Palmer Index and Palmer Drought Index have been used to describe general characteristics of the indices. The Palmer Index varies roughly between −6.0 and +6.0. Palmer arbitrarily selected the classification scale of moisture conditions based on his original study areas in central Iowa and western Kansas (5). Ideally, the Palmer Index is designed so that a −4.0 in South Carolina has the same meaning in terms of the moisture departure from a climatological normal as a −4.0 in Idaho (7). The Palmer Index has typically been calculated on a monthly basis, and a long-term archive of the monthly PDSI values for every climate division in the United States exists with the National Climatic Data Center from 1895 through the present. In addition, weekly Palmer Index values (actually modified PDSI values) are calculated for the climate divisions during every growing season and are available in the Weekly Weather and Crop Bulletin. These weekly Palmer Index maps are also available on the World Wide Web from the Climate Prediction Center at http://www.cpc.ncep.noaa.gov/products/analysis monitoring/regional monitoring/palmer.gif. The Palmer Index is popular and has been widely used for a variety of applications across the United States. It is most effective measuring impacts sensitive to soil moisture conditions, such as agriculture (1). It has also been useful as a drought monitoring tool and has been used to trigger actions associated with drought contingency plans (1). Alley (7) identified three positive characteristics of the Palmer Index that contribute to its popularity: (1) it provides decision makers with a measurement of the abnormality of recent weather for a region; (2) it provides an opportunity to place current conditions in historical perspective; and (3) it provides spatial and temporal representations of historical droughts. Several states, including New York, Colorado, Idaho, and Utah, use the Palmer Index as one part of their drought monitoring systems.

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There are considerable limitations when using the Palmer Index, and these are described in detail by Alley (7) and Karl and Knight (6). Drawbacks of the Palmer Index include: • The values quantifying the intensity of drought and signaling the beginning and end of a drought or wet spell were arbitrarily selected based on Palmer’s study of central Iowa and western Kansas and have little scientific meaning. • The Palmer Index is sensitive to the AWC of a soil type. Thus, applying the index for a climate division may be too general. • The two soil layers within the water balance computations are simplified and may not be accurately representative of a location. • Snowfall, snow cover, and frozen ground are not included in the index. All precipitation is treated as rain, so that the timing of PDSI or PHDI values may be inaccurate in the winter and spring months in regions where snow occurs. • The natural lag between when precipitation falls and the resulting runoff is not considered. In addition, no runoff is allowed to take place in the model until the water capacity of the surface and subsurface soil layers is full, leading to an underestimation of runoff. • Potential evapotranspiration is estimated using the Thornthwaite method. This technique has wide acceptance, but it is still only an approximation. Several other researchers have presented additional limitations of the Palmer Index. McKee et al. (4) suggested that the PDSI is designed for agriculture but does not accurately represent the hydrological impacts resulting from longer droughts. Also, the Palmer Index is applied within the United States but has little acceptance elsewhere (9). One explanation for this is provided by Smith et al. (10), who suggested that it does not do well in regions where there are extremes in the variability of rainfall or runoff. Examples in Australia and South Africa were given. Another weakness in the Palmer Index is that the ‘‘extreme’’ and ‘‘severe’’ classifications of drought occur with a greater frequency in some parts of the country than in others (1). ‘‘Extreme’’ droughts in the Great Plains occur with a frequency greater than 10%. This limits the accuracy of comparing the intensity of droughts between two regions and makes planning response actions based on a certain intensity more difficult. CROP MOISTURE INDEX (CMI) Description: A Palmer derivative, the CMI reflects moisture supply in the short term across major crop-producing regions and is not intended to assess long-term droughts. Pros: Identifies potential agricultural droughts. Developed by: W.C. Palmer, 1968. Weekly maps: http://www.cpc.ncep.noaa.gov/products/ analysis monitoring/regional monitoring/cmi.gif.

The Crop Moisture Index (CMI) uses a meteorological approach to monitor week-to-week crop conditions. It was developed by Palmer (11) from procedures within the calculation of the PDSI. Whereas the PDSI monitors long-term meteorological wet and dry spells, the CMI was designed to evaluate short-term moisture conditions across major crop-producing regions. It is based on the mean temperature and total precipitation for each week within a climate division, as well as the CMI value from the previous week. The CMI responds rapidly to changing conditions, and it is weighted by location and time so that maps, which commonly display the weekly CMI across the United States, can be used to compare moisture conditions at different locations. Weekly maps of the CMI are available as part of the USDA/JAWF Weekly Weather and Crop Bulletin (http://www.usda.gov/oce/waob/jawf/wwcb.html). Because it is designed to monitor short-term moisture conditions affecting a developing crop, the CMI is not a good long-term drought monitoring tool. The CMI’s rapid response to changing short-term conditions may provide misleading information about long-term conditions. For example, a beneficial rainfall during a drought may allow the CMI value to indicate adequate moisture conditions, while the long-term drought at that location persists. Another characteristic of the CMI that limits its use as a long-term drought monitoring tool is that the CMI typically begins and ends each growing season near zero. This limitation prevents the CMI from being used to monitor moisture conditions outside the general growing season, especially in droughts that extend over several years. The CMI also may not be applicable during seed germination at the beginning of a specific crop’s growing season. SURFACE WATER SUPPLY INDEX (SWSI; PRONOUNCED ‘‘SWAZEE’’) Description: The SWSI is designed to complement the Palmer in the state of Colorado, where mountain snowpack is a key element of water supply; calculated by river basin, based on snowpack, streamflow, precipitation, and reservoir storage. Pros: Represents water supply conditions unique to each basin. Cons: Changing a data collection station or water management requires that new algorithms be calculated, and the index is unique to each basin, which limits interbasin comparisons. Developed by: Shafer and Dezman, 1982. The Surface Water Supply Index (SWSI) was developed by Shafer and Dezman (12) to complement the Palmer Index for moisture conditions across the state of Colorado. The Palmer Index is basically a soil moisture algorithm calibrated for relatively homogeneous regions, but it is not designed for large topographic variations across a region and it does not account for snow accumulation and subsequent runoff. Shafer and Dezman designed the SWSI to be an indicator of surface water conditions and described the index as ‘‘mountain

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water dependent’’, in which mountain snowpack is a major component. The objective of the SWSI was to incorporate both hydrological and climatological features into a single index value resembling the Palmer Index for each major river basin in the state of Colorado (Shafer and Dezman 1982). These values would be standardized to allow comparisons between basins. Four inputs are required within the SWSI: snowpack, streamflow, precipitation, and reservoir storage. Because it is dependent on the season, the SWSI is computed with only snowpack, precipitation, and reservoir storage in the winter. During the summer months, streamflow replaces snowpack as a component within the SWSI equation. The procedure to determine the SWSI for a particular basin is as follows: monthly data are collected and summed for all the precipitation stations, reservoirs, and snowpack/streamflow measuring stations over the basin. Each summed component is normalized using a frequency analysis gathered from a long-term data set. The probability of non-exceedence—the probability that subsequent sums of that component will not be greater than the current sum—is determined for each component based on the frequency analysis. This allows comparisons of the probabilities to be made between the components. Each component has a weight assigned to it depending on its typical contribution to the surface water within that basin, and these weighted components are summed to determine a SWSI value representing the entire basin. Like the Palmer Index, the SWSI is centered on zero and has a range between −4.2 and +4.2. The SWSI has been used, along with the Palmer Index, to trigger the activation and deactivation of the Colorado Drought Plan. One of its advantages is that it is simple to calculate and gives a representative measurement of surface water supplies across the state. It has been modified and applied in other western states as well. These states include Oregon, Montana, Idaho, and Utah. Monthly SWSI maps for Montana are available from the Montana Natural Resource Information System (http://nris.state.mt.us/wis/SWSInteractive/). Several characteristics of the SWSI limit its application. Because the SWSI calculation is unique to each basin or region, it is difficult to compare SWSI values between basins or regions (13). Within a particular basin or region, discontinuing any station means that new stations need to be added to the system and new frequency distributions need to be determined for that component. Additional changes in the water management within a basin, such as flow diversions or new reservoirs, mean that the entire SWSI algorithm for that basin needs to be redeveloped to account for changes in the weight of each component. Thus, it is difficult to maintain a homogeneous time series of the index (8). Extreme events also cause a problem if the events are beyond the historical time series, and the index will need to be reevaluated to include these events within the frequency distribution of a basin component.

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RECLAMATION DROUGHT INDEX Description: Like the SWSI, the RDI is calculated at the river basin level, incorporating temperature as well as precipitation, snowpack, streamflow, and reservoir levels as input. Who uses it: The Bureau of Reclamation; the State of Oklahoma as part of their drought plan. Pros: By including a temperature component, it also accounts for evaporation. Cons: Because the index is unique to each river basin, interbasin comparisons are limited. Developed by: The Bureau of Reclamation, as a trigger to release drought emergency relief funds. RDI Classifications 4.0 or more 1.5 to 4.0 1 to 1.5 0 to −1.5 −1.5 to −4.0 −4.0 or less

extremely wet moderately wet normal to mild wetness normal to mild drought moderate drought extreme drought

The Reclamation Drought Index (RDI) was recently developed as a tool for defining drought severity and duration, and for predicting the onset and end of periods of drought. The impetus to devise the RDI came from the Reclamation States Drought Assistance Act of 1988, which allows states to seek assistance from the Bureau of Reclamation to mitigate the effects of drought. Like the SWSI, the RDI is calculated at a river basin level, and it incorporates the supply components of precipitation, snowpack, streamflow, and reservoir levels. The RDI differs from the SWSI in that it builds a temperature-based demand component and a duration into the index. The RDI is adaptable to each particular region and its main strength is its ability to account for both climate and water supply factors. Oklahoma has developed its own version of the RDI and plans to use the index as one tool within the monitoring system designated in the state’s drought plan. The RDI values and severity designations are similar to the SPI, PDSI, and SWSI. DECILES Description: Groups monthly precipitation occurrences into deciles so that, by definition, ‘‘much lower than normal’’ weather cannot occur more often than 20% of the time. Who Uses It: Australia. Decile Classifications Deciles 1–2: lowest 20% Deciles 3–4: next lowest 20% deciles 5–6: middle 20% deciles 7–8: next highest 20% deciles 9–10: highest 20%

much below normal below normal near normal above normal much above normal

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THE EARTH OBSERVING SYSTEM: AQUA

Pros: Provides an accurate statistical measurement of precipitation. Cons: Accurate calculations require a long climatic data record. Developed by: Gibbs and Maher, 1967. Arranging monthly precipitation data into deciles is another drought-monitoring technique. It was developed by Gibbs and Maher (14) to avoid some of the weaknesses within the ‘‘percent of normal’’ approach. The technique they developed divided the distribution of occurrences over a long-term precipitation record into tenths of the distribution. They called each of these categories a decile. The first decile is the rainfall amount not exceeded by the lowest 10% of the precipitation occurrences. The second decile is the precipitation amount not exceeded by the lowest 20% of occurrences. These deciles continue until the rainfall amount identified by the tenth decile is the largest precipitation amount within the long-term record. By definition, the fifth decile is the median, and it is the precipitation amount not exceeded by 50% of the occurrences over the period of record. The deciles are grouped into five classifications. The decile method was selected as the meteorological measurement of drought within the Australian Drought Watch System because it is relatively simple to calculate and requires less data and fewer assumptions than the Palmer Drought Severity Index (10). In this system, farmers and ranchers can only request government assistance if the drought is shown to be an event that occurs only once in 20–25 years (deciles 1 and 2 over a 100-year record) and has lasted longer than 12 months (15). This uniformity in drought classifications, unlike a system based on the percent of normal precipitation, has assisted Australian authorities in determining appropriate drought responses. One disadvantage of the decile system is that a long climatological record is needed to calculate the deciles accurately. BIBLIOGRAPHY 1. Willeke, G., Hosking, J.R.M., Wallis, J.R., and Guttman, N.B. (1994). The National Drought Atlas. Institute for Water Resources Report 94–NDS–4, U.S. Army Corps of Engineers. 2. McKee, T.B., Doesken, N.J., and Kleist, J. (1993). The relationship of drought frequency and duration to time scales. Preprints, 8th Conference on Applied Climatology, pp. 179–184. January 17–22, Anaheim, CA. 3. Edwards, D.C. and McKee, T.B. (1997). Characteristics of 20th century drought in the United States at multiple time scales. Climatology Report Number 97–2, Colorado State University, Fort Collins, CO. 4. McKee, T.B., Doesken, N.J., and Kleist, J. (1995). Drought monitoring with multiple time scales. Preprints, 9th Conference on Applied Climatology, pp. 233–236. January 15–20, Dallas, TX. 5. Palmer, W.C. (1965). Meteorological drought. Research Paper No. 45, U.S. Department of Commerce Weather Bureau, Washington, DC. 6. Karl, T.R. and Knight, R.W. (1985). Atlas of Monthly Palmer Hydrological Drought Indices (1931–1983) for the Contiguous United States. Historical Climatology Series 3–7, National Climatic Data Center, Asheville, North Carolina.

7. Alley, W.M. (1984). The palmer drought severity index: limitations and assumptions. Journal of Climate and Applied Meteorology 23: 1100–1109. 8. Heddinghaus, T.R. and Sabol, P. (1991). A review of the palmer drought severity index and where do we go from here? In: Proc. 7th Conf. on Applied Climatology, pp. 242–246. American Meteorological Society, Boston. 9. Kogan, F.N. (1995). Droughts of the late 1980s in the United States as derived from NOAA polar-orbiting satellite data. Bulletin of the American Meteorological Society 76(5): 655–668. 10. Smith, D.I., Hutchinson, M.F., and McArthur, R.J. (1993). Australian climatic and agricultural drought: Payments and policy. Drought Network News 5(3): 11–12. 11. Palmer, W.C. (1968). Keeping track of crop moisture conditions, nationwide: The new Crop Moisture Index. Weatherwise 21: 156–161. 12. Shafer, B.A. and Dezman, L.E. (1982). Development of a Surface Water Supply Index (SWSI) to assess the severity of drought conditions in snowpack runoff areas. In: Proceedings of the Western Snow Conference. Colorado State University, Fort Collins, CO, pp. 164–175. 13. Doesken, N.J., McKee, T.B., and Kleist, J. (1991). Development of a surface water supply index for the western United States. Climatology Report Number 91–3, Colorado State University, Fort Collins, CO. 14. Gibbs, W.J. and Maher, J.V. (1967). Rainfall deciles as drought indicators. Bureau of Meteorology Bulletin No. 48, Commonwealth of Australia, Melbourne. 15. White, D.H. and O’Meagher, B. (1995). Coping with exceptional droughts in Australia. Drought Network News 7(2): 13–17.

READING LIST Gommes, R. and Petrassi, F. (1994). Rainfall variability and drought in Sub-Saharan Africa since 1960. Agrometeorology Series Working Paper No. 9, Food and Agriculture Organization, Rome, Italy. Le Hou´erou, H.N., Popov, G.F., and See, L. (1993). Agrobioclimatic classification of Africa. Agrometeorology Series Working Paper No. 6, Food and Agriculture Organization, Rome, Italy. Wilhite, D.A. (1995). Developing a precipitation-based index to assess climatic conditions across Nebraska. Final report submitted to the Natural Resources Commission, Lincoln, NE. Wilhite, D.A. and Glantz, M.H. (1985). Understanding the drought phenomenon: The role of definitions. Water International 10(3): 111–120.

THE EARTH OBSERVING SYSTEM: AQUA NASA—Goddard Space Flight Center

EARTH SYSTEM SCIENCE Beginning in the 1960s, NASA pioneered the study of the atmosphere from the unique perspective of space This article is a US Government work and, as such, is in the public domain in the United States of America.

THE EARTH OBSERVING SYSTEM: AQUA

with the launch of its Television Infrared Observation Satellite (TIROS-1). Thanks to new satellite and computer technologies, it is now possible to study the Earth as a global system. Through their research, scientists are better understanding and improving their forecasting of short-term weather phenomena. Long-term weather and climate prediction is a greater challenge that requires the collection of better data over longer periods. Since climate changes occur over vast ranges of space and time, their causes and effects are often difficult to measure and understand. Scientists must obtain long-term data if they are to reach a full understanding of the interactions among the Earth’s physical and biological systems. NASA’s Earth Observing System (EOS) will help us to understand the complex links among air, land, water and life within the Earth system. WHAT IS AQUA? NASA’s commitment to studying the Earth as a global system continues with the Aqua spacecraft (originally called EOS PM-1), representing a key contribution by NASA to the U.S. Global Change Research Program. Aqua carries six state-of-the-art instruments to observe the Earth’s oceans, atmosphere, land, ice and snow covers, and vegetation, providing high measurement accuracy, spatial detail, and temporal frequency. This comprehensive approach to data collection enables scientists to study the interactions among the four spheres of the Earth system—the oceans, land, atmosphere, and biosphere. Aqua, Latin for ‘‘water,’’ is named for the large amount of information that the Aqua spacecraft will collect about the Earth’s water cycle. In particular, the Aqua data will include information on water vapor and clouds in the atmosphere, precipitation from the atmosphere, soil wetness on the land, glacial ice on the land, sea ice in the oceans, snow cover on both land and sea ice, and surface waters throughout the world’s oceans, bays, and lakes. Such information will help scientists improve the quantification of the global water cycle and examine such issues as whether or not the cycling of water might be accelerating. In addition to information about the water cycle, Aqua also provides information on many additional elements of the Earth system. For instance, Aqua enables studies of the fluxes of radiation from the Sun and from the Earth that combine to constitute the Earth’s radiation

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balance. It also enables studies of small particles in the atmosphere termed ‘‘aerosols’’ and such trace gases in the atmosphere as ozone, carbon monoxide, and methane. The trace gases each have a potential contribution to global warming, whereas the aerosols are more likely to have a cooling effect. Aqua also provides observations on vegetation cover on the land, phytoplankton and dissolved organic matter in the oceans, and the temperatures of the air, land, and water. All of these measurements have the potential to contribute to improved understanding of the changes occurring in the global climate and the role of the interactions among the various elements of the climate system. One of the most exciting of the potential practical benefits likely to derive from the Aqua data is improved weather forecasting. Aqua carries a sophisticated sounding system that allows determination of atmospheric temperatures around the world to an accuracy of 1◦ Celsius in 1-km-thick layers throughout the troposphere, the lowest portion of the atmosphere. The troposphere extends to an altitude of about 10–15 km, depending on location, and contains most of the global cloud cover. The anticipated 1◦ Celsius accuracy far exceeds current accuracies from satellite observations and, in conjunction with the moisture profiles also obtainable from the Aqua sounding system, offers the potential of improved weather fore-casting. NASA is working with the U.S. National Oceanic and Atmospheric Administration and the European Centre for MediumRange Weather Forecasts to facilitate the incorporation of the Aqua data in their weather forecasting efforts. INTERNATIONAL COLLABORATION Aqua is a joint project of the United States, Japan, and Brazil. THE SPACECRAFT The spacecraft was designed and built by TRW in Redondo Beach, California. Aqua is based on TRW’s modular, standardized AB1200 common spacecraft bus. This design features common subsystems scalable to the mission-specific needs of Aqua as well as future missions. Instrument payloads can be attached on a ‘‘mix and match’’ basis without changing the overall design or subsystem support requirements. THE INSTRUMENTS The Atmospheric Infrared Sounder (AIRS), built by BAE Systems, was provided by NASA’s Jet Propulsion Laboratory in Pasadena, California. AIRS is the highlighted instrument in the AIRS/AMSU-A/HSB triplet centered on measuring humidity, temperature, cloud properties, and the amounts of greenhouse gases throughout the atmosphere. AIRS/AMSU-A/HSB will improve weather forecasting, establish the connection between severe weather and climate change, examine whether the global water cycle is accelerating, and detect the effects of greenhouse gases.

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THE EARTH OBSERVING SYSTEM: AQUA

The Advanced Microwave Scanning Radiometer for EOS (AMSR-E), built by Mitsubishi Electronics Corporation, was provided by Japan’s National Space Development Agency. AMSR-E measures precipitation rate, cloud water, water vapor, sea surface winds, sea surface temperature, ice, snow, and soil moisture. The Advanced Microwave Sounding Unit (AMSU-A), built by Aerojet and provided by NASA’s Goddard Space Flight Center (GSFC) in Greenbelt, Maryland, obtains temperature profiles in the upper atmosphere (especially the stratosphere) and will provide a cloud-filtering capability for tropospheric temperature observations. The EOS AMSU-A is part of the closely coupled AIRS/AMSUA/HSB triplet. The Clouds and the Earth’s Radiant Energy System (CERES), built by TRW, was provided by NASA’s Langley Research Center in Hampton, Virginia. This instrument measures the Earth’s total thermal radiation budget, and, in combination with Moderate Resolution Imaging Spectro-radiometer (MODIS) data, provides detailed information about clouds. The first CERES instrument was launched on the Tropical Rainfall Measuring Mission (TRMM) satellite in November 1997; the second and third CERES instruments were launched on the Terra satellite in December 1999; and the fourth and fifth CERES instruments was on board the Aqua satellite launched in May 2002. The pairs of CERES on both Terra and Aqua satellites allow coincident measurements by one CERES scanning in lines perpendicular to the path of the satellite and by the other CERES scanning in lines at various angles with respect to the satellite’s path. The Humidity Sounder for Brazil (HSB), built by MatraMarconi, was provided by Brazil’s Instituto Nacional de Pesquisas Espaciais, the Brazilian Institute for Space Research. The HSB obtains humidity profiles throughout the atmosphere. The HSB is the instrument in the AIRS/AMSU-A/HSB suite that allows humidity measurements even under conditions of heavy cloudiness and haze. MODIS, built by Raytheon Santa Barbara Remote Sensing, was provided by GSFC. MODIS is a 36band spectroradiometer measuring visible and infrared radiation and obtaining data that is used to derive products ranging from vegetation, land surface cover, and ocean chlorophyll fluorescence, to cloud and aerosol properties, fire occurrence, snow cover on the land, and sea ice cover on the oceans. The first MODIS instrument was launched on board the Terra satellite. Aqua was launched in May 2002 aboard a Delta 792010L launch vehicle from Vandenberg Air Force Base, California. The stowed spacecraft is 8.8 ft (2.68 m) × 8.2 ft (2.49 m) × 21.3 ft (6.49 m). Deployed, Aqua is 15.8 ft (4.81 m) × 54.8 ft (16.70 m) × 26.4 ft (8.04 m). The spacecraft, at launch, weighed 6784 lbs with a full propellant load of 508 lbs and is powered by 4.6 kilowatts of electric power from its solar array. Aqua was launched into a circular 680-km orbit. Over a period of days after separation from the launch vehicle, it was commanded by the ground to raise its orbit to the prescribed 705-km (438-mile) orbit. This was necessary in order to allow for proper phasing of Aqua with other

spacecraft in orbit and the polar ground stations used for communications. The spacecraft was ultimately be positioned in a near-polar (98◦ ) orbit around the Earth in synchronization with the Sun, with its path over the ground ascending across the equator at the same local time every day, approximately 1:30 p.m. The early afternoon observation time contrasts with the 10:30–10:45 a.m. equatorial crossing time (descending in this case) of the Terra satellite. The two daytime crossing times account for why the Terra and Aqua satellites were originally named ‘‘EOS AM’’ and ‘‘EOS PM,’’ respectively. The combination of morning and afternoon observations allows studies concerning the diurnal variability of many of the parameters discussed above. MANAGEMENT Overall management of the Aqua mission is located at GSFC, which is managing the integration and testing of the spacecraft. The Aqua data is processed, archived, and distributed using distributed components of the Earth Observing System Data and Information System (EOSDIS). EOSDIS also provides the mission operations systems that perform the functions of command and control of the spacecraft and the instruments. NASA’s Kennedy Space Center is responsible for the launch operations, including Boeing’s Delta launch vehicle and the prelaunch integrated processing facility. The U.S. Air Force is responsible for all range-related matters. GSFC manages EOS for NASA’s Earth Science Enterprise (ESE), headquartered in Washington, DC. DATA PROCESSING AND DISTRIBUTION Aqua provides a major part of a 15-year environmental dataset focusing on global change. The Aqua instruments produce more than 750 gigabytes of data per day, which is equivalent to 75 personal computer hard disks at 10 gigabytes each per day. This massive amount of information is handled using EOSDIS, in addition to its present accumulation of nearly 3000 gigabytes per day. EOSDIS provides the high-performance computing resources needed to process, store, and rapidly transmit terabytes (thousands of gigabytes) of the incoming data every day. EOSDIS has several distributed sites that perform these functions: Distributed Active Archive Centers (DAACs) that process, store and distribute the data, and Science Investigator-led Processing Systems that process the data and send them to the DAACs for storage and distribution. EOSDIS uses an ‘‘open’’ architecture to allow insertion of new technology while enabling the system to support the changing mission and science needs throughout the EOS Program. GOALS AND OBJECTIVES NASA’s ESE identified several high-priority measurements that EOS should make to facilitate a better understanding of the components of the Earth system—the atmosphere, the land, the oceans, the polar ice caps,

ENTROPY THEORY FOR HYDROLOGIC MODELING

and the global energy budget. The specific objectives of Aqua include: • producing high-spectral resolution obtaining 1 K/ 1 km global root-mean-square temperature profile accuracy in the troposphere by 1 year after launch; • extending the improved TRMM rainfall characterization to the extra tropics; • producing global sea surface temperature daily maps under nearly all sky conditions for a minimum of 1 year; • producing large-scale global soil moisture distribution for regions with low vegetation; • producing calibrated global observations of the Earth’s continents and ocean surfaces 150 days after the mission is declared operational; • capturing and documenting three seasonal cycles of terrestrial and marine ecosystems and atmospheric and cloud properties; • producing three sets of seasonal/annual Earth radiation budget records; • producing improved measurements of the diurnal cycle of radiation by combining Aqua measurements with Terra measurements for months of overlap; • producing combined cloud property and radiation balance data to allow improved studies of the role of clouds in the climate system; and, • capturing, processing, archiving, and distributing Aqua data products, by 150 days after the mission is declared operational. A NEW PERSPECTIVE Complemented by Terra, aircraft and ground-based measurements, Aqua data enable scientists to distinguish between natural and human-induced changes. The EOS series of spacecraft are the cornerstone of NASA’s ESE, a

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long-term research effort to study the Earth as a global environment. More information on EOS and the science related to it can be found at the EOS Project Science Office website at http://eospso.gsfc.nasa.gov and at the Earth Observatory website at http://earthobservatory.nasa.gov. Further information on Aqua can be found at http://aqua.nasa.gov

ENTROPY THEORY FOR HYDROLOGIC MODELING VIJAY P. SINGH Louisiana State University Baton Rouge, Louisiana

INTRODUCTION Entropy theory has recently been employed in a broad range of applications in hydrology, and new applications continue to unfold. This paper revisits entropy theory and its application to hydrologic modeling. Hydrologic systems are inherently spatial and complex, and our understanding of these systems is less than complete. Many of the systems are either fully stochastic or part stochastic and part deterministic. Their stochastic nature can be attributed to randomness in one or more of the following components that constitute them: (1) system structure (geometry), (2) system dynamics, (3) forcing functions (sources and sinks), and (4) initial and boundary conditions. As a result, a stochastic description of these systems is needed, and entropy theory enables the development of such a description. Engineering decisions concerning hydrologic systems are frequently made with less than adequate information. Such decisions may often be based on experience, professional judgment, rules of thumb, crude analyses, safety factors, or probabilistic methods. Usually, decision-making under uncertainty tends to be relatively conservative. Quite often, sufficient data are not available to describe the random behavior of such systems. Although probabilistic methods allow for a more explicit and quantitative accounting of uncertainty, their major difficulty occurs due to the lack of sufficient or complete data. Small sample sizes and limited information render estimation of probability distributions of system variables by conventional methods difficult. This problem can be alleviated by using entropy theory that enables determining the least biased probability distributions based on limited knowledge and data. Where the shortage of data is widely rampant, as is normally the case in many countries, entropy theory is particularly appealing. Since the development of entropy theory by Shannon in the late 1940s and of the principle of maximum entropy (POME) by Jaynes in the late 1950s, there has been a proliferation in applications of entropy. The real impetus to entropy-based modeling in hydrology was, however, provided in the early 1970s, a great variety of entropybased applications have since been reported, and new applications continue to unfold. This article aims to revisit

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ENTROPY THEORY FOR HYDROLOGIC MODELING

entropy theory and to underscore its usefulness for both modeling and decision-making in hydrology. ENTROPY THEORY Entropy theory is comprised of three (1) Shannon entropy, (2) the principle entropy, and (3) the principle of minimum Before discussing these parts, it will be discuss briefly the meaning of entropy.

main parts: of maximum cross entropy. instructive to

Meaning of Entropy Entropy originated in physics. It is an extensive property like mass, energy, volume, momentum, charge, or number of atoms of chemical species, but unlike these quantities, it does not obey a conservation law. The entropy of a system is an extensive property, so the total entropy of the system equals the sum of the entropies of individual parts. The most probable distribution of energy in a system is the one that corresponds to the maximum entropy of the system. This occurs under the condition of dynamic equilibrium. During evolution toward a stationary state, the rate of entropy production per unit mass should be minimum, compatible with external constraints. This is the Prigogin principle. In thermodynamics, entropy is decomposed into two parts: (1) entropy exchanged between the system and its surroundings and (2) entropy produced in the system. According to the second law of thermodynamics, the entropy of a closed and isolated system always tends to increase. In hydraulics, entropy is a measure of the amount of irrecoverable flow energy that the hydraulic system expands to overcome friction. The system converts a portion of its mechanical energy to heat energy which then is dissipated to the external environment. Thus, the process equation in hydraulics expressing the energy (or head) loss, it can be argued, originates in the entropy concept. Entropy has been employed in thermodynamics as a measure of the degree of ignorance about the true state of a system. If there were no energy loss in a hydraulic system, the system would be orderly and organized. The energy loss and its causes make the system disorderly and chaotic. Thus, entropy can be interpreted as a measure of the amount of chaos within a system. Algebraically, it is proportional to the logarithm of the probability of the state of the system. The constant of proportionality is the Boltzmann constant, and this defines Boltzmann entropy. Shannon Entropy Shannon (1) developed the entropy theory for expressing information or uncertainty. To understand the informational aspect of entropy, we perform an experiment on a random variable X. There may be n possible outcomes x1 , x2 , . . . , xn , whose probabilities are p1 , p2 , . . . , pn ; P(X = x1 ) = p1 , P(X = x2 ) = p2 , . . . , P(X = xn ) = pn . These outcomes can be described by P(X) = ( p1 , p2 , . . . , pn );

n 

pi = 1; pi ≥ 0, i = 1, 2, . . . , n

i

(1)

If this experiment is repeated, the same outcome is not likely, implying that there is uncertainty as to the outcome of the experiment. Based on one’s knowledge about the outcomes, the uncertainty can be more or less. For example, the total number of outcomes is a piece of information, and the number of those outcomes of nonzero probability is another piece of information. The probability distribution of the outcomes, if known, provides a certain amount of information. Shannon (1) defined a quantitative measure of uncertainty associated with a probability distribution or the information content of the distribution in terms of entropy, H(P) or H(X), called Shannon entropy or informational entropy as H(X) = H(P) = −

n 

pi ln pi = E[− ln p]

(2)

i=1

If the random variable X is continuous, then Shannon entropy is expressed as


∞

H(X) = −

f (x) ln[ f (x)] dx


=−

0

ln[ f (x)] dF(x) = E[− ln f (x)]

(3)

where f (x) is the probability density function (PDF) of X, F(x) is the cumulative probability distribution function of X, and E [.] is the expectation of [.]. Thus, entropy is a measure of the amount of uncertainty represented by the probability distribution and is a measure of the amount of chaos or of the lack of information about a system. If complete information is available, entropy = 0; otherwise, it is greater than zero. The uncertainty can be quantified using entropy by taking into account all different kinds of available information. Shannon entropy is the weighted Boltzmann entropy. Principle of Maximum Entropy In search of an appropriate probability distribution for a given random variable, entropy should be maximized. In practice, however, it is common that some information about the random variable is available. The chosen probability distribution should then be consistent with the given information. There can be more than one distribution consistent with the given information. From all such distributions, we should choose the distribution that has the highest entropy. To that end, Jaynes (2) formulated the principle of maximum entropy (POME), a full account of it is presented in a treatise by Levine and Tribus (3). According to POME, the minimally prejudiced assignment of probabilities is that which maximizes entropy subject to the given information, that is, POME takes into account all of the given information and at the same time avoids considering of any information that is not given. If no information about the random variable is available, then all outcomes are equally likely, that is, pi = 1/n, i = 1, 2, 3, . . . , n. It can be shown that Shannon entropy is maximum in this case and may serve as an upper bound of entropy for all cases involving some

ENTROPY THEORY FOR HYDROLOGIC MODELING

information. In a more general case, let the information available about P or X be pi ≥ 0,

n 

pi = 1

(4)

where the cross entropy D is minimized. If no a priori distribution is available and if, according to Laplace’s principle of insufficient reason, Q is chosen as a uniform distribution U, then Eq. 7 takes the form

i=1

and

n 

D(P, U) = gr (xi )pi = ar

r = 1, 2, . . . , m

(5)

i=1

where m is the number of constraints, m + 1 ≤ n and gr is the rth constraint. Equations 4 and 5 are not sufficient to determine P uniquely. Therefore, there can be many distributions that satisfy Eqs. 4 and 5. According to POME, there will be only one distribution that will correspond to the maximum value of entropy, and this distribution can be determined using the method of Lagrange multipliers which will have the following form: pi = exp[−λ0 − λ1 g1 (xi ) − λ2 g2 (xi ) . . . − λm gm (xi )] i = 1, 2, . . . , n

(6)

where λi , i = 0, 1, 2, . . . , m, are Lagrange multipliers that are determined by using the information specified by Eqs. 4 and 5. Because the POME-based distribution is favored over those with less entropy among those that satisfy the given constraints, according to Shannon entropy as an information measure, entropy defines a kind of measure on the space of probability distributions. Intuitively, distributions of higher entropy represent more disorder, are smoother, are more probable, are less predictable, or assume less. The POME-based distribution is maximally noncommittal with regard to missing information and does not require invoking ergodic hypotheses. Principle of Minimum Cross Entropy According to Laplace’s principle of insufficient reason, all outcomes of an experiment should be considered equally likely, unless there is information to the contrary. On the basis of intuition, experience, or theory, a random variable may have an a priori probability distribution. Then, Shannon entropy is maximum when the probability distribution of the random variable is one which is as close to the a priori distribution as possible. This is called the principle of minimum cross entropy (POMCE) which minimizes Bayesian entropy (4). This is equivalent to maximizing Shannon entropy. To explain POMCE, let us suppose that we guess a probability distribution for a random variable x as Q = {q1 , q2 , . . . , qn } based on intuition, experience, or theory. This constitutes the prior information in the form of a prior distribution. To verify our guess, we take some observations X = (x1 , x2 , . . . , xn ) and compute some moments of the distribution. To derive the distribution P = {p1 , p2 , . . . , pn } of X, we take all the given information and make the distribution as near to our intuition and experience as possible. Thus, POMCE is expressed as D(P, Q) =

n  i=1

pi ln

pi qi

(7)

219

n  i=1

 n    pi pi ln pi ln pi = ln n 1/n 

(8)

i=1

Hence, minimizing D(P, U) is equivalent to maximizing Shannon entropy. Because D is a convex function, its local minimum is its global minimum. Thus, a posterior distribution P is obtained by combining a prior Q with the specified constraints. The distribution P minimizes the cross (or relative) entropy with respect to Q, defined by Eq. 7, where the entropy of Q is defined as in Eq. 2. Cross-entropy minimization results asymptotically from Bayes’ theorem. JOINT ENTROPY, CONDITIONAL ENTROPY, AND TRANSINFORMATION If there are two random variables X and Y whose probability distributions are P(x) = {p1 , p2 , . . . , pn } and Q(y) = {q1 , q2 , . . . , qn }, which are independent, then Shannon entropy of the joint distribution of X and Y is the sum of the entropies of the marginal distributions expressed as H(P, Q) = H(X, Y) = H(P) + H(Q) = H(X) + H(Y)

(9)

If the two random variables are dependent, then Shannon entropy of the joint distribution is the sum of the marginal entropy of one variable and the conditional entropy of the other variable conditioned on the realization of the first. Expressed algebraically, H(X, Y) = H(X) + H(Y | X)

(10)

where H(Y|X) is the conditional entropy of Y conditioned on X. The conditional entropy can be defined as H(X|Y) = −

m n  

p(xi , yj ) ln( p(xi | yj )

(11)

i=1 j=1

It is seen that if X and Y are independent, then Eq. 10 reduces to Eq. 9. Furthermore, the joint entropy of dependent X and Y will be less than or equal to the joint entropy of independent X and Y, that is, H(X, Y) ≤ H(X) + H(Y). The difference between these two entropies defines transinformation T(X, Y) or T(P, Q) expressed as T(X, Y) = H(X) + H(Y) − H(X, Y)

(12)

Transinformation represents the amount of information common to both X and Y. If X and Y are independent, T(X, Y) = 0. Substitution of Eq. 10 in Eq. 12 yields T(X, Y) = H(Y) − H(Y|X)

(13)

220

ENTROPY THEORY FOR HYDROLOGIC MODELING

Equation 13 states that stochastic dependence reduces the entropy of Y. ENTROPY AS A MODELING TOOL Although entropy theory has been applied in recent years to a variety of problems in hydrology, its potential as a decision-making tool has not been fully exploited. A brief discussion follows highlighting this potential. Fundamental to the concepts presented below is the need for probability distributions that can be derived by using entropy theory. Information Content of Data One frequently encounters a situation in which one exercises freedom of choice, evaluates uncertainty, or measures information gain or loss. The freedom of choice, uncertainty, disorder, information content, or information gain or loss has been variously measured by relative entropy, redundancy, and conditional and joint entropies employing conditional and joint probabilities. As an example, in the analysis of empirical data, the variance has often been interpreted as a measure of uncertainty and as revealing gain or loss in information. However, entropy is another measure of dispersion—an alternative to variance. This suggests that it is possible to determine the variance whenever it is possible to determine entropy measures, but the reverse is not necessarily true. However, variance is not the appropriate measure if the sample size is small. To measure correlation or dependence between any two variables, an informational coefficient of correlation r0 is defined as a function of transinformation, T0 , as r0 = [1 − exp(−2T0 )]0.5

(14)

The transinformation, given by Eq. 14, expresses the upper limit of common information between two variables and represents the level of dependence (or association) between the variables. It represents the upper limit of transferable information between the variables, and its measure is given by r0 . The ordinary correlation coefficient r measures the amount of information transferred between variables under specified assumptions, such as linearity and normality. An inference similar to that of the ordinary correlation coefficient, r, can be drawn by defining the amount (in percent) of transferred information by the ratio T/T0 , where T can be computed in terms of ordinary r. Criteria for Model Selection Usually, there are more models than one needs, and so a model has to be chosen. Akaike (5) formulated a criterion, called the Akaike information criterion (AIC), for selecting the best model from amongst several models as AIC = 2 log(maximized likelihood) + 2k

(15)

AIC provides a method of model identification and can be expressed as minus twice the logarithm of the maximum

likelihood plus twice the number of parameters used to find the best model. The maximum likelihood and entropy are uniquely related. When there are several models, the model that gives the minimum value of AIC should be selected. When the maximum likelihood is identical for two models, the model that has the smaller number of parameters should be selected, for that will lead to a smaller AIC and comply with the principle of parsimony. Hypothesis Testing Another important application of entropy theory is testing of hypotheses (6). By using Bayes’ theorem in logarithmic form, an evidence function is defined for comparing two hypotheses. The evidence in favor of a hypothesis over its competitor is the difference between the respective entropies of the competition and the hypothesis under test. Defining surprisal as the negative of the logarithm of the probability, the mean surprisal for a set of observations is expressed. Therefore, the evidence function for two hypotheses is obtained as the difference between the two values of the mean surprisal multiplied by the number of observations. Risk Assessment In common language, risk is the possibility of loss or injury and the degree of probability of such loss. Rational decision-making requires a clear and quantitative way of expressing risk. In general, risk cannot be avoided, and a choice has to be made between risks. There are different types of risk, such as business risk, social risk, economic risk, safety risk, investment risk, and occupational risk. To put risk in proper perspective, it is useful to clarify the distinction between risk, uncertainty, and hazard. The notion of risk involves both uncertainty and some kind of loss or damage. Uncertainty reflects the variability of our state of knowledge or state of confidence in a prior evaluation. Thus, risk is the sum of uncertainty plus damage. Hazard is commonly defined as a source of danger that involves a scenario identification (e.g., failure of a dam) and a measure of the consequence of that scenario or a measure of the ensuing damage. Risk encompasses the likelihood of converting that source into the actual delivery of loss, injury, or some form of damage. Thus, risk is the ratio of hazard to safeguards. By increasing safeguards, risk can be reduced, but it is never zero. Awareness of risk reduces risk, so awareness is part of safeguards. Qualitatively, risk is subjective and is relative to the observer. Risk involves the probability of a scenario and its consequence resulting from the occurrence of the scenario. Thus, one can say that risk is probability and consequence. Kaplan and Garrick (7) analyzed risk using entropy. HYDROLOGIC MODELING USING ENTROPY THEORY A historical perspective on entropy applications in environmental and water resources is given in Singh and Fiorentino (8) and Singh (9). Harmancioglu and Singh (10) discussed the use of entropy in water resources. A brief synopsis of entropy-based applications follows.

ENTROPY THEORY FOR HYDROLOGIC MODELING

Derivation of Probability Distributions Frequency distributions that satisfy the given information are often needed. Entropy theory is ideally suited to that end. POME has been employed to derive a variety of distributions; some have found wide applications in environmental and water resources. Many of these distributions have been summarized in Singh and Fiorentino (8) and by Singh (9). Let p(x) be the probability distribution of X that is to be determined. The information on X is available in terms of constraints given by Eq. 2. Then, the entropy-based distribution is given by Eq. 6. Substitution of Eq. 5 in Eq. 2 yields exp(λ0 ) = Z =

n 

 exp −

i=1

m 

spectra that have high-degree resolutions (13). The statistical characteristics that are used in stochastic model identification can also be estimated using MESA, thus permitting integration of spectral analysis and computations related to stochastic model development. Ulrych and Clayton (14) reviewed the principles of MESA and the closely related problem of autoregressive time series modeling. Shore (15) presented a comprehensive discussion of minimum cross-entropy spectral analysis. The relationship between spectrum W(f ) with frequency f of a stationary process x(t) and entropy H(f ) can be expressed as

 λj gj (xi )

H(f ) = (16)

j=1

where Z is called the partition function and λ0 is the zeroth Lagrange multiplier. The Lagrange parameters are obtained by differentiating Eq. 16 with respect to the Lagrange multipliers:

221

1 1 ln(2 w) + 2 4w




+w

ln[W(f )] df

(18)

w

where w is the frequency band. Equation 18 is maximized subject to the constraint equations given as autocorrelations until log m:
ρ(n) =

+w

W(f ) exp(i2 π fn
t) df ,

m ≤ n ≤ +m (19)

w

∂λ0 = −aj = E[ gj ], ∂λj

j = 1, 2, 3, . . . , m

where
t is the sampling time interval and i = (−1)1/2 . Maximization of Eq. 19 is equivalent to maximizing

∂ 2 λ0 = Var[gj ] ∂λ2j 2

∂ λ0 = Cov[gj , gk ] ∂λj ∂λk


(17)

3

∂ λ0 = −µ3 [gj ] ∂λ3j where E[.] is the expectation, Var [.] is the variance, Cov [.] is the covariance, and is the third moment about the centroid, all for gj . When there are no constraints, then POME yields a uniform distribution. As more constraints are introduced, the distribution becomes more peaked and possibly skewed. In this way, entropy reduces from a maximum for the uniform distribution to zero when the system is fully deterministic. Parameter Estimation It is desirable to estimate parameters of a distribution in terms of the given constraints. Entropy theory accomplishes precisely that. Singh (11) described POME-based estimation for a number of probability distributions used in hydrology and environmental and water resources. He also discussed a comparison of the POME-based method with the methods of moments, maximum likelihood estimation, and some others. The comparison shows that the POME-based method is comparable to some methods and is better than others.

H(f ) =

+w

ln[W( f )] df

(20)

w

which is known as Burg entropy. The spectrum W(f ) can be expressed in terms of the Fourier series as W(f ) =

∞ 1  ρ(n) exp(i2π nf
t) 2 w n=∞

(21)

Substituting Eq. 21 in Eq. 20 and maximizing lead to MESA. Jaynes (16) has shown that MESA and other methods of spectral analysis, such as Schuster, Blackman–Tukey, maximum likelihood, Bayesian, and autoregressive (AR, ARMA, or ARIMA) models are not in conflict and that AR models are a special case of MESA. Krstanovic and Singh (17,18) employed MESA for long-term stream flow forecasting. Krstanovic and Singh (19,20) extended the MESA method to develop a real-time flood forecasting model. Padmanabhan and Rao (21,22) applied MESA to analyze rainfall and river flow time series. Rao et al. (23) compared a number of spectral analysis methods with MESA and found that MESA is superior. Eilbert and Christensen (24) analyzed annual hydrologic forecasts for central California and found that dry years might be more predictable than wet years. Dalezios and Tyraskis (25) employed MESA to analyze multiple precipitation time series. Regional Precipitation Analysis and Forecasting

Entropy-Spectral Analysis for Flow Forecasting Maximum entropy spectral analysis (MESA) was introduced by Burg (12). It has several advantages over conventional spectral analysis methods. It has short and smooth

The Burg algorithm or MESA can be applied to identify and interpret multistation precipitation data sets and to explore spectral features that lead to a better understanding of rainfall structure in space and time (25).

222

ENTROPY THEORY FOR HYDROLOGIC MODELING

Then, multistation rainfall time series can be extrapolated to develop regional forecasting capabilities. Grouping of River Flow Regimes An objective grouping of flow regimes into regime types can be employed as a diagnostic tool for interpreting the results of climate models and flow sensitivity analyses. By minimizing an entropy-based objective function (such as minimum cross entropy), a hierarchical aggregation of monthly flow series into flow regime types can, therefore, be effectively performed, which will satisfy chosen discriminating criteria. Such an approach was developed by Krasovskaia (26) who applied it to a regional river flow sample for Scandinavia for two different formulations of discriminating criteria. Basin Geomorphology Entropy plays a fundamental role in characterizing landscape. Using entropy theory for the morphological analysis of river basin networks, Fiorentino et al. (27) found that the connection between entropy and the mean basin elevation is linearly related to basin entropy. Similarly, the relation between the fall in elevation from the source to the outlet of the main channel and the entropy of its drainage basin, it was found, is linear and so also was the case between the elevation of a node and the logarithm of its distance from the source. When a basin was ordered following the Horton–Strahler ordering scheme, a linear relation was found between drainage basin entropy and basin order. This relation can be characterized as a measure of basin network complexity. Basin entropy, it was also found, is linearly related to the logarithm of the magnitude of the basin network. This relation led to a nonlinear relation between the network diameter and magnitude where the exponent, it was found, is related to the fractal dimension of the drainage network. Design of Hydrologic Networks The purpose of measurement networks is to gather information in terms of data. Fundamental to evaluating these networks is the ability to determine if the networks are gathering the needed information optimally. Entropy theory is a natural tool for that determination. Krstanovic and Singh (28,29) employed the theory for space and time evaluation of rainfall networks in Louisiana. The decision whether to keep or to eliminate a rain-gauge was based entirely on the reduction or gain of information at that gauge. Yang and Burn (30) employed a measure of information flow, called directional information transfer index (DIT), between gauging stations in the network. The value of DIT varies from zero, where no information is transmitted and the stations are independent, to one where no information is lost and the stations are fully dependent. Between two stations of one pair, the station that has the higher DIT value should be retained because of its greater capability of inferring information at the other side. Rating Curve Moramarco and Singh (31) employed entropy theory to develop a method for reconstructing the discharge

hydrograph at a river section where only water level is monitored and discharge is recorded at another upstream section. The method, which is based on the assumption that lateral inflows are negligible, has two parameters linked to remotely observed discharge and permits, without using a flood routing procedure and without the need of a rating curve at a local site, relating the local river stage to the hydraulic condition at a remote upstream section. CONCLUDING REMARKS Entropy theory permits determining of the least biased probability distribution of a random variable, subject to the available information. It suggests whether or not the available information is adequate and, if not, then additional information should be sought. In this way, it brings the model, the modeler, and the decisionmaker closer. As an objective measure of information or uncertainty, entropy theory allows communicating with nature, as illustrated by its application to the design of data acquisition systems, the design of environmental and hydrologic networks, and the assessment of the reliability of these systems or networks. In a similar vein, it helps better understand the physics or science of natural systems, such as landscape evolution, geomorphology, and hydrodynamics. A wide variety of seemingly disparate or dissimilar problems can be meaningfully solved by using entropy. BIBLIOGRAPHY 1. Shannon, C.E. (1948). A mathematical theory of communications, I and II. Bell System Technical Journal 27: 379–443. 2. Jaynes, E.T. (1957). Information theory and statistical mechanics, I. Physical Review 106: 620–630. 3. Levine, R.D. and Tribus, M. (Eds.). (1978). The Maximum Entropy Formalism. The MIT Press, Cambridge, MA, p. 498. 4. Kullback, S. and Leibler, R.A. (1951). On information and sufficiency. Annals of Mathematical Statistics 22: 79–86. 5. Akaike, H. (1973). Information theory and an extension of the maximum likelihood principle. Proceedings, 2nd International Symposium on Information Theory. B.N. Petrov and F. Csaki. (Eds.). Publishing House of the Hungarian Academy of Sciences, Budapest, Hungary. 6. Tribus, M. (1969). Rational Description: Decision and Designs. Pergamon Press, New York. 7. Kaplan, S. and Garrick, B.J. (1981). On the quantitative definition of risk. Risk Analysis 1(1): 11–27. 8. Singh, V.P. and Fiorentino, M. (Eds.). (1992). Entropy and Energy Dissipation in Water Resources. Kluwer Academic Publishers, Dordrecht, the Netherlands. 9. Singh, V.P. (1998). Entropy-Based Parameter Estimation in Hydrology. Kluwer Academic Publishers, Boston, MA. 10. Harmancioglu, N.B. and Singh, V.P. (1998). Entropy in environmental and water resources. In: Encyclopedia of Hydrology and Water Resources. R.W. Hershey and R.W. Fairbridge (Eds.). Kluwer Academic Publishers, Boston, MA, pp. 225–241. 11. Singh, V.P. (1998). The use of entropy in hydrology and water resources. Hydrological Processes 11: 587–626.

EVAPORATION 12. Burg, J.P. (1975). Maximum entropy spectral analysis. Unpublished Ph.D. thesis, Stanford University, Palo Alto, CA, p. 123. 13. Fougere, P.F., Zawalick, E.J., and Radoski, H.R. (1976). Spontaneous life splitting in maximum entropy power spectrum analysis. Physics of the Earth and Planetary Interiors 12: 201–207. 14. Ulrych, T. and Clayton, R.W. (1976). Time series modeling and maximum entropy. Physics of the Earth and Planetary Interiors 12: 188–199. 15. Shore, J.E. (1979). Minimum cross-entropy spectral analysis. NRL Memorandum Report 3921, Naval Research Laboratory, Washington, DC. 16. Jaynes, E.T. (1982). On the rationale of maximum entropy methods. Proceedings of the IEEE 70: 939–952. 17. Krstanovic, P.F. and Singh, V.P. (1991). A univariate model for long-term streamflow forecasting: 1. Development. Stochastic Hydrology and Hydraulics 5: 173–188. 18. Krstanovic, P.F. and Singh, V.P. (1991). A univariate model for long-term streamflow forecasting: 2. Application. Stochastic Hydrology and Hydraulics 5: 189–205. 19. Krstanovic, P.F. and Singh, V.P. (1993). A real-time flood forecasting model based on maximum entropy spectral analysis: 1. Development. Water Resources Management 7: 109–129. 20. Krstanovic, P.F. and Singh, V.P. (1993). A real-time flood forecasting model based on maximum entropy spectral analysis: 2. Application. Water Resources Management 7: 131–151. 21. Padmanabhan, G. and Rao, A.R. (1986). Maximum entropy spectra of some rainfall and river flow time series from southern and central India. Theoretical and Applied Climatology 37: 63–73. 22. Padmanabhan, G. and Rao, A.R. (1988). Maximum entropy spectral analysis of hydrologic data. Water Resources Research 24(9): 1591–1533. 23. Rao, A.R., Padmanabhan, G., and Kashyap, R.L. (1980). Comparison of recently developed methods of spectral analysis. Proceedings, Third International Symposium on Stochastic Hydraulics. Tokyo, Japan, pp. 165–175. 24. Eilbert, R.F. and Christensen, R.O. (1983). Performance of the entropy minimax hydrological forecasts for California water years 1948–1977. Journal of Climate & Applied Meteorology 22: 1654–1657. 25. Dalezios, N.R. and Tyraskis, P.A. (1989). Maximum entropy spectra for regional precipitation analysis and forecasting. Journal of Hydrology 109: 25–42. 26. Krasovskaia, R. (1997). Entropy-based grouping of river flow regimes. Journal of Hydrology 202: 173–1191. 27. Fiorentino, M., Claps, P., and Singh, V.P. (1993). An entropybased morphological analysis of river basin networks. Water Resources Research 29(4): 1215–1224. 28. Krstanovic, P.F. and Singh, V.P. (1992). Evaluation of rainfall networks using entropy: 1. Theoretical development. Water Resources Management 6: 279–293. 29. Krstanovic, P.F. and Singh, V.P. (1992). Evaluation of rainfall networks using entropy: II. Application. Water Resources Management 6: 295–314. 30. Yang, Y. and Burn, D.H. (1984). An entropy approach to data collection network design. Journal of Hydrology 157: 307–324. 31. Moramarco, T. and Singh, V.P. (2001). Simple method for relating local stage and remote discharge. Journal of Hydrologic Engineering, ASCE 6(1): 78–81.

223

EVAPORATION THEODORE A. ENDRENY SUNY-ESF Syracuse, New York

Evaporation of water is a solar energy driven phase change from liquid to vapor that maintains the hydrologic cycle by transferring liquid water at the earth’s surface to water vapor in the atmosphere, where it may lift, condense, and precipitate to earth as liquid water. In this discussion, evaporation is defined to include the closely associated transpiration process as a subcategory. The coupled processes are often called as evapotranspiration, or ET, where transpiration focuses on the transport of liquid water through the plant roots, stem, and leaf prior to evaporation through the leave’s stomata. Further, when the discussion does not make a clear distinction between potential and actual evaporation, potential should be assumed. Potential evaporation refers to the amount of water that available energy and diffusion processes can transfer into atmospheric vapor, which is typically greater then the amount actually transferred due to limits on soil water volumes and resistances in the path.

PHYSICAL CONTROLS ON EVAPORATION Evaporation is effectively a two-step process that first, requires that water changes phase to a vapor state and second, requires that the vapor is transported by advection and/or diffusion into unsaturated air. The phase change alone does not completely satisfy the requirement for evaporation when dynamic equilibrium exists at the boundary between liquid and vapor and condensation of the vapor saturated air returns liquid water to the surface to maintain no net water loss. Transport, therefore, ensures removal of the water vapor and a net loss of heat and mass from the liquid surface. In the vapor state, evaporated water is invisible to the human eye, which detects wavelengths between 0.4 and 0.7 µm, but is detectable in other areas of the electromagnetic spectrum. As such, though clouds are derived from and contain evaporated water, they are not vapor and instead reveal condensed water droplets that geometrically scatter light. The first step in changing the phase of water from liquid to vapor requires an input of solar energy, which is stored at the surface and supports nighttime evaporation. Nearly 52% of solar energy absorbed at the earth’s surface is used for vaporization. This energy is called latent heat of vaporization, or λ, and is a function of water temperature. When the phase changes directly from frozen water to vapor, known as sublimation, a greater amount of energy is required. Phase changes require energy to separate the hydrogen bond based, attractive intermolecular forces holding water molecules in an organized pattern and close proximity. Latent heat is a name that suggests dormant, or invisible, and is used to indicate that, unlike the measurable effect of sensible heat on air temperature, the

224

EVAPORATION

solar energy used in evaporation remains ‘hidden’ from thermometer measurement. As an example, at the standard pressure of 1013.3 milibar (mb), heating liquid water from 0 to 100 ◦ C requires 4186.8 joules (J) of heat per kilogram of water per degree C temperature change and is detectable with a thermometer. In contrast, the latent heat of vaporization converting 100 ◦ C water into 100 ◦ C vapor does not have a measurable impact on water temperature. The latent heat, in megajoules per kilogram (MJ kg−1 ), required per ◦ C of water temperature, T, is given as λ = 2.501 − 0.002361 · T It is apparent that the majority of heat input to evaporate water, whether at 0 ◦ C or 99 ◦ C, is that for the phase change. A relatively constant 1350 Js−1 m−2 stream of solar energy entering the earth’s upper atmosphere provides energy to evaporate water just like a stove can boil and evaporate water. When 50% of this solar constant strikes the earth’s surface after atmospheric attenuation (e.g., scattering and reflection), it requires approximately 1 hour and 10 minutes to generate the latent heat needed to evaporate 1 kg of water at 20 ◦ C. Latent heat, though not detectable as affecting air temperature, is stored with the vapor in greater vibrational, rotational, and translational movement of vapor molecules. The removal of this vapor from the liquid, therefore, causes a measurable loss of energy and temperature from the remaining liquid, producing a cooling effect. A familiar example of this process is that wet skin, from sweat or a shower, cools faster in moving rather than still air because the wind speeds evaporation that takes heat from the body. Knowing that heat is used for evaporation, it is now clear that a covered pot will boil faster than a counterpart uncovered pot do to a lower loss of heat to the net evaporation of water. Heat stored in vapor is later released back into the environment when the vapor vibrational speeds slow, and it condenses into water, called latent heat of condensation. When vapor passes directly to solid frozen water, called deposition, a greater amount of heat stored in rotational, vibrational, and translational molecular movement is released into the environment. The second physical step in evaporation is transport by advection and/or diffusion, which provides the net movement of water molecules from the liquid water surface of soil, plants, or lakes to atmospheric vapor. Vapor and wind gradients exert the principal controls on removal of water vapor beyond the saturated layer of air that maintains condensation–evaporation equilibrium dynamics. Fundamental barriers to transport beyond the layer of dynamic equilibrium include stagnant air and saturated air above the evaporating surface, conditions readily created within and at the surface of soil and plant systems. Net evaporation therefore increases with steeper wind and saturation gradients, which are defined as the change in wind or saturation with distance above the evaporating surface. Work on fluid velocity and turbulence gradients in the mid-1900s by Prandtl and von Karman has been used to estimate momentum, sensible heat, and vapor transport from wind speed measurements.

Estimates of atmospheric wind and vapor conditions above the evaporating surface provide important data for estimating wind and saturation gradients and predicting barriers to vapor transport and net evaporation. Meteorological stations are frequently equipped with anemometers and thermometers at 2 meters (m) above the ground surface to help establish the wind and vapor gradients controlling evaporation. Wind profiles, it is assumed, begin with stagnant air at the noslip boundary, or zero-plane displacement height, and increase logarithmically. In a landscape broken by tree canopies, a 2 m wind measurement may be inadequate to represent observations. Research has shown that dynamic turbulence and eddies created by such forested heterogeneity result in increased wind and evaporation rates that exceed the estimated atmospheric potential. Based on Dalton’s work on individual pressures of multiple atmospheric gases summing to the observed atmospheric pressure, vapor is often reported as a partial pressure and can be derived from measurements of temperature. In the following equation, vapor pressure is reported in kilopascals (kPa) and temperature in ◦ C. Initially dry warm air can absorb more water than initially dry cold air before reaching saturation.  e = 0.6108 exp

17.27 T 237.3 + T



The dry-bulb temperature is used to estimate the total amount of vapor the air could absorb prior to saturation; the dew point temperature represents the temperature to which the air must cool for total saturation. When the dry-bulb and dew point temperatures are equal, the air is fully saturated. The ratio of actual to saturated vapor pressures is the relative humidity. The dew point temperature is estimated by using a psychrometer that measures the difference between dry-bulb and wet-bulb thermometers, called the wet-bulb depression, together with lookup tables relating wet-bulb depression to dew point temperature. The dry-bulb is a normal thermometer measuring air temperature, but a moist piece of cloth typically covers the wet-bulb, and evaporation of the water from the cloth causes the temperature to drop. Chilled mirrors hygrometers, hair hygrometers, and vapor pressure sensors are also used to detect the vapor content in the air. EVAPORATION MEASUREMENT AND ESTIMATION Evaporation is fundamental to both energy and water balances, yet despite the importance of evaporation to hydrologic assessment of the paths, quantities, and quality of water in the lithosphere, biosphere, and atmosphere, the complexity of the process has prevented easy or exact techniques for measuring and estimating it. The relative accuracy of yearly river basin evaporation estimates is high, as the estimate time frame and spatial area become smaller, but the simple application of energy and water balance models to solve for evaporation becomes less tractable. A variety of measurement and estimation techniques have been developed for these smaller scales,

EVAPORATION

such as hourly, daily, and monthly evaporation from reservoirs, farm fields, and single plants. In general, evaporative fluxes from the land surface are more difficult to measure than from open water, given that an immeasurable number of irregular, tiny and unique soil and leaf surfaces are involved in this phase change and that suction gradients draw water to this evaporating interface from unobserved reserves of unknown volume. Fluxes from open water, though relatively homogeneous, still provide challenges when subsurface inflows and outflows are poorly understood and significant and wind and water advected energy influencing evaporation is heterogeneously distributed. Hence, the numerous methods developed for estimating evaporation are categorized based on the type of surface, availability of water, and the importance of stored energy, water-advected energy, and air-advected energy. Actual evaporation can be measured by using a water balance approach, a turbulent-transfer approach, a potential evaporation approach, or a water quality approach. The water balance approach can function with measurements of a mass balance being kept for a water pan, such as a the Class-A Pan of the National Weather Service, a soil and plant system, such as in a weighing lysimeter, or of a small, enclosed atmosphere. Turbulent-transfer methods, which derive evaporation from estimates of momentum or heat flux, can provide estimates for larger heterogeneous areas, but assume that the air sampled by the field instrumentation used for the Bowen ratio or eddy-correlation method is well mixed to represent the upwind land area by a given fetch. Water quality methods include techniques that track concentrations of dissolved solids, which enrich when evaporation removes the water solvent, and isotope tracer studies that show heavier isotopes are enriched by preferred evaporation of lighter isotopes. Sap flow monitoring in trees provides another technique to measure the flux of water from the ground to the atmosphere. Mathematical estimates of evaporation rates have been approached by using an equally wide variety of techniques and include temperature-based, aerodynamicbased, radiation-based, and combination-based methods. Temperature-based methods, such as the monthly timestep Thornthwaite equation, use air temperature and length of day, as well as an assumed humidity, to compute the potential evaporation, and they have been adapted to suit several different climates and regions. Aerodynamic methods assume that solar radiation is not limiting and consider only wind speed and turbulence as controls on the transport of water vapor away from the surface. Radiation-based methods likewise assume that wind turbulence and eddies are not limiting and use measured incoming radiation and the latent heat of vaporization to compute evaporation flux. A popular form of this equation is the Priestly–Taylor, which increased it by a factor of 30% to account for added aerodynamic transfer. The combination method, known most extensively for the Penman–Monteith equation, uses air temperature, net radiation, wind speed, and relative humidity vapor gradients to derive minute by minute and daily evaporation rates.

225

HYDROLOGIC IMPACTS OF EVAPORATION Observations of terrestrial river discharge reveal that more water precipitates on land than evaporates from land and that more water evaporates from oceans than precipitates on oceans. Precipitation totals may vary from year to year, but evaporative demands are rather steady, which creates a greater relative fluctuation in river discharge than in precipitation. This is illustrated by considering a 20% decrease in annual precipitation from 100 to 80 cm, where 50 cm went to evaporation in both years, and discharge dropped by 40% from 50 to 30 cm. Evaporation and its impact on liquid and vapor water volumes and the partitioning of solar energy into latent and sensible heat create and maintain a range of climatic conditions, from microclimates on the scale of a tree canopy to macroclimates that describe the global distribution of plants. The volume of water evaporated from the ocean and land surface is greatest at the meteorological equator, or intertropical convergence zone (ITCZ), and smallest at the poles, which is the result of a similar longitudinal distribution of insolation intensity. Sinking Hadley cell air at the 30◦ latitude belts, which warms to absorb greater amounts of water vapor, is the cause of a belt of deserts in this region. The distribution of incoming solar radiation, which is greatest at the equator and smallest at the poles, is the driving force explaining the meridional (across lines of latitude) distribution of evaporation. Wind transport of this evaporated vapor from the equatorial region to the midlatitudes and poles, where latent energy is released to the atmosphere as sensible heat during condensation, is one of only a few processes that help to maintain the earth’s energy balance. Global water balance numbers reveal that a relatively small volume of evaporated vapor resides in the atmosphere. The earth’s atmosphere has a volume of 12,900 km3 , and contains just 0.001% of all global water. As evaporated vapor, it receives 71,000 km3 yr−1 from land, 1000 km3 yr−1 from lakes, and 505,000 km3 yr−1 from oceans, and this flux rate into its total volume equals a residence time of 8.2 days, or just over a week before evaporated water precipitates. The atmospheric vapor precipitates at a volumetric rate of 577 km3 yr−1 , of which 119,000 km3 yr−1 falls on land. Observation and estimation of river discharge at 47,000 km3 yr−1 was used to deduce the amount evaporated from land, which is 61%

Table 1. Continental Average Estimated Evaporation Continent

Area, km2

Antarctica Europe Asia South America North America Africa Australia Total Land

14,100,000 10,000,000 44,100,000 17,900,000 24,100,000 29,800,000 7,600,000 148,900,000

Evaporation, mm yr−1

Evaporation, %

28 375 420 946 403 582 420 480

17 57 60 60 62 84 94 64

226

EVAPOTRANSPIRATION Table 2. Average Estimated Evaporation Based on Studies of Precipitation and Runoff

Watershed by River Name

Continent [Nation(s)]

Brahmaputra Irrawaddy Yangtzekiang Amazon Orinoco Lena Mekong Yenesei Ganges Saint Lawrence Amur Congo Ob Mississippi La Plata Average a

Asia (Tibet/Bangladesh) Asia (Burma) Asia (China) South America (six nations) South America (Venezuela) Asia (Russia) Asia (China) Asia (Russia) Asia (China) North America (Canada, U.S.) Asia (Russia) Africa (7 nations) Asia (Russia) North America (U.S.) South America (five nations) –

Basin Size, km2

Evaporationa , %

589,000 431,000 1,970,000 7,180,000 1,086,000 2,430,000 795,000 2,599,000 1,073,000 1,030,000 1,843,000 3,822,000 2,950,000 3,224,000 2,650,000 2,224,800

35 40 50 53 54 54 57 58 58 67 68 75 76 79 80 60

Evaporation may include basin transfers because it was not directly measured but derived from precipitation and runoff measurements.

evaporating, and this differs from the continental average of 64% reported in Table 1. For the continental United States, which contains deserts in Arizona and rain forests in Washington, the average annual percentage of precipitation converted to evaporation is approximately 62%. This value is similar to global patterns but varies considerably from that measured on other continents (see Table 1) and larger watersheds (see Table 2). The agricultural impact of evaporation is both the cause of nutrient uptake and growth in plants, as well as the loss of soil water and plant stress. Maintenance of optimal water levels in the soil, called field capacity, when gravitational water has drained, is the goal of many irrigation projects. If irrigation causes evaporation to exceed local precipitation, then salts will be drawn to the soil surface, which often creates osmotic gradients at the root interface that kill the agricultural crop. Agricultural irrigation to satisfy the high evaporation demand of sunny agricultural land, such as California’s Central Valley, has become a direct competitor for use as a public water supply. READING LIST Shuttleworth, W.J. (1992). Evaporation. In: Handbook of Hydrology. D.R. Maidment (Ed.). McGraw Hill, New York. Dingman, S.L. (1994). Physical Hydrology. Prentice-Hall, Englewood Cliffs, NJ.

EVAPOTRANSPIRATION JOSE O. PAYERO University of Nebraska-Lincoln North Platte, Nebraska

Liquid water from a surface can be transformed into water vapor by either evaporation or by transpiration.

Evaporation is the process for converting liquid water to water vapor and removing it from the evaporating surface. Transpiration is the vaporization of water contained in plant tissues and the removal of vapor to the atmosphere through leaf stomata (1,2). Evaporation and transpiration occur simultaneously in cropped surfaces, and there is no easy way of quantifying the magnitude of each component. For practical applications, such as irrigation scheduling and irrigation system design, it has been more useful to consider both processes combined. The combination of these two processes is called evapotranspiration. The proportion of each component in a cropped surface is affected by factors such as vegetative cover, available water in the soil, and surface wetness. For an annual crop like corn, evaporation is the dominant component of evapotranspiration at the beginning of the season when the soil surface is exposed to solar radiation. As the crop grows and the canopy covers the surface, evaporation is minimized, and transpiration becomes the dominant component. IMPORTANCE OF EVAPOTRANSPIRATION Evapotranspiration is an important process in agriculture and other natural sciences, such as hydrology, because it is an important component of the hydrologic cycle. It represents the water that is effectively lost from the earth’s surface, and can no longer be controlled by humans. This type of water loss is often called consumptive use. Other types of processes that usually cause water losses from a given area on the earth’s surface, such as runoff and deep percolation, do not involve a change in the physical state of water and therefore, water can still be controlled to some degree by humans. Plants use water as a solvent and transport mechanism for nutrients and other chemicals, as a reagent for the chemical reactions involved in their physiological processes (such as photosynthesis), and as a component of cell cytoplasm, which allows plant tissues to stay

EVAPOTRANSPIRATION

turgid. Most of the water consumed by plants, however, is used in evapotranspiration. Evapotranspiration has the important function of regulating the temperature of plants, keeping them cool within a temperature range that favors growth. When the water supply in the soil is limited, for instance, plants respond by closing their stomata. This restricts the rate of evapotranspiration, and the temperature of the canopy tends to increase (3). This increase in canopy temperature has been used to estimate the rate of evapotranspiration of crops and as a way to detect crop water stress for irrigation scheduling (4). Because most of the water consumed by plants is lost in evapotranspiration, it takes a considerable amount of water for a crop to produce one unit weight of dry matter, as shown for different crops in (Fig. 1). Researchers have shown that crop yield is often linearly related to crop evapotranspiration, up to the point where yield is limited by factors other than water (6). Therefore, if evapotranspiration is limited, yield is usually reduced. For this reason, in regions where rainfall is not sufficient to provide enough water for crops to keep evapotranspirating at a nonlimited rate, irrigation is required to obtain adequate crop yields. The nonlimited rate of evapotranspiration, however, can also be maintained by applying excess water. Application of excess water, however, has been linked to undesired side effects such as drainage problems, salinization of soils, soil erosion, and pollution of surface and groundwaters. In places where irrigation water needs to be pumped, pumping excess water also represents higher production cost. Therefore, it is considered ideal to schedule irrigation according to crop water needs. This requires, among other things, good knowledge of crop evapotranspiration rates. MEASURING EVAPOTRANSPIRATION

Units of water per unit dry matter

Measuring the rate of evapotranspiration of crops and other surfaces is complex and is a subject that has attracted considerable research. Many methods have been devised to measure and estimate evapotranspiration. Methods for measuring evapotranspiration include the use of lysimeters, scintillometers, micrometeorological techniques such as the Bowen ratio and eddy covariance

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Figure 1. Unit weight of evapotranspirated water needed per unit weight of dry matter produced by different crops (adapted from Ref. 5).

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methods, and measurement of sap flow. Methods to measure evapotranspiration usually require expensive instrumentation and specialized training, and therefore have been confined mostly to research applications. Because of this, considerable effort has been devoted to develop methods for estimating evapotranspiration.

ESTIMATING EVAPOTRANSPIRATION Methods for estimating evapotranspiration include the use of different kinds of devices and a variety of equations. Devices such as evaporation pans, atmometers, and soil moisture monitoring devices have long been used to estimate evapotranspiration. The use of equations, however, is by far the most common method to estimate evapotranspiration. A great variety of equations have been proposed though the years, ranging from very simple to very complex models (1,7). Simple models usually try to estimate evapotranspiration based on empirical relationships involving one or several meteorological variables. Complex models consider the physics governing the evapotranspiration process and try to include all factors that significantly contribute to the process (1,8). Complex models can be further divided into single-layer models and multiple-layer models. Singlelayer models, such as the Penman–Monteith model, consider the crop canopy as a ‘‘big leaf,’’ taking into account only processes that occur between the top of the canopy and the atmosphere. Multiple-layer models, on the other hand, also take into account those processes that take place below the crop canopy. Multiple-layer models are more theoretically sound than single-layer models, but their complexity makes them impractical for widespread use. Because a single-layer model is sufficiently accurate for most practical applications and relatively simple to apply, it has been proposed as the recommended method for estimating evapotranspiration (9). Considerable effort has been made to estimate evapotranspiration using inputs obtained from remote sensing platforms, such as satellites or airplanes (1,10,11), and others have even tried to estimate evapotranspiration by measuring the flux of stable isotopes (12). Evapotranspiration, however, is more often estimated from equations that use meteorological data as input, as well as inputs that describe the characteristics of the evaporating surface. This is a convenient method because meteorological data are readily available in most places. Most nations and states support a network of meteorological stations and offer this information to the public in various ways. Meteorological data commonly used to estimate evapotranspiration include solar radiation, air temperature, relative humidity, and wind speed. A detailed procedure for calculating evapotranspiration has been described by Allen et al. (1). The method involves a two-step process. One step consists of calculating the evapotranspiration rate of a reference crop, either clipped grass or alfalfa, which is usually known as reference evapotranspiration. In older literature, this was also called potential evapotranspiration. It represents a measure of the evaporating demand of the atmosphere for a short,

cropped surface that effectively covers the ground, is growing healthily, and is not short of water, that is, a condition in which transpiration is not limited by stress, and the evapotranspiration demand of the atmosphere is met. The second step involves adjusting the reference evapotranspiration to match the conditions of the specific surface or crop being considered. This is done by calculating an adjustment factor, usually known as the crop coefficient. Depending on the accuracy required, calculating crop coefficients can also be a simple or complex process (7,13). Multiplying the reference evapotranspiration by the crop coefficient then results in the evapotranspiration rate for the crop or surface in question. REQUIREMENTS FOR EVAPOTRANSPIRATION Procedures used to estimate evapotranspiration try to simplify the complexities of the physical and physiological processes that affect evapotranspiration to a manageable number of quantifiable variables. They try to recognize that for the evapotranspiration process to take place, it is necessary to have 1. 2. 3. 4.

energy water space in the atmosphere to hold the water vapor a transport mechanism for the water vapor to move from the surface to the atmosphere.

Evapotranspiration is an energy-driven process. It takes approximately 2.45 megajoules of energy to evaporate 1 kilogram of water at 20 ◦ C. The sun supplies the energy needed for evapotranspiration from the earth’s surface. Part of the solar energy that reaches the evaporating surface, however, is reflected back to the atmosphere and cannot be used for evapotranspiration. Of the energy that stays on the evaporating surface, known as net radiation, not all is used in evapotranspiration. The energy balance of a surface also includes energy that is absorbed or released by the soil (soil heat flux), by the air (latent heat flux), and that used in evapotranspiration (latent heat flux). All of these types of heat fluxes take place simultaneously from a given surface, and their proportions depend on the characteristics of the surface and weather conditions. The amount of energy used for evapotranspiration includes short-wave radiation that comes directly from the sun, long-wave radiation or heat that comes from the surface, and advective heat that is transported horizontally by hot wind to the evaporating surface. The amount of energy available for evapotranspiration varies with latitude, day of the year, time of day, atmospheric conditions, and the characteristics of the surface itself. Figure 2 shows how the theoretical clear-sky solar radiation for different latitudes varies throughout the year. Figure 3 shows how measured solar radiation varies during the day and the effect of cloudiness in reducing the amount of solar energy that reaches the surface. As a general rule, the higher the amount of energy available at the evaporating surface, the higher

Clear-sky solar radiation, mm/day

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Figure 2. Theoretical clear-sky solar radiation values for different northern latitudes and day of the year at sea level.

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Figure 3. Measured solar radiation in North Platte, NE, on July 15 and 16, 2001.

the evapotranspiration rate, assuming that all other conditions required for evapotranspiration are met. Water is the next essential requirement for evapotranspiration. If water is not available, the energy available to evaporate water is then used to heat the air and the soil. Under desert conditions, for instance, where energy is usually plentiful, but water is normally very limited, the evapotranspiration rate can be very small. Most of the energy is converted to sensible heat flux, soil heat flux, and a very small portion or none of the energy is converted to latent heat flux. Similarly, when crops are under water stress, the evapotranspiration rate is reduced. Space in the atmosphere to hold the water vapor is also needed for evapotranspiration to proceed. When water is evaporated from a surface, the water vapor will travel to the atmosphere where it will be stored, provided that the atmosphere is not saturated. If the atmosphere is already saturated, it will not be able to store the additional water vapor, and evapotranspiration will be restricted. The drier the air in contact with the evaporating surface, the more evapotranspiration is enhanced. For this reason, the humidity of the air, usually expressed as relative humidity, is an important factor to consider when estimating evapotranspiration. The last requirement for evapotranspiration is a transport mechanism for moving water vapor from the surface to the atmosphere. There are two basic transport mechanisms for water vapor. The first is turbulence.

FOG

Turbulence is created when air moves horizontally over a rough surface. The friction created by the contact of the air with the rough elements of the surface creates a vertical component of the wind speed, which creates eddies of different sizes. The size of these eddies depends on the roughness of the surface and on the magnitude of the wind speed. These eddies carry the water vapor to the atmosphere. The other transport mechanism is buoyancy. Hot air is less dense that cold air; as the air close to the surface becomes hotter than the air above, it tends to ascend, carrying with it the water vapor. Turbulence is the most important transport mechanism in most instances, so wind speed is a very important factor in determining evapotranspiration. The higher the wind speed and the rougher the surface, the more turbulence, and the higher the evapotranspiration rate will be.

Evapotranspiration rates are most commonly expressed in units of water depth per unit time, such as millimeters per day (mm day−1 ) or inches per month (in months−1 ). Since it takes energy to evaporate water, water depths can also be expressed in terms of energy received per unit area. Therefore, evaportranspiration is often expressed in units of energy per unit area per unit time, such as watts per squared meter (w m−2 ), or megajoules per squared meter per day (MJ m−2 day−1 ). It can, however, also be expressed in units of energy per unit area, such as watts m−2 , or MJ m−2 day−1 . Many of the factors affecting evapotranspiration are so dynamic that the magnitude of the evapotranspiration rates for a given surface will vary from day to day, from place to place, and throughout the day. Figure 4 shows the calculated daily evapotranspiration rate for corn in North Platte, Nebraska, during the 2000 growing season. It shows the typical large variations in evapotranspiration rate that can be expected from day to day, as a result of normal daily changes in weather conditions. It also shows a seasonal pattern, as a response to the seasonal changes in available energy and to the changing water demand of the crop during its growing cycle.

Evapotranspiration, mm/day

Corn, North Platte, NE, 2000

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BIBLIOGRAPHY 1. Allen, R.G., Pereira, L.S., Raes, D., and Smith, M. (1998). Crop evapotranspiration—guidelines for computing crop water requirements. Irrigation and Drainage Paper No. 56, FAO, Rome, Italy. 2. Allen, R.G. et al. (2002). Evapotranspiration from a satellitebased surface energy balance for the Snake Plain Aquifer in Idaho. Proceedings of the 2002 USCID/EWRI Conference. July 9–12, San Luis Obispo, CA, pp. 167–178. 3. Gates, D.M. (1964). Leaf temperature and transpiration. Agronomy Journal 56(3): 273–278. 4. Yazar, A., Howell, T.A., Dusek, D.A., and Copeland, K.S. (1999). Evaluation of Crop Water Stress Index for LEPA irrigated corn. Irrigation Science 18: 171–180. 5. Hoeft, R.G., Nafziger, E.D., Johnson, R.R., and Aldrich, S.R. (2000). Modern Corn and Soybean Production, 1st Edn. MCSP Publications, Champaign, IL. 6. Schneekloth, J.P. et al. (1991) Crop rotations with full and limited irrigation and dryland management. Transactions of the ASAE 34(6): 2372–2380.

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7. Doorenbos, J. and Pruitt, W.O. (Eds.). (1977). Crop Water Requirements. Irrigation and Drainage Paper No. 24, FAO, Rome, Italy. 8. Suttleworth, W.J. and Wallace, J.S. (1985). Evaporation from sparse crops—an energy combination theory. Quarterly Journal of the Royal Meteorological Society 111: 839–853. 9. EWRI. (2001). The ASCE Standardized Reference Evapotranspiration Equation. Environmental and Water Resource Institute of the American Society of Civil Engineers. Standardization of Reference Evapotranspiration Task Committee Report. Available: http://www.kimberly.uidaho. edu/water/asceewri/. 10. Moran, M.S. et al. (1989). Mapping surface energy balance components by combining Landsat Thematic Mapper and ground-based meteorological data. Remote Sensing of Environment 30: 77–87. 11. Payero, J.O. (1997). Estimating Evapotranspiration of Reference Crops Using the Remote Sensing Approach. Ph.D. Dissertation, Utah State University, Logan, UT. 12. Yakir, D. and Sternberg, L.daS.L. (2000). The use of stable isotopes to study ecosystem gas exchange. Oecologia 123: 297–311. 13. Wright, J.L. (1982). New evapotranspiration crop coefficients. Proceedings of the ASCE 108: 57–75.

FOG

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ARTHUR M. HOLST

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Philadelphia Water Department Philadelphia, Pennsylvania

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Figure 4. Calculated daily evapotranspiration for corn in North Platte, NE, during the year 2000.

Fog is a cloud that materializes near the ground or over water. It is a reaction that occurs when the temperature near the ground cools to the temperature required to produce dew and causes the water vapor in the air to become visible in the form of a cloud of precipitation. There are different types of fog, which occur when different variables are involved. The two conditions necessary for fog formation are mild or no winds and air

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temperatures that are equivalent to the temperature at the dew point. Mild or no winds are important because obstructions in the way of a strong wind can cause it to blow in an up and down motion, which brings warmer air down and pushes the colder air up. Cool air is a requirement for fog formation so wind prevents it from happening. Fog, like most weather hazards, can cause serious problems for society. The only time this condition is not applicable is when fog forms over water due to the absence of obstructions. In this case, the reverse reaction occurs, and the fog worsens as the wind grows faster. Fall nights are often said to contain the perfect conditions for producing fog. At night, the ground cools which also cools the air directly above it. This reaction causes droplets of water to become suspended in the air resulting in the creation of fog. However, these perfect conditions don’t last forever. By morning, the heat from the sun begins to warm the ground, and the fog evaporates. There are many types of fog, but the four most common are evaporation fog, upslope fog, precipitation fog, and radiation fog which is commonly called ground fog. Evaporation fog is found mostly out at sea, which is why it is also known as ‘‘sea fog.’’ It is the reaction of moist air moving over colder water. The fog conditions worsen as the wind speed increases. Upslope fog is produced by moist air that is sent by a strong wind up into mountainous regions. Precipitation fog is created when precipitation in the form of rain or snow hits drier air and causes water vapor to materialize instantly. Radiation fog is seen at night when the ground is cool. The air that comes in contact with the ground also becomes cool and creates water vapor which soaks the ground and causes the formation of dew. Like most weather hazards, fog can have severe impacts on society. Fog can cause health problems in polluted areas because the water vapor produced can become acidic. Driving through a thick cloud severely reduces a person’s visibility and is a major contributor to car accidents. Even worse than the occasional car accident due to fog are the aircraft and boating accidents that can occur. Throughout history, fog has been known as the silent murderer that has taken hundreds of lives at sea in catastrophic boating accidents.

READING LIST Joseph, P. Weather Wisdom. (2002). Available: http://www.touchtmj4.com/4weather/wxwisdom/fog/fogtypes.asp. (March 19). Malan, J. Weather Wisdom. (2002). Available: http://www.touchtmj4.com/4weather/wxwisdom/fog/fog.asp. (March 19). Malan, J. Weather Wisdom. (2002). Available: http://www.touchtmj4.com/4weather/wxwisdom/fog/fogdisasters.asp. (March 19). Parsons, L. Fog. (2002). Available: http://www.activeangler.com/ articles/safety/articles/lee parsons/fog.asp. (March 19). Understanding Clouds and Fog. USA Today. (2002). Available: http://www.usatoday.com/weather/wcloud0.htm. (March 19). Williams, J. Fall nights often perfect for forming fog. USA Today. (2002). Available:http://www.usatoday.com/weather/wrfog.htm. (March 19).

COASTAL FOG ALONG THE NORTHERN GULF OF MEXICO TIMOTHY ERICKSON National Weather Service New Orleans/Baton Rouge Forecast Office Slidell, Louisiana

INTRODUCTION Coastal fog is a major problem for all traffic along the United States’ coastlines. Lives and large monetary losses have occurred because of coastal fog. These losses have been realized over land, in the air, and on the water. Huge strides have been taken to understand and combat the coastal fog forecasting problem by the National Weather Service. Studies and research continue to improve these fog forecasting techniques. Tens of thousands of people use the Gulf of Mexico as their home or as part of their occupation in many different ways, and billions of dollars in products and property are carried and moved through the channels and river systems to and from the Gulf of Mexico each year. All are affected by coastal fog many times throughout the fall, winter, and spring. This research project took place along the northern Gulf of Mexico from the upper Texas coast to the Mississippi coastline. METHODOLOGY The synoptic and mesoscale patterns used in this research were from the fall, winter, and spring of 1998, 1999, 2000, and 2001. Fog was defined as water droplets suspended in the air reducing visibility to one half-mile or less. Shower and thunderstorm activity reducing visibility to these levels was not used. No records exist for fog contributions to the hydrologic cycle. Water contributions caused by fog are dismissed as false tips in rain gauges. The reference to ‘‘boundary layer’’ in this article will be the layer of atmosphere from the surface to the base of the lowest inversion. Variables used in this research were surface pressure, rainfall, moisture advection, wind direction, wind speed, water temperatures, ambient temperature, and ambient dew point temperature from land observations and production platforms in the northern Gulf of Mexico. Parameters used from previous studies by Johnson and Graschel (1) were air temperature, dew point temperature, wind direction, wind speed, ceiling heights, and visibilities from oil and gas production platforms at an average altitude of 35 ms over the northern Gulf of Mexico. Gulf of Mexico sea surface temperatures were provided by the Tropical Prediction Center oceanographer in Miami, Florida. SCALES OCCURRING WITH COASTAL FOG Mesoscale—Widespread Horizontal range would normally be from 50 to several hundred miles.

COASTAL FOG ALONG THE NORTHERN GULF OF MEXICO

Microscale—Locally Horizontal range could be at a point close to 50 miles. CONDITIONS THAT HELP PRODUCE COASTAL FOG

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there will be plenty of moisture and a strong inversion that may develop some dense coastal fog. Even without a prefrontal trough or rainfall, the subsidence inversion and gulf moisture ahead of a slow-moving cold front will sometimes be enough to provide conditions promoting dense coastal fog.

Radiative Heat is transmitted via long wave radiation (infrared) away from Earth. This radiation causes nocturnal cooling of the ground and subsequent atmospheric boundary layer. Radiational cooling is responsible for developing a stable layer near the surface by cooling the layer of air in contact with the ground while leaving a relatively warmer layer immediately above this cooler layer. This stability by warmer air above cooler air with increasing altitude is called an inversion. This feature is important in forecasting the depth of coastal fog. As the radiative process continues, the layer will cool further until saturation is achieved and fog can be produced, given the pressure does not change much during the process. Coastal Interface. Temperature and moisture gradients are strong where the ocean meets land, especially at night during the fall, winter, and spring. During prime radiational cooling processes, this interface can become moist and stable horizontally from land to water when there is no forcing. These conditions were found when a thin ribbon of fog developed at this interface of land and water. The fog stretched linearly for more than 100 miles and vertically for hundreds of feet. Studies on how this type of fog develops are ongoing but horizontal skew-t sounding-like profiles may be a tremendous help in what is occurring at this land–water junction when this fog forms. No horizontal sounding-like profiles exist, and this is the first mention of such a profile. But if they were available, they could give some valuable insight not only into the coastal ‘‘ribbon’’ fog formation but also where thunderstorms are more likely to develop along with many other variables. Frontal This is another inversion producing process. Coastal fog can be developed with the assistance of cold, warm, or stationary fronts. Cold Frontal. An inversion develops by cooler air displacing warmer air at the surface. The interface where these air masses meet in the vertical is known as a frontal inversion. Coastal fog formation under a frontal inversion is nonexistent with fast-moving cold fronts, because of stronger winds that cause a deeply mixed layer, along with dry, cold air advection behind the front. When warm air is lifted by a cold front, it causes displacement and mid-altitude air ahead of the front becomes subsident. This process causes compressional warming, which forms another inversion well ahead of the cold front. As the surface cools at night, the boundary layer becomes stable well ahead of the front. If there has been rain ahead of the cold front with a prefrontal trough,

Warm Frontal. Relatively cool saturated or nearly saturated air ahead of a warm front is gently displaced by warmer air. The cool air is normally not very deep (100–300 ms) and is topped by warmer air at relatively low altitudes, which causes an inversion to form where the two air masses meet in the vertical. Cooler air cannot hold as much moisture as warm air. As the air ahead of the warm front is already moisture laden, cooler, and stable, water vapor condenses to small particles, which causes fog to form. This result happens frequently along the coast during the cooler months of the year and sometimes far inland. A similar scenario along the coast was also described by Hsu (2). The dynamics of warmer, higher dew point air flowing over cooler waters was described by Kotsch (3) and Mullan (4). Coastal fog can occur behind warm fronts but is not as common. The air is, by definition of a warm front, warmer than the air ahead of it. The warmer air expands and therefore can hold more water vapor, which will normally dissipate any fog behind a warm front if no overwhelming positive moisture flux is occurring in the warm sector. When this flux occurs, it can be far too great for even the warm air to hold and the water vapor will condense, which causes fog to form. During research, this result occurred only when an inversion was still present behind a warm front. Stationary and Slow Moving Fronts. These types of fronts are aggressive at developing coastal fog, which can occur in two ways ahead of a cold front that becomes stationary or slow moving. One is when moisture is not displaced and wind speeds are very light behind the trough preceding the cold front. The second is when the wind fetch is well over the marine environment and it brings warm moist air back over cooler waters or land. Warm frontal fog conditions were explained above. Two unusual fog days occurred as post-cold frontal events. A cold front passed through the southern Louisiana coastal region, and dew points and dry bulb temperatures cooled. Coastal fog then began to develop over many sites from east central Texas to south central Louisiana. The air mass change was more negative for the dry bulb temperature than for the dew point temperature, and the cold front slowed from about 10 knots to less than 5 knots as it reached the coastline. This decrease caused the air behind the front to become saturated. Frontal forcing weakened, and consequently northerly winds weakened and created perfect conditions for coastal fog development. Post-cold frontal fog is rare, but when it does occur, visibilities can plummet to less than one quarter of a mile quickly. Hsu’s (2) description on frontal fog production is similar to this research. Johnson and Gracshel (1) called frontal fog ‘‘mixing fog.’’

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Maritime (Sea Fog) When fog develops via any process over any large marine area, it is called marine or sea fog. Marine fog may be stationary or be advected elsewhere. It is during the advection process that coastal fog occurs. Dormant. As the cooler boundary layer air over land ‘‘drains’’ offshore, it causes stability to decrease, which makes mixing possible. The warm gulf gives off water vapor to the cooler air, which causes the vapor to condense, making water droplets. This process quickly deteriorates when only little ‘‘drainage’’ is occurring. Latent heat released because of condensation warms the marine boundary layer too fast, which makes the stability in this layer neutral. Fog will form in the near shore waters if this latent heating can be exhausted through the top of the maritime boundary layer. Fog development caused by this process does not occur often. When it does, the fog stays near the coast, rarely moving anywhere because of the very slow, almost laminar flow of air moving off land. When this type of fog affects the shoreline, it is also called coastal fog. The other processes mentioned above may produce nonadvective marine fog that stays in the near shore waters. Advancing. This process in which coastal fog is achieved is the most common. The air in the boundary layer is well mixed but does not entrain air from above the inversion. Fog that forms either onshore or offshore by any process and moves elsewhere is an advective fog. When fog advects from one offshore location to another or to the coast, it is said to be advancing. The most common type of advancing fog is found with post-strong cold front events. Winds from these fronts move the top layer of warm shelf water away from the coast. This water is replaced by cooler upwelled water. A strong temperature gradient exists where the cooler upwelled water meets the warm water offshore. When the next cold front produces return flow from the warmer water offshore, moisture-laden air saturates while moving over the cooler water. Extensive areas of fog or very low cloud ceilings form before moving inland, which causes extremely low visibilities stretching for miles along the coast. When the moisture came in the form of low clouds during research, the cloud deck would often descend to the surface producing very dense coastal fog as well. As warm air overlays cooler water, it produces a stable boundary layer. No vertical mixing occurs by air moving up or down. Mixing was produced solely by wind advancing the fog. The wind did not reach through the inversion layer, and therefore, no entrainment occurred. A similar scenario was recorded by Binhau (5) and Hsu (2). Their contributions showed this environment to be stable, with surface winds from the southeast through southwest. This scenario holds true for all locations along the northern Gulf of Mexico. Advective Anytime fog forms elsewhere and is forced to another location, it is a type of advection fog. A well-mixed

boundary layer also exists with advective fog, but again, air from above the inversion is not mixed downward. In some cases, the atmosphere will be capable of supporting fog but it cannot produce it. When this happens, fog may be brought into the area by wind or by the slow movement of an entire layer of air. It is important that this air mass does not mix with air from above the inversion. If this mixture occurs, dry, relatively warmer air will be mixed into the boundary layer, causing the fog to erode. Conglomerate of Two or More Types from Above This fog producer is the most common. Two or more of the above processes usually develop fog that forms almost anywhere on Earth. One major variable may occur at the time of fog formation, but most of the time fog is supported by another equal or weaker variable, for example, frontalinduced marine fog. That is, fog develops over the marine environment by a cold front, which brings warm, moist air back over cooler waters. The fog develops mainly because of the marine environment, where it derives its moisture, but it could only do so as a result of forcing by the cold front. VARIABLES PRESENT DURING COASTAL FOG PRODUCTION Variables Always Present During Coastal Fog Formation Negative or Neutral Omega Within or Just Above the Boundary Layer. When lift occurred during the research, fog would dissipate or simply not form. The lift causes mixing through the inversion, which brings dry, relatively warm air into the boundary layer. Weak or No Positive Vorticity in the Boundary Layer. Coastal fog developed under boundary layernegative vorticity regimes, but it would not develop under moderate-to-strong positive vorticity in the boundary layer. When weak vorticity occurred within or at the top of the boundary layer, fog would lift and become a low-level cloud ceiling. As the vorticity center moved past, the ceiling would once again descend to the surface, which caused coastal fog to form. Inversion. An inversion is always present during any and all fog development and duration. Variables Present During Research for Each Condition Radiation Fog Variables

Winds 0 to 3 Knots. Radiation fog events are created without mixing through the inversion. Wind greater than 3 knots was found to create too much mixing and dissipated coastal fog during radiative events. Moisture Advection or Rainfall Within 36 Hours. Moisture is needed for any fog to develop. With no moisture advection, it was found that moisture input from rainfall would be sufficient inside 36 hours. This field was dependant on several variables. These variables included amount of

COASTAL FOG ALONG THE NORTHERN GULF OF MEXICO

rainfall, ground moisture, insolation, and amount of rainfall coupled with timing. The best results were found with light rain episodes during the early morning with strong insolation during the day.

Neutral or Negative Omega. Neutral or negative lift from some height above down to the top of the boundary layer was always found when coastal fog formed during this research. Positive lift caused mixing through the inversion, which dissipated the fog. Weak-to-No Boundary Layer Positive Vorticity. Coastal fog formed under negative and neutral vorticity regimes. Coastal fog was either displaced or was not present when moderate-to-strong positive vorticity was found within the boundary layer. Coastal fog was also present with weak positive vorticity, but a few interesting findings occurred. As a weak positive vorticity center moved through the boundary layer, coastal fog would lift, which created a low-level ceiling from 100 to 400 feet. When the vorticity maxima passed, the low-level ceiling would descend to the ground, which caused coastal fog to return. This phenomenon was known for creating ‘‘bouncy’’ fog conditions where visibilities would swing wildly from as little as 0 to as much as 4 miles. Clear Skies. Clear skies were the overwhelming majority of sky conditions experienced during radiational coastal fog events. The minority sky condition consisted of very high thin cirrus clouds. No radiational coastal fog events occurred during any other cloud conditions. Outside Downtown Areas. Radiational coastal fog events during this study were found outside the downtown areas of cities. The heat island effect was enough to dissipate any fog trying to form inside these areas. Moisture Advection or Rain Within 36 Hours. The highest frequency of coastal fog during frontal regimes was found when a prefrontal trough passed. Moisture was input by both the front causing return flow from the marine environment and the prefrontal trough causing rainfall. Clear Skies or Very High Cloud Ceiling. High cloud ceilings were noted several times throughout the research when frontal-induced coastal fog developed. Winds of 10 to 20 knots were noted above a shallow inversion, which allowed heat from the boundary layer to escape and be carried away. Surface Winds of 0 to 12 Knots. During radiative conditions, coastal fog only formed when wind speeds were 0 to 3 knots. During advective or advancing conditions, coastal fog was carried to or along the coast when wind speeds were 4 to 12 knots. No coastal fog formed during this research when wind speeds were greater than 12 knots. Onshore Winds of At Least 4 Knots and Not More Than 12 Knots. During marine-induced coastal fog, winds were

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necessary to advance the fog from the marine environment to the coast. But wind speeds greater than 12 knots caused the fog to dissipate by mixing with air above the inversion. When these winds did not reach the top of the boundary layer, no mixing occurred through the inversion and the fog would simply lift, developing a low-level cloud ceiling.

No Moisture Advection. Surprisingly, moisture preconditioning was not necessary when marine fog caused coastal fog. The marine fog was the moisture advection. Clear or Very High Ceilings. Several cases showed coastal fog production via a low-level cloud deck. When heat was capable of radiating through the inversion at the top of the low-level clouds, the cloud deck lowered or deepened toward the surface through the night, finally reaching the ground, which caused coastal fog to form. It was also noted that the lower the cloud deck, the shorter the time frame coastal fog would develop. A very general rule of thumb was realized when the boundary layer was capable of supporting fog. On average, it took about an hour for the cloud ceiling to descend 100 feet. Cloud bases higher than 1000 feet never reached the surface during marine-induced coastal fog events. Temperature. Johnson and Graschel (1) found temperature differences of several variables to be important when maritime fog developed. As indicated in their article, ‘‘Sea Fog and Stratus: A Major Aviation and Marine Hazard In The Northern Gulf Of Mexico,’’ the differences between water and air temperatures, as well as between water and dew point temperatures, were the most important variables producing marine fog when the right atmospheric conditions were in place. The graphs below show these parameters versus relative humidity (RH) values. Johnson and Graschel’s study (1) as well as this project found RH values of 98% or greater always present with coastal fog. Figure 1 shows the continental shelf region along the northern Gulf of Mexico, which is shown at the 200-m depth contour. It is also the region where a cool water temperature of 20 ◦ C (68 ◦ F) or less was found to be critical for marine fog development in the northern gulf during the right atmospheric conditions. These findings may not be the same at other locations around the globe because fog development depends on temperature gradients, over water and/or land, which are relative. Water temperature findings close to these were also accomplished by Binhau (5). Figure 2 shows RH versus (Ta - Tw), where RH is relative humidity and (Ta - Tw) is the difference between the ambient air temperature and the water temperature. In Fig. 3, in regard to the positive (+) numbers, when the water temperature is cooler than the air temperature, the air must be moving between 4 knots to as high as 12 knots for coastal fog to form. During this process, the moisture-laden warmer air loses its heat to the cooler water. As the air cools to its dew point, condensation takes place. Latent heat is released during the condensation process. The air can lose this added heat to the cooler water below and through the top of the boundary layer

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via long wave radiation, so fog can begin to form. This process is efficient and is the main reason for coastal fog developing from advancing marine fog. The positive numbers in Fig. 3 show a stable boundary layer also described by Binhau (5) and Hsu (2). In Fig. 4, in regard to the negative (−) numbers, when the water temperature is warmer than the air temperature, the air must be calm or moving no faster than 3 knots for coastal fog to form. During this process, when coastal tidal marshes exist, a column of cooler dry air will have a net gain of water vapor, which causes the dew point temperature to rise. When the air and dew point temperature are very close, condensation will begin. As a result of condensation, latent heat is released. The column will need to lose this heat through the top of the boundary layer before fog will begin to form. Even though marine fog occurs by this process, it is not an efficient coastal fog producer unless fog initially forms along the coast. Most of the time, fog will develop

over the marine environment and be stationary or drift farther offshore. Negative numbers in Fig. 4 show an unstable boundary layer as described by Binhau (5). Binhau showed that this type of fog may be produced over the marine environment even with northerly winds of 30 knots. But with winds of this magnitude, coastal fog will never develop along the northern gulf coast because the winds would force it out to sea. Figure 5 shows RH versus (Td − Tw), where RH is relative humidity and (Td − Tw) is the difference between the ambient dew point temperature and the water temperature. In Fig. 6, in regard to the positive (+) numbers, it is important to remember that the atmosphere as well as the ocean is always trying to reach a state of equilibrium. When the ambient dew point temperature is higher than the water temperature, water molecules can easily move to and become a part of the ocean surface. Water molecules find it hard to break away from the waters’ surface during these conditions, and therefore, a net moisture flux from air to water occurs. The air temperature is always equal to or greater than the dew point. Hence, there will be a transfer of heat from the air to the water as well. This process cools the air temperature, but saturation is difficult to achieve because there is a net loss of water vapor to the water surface. The air continues to cool and dry until temperatures of the water, the air, and the dew point equal or become very close. This process eventually causes saturation and can, but rarely does, cause coastal fog to form when atmospheric conditions are right and (Td—Tw) is zero or very close. Normal occurrences of fog on the positive side of Fig. 6 are when fog develops elsewhere and advances into the area. In Fig. 7, in regard to the negative(−) numbers, when the water temperature is warmer than the dew point, water molecules can easily break away from the water surface to the air. Regardless of the air temperature,

COASTAL FOG ALONG THE NORTHERN GULF OF MEXICO

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Boundary layer top

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the water will try to modify it until the water and air temperatures are equal. Even if saturation is achieved, moisture will continue to be added to the air until the dew point is equal to the water temperature. As moisture is added to the air at saturation, condensation begins and latent heat is released to the air until the temperatures of the water, dew point, and air are equal. This saturation can easily develop fog near or offshore. Normally, marine fog produced by this process occurs with little airflow (0–3 knots). After marine fog forms, it can be easily forced to the coast as winds increase ahead of the next cold front. This coastal fog producer is aggressive. As Fig. 7 depicts, this process occurs with a minimal separation of air, dew point, and water temperatures. When separations are too large, there may be too much

Figure 4. This picture shows a light offshore wind, dew point temperature (Td), air temperature (Ta), and water temperature (Tw).

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Boundary layer top

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Figure 6. This picture shows winds of 12 knots or less, air temperature (Ta), dew point temperature (Td), and water temperature (Tw).

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Figure 7. This picture shows wind speeds less than or equal to 12 knots, dew point temperature (Td), and water temperature (Tw).

dry air to modify before the next system brings this air mass back to shore. If the air is moving more than 3 knots, there will be a continual replacement of dry air and/or air from above the inversion may be mixed into the boundary layer and saturation will not be achieved. Advective Fog Variables

Winds of At Least 4 Knots and Not More Than 12 Knots. Advective fog, which causes coastal fog, is different from advancing marine fog in a couple of ways, i.e., where the fog developed and the wind fetch. Fog that develops onshore and moves to the coast is only known as advective fog. Whenever fog moves to the coast, the wind direction will always be in a direction from the fog to the coastal

Tw = 20 °C

location. All other variables that are needed for marine fog to advance to the coast are also needed for advective fog.

No Moisture Advection. Even though moisture may be in place, advective fog does not need to have a premoistened atmosphere for coastal fog to be produced. Moisture advection can be induced by the fog moving to the coast. Clear or Very High Cloud Ceilings. Cloud conditions were found to be the same as for marine fog. Conglomerate Fog Variables. The variables for each condition associated with coastal fog development have to be present when coastal fog forms under more than one condition.

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SOUNDING PROFILES PRESENT DURING COASTAL FOG A sounding profile similar to Fig. 8 was always present shortly before and during radiative coastal fog events. This actual sounding shows a very shallow (∼100 feet) moist layer capped by three strong inversions. A sounding profile similar to Fig. 9 was always present shortly before and during advective and maritime coastal fog events. This actual sounding shows a very deep (∼1000 feet) moist layer capped by a moderate inversion. Fog rarely develops beneath the type of inversion in Fig. 10. This actual sounding shows a deep moist layer capped by a frontal inversion; coastal fog occurred before the passage of this cold front, but it quickly dissipated by the time this sounding was taken. Frontal-induced and conglomerate fog was found with all three types of sounding profiles. Frontal-induced coastal fog occurred more often under the first and second profile types. Sounding profiles are important for forecasting the depth of coastal fog or any type of fog. Maritime, advective, and frontal-induced coastal fog can only be as deep as the

height from the surface to the base of the inversion. These types of fog produced the deepest fog along the coast as well as inland, and consequently they took longer to dissipate. During research, most radiation-induced coastal fog formed only under a low-level inversion. The inversion could be as high as 100 feet or as low as a few feet from the ground. The inversion was frequently strong. Air parcels were not able to penetrate the inversion, but it would not stop radiative heat transfer from the surface via longwave radiation. Radiative fog did not always follow the depth rule. It could be from the surface to the base of the inversion deep or as shallow as a few inches, even when the inversion was much higher. An interesting find during the project was horizontal stability and moisture profiles during radiative fog formation conditions along the coast. The horizontal profile shown below from the coast along with the radiative vertical profile from inland locations were both present when a thin ‘‘ribbon’’ of fog formed along the coast (see Fig. 11). This fog was found to run linearly along the coast and stretched upward for hundreds of feet. The inland vertical sounding did not support such a high

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Figure 11. This illustration shows a hypothetical representation on how a very thin ribbon of fog develops at the coast.

depth of coastal fog, but this vertical profile may have been different along the coast because strong moisture and temperature gradients existed. Research into this phenomenon is ongoing. CONCLUSION

2. Hsu, S.A. (1988). Coastal Meteorology. Academic Press, New York, p. 52. 3. Kotsch, W.J. (1983). Weather for the Mariners. Naval Institute Press, Annapolis, MD. 4. Mullan, D.S. (1984). Low stratus and sea fog in the North Irish Sea. Weather (GB) 39: 235–239. 5. Binhau, W. (1985). Sea Fog. Springer-Verlag, New York.

Fog has never received the attention it deserves because it does not make for an explosive story like hurricanes and tornadoes. But fog is blamed for large monetary losses as well as property losses. Each year, fog is blamed for indirectly taking more lives than hurricanes in the United States. Fog also donates a small but significant amount of water to the hydrologic cycle. In some places, such as the western high coast region of South America, fog is the only way insects and grasses receive water. When fog develops, there is always an inversion in place, which means the boundary layer is disconnected from the remaining atmosphere above with respect to mixing. This process is called decoupling. When the inversion erodes and mixing resumes through this layer, it is said to be coupled. As a result, pollutants released to the environment will remain in the boundary layer during decoupled conditions and will mix out during coupled conditions. Petrochemical plants and other facilities-producing pollutants that are dispersed to the environment can use fog as an indicator for when not to release waste products. Results during research show there are numerous variables and observations from the microscale environment to consider when forecasting coastal fog conditions. Current National Weather Service numerical models do not solve for microscale conditions, and therefore, forecasters must rely on pattern recognition to resolve these issues. BIBLIOGRAPHY 1. Johnson, G.A. and Graschel, J. (1990). Sea Fog and Stratus: A Major Aviation and Marine Hazard in the Northern Gulf of Mexico.

RAIN FORESTS ALDO CONTI Frascati (RM), Italy

Rain forests are, by definition, those forests that receive more than 2500 mm of rain each year. Rain forests are characterized by very dense vegetation dominated by tall trees and huge biodiversity. Rain forests exist in many parts of the planet, but most of them are along the equator, where the weather is stable throughout the year and there is never a dry season. Rain forests do not have seasons at all. The amount of rain is almost constant during the year, and the temperature seldom dips below 16 ◦ C. Rain forests cover 7% of the earth’s land surface and 2% of its total surface, but are home to more than half of all animal and plant species. Despite the fact that rain forests cover less than 10% of the earth, they support a third of its plant matter. The largest tropical rain forest in the world is the Amazon Rain Forest, which lies in the countries of Brazil, Bolivia, Peru, Ecuador, and Colombia. There are rain forests in Africa, mainly in the Congo, and in Oceania. The large amount of rain also creates some of the biggest rivers and flood plains of the world. Unfortunately, rain forests are in danger. They lie mainly in poor countries, where the economic situation forces people to use all resources, which is why most rain forests have shrunk dramatically in size over the last few decades. What is in danger is their huge biodiversity. Some

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researchers estimate that the Amazon Forest alone might host up to 10 million species of animals, mainly insects. At the rate the forest is disappearing, most of the estimates say that we are losing something like 100 of these species every day, before even being described by science. Most of the animals in rain forests have adapted to live in the upper level of the canopy, where food is really plentiful. The constant weather means that there are flowers and fruits at any time of the year. The largest group of animals is that of insects that can easily climb trees and often have a highly symbiotic relationship with the plant life of the forest. Ants and termites are the most abundant animal of every rain forest. Animals are very important for the survival of the forest. Underneath the canopy, the wind is not strong enough to disperse efficiently the seeds produced by plants. Plants rely therefore on insects, that very often pick up the seeds and drop them some distance away. But the most important animals for seed dispersal are birds. Birds in rain forests eat mainly fruits, and the seeds can pass through their digestive system unscathed. By the time a bird excretes its load of seeds, it has normally flown a long distance from the plant, which ensures a high level of genetic mixing, healthy and helpful for plants. There are even some seeds that do not germinate unless they have gone through the digestive system of birds. More than one-fourth of all bird species in the world today live in tropical rain forests. To have an idea of the importance of rain forests, some figures might be interesting. A single square meter of rain forest supports between 45 and 80 kg of biomass, far more than any other biome. This biomass gains more importance considering that most of it is made of carbon removed from our atmosphere. One hectare of rain forest can contain 200 species of trees and more than 40,000 species of insects. In Panama, scientists discovered fully 80% of the world’s currently known beetle species on only 19 trees. Once, researchers discovered over 600 new species of beetle by studying a single species of tree. Although not as rich in species as their Asian or Amazonian counterparts, African rain forests contain more than half of that continent’s animal and plant species, even though they cover less than 7% of its total land area. Because of the huge biodiversity, most species have evolved to occupy very specialized niches of the environment, which means that many species depend on each other and cannot survive without each other. Deforestation and many other human activities disrupt these complicated relationships. Rain forests could play a crucial role in feeding the whole world’s population. Many vegetables and fruits such as bananas and peppers that we consume come from rain forests. Peanuts come from rain forests, as do many drinks (coffee, tea, cola), oils (palm, coconut), flavorings (cocoa, vanilla), and other foods (beans, grains, fish). And many more vegetables are still there ready to be discovered. Moreover, researchers have identified over 200 plants that produce potential cancer fighting substances. And this considering that only 1% of plants have been intensively screened for such properties. Tropical rain forests do not offer only goods. They are a vital part of the hydrogeologic

cycle of the planet and act as a global air purifier, absorbing huge amounts of carbon dioxide and releasing oxygen. Despite their importance, rain forests everywhere are exposed to huge threats. Often, forests are cleared with fire to make room for cultivation. One plot is used for a few years until the soil is exhausted, and then farmers move on to clear another patch, putting the lushest forests in danger of desertification. But other industrial interests, including timber and mining, are taking advantage of rain forests. Part of the danger comes also from animal and plant species introduced from other environments. All these activities result, every year, in a rather large loss of rain forest. It is difficult to estimate the extent of the damage, as data are not plentiful or reliable. Nevertheless, it is true that in some countries, like Madagascar, the whole forest has almost disappeared in a few decades. As a result, human activities might be inducing the most important mass extinction since the fall of dinosaurs 65 million years ago. According to some research, up to 10% of the world’s species might disappear in the next 25 years. But the truth is that over 50% of the earth’s plants and animal are in danger. Nearly 20% of known endangered vertebrates are threatened by introduced species. Cultures are going extinct, too. Since the turn of the century, 90 tribes of indigenous peoples have been wiped out in Brazil alone. The pace of annihilation is increasing; 26 of those tribes were killed or scattered in the past decade. Everything should be done to halt this loss that many scientists think might affect the earth’s climate, too, on a global scale.

FROST ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

Essentially, a type of dew, frost is ice formed by the condensation of atmospheric water vapor on a surface. It generally occurs at night, and when frost forms, it can sometimes be seen in patterns of ice crystals. Frost can be extremely damaging to outdoor crops and plants. Researchers and weather services throughout the world monitor the effects of damaging frosts. Frost forms on any surface, including cars, grass, and buildings. Frost forms through one of two processes: the formation of dew that subsequently freezes, or deposition, which is the process wherein a gas changes to a solid. In frost formation, this gas is water vapor, and the solid state is ice crystals. These two processes will occur when the air is saturated. Saturation occurs when the air holds as much water vapor as is possible at its temperature and pressure. The temperature at which frost is deposited is known as the frost point temperature. Condensing water vapor must have something upon which to condense. If the temperature of the ground is below 0 ◦ C, then the deposit will initially be as dew. Over time, this dew will eventually freeze, forming frost. However, if the dew point of the air is below 0 ◦ C, the

FROST DAMAGE

deposit will be hoarfrost, which is ice that forms directly through deposition, without initially forming as dew. Hoarfrost is also known as black frost, because unlike regular frost, it is not visible as white crystals (normal frost is called white frost). It is possible for frost to form even if the air temperature is above freezing. Frost formation depends solely on the air’s dew point. However, the temperature of a surface affects whether or not dew, frost, or hoarfrost form because colder objects radiate less heat into the air surrounding them, keeping that air’s dew point down. The formation of frost is also governed by a process known as radiational cooling. Frost formation requires a surface temperature below 0 ◦ C, so cold surfaces are necessary. At night, certain surfaces will cool much faster than the surrounding air and other surfaces because all objects radiate heat at all times. During the day, objects generally recoup any lost energy from energy received from the Sun. However, at night, objects no longer receive heat radiated by the Sun, and so less energy is generally received, resulting in a rapidly decreasing nighttime temperature. Frost is more likely to form on clear nights because surfaces cool faster when no clouds radiate heat. On cloudy nights, the clouds may radiate enough heat to surfaces to prevent frost from forming. Frost typically forms under conditions of light or no wind and sufficiently cold temperature. Winds cause air turbulence, and this turbulence mixes the air, which inhibits frost formation. Typically, frost will form more easily overnight because temperatures tend to be lower and the air moves more slowly than it does during the day. Due to radiational cooling, frost forms less often in areas where many buildings, trees, and other objects are; it also forms less often near bodies of water. Multiple factors are used by scientists and meteorologists to determine whether or not frost will occur on a given night. One is the general weather of that day/night. The situation most favorable to frost formation is a cloudy day followed by a clear night because clouds prevent the Sun from adequately heating the soil. Humidity is also used. If the dew point is over 5.5 ◦ C at night, frosts are unlikely. If it is below 2.2 ◦ C, a frost is highly probable. In areas where frost forms, local weather services will designate the type of frost that might be deposited in the region. These designations are light, heavy, and killing. A light frost will have no destructive effects on vegetation. A heavy frost is a significant deposit but is not likely to affect the staple vegetation of a region. A killing frost is severely destructive to vegetation and can decimate an entire crop. In the United States, frost warnings will be issued only until October 15th west of a line from Frederick, Maryland, to Charlottesville, Virginia. East of this line, warnings are not issued past November 1st. In the spring, frost warnings are issued only if there is a possibility that crops and other plants could be damaged. Killing frosts are monitored by weather services, as regions attempt to predict possible arrival dates, so that crop producers can better prepare for their arrival. Various methods can be employed to diffuse the effects of a harmful frost, such as placing small heating systems throughout a crop area, continuously sprinkling water on crops

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throughout the night, or, on a smaller scale, placing simple screening cloths around a home garden. Farmers and scientists are continuing to experiment with new methods to avoid the harmful effects of black frost on vegetation. Frost harms plants by forming ice in and around cells. The water around the cells is purer than that inside the cell, so it will usually freeze first. Those plants that have greater quantities of solutes within their cells are more resistant to frost and can withstand cold temperatures more easily than those plants with little solute in their water. Plant damage from frost is determined by the type of plant, the stage of growth of the plant, and the length of time the temperature is low enough for frost formation. READING LIST The Encyclopedia of the Atmospheric Environment on-line. Available: http://www.doc.mmu.ac.uk/aric/eae/Weather/Older/Frost. html. The Weather Doctor. Available: http://www.islandnet.com/∼see/ weather/whys/frost.htm. About Frost by Steve Horstmeyer. Available: http://www.shorstmeyer.com/wxfaqs/frost/frost.html. UCLA Department of Atmospheric Science on-line. Available: http://www.atmos.ucla.edu/web/. Iowa State University—On-line Case Studies. Available: http://www.iitap.iastate.edu/jhodson/idot/frost/jan1/frost1.html. Dew and Frost. Available: http://www.nssl.noaa.gov/∼cortinas/ 1014/112 2.html. ASK—A—SCIENTIST Archive, Argonne National Laboratory. Available: http://newton.dep.anl.gov/askasci/wea00.htm. When Does The Last Frost Usually http://www.crh.noaa.gov/gjt/frost.htm.

Occur?

Available:

How Does Frost Injury Occur? Available: http://www.gov.on.ca/ OMAFRA/english/crops/field/news/croppest/cp0798 w.htm. Sterling Watch/Warning Definitions. Available: http://www.erh. noaa.gov/er/lwx/Defined/. Cooling Air. Available: http://www.doc.mmu.ac.uk/aric/eae/ Weather/Older/Cooling Air.html. Frosty Cars. Available: http://www.weatherwise.org/qr/qry.frostycar.html. Frost Damage, Control, and Prevention—Fruit and Vines. Available: http://www.pir.sa.gov.au/pages/agriculture/horticulture/ frostdamagefs.pdf. Frost Risk Assessment and Damage. Available: http://www.pir.sa. gov.au/pages/agriculture/horticulture/frost risk.htm:sectID =445&tempID =11.

FROST DAMAGE ARTHUR M. HOLST Philadelphia Water Department Philadelphia, Pennsylvania

Frost is ice formed by the condensation of atmospheric water vapor on a surface. When low temperatures are present in a certain region, the potential for frost damage to plant life exists in that region. Frost damage can injure plants permanently or slow plant development. Several

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factors dictate the extent and severity of frost damage to plants in any given region. Frost damage can have profound effects on agriculture and ecosystems. Frost is most likely to form on a cold, clear night that has been preceded by a cold and cloudy day. Lack of wind is vital to a developing frost. The little heat that was stored in the ground during the day will dissipate more quickly overnight. If the dew point has dropped below freezing, the result, is a heavy frost. As conditions change from day to night, an air temperature drop of 1 ◦ C per hour is a signal that a frost is likely to develop. Even if the conditions of freezing temperatures and a calm night are present, some areas will be more susceptible to frost damage than others. Fields that have lighter soils, which dry out faster, fail to insulate the soil below. This will prevent the natural warming by soil radiation of the air directly above the soil surface and increase the likelihood that a heavy frost will form. Recently cultivated fields will suffer the same fate. Low-lying valleys, where cold dense air cannot be affected by winds, can be heavily damaged. Areas of a field that are close to the edge of a crop formation are also susceptible. In these areas, grass near a crop formation acts as a blanket or insulator, preventing warm soil from heating cold air directly above it. Finally, areas that have recently been treated with herbicides are more susceptible to frost damage. Herbicide stress on a plant can be compounded by cold stress from the weather, increasing the possibility of frost damage. Frost damage is evidenced by a variety of symptoms on plants. Plant leaves are the best indicators of frost damage. Frost damage on the youngest leaves of a plant (the top leaves) is often seen as a burn on the tips of the leaves. More severe damage is a darkening of the entire leaf. In the most extreme form, the entire plant takes on a black appearance. This darkening is evidence that the frost penetrated and destroyed cell membranes of the plant. Severe darkening of this kind is sometimes referred to as a ‘‘killing frost.’’ Whether a killing frost has set in on plant life may not be noticeable until a day or two after the frost. If the plant has turned almost completely brown, chances of recovery are not good. However, a closer look may reveal that the lower part of the plant, or the pseudostem, is still green, a good sign that some recovery from the damage may be possible. The three most important factors in determining the ability of plant life to withstand damaging frost are the plant’s maturity, its health prior to a frost, and the weather immediately following the frost. Susceptibility to frost damage increases as plant development increases. As the growing season progresses, the chances of weather conditions conducive to a damaging frost decrease. Young plants are less susceptible to frost damage that will lead to a plant’s death because their growing points are still below ground, insulated from freezing temperatures. However, should frost injury occur at this young stage, it could severely delay growth as the season progresses and affect the overall harvest. More mature plants present more opportunities for frost damage. Mature plants have more exposed leaves and may have growing points above the earth’s surface. Damage to the outside of a mature plant can constrict future growth. Plant health prior to a frost also determines the ability of a plant to recover. If plants

have been continually exposed to cold stress, herbicides, excessive moisture, or disease, even the most minimal frost can be debilitating. Finally, the weather following frost damage plays an important role in the plant’s recovery. If warm temperatures follow frost damage, a plant’s ability to recover increases. Worldwide, 5–15% of all agricultural production is lost to frost damage each year. Frost can also cause a loss of food supplies for an animal species by killing leaves, seeds, and fruits. READING LIST www.epa.gov—United States Environmental Protection Agency—Office of Pesticide Programs. www.sciencedaily.com/releases/2000/10/001017073120.htm— Science Daily Magazine—Climate Change Shifts Frost Seasons & Plant Growth. www.agric.gov.ab.ca/crops—Alberta Agriculture, Food, And Rural Development—Frost Damage to Cereals. www.pioneer.com/usa/crop—Pioneer Hi-Bred International, Inc.—Microclimatic Effects on Frost Damage, Early Season Frost Damage to Corn. www.pir.sa.gov.au/pages/agriculture—PIRSA Agriculture— Frost Risk Assessment and Damage. Carter, P.R. (1995). Late spring frost and post-frost clipping effect on corn growth and yield. J. Prod. Agric.. Kunkel, K.E. and Hollinger, S.E. (1995). Late spring freezes in the central USA: Climatological aspects. J. Prod. Agric. Nielsen, R.L. (1999). Assessing frost damage to young corn. Purdue pest management and crop production newsletter. Purdue Univ., 27 May 1999.

THE GLOBAL WATER CYCLE U.S. Global Change Research Program

USGCRP-supported research on the global water cycle focuses on: (1) the effects of large-scale changes in land use and climate on the capacity of societies to provide adequate supplies of clean water; and (2) how natural processes and human activities influence the distribution and quality of water within the Earth system and to what extent the resultant changes are predictable. Specific areas include:

This article is a US Government work and, as such, is in the public domain in the United States of America.

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identifying trends in the intensity of the water cycle and determining the causes of these changes (including feedback effects of clouds on the global water and energy budgets as well as the global climate system); predicting precipitation and evaporation on timescales of months to years and longer; and modeling physical/biological processes and human use of water, to facilitate efficient water resources management. The USGCRP budget includes $311 million in FY 2003 for research and observations related primarily to the Global Water Cycle. The Global Water Cycle program studies the movements and transformations of water, energy, and water-borne materials through the Earth system and their interactions with ecosystems. The movements and transformations of water are important because they appear to control the variability of the Earth’s climate and they provide an essential resource for the development of civilization and the Earth’s environment. Figure 1 schematically illustrates the movements and transformations. This cycling involves water in all three of its phases—solid, liquid, and gaseous—and exchanges large amounts of energy as water moves and undergoes phase changes. Therefore, the water cycle operates necessarily on a broad continuum of time and spatial scales.

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Water vapor is a greenhouse gas that maintains temperatures in a range required by life on Earth. Many of the uncertainties in the current projections of the effects of the atmospheric buildup of carbon dioxide are related to the feedbacks between the climate and the water cycle. While warmer temperatures enable the atmosphere to hold more water leading to further warming, the complex interactions among changing cloudiness, precipitation patterns, land cover, and decreasing snow and ice cover have limited the quantitative understanding of the links between water and climate warming. Water is not evenly distributed over the globe, nor is it always accessible for human use. Society is becoming more vulnerable to variations in the water cycle as a result of expanding populations and increasing water use. The increasing demands for water accompanied by the growing economic losses from droughts and floods place pressure on the science community to develop the knowledge and tools needed to manage our limited water resources more effectively. There are large potential paybacks from increased investments in scientific research to improve the monitoring and prediction of the global water cycle variations and in water management applications.

Figure 1. Conceptualization of the water cycle.

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On a national basis, near-crisis situations have occurred in several dry southwestern river basins, including the Colorado and Rio Grande, where overallocation has taken place. Recent drought conditions and rapid development in these basins have exposed the intensity of competition that exists over the available water resources. The development of a capability to predict where water management crises will emerge due to a drought or extended flood conditions is a priority for the Global Water Cycle program. The ability to provide probabilistic forecasts of rainfall and snowfall at various time and space scales is at the center of all potential applications of climate change science and climate information systems. The program has research activities directed at developing experimental predictions that will ultimately benefit society through better protection of human health and assets, and more efficient water system management and infrastructure planning. Human activity is an integral part of the water cycle. A recent USGCRP-commissioned report, A Plan for a New Science Initiative on the Global Water Cycle, issued in 2001, concluded that, among other priorities, there is a pressing need to determine the causes of water cycle variations on both global and regional scales, and to what extent these variations are induced by human activities. In view of this emerging link between water science and water resource issues, the USGCRP global water cycle strategic plan addresses two major questions: (1) What are the effects of large-scale changes in land use and climate on the capacity of societies to provide adequate supplies of clean water, and (2) how do natural processes and human activities influence the distribution and quality of water within the Earth system and to what extent are resultant changes predictable? Stakeholders are helping to define the Global Water Cycle program at the catchment and larger river basins scales. Users are interested in better forecasts of precipitation, runoff, and soil moisture. Reservoir management decisions require forecast lead times of up to seasons and, in some cases, years. For planning reservoirs, dam recommissioning, and water control infrastructure, and developing new proposals for water law, projections of water variability are required on the decadal to century timescales. The USGCRP Global Water Cycle program focuses on characterizing, explaining, and predicting variability and long-term changes in the global water cycle and their impacts. To address the issues arising from the intimate role of the water cycle in controlling climate variability on seasonal to multidecadal timescales, the program investigates the pathways of water movement between the biosphere and surface hydrologic systems, the atmosphere, and the oceans, as well as feedback processes between climate, weather, and biogeochemical cycles. Because the biosphere is a substantial regulator of the Earth’s carbon cycle, the global water cycle maintains a considerable influence upon the global pathways of carbon. Globally, the cycling of water and its associated energy and nutrient exchanges among the atmosphere, ocean, and land determine the Earth’s climate and cause much of climate’s natural variability.

A critical contribution of the USGCRP to Federal water activities lies in the benefits that come from drawing together the wide range of programs and expertise from different agencies with the capabilities of the academic community to address these complex issues. The elements of the management structure that the USGCRP has put in place during the past year include: (1) Interagency Global Water Cycle working group, (2) Global Water Cycle scientific steering group, and (3) Global Water Cycle program office. The linkages between the global water cycle, the global carbon cycle, and climate will be explored in the coming year through this strengthened program management structure. SEE ALSO: Water Cycle [also available: PDF Version]. Chapter 5 from the Strategic Plan for the Climate Change Science Program (July 2003). See also the draft white paper, The Global Water Cycle and Its Role in Climate and Global Change [PDF] (posted 27 Nov 2002). Water Cycle. Presentation from Breakout Session 8 of the US Climate Change Science Program: Planning Workshop for Scientists and Stakeholders, 3–5 December 2002, Washington, DC. Climate Variability—Atmospheric Composition— Water Cycle. Presentation from Breakout Session 19 of the US Climate Change Science Program: Planning Workshop for Scientists and Stakeholders, 3–5 December 2002, Washington, DC.

GROUND-BASED GPS METEOROLOGY AT FSL SETH I. GUTMAN KIRK L. HOLUB NOAA Forecast Systems Laboratory

INTRODUCTION Water vapor is one of the most important components of the Earth’s atmosphere. It is the source of precipitation, and its latent heat is a critical ingredient in the dynamics of most major weather events. As a greenhouse gas, water vapor also plays a critical role in the global climate system: it absorbs and radiates energy from the sun and affects the formation of clouds and aerosols and the chemistry of the lower atmosphere. Despite its importance in climate and weather prediction, water vapor has been one of the most poorly measured and least understood components of the Earth’s atmosphere. Researchers at FSL and elsewhere are utilizing recent technology to reverse this situation. The ability to use the Global Positioning System (GPS) to make accurate refractivity measurements under all weather conditions has led to the development of a promising new meteorological observing system for NOAA. The This article is a US Government work and, as such, is in the public domain in the United States of America.

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first and most mature application of ground-based GPS meteorology involves the measurement of integrated (total column) precipitable water vapor (IPW) in the atmosphere. The GPS-IPW technique is more advantageous than conventional water vapor observing systems because of its low-cost, high-measurement accuracy, all weather operability, and long-term measurement stability. Further, GPS-IPW requires no external calibration, operates unattended for long periods with high reliability, and is easily maintained. Since GPS-IPW measurements are compatible with satellite data retrievals, they provide an independent method for calibrating and validating global satellite observations. These positive attributes, however, are accompanied with one major disadvantage: GPS-IPW provides no direct information about the vertical distribution of water vapor in the atmosphere. In an attempt to mitigate this deficiency, researchers at government laboratories and universities around the world are investigating the best ways to use GPS-IPW as a ‘‘proxy quantity’’ for moisture profiles in weather forecasting. In this article we discuss how IPW is now calculated from GPS signal delays and the potential use of slant-path measurements in numerical weather prediction models. Preliminary results of the effect of GPS-IPW on numerical weather prediction, the demonstration network, data and product availability, and plans for the operational network are also described. CALCULATING IPW FROM GPS SIGNAL DELAYS GPS signals are delayed as they pass through the Earth’s atmosphere (Fig. 1). The signal delay caused by the presence of free electrons in the ionosphere makes the largest contribution to the total atmospheric delay. Because the ionosphere is a dispersive medium, the velocity of the GPS signals is frequency dependent and its impact can be effectively eliminated by using dual frequency receivers. Below the ionosphere, in the electrically neutral portion of the atmosphere, refraction (that is, slowing and bending) of the GPS signal is caused by changes in temperature,

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pressure, and water vapor. Most of this delay occurs in the troposphere, which extends from about 9 km at the poles to about 13 km at the equator. The primarily tropospheric delay consists of a hydrostatic (or dry) component caused by the mass of the atmosphere and a wet component (the wet delay) caused by the dipole moment of the water vapor molecule. The contributions of the wet and dry components of the tropospheric signal delay are in the same proportion as the wet and dry components of the atmosphere. FSL currently collects GPS observations from a demonstration network of 55 sites (Fig. 2) and processes them to produce IPW measurements every 30 minutes using the scheme shown in Fig. 3. The first step in obtaining IPW from GPS observations is to determine the zenith-scaled delay caused by the neutral atmosphere. This delay is commonly referred to as the zenith tropospheric delay (ZTD), and is calculated from carrier phase and range observations made by networks of GPS receivers. The calculation is made using GPS analysis software such as GAMIT (GPS At MIT), which in addition to the GPS observations, requires improved satellite orbits and parameters describing the orientation of the Earth in space and time. Next, the ZTD is separated into its wet and dry components using additional observations made by collocated surface meteorological sensors. The zenith- scaled hydrostatic delay (ZHD) is caused by the mass of the atmosphere directly above the site and can be estimated with great accuracy from a surface pressure measurement. The wet signal delay (ZWP) is caused by water vapor along the paths of the radio signals to all satellites in view, about 6 to 8 with the current GPS satellite constellation. ZWP is calculated simply by subtracting the hydrostatic delay from the tropospheric delay. The resulting wet delay can be mapped into IPW with an error of about 5 degrees using a quantity that is proportional to the mean vapor pressure-weighted temperature of the atmosphere (Tm). Tm may be estimated from a climate model, the surface temperature derived from a numerical weather prediction model, or measured directly using remote sensing techniques. FSL is planning to utilize modelderived Tm estimates operationally.

Figure 1. Signal delays caused by the atmosphere.

Figure 2. A map of the NOAA-FSL Global Positioning System Integrated Precipitable Water (GPS-IPW) Demonstration Network (55 sites) as of October 1999.

Figure 3. The FSL-developed data processing scheme used to produce IPW measurements. 246

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Figure 4. Scatterplot of GPS and rawinsonde observations of integrated precipitable water vapor at the ARM CART site near Lamont, Oklahoma, between January 1996 and September 1999.

Integrated precipitable water calculated from GPS signal delays is physically identical to integrated measurements or retrievals made by other upper-air observing systems including rawinsondes, ground-based microwave water vapor radiometers, or satellite microwave and infrared instruments including sounders and interferometers. Comparisons of GPS and radiosonde-derived total column water vapor have been carried out continuously since 1996 under all weather conditions at the DOE Atmospheric Radiation Measurement site near Lamont, Oklahoma. Results from 3600 comparisons (to September 1999) indicate a mean difference of 0.08 mm and a standard deviation of 2.1 mm (Fig. 4). SLANT-PATH SIGNAL DELAY MEASUREMENTS Recent investigations by FSL director A.E. MacDonald and Yuanfu Xie (of the Forecast Research Division) of the potential use of line-of sight estimates of path-integrated water vapor (derived from slant-path GPS signal delay

measurements) to retrieve the 3-D moisture field have been very interesting and potentially significant. The experiments involve assimilating simulated slant-path moisture measurements from a wide area network of closely spaced stations into the Quasi-Nonhydrostatic (QNH) model using variational techniques. In recent research, their simulations indicate that it may be possible to recover the three- dimensional structure of the moisture field from a densely spaced network of ground-based GPS receivers making a single line-of- sight, or slant path, measurement of the signal delay to all satellites in view. The configuration of the GPS satellite constellation as seen from Boulder, Colorado, between 1200 and 1300 UTC on 28 September 1999 is shown in Fig. 5. A GPS satellite moves across the sky at the rate of about 30 degrees per hour. Although 10 satellites are visible above the horizon in this example, six to eight would be more typical at any one time. Making a slant-path signal delay measurement with the same accuracy as a zenith-scaled measurement is

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Figure 5. Configuration of the GPS satellite constellation as seen from Boulder between 1200 and 1300 UTC 28 September 1999.

not trivial. The sources of measurement error that are successfully managed through geodetic modeling of the zenith tropospheric signal delay will have to be dealt with in other ways. Although some of the most important information about the structure of the atmosphere can be obtained from low-angle observations, measurement errors increase significantly along with the negative impact of multipath reflections from nearby obstacles as satellites approach the horizon. One way to mitigate these problems is to utilize advanced GPS receivers and antennas that maximize the ability to track satellites under all conditions and reject multipath reflections. Unfortunately, not all problems can be eliminated through the selection of hardware, and advanced data processing techniques will be needed as well. Research at Scripps Institution and the University of Hawaii into ways to monitor the accuracy of GPS orbit

predictions suggests that these techniques can also be used to reduce systematic errors in slant-path signal delay or refractivity measurements to individual satellites. EFFECT OF GPS-IPW DATA ON THE ACCURACY OF NUMERICAL WEATHER PREDICTION Since 1997, parallel runs with and without GPS have been carried out using the research version of the Rapid Update Cycle (RUC-2) model to assess how GPSIPW data affect the accuracy of numerical weather prediction. Results from the first two years using optimal interpolation techniques have been encouraging despite the fact that the observations came from only a limited number of widely spaced sites. Model runs using data acquired from more sites over a larger area through September 1999 confirm improvements

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Figure 6. GPW-IPW installation at the NOAA Profiler Network site at Platteville, Colorado.

in forecast accuracy, especially under conditions of active weather. Therefore, NOAA meteorologists expect significant improvements in short-term forecasts of clouds, precipitation, and severe weather when high-resolution numerical weather prediction models routinely use data from a nationwide network of GPS-IPW systems in conjunction with data from other observing systems and advanced data assimilation techniques. The decision to implement ground-based GPS Meteorology (GPS-Met) as a next-generation upper-air observing system will be supported in part by promising assessments such as this one. In anticipation of a favorable decision, network design and implementation options for a national network of ground-based GPS- IPW systems are being evaluated at FSL. GPS-IPW DEMONSTRATION NETWORK The rapid development of the GPS-IPW Demonstration Network for meteorological remote sensing has been made possible by a fortuitous synergy with the positioning and

navigational applications of GPS by the U.S. Coast Guard and U.S. Department of Transportation. As of October 1999, the data acquisition component of the demonstration network consisted of 55 GPS-IPW systems operating in the continental United States and Alaska. Thirty-four systems are currently installed at NOAA Profiler Network (NPN) sites, seven at sites belonging to other NOAA organizations or institutions affiliated with NOAA, 11 belong to the U.S. Coast Guard Maritime Differential GPS (DGPS) system, and three are at the Department of Transportation Nationwide Differential GPS facilities. Typical sites from each organization are illustrated in Figs. 6–9. In addition to supporting the assessment of GPS as a possible next-generation upper-air observing system, the GPS-IPW Demonstration Network is designed to help NOAA accomplish the following tasks: • Evaluate the engineering and scientific bases of ground- based GPS-Met, including advanced data acquisition and processing techniques.

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Figure 7. GPS-IPW installation at the Scripps Institution of Oceanography at La Jolla, California.

• Develop and test strategies to build, monitor, operate and maintain large networks of GPS reference stations for meteorological remote sensing. • Develop techniques to acquire, process, and quality control GPS observations and data products. • Provide observations and derived meteorological products to users (such as forecasters, modelers, researchers) and data archives. • Transfer ground-based GPS-Met technologies to operational use in weather forecasting and climate monitoring. • Other possible applications under investigation include calibration and validation of environmental satellite data and improved positioning and navigation services. All ground-based observing systems in the GPSIPW Demonstration Network consist of dual-frequency GPS receivers and antennas, and collocated surface meteorological sensors. These systems are located at sites where shelter, power, and communications are available to operate and collect

data from the instruments, and transmit these data in real or near real time to one of two locations. The generalized flow of data and products from the network is illustrated in Fig. 10. DATA AND PRODUCT AVAILABILITY GPS and surface meteorological observations from the GPS-IPW Demonstration Network sites are available to the general public in near real time through the NOAA National Geodetic Survey. Information and raw data may be acquired via the Web. Processed data, including GPS signal delays and integrated precipitable water vapor, are available shortly after improved NAVSTAR GPS satellite orbits and Earth Orientation Parameters are available from one of the International GPS Service (IGS) tracking stations. This usually occurs within 24 hours of the close of the day, but efforts to accelerate the process and make improved orbits available within 1–3 hours are well underway. IPW and other products may be acquired from the FSL Demonstration Division, GPS-Met Observing Systems Branch.

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Figure 8. GPS-IPW installation at the U.S. Coast Guard Differential GPS site at Cape Canaveral Air Force Station.

PLANS FOR NETWORK EXPANSION AND IMPROVEMENT Our primary goals in 2000 are to continue to expand the demonstration network, demonstrate distributed data processing using low-cost PCs instead of high-end workstations, and implement real-time data processing. Expansion of the Demonstration Network. Now that all NOAA Profiler Network sites have been equipped with GPS-IPW systems, expansion of the network will mostly proceed by installing GPS Surface Observing System (GSOS) packages at U.S. Coast Guard Maritime DGPS and Department of Transportation NDGPS sites. Depending on the availability of funds and the status of interagency agreements under review, during the next year we hope to install 21 new systems at DGPS sites (mostly in the Mississippi Valley and Great Lakes regions), 11 NDGPS sites, and one at the Department of Energy ARM facility at Point Barrow, Alaska (Fig. 11). Implementation of Real-time Data Processing. We define real-time data processing as acquiring and processing GPS and ancillary observations to

yield signal delay or IPW calculations within a single numerical weather prediction assimilation cycle. In the case of the Rapid Update Cycle, running operationally at the National Centers for Environmental Prediction, this is approximately 75 minutes. Real-time data processing techniques are being tested and evaluated in a collaborative effort involving FSL, the Scripps Permanent Orbit Array Center, and the University of Hawaii at Manoa. Techniques involve acquiring data from a subset of the IGS global tracking network and using these observations to produce an improved retrospective orbit with only about 2-hour latency. An orbit prediction that covers the data gap is also made, and it is this short-term prediction that is used to calculate IPW. In theory, the error in a prediction that spans only 2 or 3 hours will be proportionally less than an error made over an interval of 36–48 hours. Real-time quality control techniques are also under evaluation. The most promising involve continuous monitoring of the relative positions of a number of sites, and using these data to infer changes in orbit

Figure 9. GPS-IPW installation at the DOT National Differential GPS site at Whitney, Nebraska.

Figure 10. Flow of data and products from the GPS-IPW Demonstration Network.

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Figure 11. Planned expansion of the Demonstration Network to about 92 sites during 2000.

Figure 12. Expected configuration of the NOAA/FSL GPS-IPW Demonstration Network by 2005. Sites in Hawaii and the Caribbean Sea are not shown.

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Figure 13. An Automated Surface Observing System (ASOS) installation at Cape Hatteras, North Carolina.

Figure 14. GPS receiver placement on top of FSL’s new office building, the David Skaggs Research Center.

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Figure 15. A map of the AWIPS offices.

accuracy for specific satellites in the constellation. When a problem is encountered, the satellites are removed temporarily from the ephemerides until an updated orbit can be produced. Distributed Data Processing. Recent advances in lowcost PC processor speed and memory will be utilized to perform data processing in a fully distributed environment. During this year we have demonstrated the ability to partition a large network into smaller subnetworks, process each independently in substantially less time with no significant loss of accuracy and precision. Operational GPS-IPW Network Implementation Strategy. The expansion of the GPS-IPW Demonstration Network to about 200 sites, and the transition from retrospective to real-time data processing will enable us to assess the impact of these data on weather forecast accuracy. Based on the results of these studies, a decision to implement ground-based GPS-IPW as a next-generation upper-air observing system for NOAA is expected. The following plan has been developed to expand the demonstration network to an operational network of about 1000 sites with an average station spacing of somewhat less than 100 km (Fig. 12). • Once transition of the GPS-IPW Demonstration Network to operational status has become a reality, receivers and antennas will be upgraded. Communications will transfer from FTS-2000 to the AWIPS [Advanced Weather Interactive Processing System] communications systems via the Internet. • GPS receivers will be added to about 800 Automated Surface Observing (ASOS) sites (sample site shown in

Fig. 13). The reason for collocating GPS at ASOS sites is to take advantage of the surface meteorological data and site infrastructure, including shelter, power, data communications, field maintenance, and logistics support. This will minimize implementation time and life-cycle cost. The GPS antenna installation at a typical ASOS site will resemble the one at FSL’s new location, the David Skaggs Research Center (Fig. 14). • Data processing hardware, software, and training will be provided to all AWIPS offices (Fig. 15). [Editor’s Note: More information on the topics covered here is available by contacting Seth Gutman, who can provide copies of published articles which include a list of references.]

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII JAMES O. JUVIK D.C. SINGLETON G.G. CLARKE University of Hawaii Hilo, Hawaii

INTRODUCTION The island of Hawaii, with a surface area of only 10,455 km, exhibits a spectacular range of climatic This article is a US Government work and, as such, is in the public domain in the United States of America.

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Figure 15. A map of the AWIPS offices.

accuracy for specific satellites in the constellation. When a problem is encountered, the satellites are removed temporarily from the ephemerides until an updated orbit can be produced. Distributed Data Processing. Recent advances in lowcost PC processor speed and memory will be utilized to perform data processing in a fully distributed environment. During this year we have demonstrated the ability to partition a large network into smaller subnetworks, process each independently in substantially less time with no significant loss of accuracy and precision. Operational GPS-IPW Network Implementation Strategy. The expansion of the GPS-IPW Demonstration Network to about 200 sites, and the transition from retrospective to real-time data processing will enable us to assess the impact of these data on weather forecast accuracy. Based on the results of these studies, a decision to implement ground-based GPS-IPW as a next-generation upper-air observing system for NOAA is expected. The following plan has been developed to expand the demonstration network to an operational network of about 1000 sites with an average station spacing of somewhat less than 100 km (Fig. 12). • Once transition of the GPS-IPW Demonstration Network to operational status has become a reality, receivers and antennas will be upgraded. Communications will transfer from FTS-2000 to the AWIPS [Advanced Weather Interactive Processing System] communications systems via the Internet. • GPS receivers will be added to about 800 Automated Surface Observing (ASOS) sites (sample site shown in

Fig. 13). The reason for collocating GPS at ASOS sites is to take advantage of the surface meteorological data and site infrastructure, including shelter, power, data communications, field maintenance, and logistics support. This will minimize implementation time and life-cycle cost. The GPS antenna installation at a typical ASOS site will resemble the one at FSL’s new location, the David Skaggs Research Center (Fig. 14). • Data processing hardware, software, and training will be provided to all AWIPS offices (Fig. 15). [Editor’s Note: More information on the topics covered here is available by contacting Seth Gutman, who can provide copies of published articles which include a list of references.]

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII JAMES O. JUVIK D.C. SINGLETON G.G. CLARKE University of Hawaii Hilo, Hawaii

INTRODUCTION The island of Hawaii, with a surface area of only 10,455 km, exhibits a spectacular range of climatic This article is a US Government work and, as such, is in the public domain in the United States of America.

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diversity comparable with that found on large continents. Three major factors contribute to this climatic diversity: 1. Topographic relief. The volcanic mountains of Mauna Kea and Mauna Loa reach summit elevations of 4,205 m and 4,168 m, respectively. The attitudinal range provides for a diversity of temperatures, and the mountains themselves are barriers that induce orographic precipitation. 2. Large-scale synoptic wind field. The strong and persistent northeast trade winds interact with the island topography to produce distinctive windward and leeward climates. The associated upper-level trade wind inversion exerts a particularly strong control on mountain precipitation gradients. 3. Local circulation. Differential heating and cooling of the land, water, mountain, and lowland areas on Hawaii give rise to localized wind regimes which add to the island’s climatic diversity. KOPPEN CLIMATIC ZONES Integrating the attitudinal temperature gradients with the annual, seasonal, and spatially variable rainfall

Figure 1. Distribution contours of mean annual rainfall (mm), superimposed on topographic map of the island of Hawaii. (Redrawn from 2,3).

regimes results in a diverse combination of climatic environments. The Koppen climate classification uses monthly temperature and precipitation characteristics in a descriptive system that distinguishes broad regional and global climatic zones. The system has been often criticized for its empirical approach and lack of emphasis on ‘‘dynamic processes’’ (e.g., 1); however, as a ‘‘first approximation’’ the Koppen classification offers useful insights into regional climatic patterns. Four broad Koppen climatic zones are distinguished on the island of Hawaii. They are organized primarily as concentric attitudinal bands on the mountain slopes. Figure 2 illustrates the spatial distribution of these Koppen climatic types on the island. The map was constructed on the basis of temperature (absolute or extrapolated) and precipitation data from 55 island stations. Discussions of the zones follow. Humid Tropical Zone (A climates) Characterized by warm temperatures throughout the year and relatively high annual rainfall, humid tropical climates occupy the lower slopes of the island from sea level to about 450 m (slightly higher in warmer areas of leeward

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Kona). This tropical zone may be further differentiated on the basis of rainfall seasonality. Large-scale synoptic disturbances in winter (mid-latitude cyclonic storms) produce substantial rainfall that is to some extent independent of slope aspect or elevation, and as a result most locations on the island exhibit an absolute winter maximum in rainfall. However, windward areas of Hawaii also receive substantial orographic rainfall throughout the year, with the result that there is no distinct dry season (Af climate). Lowland areas on the island that are transitional in location between windward and leeward receive less orographic rainfall (since they are not oriented normal to trade wind flow) and exhibit a distinctive summer dry season (As climate). Humid summer-dry climates are not common anywhere in the world, since for most tropical locations rainfall is at a maximum in the summer, the result of increased convective instability in the high-sun period. Outside of Hawaii the As climate type occurs only in southern Madras (India) and adjacent northern Sri Lanka. The leeward or Kona coast of Hawaii contains the only extensive area of summer maximum rainfall in the Hawaiian archipelago (Aw, winter-dry climate). Isolated from the prevailing trade wind flow by intervening high mountains, the Kona coast’s dominant circulation pattern is formed by a localized land-sea breeze regime. Increased land surface temperatures in summer strengthen the daily sea breeze regime and increase convective instability, leading to a high frequency of afternoon thundershowers. The vertical structure necessary for thundershower

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development is further assured by the high mountains, which exclude the trade wind aloft and limit the potential for strong vertical wind shear. The presence of a strong shearing force would otherwise tend to destabilize these leeward convectional cells. Although there is generally a summer rainfall maximum throughout Kona, the Aw climate gives way to Af at elevations above 400 m, where, by virtue of general orographic position, there is adequate precipitation in all months. Arid and Semi Arid Zones (B climates) A classic rain shadow desert exists on the leeward side of the Kohala mountains. Smaller and lower (maximum elevation 1,670 m) than Mauna Kea and Mauna Loa, the Kohala mountains are incapable of blocking out trade wind flow to leeward. Having become depleted of moisture during windward ascent, the trades warm adiabatically to leeward, promoting a hot, arid zone. With only 190 mm of annual rainfall, Kawaihae on the leeward Kohala coast is the driest location in the Hawaiian archipelago. The Koppen system distinguishes two climatic subtypes, the true desert (BWh climate) and the semidesert (BSh climate) on the basis of relative aridity. In leeward Kohala the true desert gives way to semi-desert at higher elevations. 1 The Temperate (C) and Polar (E) climates as originally proposed by Koppen were not applied in high-altitude tropical environments, which, because of their orographic complexity,

258

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII

Figure 3. ‘‘Walter’’ climate diagrams for four Hawaii island stations (4).

Temperate Zone (C climates)1 Average air temperature in Hawaii decreases with altitude at the rate of about 0.55øC/100 m (Price, 1973). When the criteria of the Koppen classification are used, at elevations above 400–500 m on the mountain slopes, tropical climates grade to temperate as a result of decreasing average temperatures. As a result of the moderating influence of altitude, almost two-thirds of the ‘‘tropical’’ island of Hawaii possesses a temperate upland climate. The majority of this zone is characterized by warm summers and adequate precipitation in all months (Cfb climate). Except for the absence of a stronger season variability the upland Hawaii climates are analogous to those of similar Koppen designation in Pacific coastal areas of North America. Ascending orographic clouds lack of meteorological data, and absence of strong seasonality, were simply designated as highland climates (H). In more recent global and regional climatic maps of the tropics, highland C and E climatic areas are frequently portrayed in order to show approximate attitudinal analogs for these broad latitudinal climatic zones.

compressed between the rising mountain slope and an upper-air temperature inversion produce frequent ground level mountain fog, an important moisture source for upland vegetation (Juvik and Perreira, 1974). At still higher elevations on Mauna Kea and Mauna Loa (above 2,000 m) there is a tendency toward summer drought. The increased strength and frequency of the trade wind inversion in summer (modal elevation 1,800 m) inhibits the vertical penetration of orographic clouds and precipitation to the higher slopes. This summer-dry zone (Gsb climate) also occurs at a lower elevation in leeward Kohala, Mauna Kea, and Mauna Loa, where summer orographic precipitation is largely absent. Above 2,500 m on both Mauna Loa and Mauna Kea the summer-dry regime changes from warm to cool (Gc climate). Alpine (periglacial) Zone (E climates) Above 3,200-m level on Mauna Kea and Mauna Loa all months have a mean temperature below 10 ◦ C, and the climates are classified as periglacial (ET). Nighttime freezing is common throughout the year. Although it

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII

exhibits a winter maximum, annual rainfall is very low (200–400 mm) and variable. Above the 3,500-m level, winter snowfall accounts for a substantial portion of the seasonal precipitation. Koppen used the 10 ◦ C (warmest month) boundary to separate the C and E climates on the basis that trees will not normally grow where mean temperatures fall below this level. Hence E climates characterized the treeless arctic tundra. The upper tree line of Mauna Kea (3,000 m) corresponds fairly closely to the C/E boundary mapped in Fig. 2. On Mauna Loa the tree line is much lower for edaphic reasons (recent lava).

WATER BALANCE The preceding discussion of Koppen climatic zones on Hawaii provides a general overview of the dramatic range in regional climatic diversity found on the island. However, this descriptive approach says little about the direct linkage of climate to physical and biological processes at the earth/atmospheric interface. An integration of seasonal moisture supply (precipitation) with the evaporation and transpiration demands of the environment (determined primarily by solar energy inputs) provides an index of moisture surplus or deficit. Such indices can illuminate direct process/response relationships between climate and the terrestrial ecosystem. In an initial survey of water balance climatology on the island of Hawaii, Mueller-Dombois (4) constructed a series of climate diagrams 21 stations (see Fig. 3). This type of diagram, popularized by Walter and Lieth (5), portrays seasonal curves of mean monthly temperature and precipitation. According to Muller-Dombois (6) an index of precipitation efficiency is built into the diagrams by making one degree of temperature (Celsius) equal to two millimeters of precipitation in the scaling of the two ordinates. This is based on the assumption that monthly potential evapotranspiration (in millimeters) is roughly equal to twice the mean monthly temperature (7). Wherever the precipitation curve drops below the temperature curve, a drought season is indicated. Thus the graph is transformed into a water balance diagram with the temperature curve interpreted as an index of potential evapotranspiration.

259

A serious problem inherent in this graphing technique is the tendency to approximate evapotranspiration with a simple linear function of air temperature (i.e., the 2:1 ratio). Chang (8,9) has reviewed the problems of temperature-based estimations of potential evapotranspiration and points out that solar radiation rather than temperature is the primary forcing function in the evaporative process. Temperature-based estimates of evaporation implicitly assume a strong correlation between temperature and solar radiation. In Hawaii, as a result of advection and the buffering effect of the surrounding marine environment, there is generally poor correlation between solar radiation and temperature. In Hilo, for example, the range in mean monthly air temperature is only 1.4 ◦ C between June (24.2 ◦ C) and December (22.8 ◦ C). By contrast, the receipt of solar radiation in June (563 Ly; see solar radiation data for 1965 from Ref. 10) is more than twice that in December (263 Ly). It is obvious from the above comparison that temperature-based estimates of evapotranspiration cannot be expected to portray realistically the seasonal fluctuations implied in the radiation data. However, in the absence of a dense network of solar radiation monitoring stations on the island, upon which more sophisticated spatial modeling of evapotranspiration might be based, it is necessary to revert to some form of temperaturederived estimation in a ‘‘first approximation’’ of water balance regimes. Thornthwaite (11) has developed perhaps the most widely adopted method of estimating potential evapotranspiration. His empirical formula is based essentially on air temperature: (1) E = 1.6 (10T/I)a . Potential evapotranspiration E is computed from mean monthly temperature T and an empiric ‘‘heat index’’ I, which itself is an exponential function of temperature; a is a constant. To obtain mean monthly evapotranspiration, the values derived from eq. (1) are corrected for mean daylength and number of days in the month. The Thornthwaite equation, although subject to the general limitations of all temperature-based methods, might be expected to give better results than the Walter method in Hawaii,

Figure 4. Monthly estimated potential evapotranspiration and measured pan evaporation for Hilo and Pahala data.

260

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII

since potential evapotranspiration is expressed as an exponential rather than linear function of temperature. In Fig. 4, monthly values of potential evapotranspiration derived by both the Walter and the Thornthwaite methods have been plotted along with class ‘‘A’’ panevaporation data for Hilo and Pahala. It is evident that Walter grossly underestimates pan evaporation (here assumed to be equal to potential evapotranspiration) and also fails to detect the seasonal rhythm apparent in the pan data. Thornthwaite also underestimates pan evaporation but does so in a fairly consistent manner and achieves a strong covariation with the pan data in seasonal rhythm. This suggests that the Thornthwaite method might be useful in Hawaii if a correction factor could be derived to compensate for the consistent underestimation exhibited in Fig. 4. In Fig. 5, monthly values of the Thornthwaite potential evapotranspiration estimate are plotted against pan evaporation for Hilo and Pahala. With a regression coefficient of 0.844, approximately 71% of the observed variation in pan evaporation can be explained by variation in the Thornthwaite estimate. (The regression coefficient is significant at 0.01 level.) Potential evapotranspiration Y can thus be reasonably predicted from the Thornthwaite values X by the linear regression equation Y = 42.8 + 1.016 (X)

(2)

Before eq. (2) can be applied as a general (islandwide) correction factor for the Thornthwaite potential evapotranspiration estimate, it must be verified that the relationship established in Fig. 5 (for two lowland locations) is equally valid for mid- and high-altitude areas of the island. There are no class ‘‘A’’ pan evaporation data for inland mountain areas of Hawaii with which the lowland-derived correction factor might be compared. However, Juvik and Clarke (12) have accumulated limited experimental data on mountain evaporation gradients in Hawaii Volcanoes National Park on the east flank of Mauna Loa.

Figure 6. Constant-level pan evaporimeter. a) field installation with inner tube reservoir; b) evaporimeter detail. Note foam insulation around evaporation pan.

Figure 5. Relationship between monthly class ‘‘A’’ pan evaporation and estimated monthly potential evapotranspiration (Thornthwaite) for Hilo and Pahala data.

These data were obtained by using four constant-level pan evaporimeters (Fig. 6) situated along an attitudinal transect between sea level and 2,000 m. In Fig. 7, measured mean daily evaporative rates (averages from 133 days of simultaneous readings taken from September 1974 through May 1975) are plotted against elevation. There is a clear linear decrease in evaporation over the attitudinal range surveyed (approximately 0.72 mm/day/1,000 m). Figure 7 also shows the corrected (eq. 2) Thornthwaite potential evapotranspiration values (mean of 9 months, September to May) derived from temperature-recording stations that occur near the evaporimeter transect. There is good agreement between the Thornthwaite and the evaporimeter values (differences range from 1% to 12%), largely because air temperature also decreases linearly with elevation. On the basis of the close agreement in Fig. 7, the corrected Thornthwaite estimate was considered acceptable to use for all areas of the island

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII

Figure 7. Relationship between measured and predicted evapotranspiration along an attitudinal transect in Hawaii Volcanoes National Park on the east flank of Mauna Loa (pan data from 12).

in the derivation of monthly and annual potential evapotranspiration from temperature data. Corrected Thornthwaite estimates of monthly and annual potential evapotranspiration were computed from standard tables (13) and eq. (2), for 30 stations on the island of Hawaii. The evapotranspiration data were then integrated with monthly precipitation values to produce seasonal water balance diagrams (Fig. 8). Because some of the water surplus received in the wet season is stored as soil moisture for utilization during dry periods, the computation of seasonal water balance must incorporate a parameter describing the moisture storage capacity of the soil.

261

On the geologically youthful island of Hawaii, soils are not generally well developed except for limited areas where ash deposits occur in high-rainfall zones. Recent lava flows exhibiting little or no soil development cover substantial portions of the island. The average depth to bedrock for 75 different Hawaii island soil types and subtypes has been calculated at 0.89 m with only moderate variation (14), lending quantitative credence to this stated geological youthfulness of the island. Soil moisture storage capacity has not been well studied for most Hawaiian soil types. A value of 125 mm/m is the average moisture capacity for ten different soil types for which data are available. If this is assumed to be representative, then the average soil moisture storage capacity for all Hawaii island locations would be 111.2 mm (i.e., 0.89 × 125). In the construction of water balance diagrams for all island stations this value was rounded off to 100 mm so that standard moisture depletion tables could be employed in water balance calculations (13). The 30 water balance diagrams constructed for the island depict both steep gradients and pronounced regional differences in seasonal moisture surplus and deficit. In Fig. 8 the difference between annual precipitation and potential evapotranspiration has been mapped in four zones: 1. Annual surplus exceeding 1,000 mm. This zone comprises 20% (2,100 km2 ) of the island area and is restricted to the high-rainfall regions of windward Mauna Kea, Mauna Loa, and the summit area of Kohala. The annual moisture surplus in this zone

262

CLIMATE AND WATER BALANCE ON THE ISLAND OF HAWAII

ranges as high as 3,377 mm (station 10) at middle elevations. All stations within this zone (stations 2, 6, 10, 11, 14, and 15) exhibit an absolute winter maximum in precipitation, and a secondary summer maximum also occurs at elevated stations where summer orographic precipitation is exaggerated (e.g., stations 2 and 14). 2. Annual surplus between 0 and 1,000 mm. This zone comprises 21% (2,200 km2 ) of the island area and extends from middle to high elevations on the windward slopes down to sea level in those areas where slope aspect is not oriented perpendicular to prevailing trade wind flow, and thus the orographic rainfall component is diminished. For the windward stations there is typically a moderate summer drought (stations 5, 7, 9, 13, and 17) from 2 to 5 months long. The increased strength of the trade wind inversion in particular limits summer rainfall at higher elevations. The localized core area of high convectional rainfall in Kona also falls within this moisture zone. However, here the deficit period occurs in winter (stations 24, 28, and 29) and is not severe. 3. Annual deficit between 0 and 1,000 mm. This zone comprises 54% (5,600 km2 ) of the island area and occupies a predominantly leeward location on Kohala, Mauna Kea and Mauna Loa. The drought period may be concentrated in either the summer

(on the windward side) or the winter (on the Kona side) and is typically 6 to 12 months long. 4. Annual deficit exceeding 1,000 mm. This zone comprising 5% (550 km2 ) of the island area is restricted to leeward Kohala and north Kona. The annual moisture deficit may exceed 1,900 mm (station 1). SUMMARY The Thornthwaite water balance diagrams and map demonstrate graphically the tremendous climatic diversity on the island of Hawaii. Although the Koppen map (Fig. 2) shows only a relatively small portion of the island to be arid or semi-arid (BWh and BSh), from the water balance analysis it is evident that nearly 60% (zones 3 and 4 above) of the island experiences an annual moisture deficit. Acknowledgments This research was supported in part by grants from the Hawaii Natural History Association and the U.S. Department of Interior, Office of Water Resources Research.

BIBLIOGRAPHY 1. Carter, D.B. and Mather, J.R. (1966). Climatic classification and environmental biology. Publ. Climatol. 19(4): 305–390. 2. Taliaferro, W.J. (1959). Rainfall of the Hawaiian Islands. Hawaii Water Authority, p. 394.

HEAT OF VAPORIZATION 3. State of Hawaii (1970). An inventory of basic water resources ¨ dataAisland of Hawaii. Rep. R34, Dept. of Land and Natural Resources, Honolulu, Hawaii, p. 188. 4. Mueller-Dombois, D. (1966). Climate. Chap. IV. In: Atlas for Bioecology Studies in Hawaii Volcanoes National Park. Maxwell S. Doty and D. Mueller-Dombois. (Eds.). U.S. National Park Service, p. 507. (Republished as Hawaii Agr. Exp. Sta. Bull. 89, 1970.) 5. Walter, H. and Lieth, H. (1960). Klimadiagram-Weltatlas, Jena. 6. Mueller-Dombois, D. (1976). The Major Vegetation Types and Ecological Zones in Hawaii Volcanoes National Park and Their Application in Park Management and Research. In Proceedings of the First Conference in Natural Sciences, Hawaii Volcanoes National Park, edited by S. W. Smith, Department of Botany, University of Hawaii, Honolulu, 149–161. 7. Gaussen. (1954). 8. Chang, J.-H. (1959). An evaluation of the 1948 Thornthwaite classification. Ann. Assoc. Amer. Geogr. 49(1): 24–30. 9. Chang, J.-H. (1968). Climate and Agriculture. Aldine Publishing Co., Chicago, p. 304. 10. Lof, G.O.C., Duffie, J.A., and Smith, C.O. (1966). World Distribution of Solar Radiation. Rep. 21, Solar Energy Laboratory, Univ. of Wis., p. 59. (plus maps). 11. Thornthwaite, C.W. (1948). An approach toward a rational classification of climate. Geogr. Rev. 38: 55–94. 12. Juvik, J.O. and Clarke, G.G. (1976). Topoclimatic Gradients in Hawaii Volcanoes National Park (abstract). In Proceedings of the First Conference in Natural Sciences, Hawaii Volcanoes National Park, edited by S. W. Smith, Department of Botany, University of Hawaii, Honolulu, p. 113. 13. Thornthwaite, C.W., and J.R. Mather (1957). Instructions and tables for computing potential evapotranspiration and the water balance. Publ. Climatol. 10(3): 1–311. 14. Sato, H.H. et al. (1973). Soil Survey of Island of Hawaii, State of Hawaii, U.S.D.A. Soil Conservation Service, p. 115. (plus maps).

READING LIST Juvik, J.O. and Perreira, D.J. (1974). Fog interception on Mauna Loa, Hawaii. Proc. Assoc. Amer. Geogr. 6: 22–25. Price, S. (1973). Climate. In: Atlas of Hawaii. R.W. Armstrong. (Ed.). University of Hawaii Press, Honolulu, pp. 53–60.

Unita mass phase change H2O (liquid) H2O (vapor)

Initial equilibrium state

263

HEAT OF VAPORIZATION NARAINE PERSAUD Virginia Polytechnic Institute and State University Blacksburg, Virginia

HEAT OF VAPORIZATION The heat of vaporization (here denoted as Lv ) is defined as the heat added (or given off) when unit mass undergoes isobaric phase transformation in any closed two-phase, one-component liquid/vapor system. In engineering and meteorology, Lv is used in a restricted sense to mean the heat of vaporization of the two-phase liquid water/water vapor system. Although much of the subsequent discussion focuses on Lv , for this specific system as an example, the concepts covered are universally applicable to all fluid/vapor systems. As illustrated below (Fig. 1), for the liquid water/water vapor system, Lv represents the heat gained when unit mass of water in the system evaporates in the isobaric phase transformation H2 O (liquid) → H2 O (vapor). For the reverse phase change H2 O (vapor) → H2 O (liquid), i.e., condensation, Lv is lost from the system. This seemingly simple phase transition H2 O (liquid) ↔ H2 O (vapor) is the fundamental driving process of the earth’s hydrological cycle, the working principle of the steam engine that ushered humanity into the industrial revolution along with its (often negative) social and environmental pollution consequences, and the physical mechanism that maintains the body temperature of plants and warm-blooded animals. In general, the state of any closed two-phase, onecomponent system is defined by the state variables temperature (T in ◦ K), saturation vapor pressure (P in Pascal), and volume (V in m3 ). The behavior of any such system (generally termed as PVT systems) is usually represented as a family of experimental constant temperature curves (isotherms) on a P-V coordinate plane called an Amagat–Andrews diagram. The general shape of these experimental isotherms is illustrated below (Fig. 2). For the liquid water/water vapor system, the liquid and vapor phases coexist in equilibrium only at P-V coordinates between 2 and 3 along an isotherm, provided

Final equilibrium state

Final volume > Initial volume Water vapor, mass = m v, Pressure = Pvap, Temperature = T Liquid water, mass = m w, Temperature = T

Water vapor, mass = m v + 1 Pressure = Pvap, Temperature = T Liquid water, mass = m w − 1 Temperature = T

Heat in = L v Joules

Figure 1. Liquid Water/Water Vapor System.

264

HEAT OF VAPORIZATION

Vapor Pressure P (Pa)

4

3

2 T1 (°K) 1 Volume V (m3)

Figure 2. Amagat–Andrews Diagram.

that T1 is above the triple point temperature of water (0.01 ◦ C, i.e., the temperature at which ice, liquid water, and water vapor can coexist in equilibrium), and below the critical temperature (374 ◦ C, i.e., the temperature above which it is impossible to produce condensation by increasing the pressure). Between 2 and 1, the system can exist as vapor only, and as liquid between 3 and 4. Thus the liquid-vapor phase transition at a given temperature can only take place at constant pressure or vice-versa. Consequently, as shown in Fig. 1, isobaric liquid vaporization and condensation is necessarily an isothermal process, implying that the triple-point saturation vapor pressure is fixed (it is 611 Pa), and so is the saturation vapor pressure at the critical temperature (it is 2.21 × 107 Pa = 218.2 atm). Similar isotherms and parameters exist for all liquid/vapor systems. This observed behavior of closed two-phase, onecomponent systems is of course predicted by Gibb’s Phase Rule (1), namely F + N = C + 2, where F = degrees of freedom, i.e., the smallest number of intensive variables (such as pressure, temperature, concentration of components in each phase) that must be specified to completely describe the state of the system; N = number of phases, i.e., distinct subsystems of uniform chemical composition and physical properties; and C = the number of components, i.e., the number of independent chemical constituents meaning those constituents whose concentration can be varied independently in the different phases. In a liquid/vapor system P = 2 and C = 1, and therefore F = 1, implying only one intensive variable is needed to specify the state of the system. Therefore, temperature and pressure cannot be fixed independently. For the liquid water/water vapor system, this means physically that at a given temperature between the triple point and critical temperature, water vapor will evaporate or condense to achieve the equilibrium saturation vapor pressure as would be evidenced in a complete Amagat–Andrews diagram for water (2,3). The earth’s atmosphere and oceans can be considered as a vast closed two-phase liquid water/moist air system. Consequently, for most practical engineering and meteorological applications, one is interested in the heat of vaporization of the liquid water/moist air system rather than a pure liquid water/water vapor system. Fortunately, the presence of the other gases (collectively called dry air) in the liquid water/moist air system has negligible effect on the saturation vapor pressure. The reason is that the dry air component in the liquid water/moist air system remains unchanged and is always in the gaseous state

during phase transition at temperatures and pressures of practical interest. Therefore, it can be considered as a closed subsystem as opposed to the open liquid water and water vapor subsystems. Consequently, results obtained from an analysis of the thermodynamics of the pure system are applicable to the natural liquid water/moist air system. Energy conservation required under the first law of thermodynamics implies that heat (Q) exchanged reversibly with the surroundings between equilibrium states of any closed two-phase PVT system is consumed by any internal energy change (
U) of the liquid and vapor phases associated with the mass change from one phase to the other, and any mechanical work (±P
V) realized as the volume of the system increases (positive work) or decreases (negative work). Stated mathematically, Q =
U + P
V, or in a differential form, δQ = dU + pdV. Here, P is the saturation vapor pressure (Pvap in Fig. 1). As entropy (S) is defined as Q/T, then δQ = TdS. The first law can therefore be restated in terms of exact differentials as TdS = dU + PdV. Dividing by dV at constant T and rearranging gives dU/dV = T(dS/dV) − P. Using the Maxwell relation (∂S/∂V)T = (∂P/∂T)V (1), dS/dV can be replaced (for fixed T) by dP/dT. The equation becomes dU/dV = T(dP/dT) − P. At a given pressure and temperature, the internal energy of the system (U in Joules) can be partitioned as mw uw + mvap uvap , where mw , mvap and uw , uvap represent the masses and the specific internal energies (internal energy per unit mass in J kg−1 ) of the water and water vapor in the system. Similarly, the volume (V in m3 ) of the system can be partitioned as mw vw + mvap vvap , where vw , vvap represent the specific volume (volume per unit mass in J kg−1 ) of the water and water vapor in the system. If, as illustrated in Fig. 1, the system internal energy changes by
U from U to U +
U as a result of Lv Joules of heat absorption to convert unit mass of water to water vapor, then U +
U = (mw − 1)uw + (mw + 1)uvap , and therefore
U = (uw − uvap ). Similar reasoning shows that, if the volume changes from V to V +
V in the process, then
V = (vvap − vw ). The mechanical work because of volume change is P
V, where P (the saturation vapor pressure) is a constant at a fixed temperature. Therefore, the heat absorbed (or released) by the system for isobaric phase transition of unit mass in the liquid water/water vapor system (Lv by definition) =
U + P
V. Dividing by
V gives
U/
V = (Lv /
V) − P. Substituting
V = (vvap − vw ) gives
U/
V = [Lv /(vvap − vw )] − P, or in a differential form, dU/dV = [Lv /(vvap − vw )] − P. [It should be noted that because enthalpy (H) is defined as H = U + PV, then
U + P
V =
H, and therefore Lv is the same as the specific enthalpy change (
h =
H per unit mass) for phase transition of unit mass in the liquid water/water vapor system.] Combining the results for dU/dV from the two previous paragraphs gives T(dP/dT) − P = [Lv /(vvap − vw )] − P, and therefore T(dP/dT) = Lv /(vvap − vw ), which can be rearranged to obtain the general forms of the Clapeyron (Emile Clapeyron, 1799–1864) equation dP/dT = Lv /[T(vvap − vw )] =
h/[T(vvap − vw )] or Lv = [T(vvap − vw )]dP/dT.

HYDROLOGIC HISTORY, PROBLEMS, AND PERSPECTIVES

The Clapeyron equation can be used to obtain Lv at a given temperature T for any liquid provided one can obtain values of (vvap − vw ) and an accurate representation of dP/dT. Values of (vvap − vw ) can be obtained from tabulated measurements (3). Alternatively, because vvap vw at the low pressures, (vvap − vw ) can be taken as equal to vvap . Assuming further that at low pressures water vapor behavior closely approximates that of an ideal gas, then vvap = RT/P, where R is the specific gas constant for water = 8.314/0.018 = 461.9 J kg−1 ◦ K−1 . The Clapeyron equation becomes Lv = (RT 2 /P)dP/dT, and this form is referred to as the Clausius–Clapeyron equation (Rudolph Clausius, 1822–1888). Unfortunately, no single function has been shown to adequately represent vapor pressure data for various liquid/vapor systems over wide ranges of T. For the liquid water/water vapor system, values of dP/dT for a given value of T can be obtained by finite differencing of tabulated values. As an example, consider using the above equation to estimate Lv for water at human body temperature of 37 ◦ C. Tabulated values are vw = 1.007 cm3 g−1 = 0.01007 m3 kg−1 , vvap = 22,760 L kg−1 , = 22.760 m3 kg−1 , P = 5.940 kPa at 36 ◦ C and 6.624 kPa at 38 ◦ C, which gives (vvap − vw ) = 22.760 − 0.01007 = 22.75 m3 kg−1 . By finite differencing, dP/dT ≈
P/
T = (6.624 − ◦ 5.940)/(38 − 36) = 0.342 kPa ◦ C−1 = 0.342 kPa K−1 = ◦ −1 ◦ (a temperature difference of 1 C is the 342 Pa K same as a difference of 1◦ K). At T = 37 + 273.16 = ◦ 310.16◦ K, Lv = 310.16◦ K (22.75 m3 kg−1 ) ×342 Pa K −1 = −1 2399 kJ kg . The actual value (tabulated as
h) is 2414 kJ kg−1 , an error of 100 m high and combined waterfalls totaling a height of 300 m (19). The native shrimp, snails, and prawns (Table 1) also have the ability to climb waterfalls (20). Postlarval migration has been found to be directly related to stream discharge (22,23). OTHER ENDEMIC STREAM ANIMALS The native stream organisms have closely related marine relatives, and it is evident that the freshwater species

Amphidromy (Fish, shrimp, snails, prawns)

1. Adults live and breed in the streams 5. Some species use a ventral sucker to climb waterfalls to reach quality breeding habitat

4. Embryos develop into post-larvae in the ocean and begin to migrate back up into the stream mouths

2. Eggs deposited on stream substrate

3. Embroys hatch from eggs and are carried by stream current to the ocean for Growth and development

Figure 2. A schematic depicting the amphidromous life cycle of the native fish, snails, shrimp, and prawns found in Hawaiian streams. The life cycle requires postlarval development in the ocean and a migration back into the streams that is not immediately for breeding.

804

A HISTORY OF HAWAIIAN FRESHWATER RESOURCES

Table 1. The Freshwater Animals of Hawaiian Streams Including the Common Name, Scientific Name, Hawaiian Name, and Geographic Range Taxa Type Fish

Snails

Sponge Crustaceans

Insects

Common Name

Scientific Name

general term for freshwater fishes freshwater amphidromous gobies

Lentipes concolor

freshwater amphidromous eleotrid euryhaline-Hawaiian flag-tail fish euryhaline-striped or gray mullet freshwater amphidromous snail

Sicyopterus stimpsoni Awaous guamensis Stenogobius hawaiiensis Eleotris sandwicensis Kuhlia sandvicensis Mugil cephalus Neritina granosa

estuarine amphidromous snail

Neritina vespertina

brackish/marine neritid snail lymnaeid snails freshwater sponge general term for freshwater shrimp freshwater, mountain srhimp

Theodoxus cariosus Erinna newcombi Heteromyenia baileyi

freshwater prawn anchialine pond shrimp adult dragonfly adult damselfly immature dragonfly or damselfly immature dragonfly or damselfly immature dragonfly or damselfly immature dragonfly or damselfly many other endemic insects

Macrobrachium grandimanus Halocaridina rubra Anas spp. Megalagrion spp. Megalagrion & Anax spp. Megalagrion & Anax spp. Megalagrion & Anax spp. Megalagrion & Anax spp. various species

Atyoida bisulcata

Hawaiian Name ‘o’opu ‘o’opu ’alamo’o ‘o’opu hi’u kole ‘o’opu hi‘u’ula ‘o’opu nu’ukole ‘o’opu nopili ‘o’opu nakea ‘o’opu naniha ‘o’opu ’akupa aholehole ‘ama’ama hihiwai wi hapawai hapakai pipiwai ∗ ∗

’opae ’opaekala’ole ’opae kuahiwi ’opae kolo ’opae ’oeha’a ’opae ’ula pinao pinao ’ula lohelohe lohaloha pua’alohehole ’olopelope ∗

Range endemic

endemic indigenous endemic endemic endemic worldwide endemic endemic endemic endemic indigenous endemic

endemic endemic endemic endemic endemic endemic endemic endemic endemic

*Indicates that a Hawaiian name is not available.

evolved from marine habitats (Table 1) (10,16,24,25). Some examples of this can be found in stream insect communities consisting of midges, Telmatogeton (Chironomidae); beach flies, Procanace (Canacidae); and shore flies, Scatella (Ephydridae), which are thought to have radiated into streams from ancestral intertidal habitats (26–28). The Hawaiian Telmatogeton complex is unique with all five endemic species restricted to freshwater, the only representatives of an exclusively marine intertidal subfamily worldwide (26–29). Another important insect group is the endemic damselflies of the Megalagrion complex. This genus has shown rapid speciation in concert with specific freshwater and terrestrial habitats (30–32). Contrary to many other aquatic insects of Hawaii, the endemic damselflies are thought to have evolved from one freshwater ancestor to Hawaii (31,32). As a result of recent extinctions, numerous research initiatives have been carried out to understand the evolution, biology, and conservation of this unique group of insects (31,32). Other freshwater organisms are reported in Table 1.

fish, snails, shrimp, and some insects) and/or isolated populations that have undergone rapid speciation (e.g., the damselflies). For millennia, these systems connected the mountains to the ocean along what is called in Hawaiian the mauka—makai continuum. The physical connection was the flowing streams supporting the biological connections of the amphidromous species. However, many streams were heavily diverted in the late 1800s for irrigating an expanding sugar cane industry. Stream flow removal and riparian degradation has had the following effects: (a) destroying and eliminating breeding habitat for all stream animals, (b) preventing amphidromous egg and larval drift to the ocean, (c) obstructing postlarval recruitment back into the streams, and (d) facilitating invasive species establishment (9,22,33–35). Water is a critical and limited resource, not only to humans but also to the native stream organisms. Balancing these will take compromise and scientifically based freshwater resource management.

SURFACE WATER REMOVAL

Historic Management

The streams of Hawaii are habitat to relatively recently evolved communities with close marine ancestors (e.g.,

The ahupua’a system was the first freshwater management system in Hawaii. Developed by the ali’i (chiefs) of

FRESHWATER RESOURCE USE AND MANAGEMENT

A HISTORY OF HAWAIIAN FRESHWATER RESOURCES

the first Hawaiians 1500–1600 years ago, it maintained adequate freshwater resources for the entire population of each island. The ahupua’a was what now is considered a watershed and extended from the headwater springs into the ocean. Along this mauka—makai continuum, the land along the stream was divided into pie-like slices called ‘ili kuponos, that were designated to maka’ainana (commoners) and overseen by a chief-appointed konohiki. The commoners did not own the land or water, for they only maintained them for the chiefs. Each konohiki was appointed to a single ahupua’a, and each ‘ili kupono was divided into kalo lo’i (taro patches). Stream water was diverted by pani wai (diversions) and carried through small ‘auwai (ditches) to the lo’i. The lo’i were connected in a stair-step manner by additional ditches maintained by the responsible commoners. If the ditches were not maintained upstream, all downstream lo’i would be affected. Contrary to contemporary freshwater management, the stream water was always returned to the stream before it entered the ocean, and was never carried out of the watershed. Therefore, the water continually connected the highest mountain reaches to the lowland settlements near the sea, maintaining the mauka—makai continuum. The early Hawaiians understood the connectivity of the land, freshwater, and sea. The ahupua’a system maintained adequate freshwater resources for human use while balancing ecological function. Hawaiians also understood the power and importance of water to their culture. For example, the Hawaiian word for water, wai, is the root for wealth (waiwai) and law (kanawai) (36). Over the next 60 years, native Hawaiians would find just how important water could be to foreigners, or haoles—those without breathe, or mana (spirit). The sugar cane industry completely changed freshwater resource management: water would be removed from the ahupua’a, along with much of the native Hawaiian culture. Contemporary Management The influence of western missionaries and merchants shifted the Hawaiian perception of water from a shared resource to a commodity that provided maximum use and greatest reward to plantation owners (37). After the arrival of Captain James Cook in 1778, freshwater management changed little until the The Great Mahele of 1848, enacted by King Kamehameha III under heavy influence by the decedents of original missionaries, the new owners of the sugar cane industry (and pineapple later). The Great Mahele was the first act to privatize land and water in Hawaii. For nearly 25 years the kings of Hawaii attempted to negotiate reciprocity treaties with the United States In 1876, King Kalakaua finally established a reciprocity treaty assuring duty-free exchange between Hawaii and the United States, giving sugar companies a competitive edge in the world market and giving Pearl Harbor to the United States, which was the same year of the first water licenses, the creation of the first private water company, and the construction of the diversionditch systems, which relocated millions of gallons of water from the windward watersheds to leeward sugar cane fields. According to Wilcox (36), one pound of sugar takes approximately 4000 pounds, or 500 gallons, of water.

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The growth of sugar cane and pineapple took over and ‘‘transformed Hawaii from a traditional, insular, agrarian, and debt-ridden society into a multicultural, cosmopolitan, and prosperous one’’ (36). During this time, the number of sugar plantations rose from 5 in 1857 to 90 in 1884 (36). From the late 1870s to the mid1900s, the Hawaiian sugar cane industry continued to grow, in part because of political access to surface and groundwater resources. By the early 1900s, almost every stream on the windward sides of each island was at least partially diverted, and in 1920, the sugar cane industry was diverting >800 mgd and pumping an additional 400 mgd from aquifers (36). Almost 20 years later, the entire city of Boston was only consuming 80 mdg (36). The privately developed surface water of Hawaii was quite different from publicly controlled surface water development of the western United States, which put control of water into the hands of western businessmen, those who would eventually contribute to the overthrow of the last Hawaiian Monarchy. (Several books are available that describe the sequence of events that led to the overthrow (38–40).) With the developing agricultural industry, the population and ethnic diversity expanded as well. A larger labor force was needed to construct the ditches and diversions and work the agricultural fields. For cheap labor, the agricultural companies looked to Japan, China, and other Asian nations. This immigration resulted in several ethnic populations living and working on plantation property, creating mixed nonHawaiian communities that are the roots of the ethnic diversity found in Hawaii today. As the population and successful industries increased, the popularity of the islands grew in the mid-1900s, whereas the sugar cane industry met fierce competition from other nations, and has resulted in the slow demise of the former economic strength. With the buildup and maintained military presence of Pearl Harber from World War II, the economic focus of Hawaii saw a change from agriculture to military establishment to tourism that is now the economic base of the state. Along the southwestern coast of each island, numerous hotels, shopping malls, golf courses, and restaurants have been developed because most people prefer to vacation on the sunny, dry side of the island. However, similar to the sugar cane scenario, water is not readily available on the leeward side. Therefore, water remains diverted from streams, or pumped from dwindling freshwater aquifers, in order to accommodate tourists, water the green lawns, and fill hotel swimming pools and spas. A BRIEF INTRODUCTION TO HAWAIIAN WATER LAW Water Rights The expansion of the sugar industry resulted in many water disputes between sugar companies, various landowners, and native Hawaiians. These disputes helped recognize the following water rights: (a) appurtenant, (b) riparian, (c) prescriptive, and (d) surplus (41). Appurtenant rights developed from ancient rights and were officially declared in the earliest Hawaiian water case,

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A HISTORY OF HAWAIIAN FRESHWATER RESOURCES

Peck vs. Baily (1867). This case determined that if land was entitled by a landowner to cultivate taro, than some stream water was also. Riparian rights permit water use to those who own land next to a stream; however, use must be reasonable and not harmful to other landowners (42). In riparian law, the focus is on maintaining stream integrity while allowing landowner priority uses (e.g., domestic drinking and washing) over nonpriority uses (e.g., irrigation, mining) (41). Both appurtenant and riparian water rights consider water as a common good. Prescriptive and surplus water rights address private ownership of water sources. Prescriptive water rights are granted if a party has proven using water ‘‘belonging’’ to another party for an extended and continuous period of time (e.g., 20 years) (37). Surplus water is defined as the amount of water in a stream not covered by any prior water right (37). In court, cases of surplus water were heard independently, which precluded the establishment of defined standards for future disputes. For example, in the case of Hawaiian Commercial & Sugar Co. v. Wailuku Sugar Co. (1904), the court ruled that surplus water traveling through an ahupua’a belonged to the private land owner, but in Carter v. Territory (1917), the court concluded that surplus water was to be used reasonably according to riparian principles, thus shared among landowners. The inconsistent rulings and absence of an established in-stream flow standard left the status of surplus waters questionable until the state code in 1987. Summary of Case Law From the 1850s to the 1980s, the water rights of Hawaii were evolving according to case law, yet no standard criteria for the allocation, transport, and quantity of freshwater was resolved. There were several political, economical, and social reasons for no established criteria. First, the rise of the sugar cane industry boosted the economy, and with it came political power to the foreign owners. Second, during this period, the Hawaiian Islands shifted from a monarchy to a republic to a U.S. Territory, and finally became a U.S. state, which created an atmosphere of economic and political change that left many native Hawaiians in a state of social and political confusion (36), where they did not resist changes in water law, as it was not in their culture to protest actions of the King. As a result of this, there was difficulty in establishing a cause-effect relationship, and many of the villages most affected were too small to fight the wealthy plantation owners (36,37). In 1978, the Hawaii State Constitution mandated that a water code and commission be established to manage water resources. And in the late 1980s, a Hawaii State Water Code and the Commission on Water Resource Management was passed by the Hawaii State Legislature (37). The State Water Code became a regulatory mechanism for state agencies and counties and required the Commission to monitor and guide water allocation by establishing minimum in-stream flow standards that were to enhance, protect, and reestablish beneficial in-stream uses. Current trends have been developing during the 1990s that are testing the

utility of the Hawaii State Water Code. The first of several cases, and a landmark case for future petitions, was the Waiahole Combined Contested Case Hearing. In this case, the Hawaii Supreme Court ruled for the establishment of minimum in-stream flow standards and than any additional flow secondarily allocated for other uses. In-stream flow standards are currently being investigated. The Waiahole Ditch Case was the first test of the State Water Code, yet many similar cases are being petitioned at the time of this writing. The diminishing sugar industry is opening a new chapter in Hawaiian case law, a chapter that must not only consider economics but also ‘‘traditional and customary Hawaiian rights, protection and procreation of fish and wildlife, ecological balance, scenic beauty, public recreation, beneficial in-stream uses, and public interest’’ (36).

SUMMARY AND SYNTHESIS The history of Hawaiian water resource issues is complex. The system of surface water management went through a dramatic change with the explosion of agriculture and the associated political power of missionary descendents; however, this explosion was inherently dependent on privately developed surface waters—a feedback proliferation of each other. The surface water development from the late 1800s through the mid-1900s drained watersheds of both water and native Hawaiian culture. The efficient removal of water high in the watersheds left a trickle downstream for native taro farmers. With the construction of massive diversion-ditch systems, native Hawaiians emigrated from the ahupua’a to build ditches and work sugar cane fields. At the same time, thousands of immigrants from China, Japan, and the Philippines melted into the sugar cane plantation farms, changing the demographics of Hawaii. The push by sugar businessmen to get an extended reciprocity treaty secured rights of Pearl Harbor to the United States; it also led to the coup of the Hawaiian Monarchy. At the same time, streams were being impacted by water removal and rapid riparian destruction, negatively affecting the native biological communities of the mauka–makai continuum. The degradation of stream habitats in Hawaii has become an increasingly important issue to many native Hawaiians, scientists, private organizations, and individuals. Water resources have been, and will continue to be, a highly debated political issue. Water is necessary for agricultural, industrial, and municipal uses throughout Hawaii; however, in order to maintain the unique biodiversity of the streams, a compromise must be made between continued development and conservation. Thus, the future of the Hawaii’s freshwater ecosystems and economic security is dependent on wise freshwater management and allocation, which should not be made in a scientific vacuum. Objective and sustainable answers can only be achieved with a solid scientific base, one free from political bias that has been the history of Hawaiian freshwater resources.

A HISTORY OF HAWAIIAN FRESHWATER RESOURCES

BIBLIOGRAPHY 1. Culliney, J.L. (1988). Islands in a Far Sea: Nature and Man in Hawaii. University of Hawaii Press, Honolulu, HI. 2. Hazlett, R.W. and Hyndman, D.W. (1996). Roadside Geology of Hawaii. Mountain Press, Missoula, MT. 3. Stearns, H.T. (1985). Geology of the State of Hawaii, 2nd Edn. Pacific books, Palo Alto, CA.

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shrimp postlarvae in two Maui streams. Micronesica 34(2): 243–248. 22. Benbow, M.E., Burky, A.J. and Way, C.M. (2004). The use of two modified Breder traps to quantitatively study amphidromous upstream migration. Hydrobiologia 527: 139–151. 23. Burky, A.J., Benbow, M.E., and Way, C.M. (2004). Two modified Breder traps for quantitative studies of postlarval amphidromy, Hydrobiologia.

4. Juvik, S.P. and Juvik, J.O. (1998). Atlas of Hawaii, 3rd Edn. University of Hawaii Press, Honolulu, HI.

24. Kinzie, R.A.I. (1997). Evolution and life history patterns in freshwater gobies. Micronesica 30(1): 27–40.

5. U.S.G.S. (1997). U.S. Geological Survey Programs in Hawaii and the Pacific. Department of Interior, United States Geological Survey.

25. McDowall, R.M. (2003). Hawaiian biogeography and the islands’ freshwater fish fauna. J. Biogeogr. 30(5): 703–710.

6. Carlquist, S. (1980). Hawaii: A Natural History. Pacific Tropical Botanical Gardens, Honolulu, HI. 7. Mueller-Dombois, D., Bridges, K.W., and Carson, H.L. (1981). Island Ecosystems: Biological Organization in Selected Hawaiian Communities. Hutchinson Ross, Stroudsburg, PA. 8. Kay, E.A. (Ed.). (1994). A Natural History of the Hawaiian Islands, Selected Readings II. University of Hawaii Press, Honolulu, HI. 9. Way, C.M., Burky, A.J., Harding, J.M., Hau, S., and Puleloa, W.K.L.C. (1998). Reproductive biology of the endemic goby, Lentipes concolor, from Makamaka’ole Stream, Maui and Waikolu Stream, Moloka’i. Environ. Biol. Fish. 51: 53–65. 10. Benbow, M.E. (1999). Natural History and Bioenergetics of an Endemic Hawaiian Chironomid: Fluctuating Stability in a Stochastic Environment. Ph.D. Dissertation. Department of Biology, University of Dayton, Dayton, OH. 11. Benbow, M.E., Burky, A.J., and Way, C.M. (1997). Larval habitat preference of the endemic Hawaiian midge, Telmatogeton torrenticola Terry (Telmatogetoninae). Hydrobiologia 346: 129–136. 12. McDowall, R.M. (1996). Is there such a thing as amphidromy? Micronesica 30(1): 3–14. 13. McDowall, R.M. (1997). The evolution of diadromy in fishes (revisited) and its place in phylogenetic analysis. Rev. Fish Biol. Fish 7(4): 443–462. 14. Radtke, R.L. and Kinzie, R.A. (1996). Evidence of a marine larval stage in endemic Hawaiian stream gobies from isolated high-elevation locations. T. Am. Fish Soc. 125: 613–621. 15. Radtke, R.L., Kinzie, R.A., and Folsom, S.D. (1988). Age at recruitment of Hawaiian freshwater gobies. Environ. Biol. Fish 23(3): 205–213. 16. Benbow, M.E., Burky, A.J., and Way, C.M. (2004). Morphological characteristics and species separation of Hawaiian postlarval amphidromous fishes. Micronesica 37(1): 127–143.

26. Hardy, D.E. and Delfinado, M.D. (1980). Diptera: Cyclorrhapha III, Series Schizophora, Section Acalypterae, Exclusive of Family Drosophilidae. Vol. 13. The University Press of Hawaii, Honolulu, HI. 27. Newman, L.J. (1988). Evolutionary relationships of the Hawaiian and North American Telmatogeton (Insecta; Diptera: Chironomidae). Pacific. Sci. 42(1–2): 56–64. 28. Newman, L.J. (1977). Chromosomal evolution of the Hawaiian Telmatogeton (Chironomidae, Diptera). Chromosoma 64: 349–369. 29. Wirth, W.W. (1947). A review of the genus Telmatogeton Schiner, with descriptions of three new Hawaiian species (Diptera: Tendipedidae). Proceedings of the Hawaiian Entomological Society 13: 143–191. 30. Jordan, S., Simon, C., and Polhemus, D.A. (2003). Molecular systematics and adaptive radiation of Hawaii’s endemic damselfly genus Megalagrion (Odonata: Coenagrionidae). Syst. Biol. 52(1): 89–109. 31. Polhemus, D.A. (1997). Phylogenetic analysis of the Hawaiian damselfly genus Megalagrion (Odonata: Coenagrionidae): implications for biogeography, ecology, and conservation biology. Pacific Science 51(4): 395–412. 32. Polhemus, D.A. and Asquith, A. (1996). Hawaiian Damselflies: A Field Identification Guide. Bishop Museum Press, Honolulu, HI. 33. Englund, R.A. and Filbert, R.B. (1999). Flow restoration and persistence of introduced species in Waikele Stream, O’ahu. Micronesica 32(1): 143–154. 34. Benbow, M.E., Burky, A.J., and Way, C.M. (2003). Life cycle of a torrenticolous Hawaiian chironomid (Telmatogeton torrenticola): stream flow and microhabitat effects. Ann. Limnol. Int. J. Lim. 39(2): 103–114. 35. McIntosh, M.D., Benbow, M.E., and Burky, A.J. (2002). Effects of stream diversion on riffle macroinvertebrate communities in a Maui, Hawaii, stream. River Res. Applic. 18: 569–581. 36. Wilcox, C. (1996). Sugar Water: Hawaii’s Plantation Ditches. University of Hawaii Press, Honolulu, HI.

17. Lindstrom, D.P. (1998). Reproduction, Early Development and Larval Transport Dynamics of Amphidromous Hawaiian Gobioids. Ph.D. Dissertation, University of Hawaii, Honolulu, HI.

37. Mackenzie, M.K. (1991). Native Hawaiian Rights Handbook. Native Hawaiian Legal Corporation, Honolulu, HI.

18. McDowall, R.M. (1992). Diadromy: origins and definitions of terminology. Copeia 1: 248–251.

38. Liliuokalani. (1990). Hawaii’s Story by Hawaii’s Queen. Mutual Publishing, Honolulu, HI.

19. Kinzie, R.A., III (1990). Species Profiles: Life Histories and Environmental Requirements of Coastal Vertebrates and Invertebrates Pacific Ocean Region. Amphidromous Macrofauna of Hawaiian Island Streams. US Army Engineer Waterways Experiment Station, Vicksburg, MS.

39. Daws, G. (1968). Shoal of Time: A History of the Hawaiian Islands. University of Hawaii Press, Honolulu, HI.

20. Kinzie, R.A., III and Ford, J.I. (1982). Population Biology in Small Hawaiian Streams. Water Resources Research Center, University of Hawaii at Manoa, Honolulu, HI, p. 147. 21. Benbow, M.E., Orzetti, L.L., McIntosh, M.D., and Burky, A.J. (2002). A note on cascade climbing of migrating goby and

40. Cooper, G. and Daws, G. (1990). Land and Power in Hawaii. University of Hawaii Press, Honolulu, HI. 41. MacKenzie, M.K. (1991). Water rights, In: Native Hawaiian Rights Handbook. M.K. MacKenzie (Ed.). Native Hawaiian Legal Corporation, Honolulu, HI, pp. 149–172. 42. Schoenbaum, T.J. and Rosenburg, R.H. (1996). Environmental Policy Law: Problems, Cases and Readings. University Casebook Series, Westbury, NY.

GROUND WATER ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES

minerals oxidize to form acidic products, which then can be dissolved in water. The water containing these dissolved products often has a low pH, high amounts of dissolved metals such as iron (Fe) and aluminum (Al), and sulfate. The metal concentrations in AMD depend on the type and quantity of sulfide minerals present, and the overall water quality from disturbed areas depends on the acid-producing (sulfide) and acid-neutralizing (carbonate) minerals contained in the disturbed rock. The carbonate content of overburden determines whether there is enough neutralization potential or base to counteract the acid produced from pyrite oxidation. Of the many types of acidneutralizing compounds present in rocks, only carbonates (and some clays) occur in sufficient quantity to effectively neutralize acid-producing rocks. A balance between the acid-producing potential and neutralizing capacity of the disturbed overburden will indicate the ultimate acidity or alkalinity that might be expected in the material upon complete weathering. Approximately 20,000 km of streams and rivers in the United States are degraded by AMD, but sulfide minerals occur throughout the world causing similar problems. About 90% of the AMD reaching streams originates in abandoned surface and deep mines. No company or individual claims responsibility for reclaiming abandoned mine lands and contaminated water flowing from these sites is not treated. Control of AMD before land disturbance requires an understanding of three important factors: (1) overburden geochemistry, (2) method and precision of overburden handling and placement in the backfill during reclamation, and (3) the postmining hydrology of the site.

JEFFREY G. SKOUSEN West Virginia University Morgantown, West Virginia

Acid mine drainage (AMD) occurs when metal sulfides are exposed to oxidizing conditions. Leaching of reaction products into surface waters pollute over 20,000 km of streams in the United States alone. Mining companies must predict the potential of creating AMD by using overburden analyses. Where a potential exists, special handling of overburden materials and quick coverage of acid-producing materials in the backfill should be practiced. The addition of acid-neutralizing materials can reduce or eliminate AMD problems. Placing acidproducing materials under dry barriers can isolate these materials from air and water. Other AMD control technologies being researched include injection of alkaline materials (ashes and limestone) into abandoned underground mines and into buried acid material in mine backfills, remining of abandoned areas, and installation of alkaline recharge trenches. Chemicals used for treating AMD are Ca(OH)2 , CaO, NaOH, Na2 CO3 , and NH3 , with each having advantages under certain conditions. Under low-flow situations, all chemicals except Ca(OH)2 are cost effective, whereas at high flow, Ca(OH)2 and CaO are clearly the most cost effective. Floc, the metal hydroxide material collected after treatment, is disposed of in abandoned deep mines, refuse piles, or left in collection ponds. Wetlands remove metals from AMD through formation of oxyhydroxides and sulfides, exchange and organic complexation reactions, and direct plant uptake. Aerobic wetlands are used when water contains enough alkalinity to promote metal precipitation, and anaerobic wetlands are used when alkalinity must be generated by microbial sulfate reduction and limestone dissolution. Anoxic limestone drains are buried trenches of limestone that intercept AMD underground to generate alkalinity. Under anoxia, limestone should not be coated with Fe+3 hydroxides in the drain, which decreases the likelihood of clogging. Vertical flow wetlands pretreat oxygenated AMD with organic matter to remove oxygen and Fe+3 , and then the water is introduced into limestone underneath the organic matter. Open limestone channels use limestone in aerobic environments to treat AMD. Coating of limestone occurs, and the reduced limestone dissolution is designed into the treatment system. Alkaline leach beds, containing either limestone or slag, add alkalinity to acid water. At present, most passive systems offer short-term treatment and are more practical for installation on abandoned sites or watershed restoration projects where effluent limits do not apply and where some removal of acid and metals will benefit a stream. Acid mine drainage (AMD) forms when sulfide minerals deep in the earth are exposed during coal and metal mining, highway construction, and other large-scale excavations. Upon exposure to water and oxygen, sulfide

OVERBURDEN ANALYSES, HANDLING, AND PLACEMENT Premining analysis of soils and overburden are required by law (1). Identifying the acid-producing or acid-neutralizing status of rock layers before disturbance aids in developing overburden handling and placement plans. Acid-base accounting provides a simple, relatively inexpensive, and consistent procedure to evaluate overburden chemistry. It balances potential acidity (based on total or pyritic sulfur content) against total neutralizers. Samples containing more acid-producing than acid-neutralizing materials are ‘‘deficient’’ and can cause AMD, whereas those rock samples with the reverse situation have ‘‘excess’’ neutralizing materials and will not cause AMD. Rock layers with equal proportions of each type of material should be subjected to leaching or weathering analyses (2). Kinetic tests such as humidity cells and leach columns are important because they examine the rate of acidproducing and neutralization reactions. This information from kinetic tests can supplement information given by acid-base accounting and help regulators in permitting decisions (3). The prevailing approach to control AMD is to keep water away from pyritic material. Once overburden materials have been classified, an overburden handling and placement plan for the site can be designed. 1

2

ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES

Segregating and placing acid-producing materials above the water table is generally recommended (2,4). Where alkaline materials overwhelm acid-producing materials, no special handling is necessary. Where acid-producing materials cannot be neutralized by onsite alkaline materials, it is necessary to import a sufficient amount to neutralize the potential acidity or the disturbance activity may not be allowed. Postmining Hydrology The hydrology of a backfill and its effect on AMD are complex. Generally, the porosity and hydraulic conductivity of the materials in a backfill are greater than those of the consolidated rock overburden that existed before mining, and changes in flow patterns and rates should be expected after mining (5). Water does not move uniformly through the backfill by a consistent wetting front. As water moves into coarse materials in the backfill, it follows the path of least resistance and continues downward through voids or conduits until it encounters a barrier or other compacted layer. Therefore, the chemistry of the water from a backfill will reflect only the rock types encountered in the water flow path, and not the entire geochemistry of the total overburden (6). Diverting surface water above the site to decrease the amount of water entering the mined area is highly recommended. If it cannot be diverted, incoming water can be treated with limestone to improve water quality. Under certain conditions, pyritic material can be placed where it will be rapidly and permanently inundated, thereby preventing oxidation. Inundation is only suggested where a water table may be reestablished, such as below drainage deep mines (see WET COVERS). CONTROL OF AMD Acid mine drainage control can be undertaken where AMD exists or is anticipated. Control methods treat the acidproducing rock directly and stop or retard the production of acidity. Treatment methods add chemicals directly to acidic water exiting the rock mass. Companies disturbing land in acid-producing areas must often treat AMD, and they face the prospect of long-term water treatment and its liabilities and expense. Cost-effective methods, which prevent the formation of AMD at its source, are preferable. Some control methods are most suited for abandoned mines, and others are only practical on active operations. Other methods can be used in either setting. Land Reclamation Backfilling (regrading the land back to contour) and revegetation together are effective methods of reducing acid loads from disturbed lands (7). Water flow from seeps can be reduced by diversion and reclamation, and on some sites where flow may not be reduced, water quality can change from acid to alkaline by proper handling of overburden. Diverting surface water or channeling surface waters to control volume, direction, and contact time can minimize the effects of AMD on receiving streams. Surface diversion involves construction of drainage ditches to move

surface water quickly off the site before infiltration or by providing impervious channels to convey water across the disturbed area. Alkaline Amendment to Active Disturbances Certain alkaline amendments can control AMD from acid-producing materials (8–11). All alkaline amendment schemes rely on acid-base accounting or kinetic tests to identify the required alkalinity for neutralization of acidic materials. Special handling of overburden seeks to blend acid-producing and acid-neutralizing rocks in the disturbance/reclamation process to develop a neutral rock mass. The pit floor or material under coal is often rich in pyrite, so isolating it from groundwater may be necessary by building highwall drains (which move incoming groundwater away from the pit floor) or placing impermeable barriers on the pit floor. Acid-forming material can be compacted or capped within the spoil (12). If insufficient alkalinity is available in the spoil, then external sources of alkalinity must be imported (13,14). Limestone is often the least expensive and most readily available source of alkalinity. It has a neutralization potential of between 75% and 100%, and it is safe and easy to handle. On the other hand, it has no cementing properties and cannot be used as a barrier. Fluidized bed combustion ashes generally have neutralizing amounts of between 20% and 40%, and they tend to harden into cement after wetting (15). Other power-generation ashes, like flue gas desulfurization products and scrubber sludges, may also have significant neutralization potential, which make them suitable alkaline amendment materials (16). Other materials, like kiln dust, produced by lime and cement kilns, or lime muds, grit, and dregs from pulp and paper industries contain neutralization products (10). Steel slags, when fresh, have neutralizing amounts from 45% to 90%. Slags are produced by several processes, so care is needed to ensure that candidate slags are not prone to leaching metal ions like Cr, Mn, and Ni. Phosphate rock has been used in some studies to control AMD. It may react with Fe released during pyrite oxidation to form insoluble coatings (17), but phosphate usually costs much more than other calcium-based amendments and is needed in about the same amounts (18). Alkaline Recharge Trenches Alkaline recharge trenches (19) are surface ditches or cells filled with alkaline material, which can minimize or eliminate acid seeps through an alkaline-loading process with infiltrating water. Alkaline recharge trenches were constructed on top of an 8-ha, acid-producing coal refuse disposal site, and after 3 years, the drainage water showed 25% to 90% acidity reductions with 70% to 95% reductions in Fe and sulfate (20). Pumping water into alkaline trenches greatly accelerates the movement of alkalinity into the backfill and can cause acid seeps to turn alkaline (21). Dry Barriers Dry barriers retard the movement of water and oxygen into areas containing acid-producing rock. These ‘‘water

ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES

control’’ technologies (4) include impervious membranes, dry seals, hydraulic mine seals, and grout curtains/walls. Surface barriers can achieve substantial reductions in water flow through piles, but generally they do not control AMD completely. Grouts can separate acid-producing rock and groundwater. Injection of grout barriers or curtains may significantly reduce the volume of groundwater moving through backfills. Gabr et al. (22) found that a 1.5-m-thick grout wall (installed by pumping a mixture of Class F fly ash and Portland cement grout into vertical boreholes near the highwall) reduced groundwater inflow from the highwall to the backfill by 80%, which results in some seeps drying up and others being substantially reduced in flow. At the Heath Steele Metal Mine in New Brunswick, a soil cover was designed to exclude oxygen and water from a tailings pile (23). It consisted of a 10cm gravel layer for erosion control, 30-cm gravel/sand layer as an evaporation barrier, 60-cm compacted till (conductivity of 10−6 cm/sec), 30-cm sand, and pyritic waste rock. This barrier excluded 98% of precipitation, and oxygen concentrations in the waste rock dropped from 20% initially to around 1%. At the Upshur Mining Complex in West Virginia, Meek (12) reported covering a 20-ha spoil pile with a 39-mil PVC liner, and this treatment reduced acid loads by 70%. Wet Covers Disposal of sulfide tailings under a water cover, such as in a lake or fjord, is another way to prevent acid generation by excluding oxygen from sulfides. Wet covers also include flooding of aboveground tailings in ponds. Fraser and Robertson (24) studied four freshwater lakes used for subaqueous tailings disposal and found that the reactivity of tailings under water was low and that there were low concentrations of dissolved metals, thereby allowing biological communities to exist. Alkaline Amendment to Abandoned Mines Abandoned surface mines comprise huge volumes of spoil of unknown composition and hydrology. Abandoned underground mines are problematic because they are often partially caved and flooded, cannot be accessed, and have unreliable or nonexistent mine maps. Re-handling and mixing alkalinity into an already reclaimed backfill is generally prohibitively expensive. Filling abandoned underground mine voids with nonpermeable materials is one of the best methods to prevent AMD. Underground mine voids are extensive (a 60-ha mine with a coal bed height of 1.5 m and a recovery rate of 65% would contain about 600,000 m3 of voids), so fill material and the placement method must be cheap. Mixtures of Class F fly ash and 3–5% Portland cement control subsidence in mined-under residential areas and these slurries are generally injected through vertical boreholes at between 8- and 16-m centers. Pneumatic (air pressure) and slurry injection for placing fly ash in abandoned underground mines can extend the borehole spacing to about 30 m (25). On reclaimed surface mines still producing AMD, researchers in Pennsylvania saw small improvements in water quality after injecting coal combustion residues into buried pods of pyritic materials.

3

Remining and Reclamation ‘‘Remining’’ means returning to abandoned surface or underground mines for further coal removal. Where AMD occurs, remining reduces acid loads by (1) decreasing infiltration rates, (2) covering acid-producing materials, and (3) removing the remaining coal, which is the source of most of the pyrite. Hawkins (26) found contaminant loads of 57 discharges from remined sites in Pennsylvania to be reduced after remining and reclamation. Short-term loads were sometimes increased during the first six months after remining and reclamation, but reduction in loads after six months resulted from decreased flow rather than large changes in concentrations. Ten remining sites in Pennsylvania and West Virginia were reclaimed to current standards (which included eliminating highwalls, covering refuse, and revegetating the entire area), and all sites had improved water quality (15). CHEMICAL TREATMENT OF AMD If AMD problems develop during mining or after reclamation, a plan to treat the discharge must be developed. A water treatment system consists of an inflow pipe or ditch, a storage tank or bin holding the treatment chemical, a valve to control its application rate, a settling pond to capture precipitated metal oxyhydroxides, and a discharge point. At the discharge point, water samples are analyzed to monitor whether specified parameters are being attained. Water discharge permits (NPDES) on surface mines usually require monitoring of pH, total suspended solids, and Fe and Mn concentrations. The type and size of a chemical treatment system is based on flow rate, pH, oxidation status, and concentrations of metals in the AMD. The receiving stream’s designated use and seasonal fluctuations in flow rate are also important. After evaluating these variables over a period of time, the operator can consider the economics of different chemicals. Six chemicals treat AMD (Table 1). Each is more or less appropriate for a specific condition. The best choice depends on both technical (acidity levels, flow, and the types and concentrations of metals) and economic factors (chemical prices, labor, machinery and equipment, treatment duration, and interest rates). Enough alkalinity must be added to raise pH to between 6 and 9 so insoluble metal hydroxides will form and settle out. Treatment of AMD with high Fe (ferric) concentrations often affords coprecipitation of other metals with the Fe hydroxide, thereby removing them from AMD at a lower pH. Limestone has been used for decades to raise pH and precipitate metals in AMD. It has the lowest material cost and is the safest and easiest to handle of the AMD chemicals. Unfortunately, it is limited because of its low solubility and tendency to develop an external coating, or armor, of Fe(OH)3 when added to AMD. Fine-ground limestone may be dumped in streams directly or the limestone may be pulverized by water-powered rotating drums and metered into the stream. Limestone has also treated AMD in anaerobic (anoxic limestone drains) and aerobic environments (open limestone channels).

4

ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES Table 1. Chemical Compounds Used in AMD Treatment 2000 Costc $ per Mg or L

Common Name

Chemical Name

Formula

Conversion Factora

Neutralization Efficiencyb

Bulk

9) is required to remove ions such as Mn. Unfortunately, increasing the lime rate increases the volume of unreacted lime that enters the floc-settling pond. Pebble quicklime (CaO) is used with the Aquafix Water Treatment System using a water wheel concept (28). A water wheel is turned based on water flow, which causes a screw feeder to dispense the chemical. This system was initially used for small and/or periodic flows of high acidity because CaO is very reactive, but water wheels have been attached to large silos for high-flow/high-acidity situations. Tests show an average of 75% cost savings over NaOH systems and about 20% to 40% savings over NH3 systems. Soda Ash Soda ash (Na2 CO3 ) generally treats AMD in remote areas with low flow and low amounts of acidity and metals. This choice is usually based on convenience rather than on chemical cost. Soda ash comes as solid briquettes and is gravity fed into water through bins. The number of briquettes used per day is determined by the rate of flow and quality of the water. One problem is that the briquettes absorb moisture, expand, and stick to the corners of the bin and will not drop into the stream. For short-term treatment, some operators use a much simpler system that employs a wooden box or barrel with holes that allows water inflow and outflow. The operator simply fills the barrel with briquettes on a regular basis and places the barrel in the flowing water. This system offers less control of the amount of chemical used. Caustic Soda Caustic soda (i.e., lye, NaOH) is often used in remote low-flow, high-acidity situations, or if Mn concentrations

in the AMD are high. The system can be gravity fed by dripping liquid NaOH directly into the AMD. Caustic is very soluble, disperses rapidly, and raises the pH quickly. Caustic should be applied at the surface of ponds because the chemical is denser than water. The major drawbacks of using liquid NaOH for AMD treatment are high cost and dangers in handling. Ammonia Ammonia compounds (NH3 or NH4 OH) are extremely hazardous. NH3 is compressed and stored as a liquid but returns to the gaseous state when released. Ammonia is extremely soluble, reacts rapidly, and can raise the pH of receiving water to 9.2. At pH 9.2, it buffers the solution to further pH increases, and therefore very high amounts of NH3 must be added to go beyond 9.2. Injection of NH3 into AMD is one of the quickest ways to raise water pH, and it should be injected near the bottom of the pond or water inlet because NH3 is less dense than water. NH3 is cheap, and a cost reduction of 50% to 70% is usually realized when NH3 is substituted for NaOH (29). Major disadvantages of using NH3 include (1) the hazards; (2) uncertainty concerning nitrification, denitrification, and acidification downstream; and (3) consequences of excessive application rates, which cause toxic conditions to aquatic life. Costs of Treating AMD Costs were estimated for five treatment chemicals under four sets of flow and acid concentration conditions [Table 1 from Skousen et al. (30)]. Na2 CO3 had the highest labor requirements (10 hours per week) because the dispensers must be filled by hand and inspected frequently. Caustic had the highest reagent cost per mole of acid-neutralizing capacity, and Na2 CO3 had the second highest. Hydrated lime treatment systems had the highest installation costs of the five chemicals because of the need to construct a lime treatment plant and install a pond aerator. However, the cost of Ca(OH)2 was very low, and the combination of high installation costs and low reagent cost made Ca(OH)2 systems particularly appropriate for long-term treatment of high-flow/high-acidity conditions.

ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES

For a 5-year treatment, NH3 had the lowest annual cost for the low-flow/low-acid situation. Pebble quicklime had about the same cost as the NH3 system, but slightly higher installation costs. Caustic was third because of its high labor and reagent costs, and Na2 CO3 was fourth because of high labor costs. Hydrated lime was the most expensive because of its high installation costs. At highflow/high-acidity, the Ca(OH)2 and CaO systems were clearly the cheapest treatment systems (annual costs of about $250,000 less than NH3 , the next best alternative). After chemical treatment, the treated water flows into sedimentation ponds so metals in the water can precipitate. All AMD treatment chemicals cause the formation of metal hydroxide sludge or floc. Sufficient residence time of the water (dictated by pond size and depth) is important for adequate metal precipitation. The amount of metal floc generated depends on water quality and quantity, which in turn determines how often the ponds must be cleaned. Knowing the chemical and AMD being treated will provide an estimate of the stability of metal compounds in the floc. Floc disposal options include (1) leaving it submerged indefinitely, (2) pumping or hauling it to abandoned deep mines or to pits dug on surface mines, and (3) dumping it into refuse piles. Pumping flocs onto land and letting them age and dry is a good strategy for disposal, because they become crystalline and behave like soil material. Each AMD is unique, requiring site-specific treatment. Each AMD source should be tested with various chemicals by titration tests to evaluate the most effective chemical for precipitation of the metals. The costs of each AMD treatment system based on neutralization (in terms of the reagent cost, capital investment, and maintenance of the dispensing system) and floc disposal should be evaluated to determine the most cost-effective system. PASSIVE TREATMENT OF AMD Active chemical treatment of AMD is often an expensive, long-term proposition. Passive treatment systems have been developed that do not require continuous chemical inputs and that take advantage of natural chemical and biological processes to cleanse contaminated mine waters. Passive technologies include constructed wetlands, anoxic limestone drains, vertical flow wetlands (also known as SAPS), open limestone channels, and alkaline leach beds (Fig. 1). In low-flow and low-acidity situations, passive systems can be reliably implemented as a single permanent solution for many AMD problems. Constructed Wetlands Wetlands are of two basic types: aerobic and anaerobic. Metals are retained within wetlands by (1) formation of metal oxides and oxyhydroxides, (2) formation of metal sulfides, (3) organic complexation reactions, (4) exchange with other cations on negatively charged sites, and (5) direct uptake by living plants. Other beneficial reactions in wetlands include generation of alkalinity caused by microbial mineralization of dead organic matter, microbial dissimilatory reduction of Fe oxyhydroxides and SO4 , and dissolution of carbonates.

5

Aerobic wetlands consist of relatively shallow ponds (30 cm) with substrates of soil, peat moss, spent mushroom compost, sawdust, straw/manure, hay bales, or other organic mixtures, often underlain or admixed with limestone. Anaerobic wetlands are most successful when used to treat small flows of acidic water. Anaerobic wetlands use chemical and microbial reduction reactions to precipitate metals and neutralize acidity. The water infiltrates through a thick permeable organic subsurface that becomes anaerobic because of high biological oxygen demand. Other chemical mechanisms that occur in situ include metal exchanges, formation and precipitation of metal sulfides, microbial-generated alkalinity, and formation of carbonate alkalinity (because of limestone dissolution). As anaerobic wetlands produce alkalinity, they can be used in net acidic and high dissolved oxygen (>2 mg/L) AMD. Microbial mechanisms of alkalinity production are critical to long-term AMD treatment. Under high acid loads (>300 mg/L), pH-sensitive microbial activities are eventually overwhelmed. At present, the sizing value for Fe removal in these wetlands is 10 gs per day per meter squared (31). Sorption onto organic materials (such as peat and sawdust) can initially remove 50% to 80% of the metals in AMD (32), but the exchange capacity declines with time. Over the long term, metal hydroxide precipitation is the predominant form of metal retention in a wetland. Wieder (33) reported up to 70% of the Fe in a wetland to be composed of Fe+3 oxyhydroxides, whereas the other 30% is reduced and combined with sulfides (34). Sulfate reducing bacteria (SRB) reactors have been used to generate alkalinity by optimizing anaerobic conditions. Good success has been noted for several systems receiving high and low flows (35,36). Anoxic Limestone Drains Anoxic limestone drains are buried cells or trenches of limestone into which anoxic water is introduced. The limestone raises pH and adds alkalinity. Under anoxic conditions, the limestone does not coat or armor with Fe hydroxides because Fe+2 does not precipitate as Fe(OH)2 at pH 6.0. Faulkner and Skousen (37) reported both successes and failures among 11 anoxic drains in WV. Failures resulted when ferric iron and Al precipitate as hydroxides in the limestone causing plugging and coating.

6

ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES

Determine flow rate analyze water chemistry calculate loading

Net alkaline water

Net acidic water

Determine do, ferric iron, al

Do < 1 mg/L and ferric < 1 mg/L and al < 1 mg/L

Net alkaline water

Anoxic limestone drain

Do > 1 mg/L and ferric > 1 mg/L and al > 1 mg/L

Net acid water

Settling pond

Settling pond

Aerobic wetland

Sulfate reducing bioreactor

Anaerobic wetland

Saps

Open limestone channel

Slag or ls leach bed

Settling pond

Settling pond

Settling pond

Settling pond

Settling pond

Meet effluent standards?

No

Meet effluent standards? Re-evaluate design

No

Yes Figure 1. Diagram of possible passive treatment systems to treat mine water based on water flow and chemistry.

Longevity of treatment is a major concern for anoxic drains, especially in terms of water flow through the limestone. Selection of the appropriate water and environmental conditions is critical for long-term alkalinity generation in an anoxic drain. Eventual clogging of the limestone pore spaces with precipitated Al and Fe hydroxides, and gypsum is predicted (38). For optimum performance, no Fe+3 , dissolved oxygen, or Al should be present in the AMD. Like wetlands, anoxic limestone drains may be a solution for AMD treatment for specific water conditions or for a finite period after which the system must be replenished or replaced. Vertical Flow Wetlands In these modified wetlands [called SAPS by Kepler and McCleary (39)], 1 to 3 m of acid water is ponded over an organic compost of 0.2 to 0.3 m, underlain by 0.5 to 1 m of limestone. Below the limestone are drainage pipes

Yes

Discharge

that convey the water into an aerobic pond where metals are precipitated. The hydraulic head drives ponded water through the anaerobic organic compost, where oxygen stripping as well as Fe and sulfate reduction can occur before water entry into the limestone. Water with high metal loads can be successively cycled through additional wetlands. Iron and Al clogging of limestone and pipes can be removed by flushing the system (40). Much work is being done on these wetlands presently, and refinements are being made for better water treatment. Open Limestone Channels Open limestone channels are another means of introducing alkalinity to acid water (41). We usually assume that armored limestone ceases to dissolve, but Ziemkiewicz et al. (42) found armored limestone to be 50% to 90% effective in neutralizing acid compared with unarmored limestone. Seven open channels in the field reduced acidity

ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES

in AMD by 4% to 62% compared with a 2% acid reduction in a sandstone channel. Open limestone channels show promise for neutralizing AMD in watershed restoration projects and AML reclamation projects where there can be only a one-time installation cost, little to no maintenance is required, and water exiting the system does not have to meet water quality standards. Long channels of limestone can convey acid water to a stream or other discharge point. Cross sections of channels can be designed with calculated amounts of limestone (which will become armored) to treat the water. Open limestone channels work best on steep slopes (>20%), where flow velocities keep metal hydroxides in suspension, thereby limiting plugging. If constructed correctly, open limestone channels should be maintenance free and provide AMD treatment for decades. Alkaline Leach Beds Limestone, when placed in an open pond or leach bed, will dissolve slowly over time and continually add alkalinity to water unless the limestone gets coated with metal hydroxides, thereby reducing its dissolution rate (41). Therefore, limestone treatment in aerobic systems works best in low-pH, metal-free water, and can add alkalinity to streams before encountering acid water downstream (42). As limestone generally reacts relatively slowly under field conditions, steel slag, a byproduct of steel making and composed of hydrated amorphous silica and calcium compounds, can be used as an alkaline material to add alkalinity to water. Steel slags have high neutralization potentials (from about 50–70%), can generate exceptionally high levels of alkalinity in water, and do not armor (43). Steel slag fines can be used in leach beds. Effluents from slag leach beds attain high pH (>10) and have alkalinity concentrations in the thousands of milligrams/liter. Slag leach beds may receive AMD directly, or effluent from ‘‘fresh water’’ beds may be combined with an AMD source downstream to treat acid indirectly. SUMMARY Acid mine drainage occurs when metal sulfides are oxidized. Leaching of reaction products into surface waters pollute over 20,000 km of streams in the United States alone. Companies must predict AMD before mining by using overburden analyses. On sites where a potential exists, special handling of overburden materials and quick coverage of acid-producing materials in the backfill should be practiced. Alkaline addition with materials such as kiln dust and FBC ash can reduce or completely eliminate AMD problems. Other control techniques include dry barriers, wet barriers, injection of alkaline materials into underground mines, remining of abandoned areas, and alkaline recharge trenches. Five chemicals typically treat AMD, and each has characteristics that make it suitable for specific applications. Companies must select a chemical that treats the water adequately and costeffectively. Passive systems are low maintenance systems that are implemented on abandoned mine land and stream restoration projects. Certain systems are more suited to

7

specific water quality and show good success where the acid levels do not overwhelm the system. BIBLIOGRAPHY 1. Sobek, A., Skousen, J., and Fisher, S. (2000). Chemical and physical properties of overburdens and minesoils. In: Reclamation of Drastically Disturbed Lands, 2nd Edn. American Society of Agronomy, Madison, WI. 2. Skousen, J.G., Sencindiver, J.C., and Smith, R.M. (1987). A Review of Procedures for Surface Mining and Reclamation in Areas with Acid-Producing Materials. EWRC 871, West Virginia University, Morgantown, WV. 3. Geidel, G., Caruccio, F.T., Hornberger, R., and Brady, K. (2000). Guidelines and recommendations for use of kinetic tests for coal mining (AMD) prediction in the eastern U.S. In: Prediction of Water Quality at Surface Coal Mines. National Mine Land Reclamation Center, West Virginia University, Morgantown, WV. 4. Skousen, J., Rose, A., Geidel, G., Foreman, J., Evans, R., and Hellier, W. (1998). Handbook of Technologies for Avoidance and Remediation of Acid Mine Drainage. National Mine Land Reclamation Center, West Virginia University, Morgantown, WV. 5. Caruccio, F.T. and Geidel, G. (1989). Water management strategies in abating acid mine drainage—Is water diversion really beneficial? In: 1989 Multinational Conference on Mine Planning and Design. 16–17 Sept. 1989, University of Kentucky, Lexington, KY. 6. Ziemkiewicz, P.F. and Skousen, J.G. (1992). Prevention of acid mine drainage by alkaline addition. Green Lands 22(2): 42–51. 7. Faulkner, B.B. and Skousen, J.G. (1995). Effects of land reclamation and passive treatment systems on improving water quality. Green Lands 25: 34–40. 8. Brady, K., Smith, M.W., Beam, R.L., and Cravotta, C.A. (1990). Effectiveness of the use of alkaline materials at surface coal mines in preventing or abating acid mine drainage: Part 2. Mine site case studies. In: Proceedings, 1990 Mining and Reclamation Conference. J. Skousen et al. (Eds.). 23–26 April 1990, West Virginia University, Morgantown, WV. 9. Perry, E.F. and Brady, K.B. (1995). Influence of neutralization potential on surface mine drainage quality in Pennsylvania. In: Proceedings, Sixteenth Annual Surface Mine Drainage Task Force Symposium. 4–5 April 1995, West Virginia University, Morgantown, WV. 10. Rich, D.H. and Hutchison, K.R. (1994). Coal refuse disposal using engineering design and lime chemistry. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06A-94, Pittsburgh, PA. 11. Rose, A.W., Phelps, L.B., Parizek, R.R., and Evans, D.R. (1995). Effectiveness of lime kiln flue dust in preventing acid mine drainage at the Kauffman surface coal mine, Clearfield County, Pennsylvania. In: Proceedings, 1995 National Meeting of the American Society for Surface Mining and Reclamation. 3–8 June 1995, Gillette, WY. 12. Meek, F.A. (1994). Evaluation of acid prevention techniques used in surface mining. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06B-94, Pittsburgh, PA. 13. Skousen, J. and Larew, G. (1994). Alkaline overburden addition to acid-producing materials to prevent acid mine drainage. In: International Land Reclamation and Mine

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ACID MINE DRAINAGE: SOURCES AND TREATMENT IN THE UNITED STATES Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06B-94, Pittsburgh, PA.

24–29 April 1994, USDI, Bureau of Mines SP 06A-94, Pittsburgh, PA.

14. Wiram, V.P. and Naumann, H.E. (1995). Alkaline additions to the backfill: A key mining/reclamation component to acid mine drainage prevention. In: Proceedings, Sixteenth Annual Surface Mine Drainage Task Force Symposium. 4–5 April 1995, West Virginia University, Morgantown, WV.

27. Skousen, J. and Ziemkiewicz, P. (1996). Acid Mine Drainage Control and Treatment, 2nd Edn. National Research Center for Coal and Energy, National Mine Land Reclamation Center, West Virginia University, Morgantown, WV.

15. Skousen, J., Bhumbla, D., Gorman, J., and Sencindiver, J. (1997). Hydraulic conductivity of ash mixtures and metal release upon leaching. In: 1997 National Meeting of the American Society for Surface Mining and Reclamation. 10–15 May 1997, Austin, TX.

28. Jenkins, M. and Skousen, J. (2001). Acid mine drainage treatment costs with calcium oxide and the Aquafix Machine. Green Lands 31: 46–51. 29. Skousen, J., Politan, K., Hilton, T., and Meek, A. (1990). Acid mine drainage treatment systems: chemicals and costs. Green Lands 20(4): 31–37.

16. Stehouwer, R., Sutton, P., Fowler, R., and Dick, W. (1995). Minespoil amendment with dry flue gas desulfurization byproducts: element solubility and mobility. J. Environ. Qual. 24: 165–174.

30. Skousen, J.G., Sexstone, A., and Ziemkiewicz, P. (2000). Acid mine drainage control and treatment. In: Reclamation of Drastically Disturbed Lands, 2nd Edn. American Society of Agronomy, Madison, WI.

17. Evangelou, V.P. (1995). Pyrite Oxidation and its Control. CRC Press, Boca Raton, FL.

31. Hedin, R.S. and Nairn, R.W. (1992). Passive treatment of coal mine drainage. Course Notes for Workshop. U.S. Bureau of Mines, Pittsburgh, PA.

18. Ziemkiewicz, P.F. and Meek, F.A. (1994). Long term behavior of acid forming rock: results of 11-year field studies. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06B-94, Pittsburgh, PA. 19. Caruccio, F.T., Geidel, G., and Williams, R. (1984). Induced alkaline recharge zones to mitigate acid seeps. In: Proceedings, National Symposium on Surface Mining, Hydrology, Sedimentology and Reclamation. 7–10 Dec. 1984, Univ. of Kentucky, Lexington, KY. 20. Nawrot, J.R., Conley, P.S., and Sandusky, J.E. (1994). Concentrated alkaline recharge pools for acid seep abatement: principles, design, construction and performance. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06A-94, Pittsburgh, PA. 21. Ziemkiewicz, P.F., Donovan, J., Frazier, J., Daly, M., Black, C., and Werner, E. (2000). Experimental injection of alkaline lime slurry for in situ remediation of an acidic surface mine aquifer. In: Proceedings, Twenty-first West Virginia Surface Mine Drainage Task Force Symp. April 4–5, 2000, Morgantown, WV. 22. Gabr, M.A., Bowders, J.J., and Runner, M.S. (1994). Assessment of acid mine drainage remediation schemes on groundwater flow regimes at a reclaimed mine site. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06B-94, Pittsburgh, PA.

32. Brodie, G.A., Hammer, D.A., and Tomljanovich, D.A. (1988). An evaluation of substrate types in constructed wetlands acid drainage treatment systems. In: Mine Drainage and Surface Mine Reclamation. 19–21 April 1988, Vol. 1, Info. Circular 9183, U.S. Bureau of Mines, Pittsburgh, PA. 33. Wieder, R.K. (1993). Ion input/output budgets for wetlands constructed for acid coal mine drainage treatment. Water, Air, and Soil Pollution 71: 231–270. 34. Wieder, R.K. (1992). The Kentucky wetlands project: A field study to evaluate man-made wetlands for acid coal mine drainage treatment. Final Report to the U.S. Office of Surface Mining, Villanova Univ., Villanova, PA. 35. Canty, M. (2000). Innovative in situ treatment of acid mine drainage using sulfate-reducing bacteria. In: Proceedings, Fifth International Conference on Acid Rock Drainage. Society for Mining, Metallurgy, and Exploration, Inc., Denver, CO. 36. Gusek, J., Mann, C., Wildeman, T., and Murphy, D. (2000). Operational results of a 1200 gpm passive bioreactor for metal mine drainage, Missouri. In: Proceedings, Fifth International Conference on Acid Rock Drainage. Society for Mining, Metallurgy, and Exploration, Inc., Denver, CO. 37. Faulkner, B.B. and Skousen, J.G. (1994). Treatment of Acid Mine Drainage by Passive Treatment Systems. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06A-94, Pittsburgh, PA.

23. Bell, A.V., Riley, M.D., and Yanful, E.G. (1994). Evaluation of a composite soil cover to control acid waste rock pile drainage. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06B-94, Pittsburgh, PA.

38. Nairn, R.W., Hedin, R.S., and Watzlaf, G.R. (1991). A preliminary review of the use of anoxic limestone drains in the passive treatment of acid mine drainage. In: Proceedings, Twelfth Annual West Virginia Surface Mine Drainage Task Force Symposium. 3–4 April 1991, West Virginia University, Morgantown, WV.

24. Fraser, W.W. and Robertson, J.D. (1994). Subaqueous disposal of reactive mine waste: an overview and update of case studies-MEND/Canada. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06A-94, Pittsburgh, PA.

39. Kepler, D.A. and McCleary, E. (1997). Passive aluminum treatment successes. In: Proceedings, Eighteenth Annual West Virginia Surface Mine Drainage Task Force Symposium. 15–16 April 1997, West Virginia University, Morgantown, WV.

25. Burnett, J.M., Burnett, M., Ziemkiewicz, P., and Black, D.C. (1995). Pneumatic backfilling of coal combustion residues in underground mines. In: Proceedings, Sixteenth Annual Surface Mine Drainage Task Force Symposium. 4–5 April 1995, West Virginia University, Morgantown, WV.

40. Kepler, D.A. and McCleary, E. (1994). Successive alkalinityproducing systems (SAPS) for the treatment of acidic mine drainage. In: International Land Reclamation and Mine Drainage Conference. 24–29 April 1994, USDI, Bureau of Mines SP 06A-94, Pittsburgh, PA.

26. Hawkins, J.W. (1994). Assessment of contaminant load changes caused by remining abandoned coal mines. In: International Land Reclamation and Mine Drainage Conference.

41. Ziemkiewicz, P.F., Skousen, J., and Lovett, R. (1994). Open limestone channels for treating acid mine drainage: a new look at an old idea. Green Lands 24(4): 36–41.

AQUIFERS 42. Ziemkiewicz, P.F. et al. (1997). Acid mine drainage treatment with armored limestone in open limestone channels. J. Environ. Qual. 26: 718–726. 43. Ziemkiewicz, P.F., Skousen, J., and Simmons, J. (2001). Cost benefit analysis of passive treatment systems. In: Proceedings, 22nd West Virginia Surface Mine Drainage Task Force Symposium. 3–4 April, Morgantown, WV. 44. Ziemkiewicz, P.F. and Skousen, J. (1998). The use of steel slag in acid mine drainage treatment and control. In: Proceedings, 19th West Virginia Surface Mine Drainage Task Force Symposium. 7–8 April, Morgantown, WV.

READING LIST Rich, D.H. and Hutchison, K.R. (1990). Neutralization and stabilization of combined refuse using lime kiln dust at High Power Mountain. In: Proceedings, 1990 Mining and Reclamation Conference. 23–26 April 1990, West Virginia University, Morgantown, WV. Skousen, J.G., Hedin, R., and Faulkner, B.B. (1997). Water quality changes and costs of remining in Pennsylvania and West Virginia. In: 1997 National Meeting of the American Society for Surface Mining and Reclamation. 10–15 May 1997, Austin, TX.

AQUIFERS

9

considered to constitute an aquifer in other settings. For example, a small intermontane valley aquifer that consists of very shallow, sandy loam deposits a few feet thick may supply enough water to maintain a pumping rate of 0.5–2 gallons per minute, enough for domestic water supply wells and some stock supply wells. This may be significantly more water than would be provided by the hardrock formations underlying and bounding such a shallow aquifer. In contrast, the same sandy loam deposits of the same thickness would not be considered an aquifer if they were part of an unconsolidated sedimentary sequence in a larger alluvial basin, where gravel and sand aquifers yield from 50 to more than 1,000 gallons per minute. ROLE OF AN AQUIFER IN THE HYDROLOGIC CYCLE Aquifers are part of the hydrologic cycle. They receive water through • • • •

recharge from precipitation, recharge from irrigation return water, seepage from rivers and streams, lateral transfer of water from neighboring aquifer basins, and • leakage from aquifer formations situated either above or below the aquifer.

THOMAS HARTER University of California Davis, California

GENERAL DEFINITION An aquifer is a geologic formation or geologic unit from which significant amounts of groundwater can be pumped for domestic, municipal, or agricultural uses. The four major types of rock formations that serve as aquifers are unconsolidated sand and gravel, sandstone, carbonate rocks, and fractured volcanic rocks. Aquifers may also occur in other geologic formations, particularly in fractured zones of igneous, metamorphic, or sedimentary rocks.

Water that collects in aquifers from those sources over periods of years, decades, centuries, and even millennia is discharged back to the surface through (Fig. 1) • • • •

springs, subsurface discharge into rivers and streams, lateral outflow to downgradient aquifers, vertical leakage to overlying or underlying aquifers, and • man-made wells. LIMITATION TO FORMATIONS WITH APPROPRIATE WATER QUALITY

The word aquifer was probably adopted around the early twentieth century from the French word aquif`ere, which originates from the two Latin words aqua (water) and ferre (to carry, to bear). Hence, literally translated from Latin, aquifer means ‘that which carries water.’

The term aquifer is applied to formations that produce low to moderate salinity water (appropriate for domestic, municipal, or agricultural uses). Geologic formations containing exclusively brackish or saline groundwater (even if they are made of highly permeable material) are typically not referred to as aquifers unless the salinity was induced through human activity (e.g., in seawater intrusion, which is the advance of saline seawater into an overpumped aquifer).

FURTHER DEFINITIONS

AQUIFER SIZE

There is no strict definition of the hydrogeologic attributes or volumetric extent necessary to make a geologic formation or geologic unit an aquifer. Rather, the term aquifer is used for local formations that have relatively higher permeability than surrounding formations. Geologic units that form an aquifer in one setting may therefore not be

Aquifers can be vastly different in size: a small local aquifer in a mountainous setting may be only a few feet thick and extend over an area of a few acres to tens of acres. Other aquifers span entire regions. For example, the Ogallala aquifer in the western-central United States underlies most of the High Plains region, which extends

ORIGIN OF THE WORD

10

AQUIFERS

Recharge

Artesian well

Piezometric surface Stream Water table

Unconfined aquifer Unconfined aquifer

Aquitard gw flow

Impermeable hardrock (aquitard)

Confined aquifer

Figure 1. Schematic representation of uncontinued aquifers, confined aquifers, aquitards, and aquicludes. Blade vertical arrows indicate recharge. Black horizontal arrows indicate pumping. Light colored arrows indicate the direction of groundwater movement.

eastward from the Rocky Mountains through parts of Texas, Oklahoma, Colorado, Kansas, and Nebraska. The aquifer consists of alluvial sediments, predominantly sands and gravel. It is an important production aquifer. An aquifer is characterized by its geologic extent (regional extent and thickness), the type of geologic formations that makes up the aquifer, the hydraulic conductivity, the transmissivity (which is defined as the product of hydraulic conductivity and aquifer thickness), the specific yield (the drainable porosity), the specific storage (the amount of water and rock compressed by hydrostatic pressure in a confined aquifer, see below), and the specific capacity (specific capacity is the amount of water pumped from a well per foot of water level drawdown created by pumping). The hydraulic conductivity of aquifers typically ranges from 1 m/day to more than 100 m/day. The specific capacity of wells located in aquifers may range from less than 0.1 gpm/ft (small, low-yielding aquifers suitable for domestic water supplies) to more than 100 gpm/ft (large production aquifers suitable for municipal and irrigation pumping).

the thickness of the geologic formation. With respect to hydraulic conductivity, porosity, and specific yield or specific storage, hydrogeologists have found that small variations that occur in the geologic composition of aquifer formations often result in large localized changes in hydraulic conductivity. This latter phenomenon is referred to as ‘‘natural aquifer heterogeneity.’’ As a result of all this variability, each well within the same aquifer will have a different specific capacity. Sometimes, the specific capacity of wells can vary quite significantly from well to well, especially in fractured rock aquifers, but also in unconsolidated aquifers with sand and gravel. Hydraulic conductivity, thickness, and specific yield or specific storage of an aquifer are determined indirectly by using literature values available for specific geologic formations, by using computer models in conjunction with local observations of groundwater fluxes or groundwater table fluctuations, or directly by performing an aquifer test (pumping test).

AQUIFER CHARACTERIZATION

Aquifers are the major hydrogeologic units within the hydrogeologic framework of a region from which groundwater is or can be extracted. The description of local or regional hydrogeology centers around the description of aquifers, that is, of those geologic formations with the highest significance—locally or regionally—with respect to (potential) groundwater production. Geologic formations that bound aquifers are referred to as aquicludes or aquitards. Aquicludes are, for all practical purposes, impermeable. Important aquicludes are thick, continuous clay formations and unfractured igneous rocks. Aquitards are geologic formations that have a lower

The amount of water that can be pumped from an aquifer depends primarily on four parameters: the hydraulic conductivity (also called the permeability) of the aquifer, the thickness of the aquifer, the specific yield or specific storage of the aquifer (related mostly to its porosity), and the amount of competition for water between wells. All four of these may change from location to location. The amount of pumping wells and the rate at which wells are pumping may be different from area to area; the thickness of the aquifer naturally changes with

AQUIFERS, AQUITARDS, AND AQUICLUDES

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

hydraulic conductivity than adjacent (aquifer) formations and therefore act as a partial barrier to groundwater flow between overlying aquifers. Aquitards can consist of material similar to aquifers, but either the amount of fine sediments is much larger (in unconsolidated formations) relative to the aquifer formation, or the degree of fracturing and size of fractures is smaller than that in the aquifer formation (in hardrock formations).

separated from the aquifer below by a zone that is unsaturated or aerated. This should not be confused with an unconfined shallow aquifer that is separated from a deeper confined aquifer through thick but saturated layers of clay.

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

CONFINED AND UNCONFINED AQUIFERS Aquifers can be either unconfined or confined, depending on the existence of an overlying aquitard or aquiclude. In an unconfined aquifer, there is no overlying aquitard or aquiclude. Recharge to the aquifer from the land surface or from and to streams is not restricted. The water table moves freely up and down, depending on the water stored, added to, or removed from the unconfined aquifer. The water level in a borehole drilled into an unconfined aquifer will be the same as the water level in the aquifer (if we ignore the effects of the capillary fringe). In a confined aquifer, on the other hand, water is ‘‘sandwiched’’ between two aquitards or between an aquitard and an aquiclude above and below the aquifer. Water in a confined aquifer is under hydrostatic pressure created by the weight of the overlying geologic formations and the water pressure created by the higher water levels in the usually remote recharge area of a confined aquifer. Due to the pressure in a confined aquifer, the water level in a borehole drilled into a confined aquifer will rise significantly above the top of the aquifer. An artesian well occurs where the pressure is so large that the water level in a well drilled into the confined aquifer rises above the land surface. A confined aquifer does not have a water table—it is completely filled with groundwater. The water level in wells drilled into a confined aquifer, instead, corresponds to the hydrostatic pressure head or potentiometric surface of the aquifer, which is located higher than the upper boundary of the aquifer itself. If the hydrostatic pressure head falls below the top of the confined aquifer, it becomes unconfined. An aquifer that is confined by an aquitard rather than an aquiclude is referred to as a ‘‘leaky aquifer’’ or a ‘‘semiconfined aquifer.’’ The aquitard is not always a contiguous layer of less permeable material. Local accumulations of multiple, smaller clay lenses and other clay-rich or otherwise impermeable layers dispersed within a more permeable formation may render the entire formation an aquitard. The actual low permeable lenses are not contiguous, but the overall effect of their presence within such a heterogeneous formation on the regional aquifer below is identical to that of a continuous aquitard formation. PERCHED WATER TABLE Occasionally, water collects above an impermeable or low permeability layer within the unsaturated (aerated) zone and forms a ‘‘perched’’ water table. By definition, a ‘‘perched’’ water table is a saturated groundwater zone

11

S.N. RAI National Geophysical Research Institute Hyderabad, India

V.P. SINGH Louisiana State University Baton Rouge, Louisiana

INTRODUCTION Groundwater plays a major role in augmenting water supply to meet the ever-increasing domestic, agricultural, and industrial demands. Increasing dependence of water supply on groundwater resources is resulting in increasing use of aquifers as a source of fresh water supply and subsurface reservoir for storing excess surface water. Aquifers are the geological formations that can store water as well as allow the flow of significant amount of water through their pores under ordinary field conditions. If the aquifer is bounded by two impermeable formations from top and bottom, it is called a confined aquifer. If the upper boundary of the aquifer is the water table, it is called an unconfined aquifer. The advantage of unconfined aquifers over confined aquifers to serve as a subsurface reservoir is that the storage of groundwater in large quantity is possible only in unconfined aquifer, which is because the storativity of the unconfined aquifer is linked to the porosity and not to the elastic properties of the water and solid matrix, as in case of the confined aquifer (1). Also, the vast surface area of the unconfined aquifer above the water table is available to receive the surface applied recharge, whereas in case of the confined aquifer, only a small open area exposed near to the ground surface or leaky portion of the aquifer boundary is available to receive the recharge (Fig. 1). This article deals with the artificial recharging of unconfined aquifer and related problems. Natural replenishment of aquifers occurs very slowly. Therefore, withdrawal of groundwater at a rate greater than the natural replenishment rate causes declining of groundwater level, which may lead to decreased water supply, contamination of fresh water by intrusion of pollutant water from nearby sources, seawater intrusion into the aquifer of coastal areas, etc. To increase the natural replenishment, artificial recharging of the aquifer is becoming increasingly important in groundwater management. The artificial recharge may be defined as an augmentation of surface water into aquifers by some artificially planned operations. The source of water for recharge may be direct precipitation, imported water, or reclaimed wastewater. The purpose of artificial recharging

12

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

Recharge area Artesian well

Infiltration Ground surface Recharge Piezometric surface Water table

Unconfined aquifer Confined aquifer Leakage Impervious stratum Semipervious stratum

Figure 1. Aquifer types.

of groundwater systems has been to reduce, stop, or even reverse the declining trend of groundwater level; to protect fresh groundwater in coastal aquifers against saline water intrusion from the ocean; and store surface water, including flood or other surplus water, imported water, and reclaimed wastewater for future use. RECHARGE METHODS A variety of direct surface, direct subsurface, and indirect recharge techniques have been developed to recharge groundwater systems. The choice of a technique depends on the source of water, quality of the water, the type of aquifer, topographical condition, etc. The most widely practiced methods are direct surface techniques, which include surface flooding in basins, ponds, lakes, ditches, trenches, and furrow systems; stream and channel modification; and bunds (2–5). Trenches are constructed mostly in foothill regions to arrest the runoff and put it into the aquifers for storage. Stream channel modification involves alteration in the course of stream flow to detain stream flow and increasing the stream bed area for recharging purposes. Construction of check dams across the stream flow is one technique of stream channel modification. It enhances artificial recharge in two ways. Above the dam, impoundments enhance recharge by increasing the recharge area and detaining water for a longer period by reducing the rate of water flow. Below the dam, recharge is enhanced through exposure of a larger area than the original area of stream channel flow. Bunds, which are small earthen barriers, are constructed in agricultural lands with slopes to facilitate impounding of runoff for a longer duration, thereby increasing recharge. In indirect subsurface recharge techniques, water is injected directly into an aquifer through (a) natural openings in the aquifers, (b) pits or shafts, and (c) wells. In contrast to the direct surface techniques, groundwater recharge by indirect subsurface techniques is practiced mostly for recharging the confined aquifer and where the topography or existing land use, such as in urban areas, makes recharge by surface flooding impractical. Indirect recharge techniques involve special cases in which potable water supply is provided by river bank or sand dune filtration of generally polluted river water (6,7). In many cases, excess recharging leads to the growth of water table near the ground surface and causes several

types of environmental problems, such as water logging, soil salinity, etc. In these situations, proper management of groundwater resources is needed to overcome the shortage of water supply on one hand and to prevent the environmental problems on the other hand. In order to address the management problem, one must be able to predict the response of the aquifer system to any proposed operational policy of groundwater resources development such as artificial recharging. Such problems are referred to as forecasting problems. Their solution will provide the new state of the groundwater system. Once the new state is known, one can check whether the related recharge scheme is feasible. Then one can compare responses of different proposed recharge schemes in order to select the best scheme that can meet the preset objectives of groundwater resources development without disturbing the regional water balance and without creating any kind of environmental problems. The forecasting problems are effectively tackled by application of modeling techniques. A model is the simplified representation of a complex real physical system and the processes taking place in it. It can be physical (for example, a laboratory sand pack model), electrical analog, or mathematical. Development and applications of mathematical models are much easier than the other two types of models. Therefore, mathematical models are mostly in use today for solving groundwater management problems. MATHEMATICAL MODELING Modeling begins with a conceptual understanding of the physical problem (in this case, groundwater flow in the unconfined aquifer). The next step is translating the physical problem into a mathematical framework resulting in equation forms that describe the groundwater flow. Mathematical models may be deterministic, statistical, or some combination of the two. Deterministic models retain a good measure of physical insight while permitting any number of problems of the same class to be tackled with the same model. Here, discussion will be confined to deterministic models. Formulations of groundwater flow equations are based on the conservation principles dealing with mass and momentum. These principles require that the net quantity of mass (or momentum) entering or leaving a specified volume of aquifer during a given time interval be equal to the change in the amount of mass (or moment) stored in the volume. Groundwater flow equations are formulated by combining the equation of motion in the form of Darcy’s law, which follows principle of conservation of momentum with the mass balance equations, also known as continuity equations, which follows the principle of conservation of mass. Some mathematical models commonly used for solving the forecasting problem in the presence of recharge are discussed below: 2-D groundwater flow in an inhomogeneous anisotropic unconfined aquifer with a horizontal base is described by the following equation (1,8):

 ∂ ∂h ∂h ∂h ∂ Kx h + Ky h + N(x, y, t) = Sy ∂x ∂x ∂y ∂y ∂t

(1)

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

in which h is the variable water table height measured from the horizontal base of the aquifer, Kx and Ky are the hydraulic conductivities in x and y directions, respectively, Sy is the specific yield, t is time of observation, and N(x, y, t) is the sum of all recharge rates from distributed sources (recharge basins, ponds, streams, etc.) and withdrawal rates from distributed sinks (wells, leakage boundaries, etc.) and is represented by  n    Ni (t) for xi1 ≤ x ≤ xi2 , yi1 ≤ y ≤ yi2 (2) N(x, y, t) = i=1   o elsewhere where n is the total number of basins, Ni (t) is the timevarying recharge (or pumping) rate for the ith basin (or well, respectively), and xi1 , xi2 , yi1 , and yi2 are the coordinates of ith basin (or well). Ni (t) is positive for recharge to the aquifer and negative for pumping. For an inhomogeneous isotropic aquifer, Eq. (1) becomes

 ∂h ∂ ∂h ∂h ∂ Kh + Kh + N(x, y, t) = Sy ∂x ∂x ∂y ∂y ∂t

(3)

Equations (1 and 3) are nonlinear second-order partial differential equation. The nonlinearity is because of the presence of h as a coefficient of partial derivatives on the left-hand side. Solving these equations because of nonlinearity is possible only by numerical methods, such as finite difference, finite element, and boundary elements (1,9,10). These equations need to be linearized for their analytical solution. For homogenous isotropic aquifers (K = constant), Eq. (3) can be written in the following two forms:  K



 ∂h ∂ ∂h ∂h ∂ h + h + N(x, y, t) = Sy ∂x ∂x ∂y ∂y ∂t

(4)

Sy ∂h2 ∂ 2 h2 2N(x, y, t) ∂ 2 h2 + + = (5) ∂x2 ∂y2 K Kh ∂t Two procedures of lineraization are commonly used to facilitate analytical solutions of Eqs. (4 and 5). According to the first procedure, i.e., the Baumann procedure of lineraization, if the variation in h is much less than the initial height of the water table h0 , then the coefficient h appearing on the left-hand side of Eq. (4) can be replaced by h0 (11). Then Eq. (4) can be rewritten as
T

∂ 2h ∂ 2h + 2 ∂x2 ∂y

 + N(x, y, t) = Sy

∂h ∂t

(6)

where T = Kh0 (known as transmissivity). Now, Eq. (6) is linear in h. Sometimes the mean depth of saturation is also used in place of h0 . In the second procedure, i.e., the Hantush procedure of linearization, h appearing in the denominator on the righthand side of Eq. (5) is replaced by the weighted mean of the depth of saturation h, a constant of linearization that is approximated by 0.5[h0 + h(te )], and te is the period at

13

the end of which h is to be approximated (12). Then, Eq. (5) becomes Sy ∂h2 ∂ 2 h2 2N(x, y, t) ∂ 2 h2 + + = 2 2 ∂x ∂y K Kh ∂t

(7)

Now, Eq. (7) becomes linear in h2 . By substituting a new variable H, defined as H = h2 − h0 2 , into Eq. (7) gives ∂ 2H ∂ 2H 2N(x, y, t) Sy ∂H + + = 2 ∂x ∂y2 K Kh ∂t

(8)

GROUNDWATER FLOW EQUATIONS FOR SLOPING AQUIFER 2-D groundwater flow in a sloping unconfined aquifer is described by (13,14) ∂ 2s ∂s 2N(t) 1 ∂s ∂ 2s + 2 − 2a + = 2 ∂x ∂y ∂x K  ∂t

(9)

where s = h2 , a = θ/2D, θ = slope of the base, D = the mean depth of saturation,  = KD/Sy , x, y = space coordinates, t = time of observation, and N(t) = time varying rate of recharge. One-dimensional groundwater flow equations can be obtained by substituting zero for the derivative of y in the above equations. These equations are used to predict the water table fluctuations in response to artificial recharge from strip basins, canals, channels, etc. GROUNDWATER FLOW EQUATIONS IN CYLINDRICAL COORDINATES These types of equations are used to describe groundwater flow induced by recharging through circular-shaped recharge basins/wells and is given by (15,16). S

1 ∂ ∂h =− (rq) + N(r) ∂t r ∂r

where q is defined by Darcy’s law as
 ∂h q = −Kh ∂r

(10)

(11)

The groundwater flow equations presented here are in the form of partial differential equations having an infinite numbers of solutions. To obtain a unique solution for a particular problem, more information about the aquifer’s parameters, geometry of the flow domain and recharge basins, duration and rate of recharge rate, and initial and boundary conditions are needed. These values can be deduced from field as well as experimental methods (1,17,18). Initial Conditions Initial conditions describe distribution of h at all points of the flow domain at the beginning of the investigation, that is, at t = 0, which is expressed as h = ψ(x, y, 0)

(12)

14

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

where ψ is a known value of h for all points of the flow domain. Boundary Conditions These conditions describe the nature of interaction of the flow system with its surroundings. Three types of boundary conditions are generally encountered in groundwater flow problems. • Dirichlet boundary condition—In this case, h is prescribed for all points of the boundary for the entire period of investigation, which is expressed as h = ψ (x, y, t)

(13)

where ψ (x, y, t) are known values of h at all points on the boundary. • Neumann boundary condition—This type of boundary condition prescribes the flux across the boundary of the flow system and can be expressed as q = ψ1 (x, y, t)

(14)

where ψ1 (x, y, t) are known values at the boundary. A special case of this boundary condition is the no flow boundary condition in which flux is zero. This condition occurs at impermeable surface or at the groundwater divide, a surface across which no flow takes place. • Cauchy boundary condition—This boundary condition is encountered at the semipervious boundary layer between the aquifer and an open water body such as a river. As a result of the resistance to the flow offered by the semipervious boundary that lies between the aquifer and the river, the water level in the river differs from that in the aquifer on the other side of the semipervious boundary. In this case, the flux is defined by q=K



h − h0 b

performance of some such analytical solutions (27–31). However, the rate of recharge largely depends on the infiltration rate, which initially decreases because of swelling and dispersion of soil particles. After some time, the infiltration rate increases because of the release of air entrapped into soil pores and reaches to a maximum value. Then, it starts decreasing because of clogging of soil pores beneath the bottom of the basin. Recharge rate also follows a more or less similar pattern of variation with some time lag and less intensity. When it falls below a prescribed low value, the recharge operation is discontinued for some time. After drying and, if necessary, scrapping of the silted base of the basin, a high recharge rate closer to its initial value is rejuvenated in the next phase of recharge operation (1,32–34). Zomorodi (35) has also shown that the analytical solution of Dagan (36), which is based on the assumption of constant recharge rate, fails to predict the recession of the water table caused by decrease in the recharge rate. Therefore, it would be more appropriate to consider recharge rate as time-dependent to simulate the actual field conditions. Some solutions have been developed for the time-varying recharge cases in which the decreasing rate of recharge has been represented by two linear elements (37–39) or by exponential function (14,16,40–44). However, approximation of time-varying recharge by two linear elements or exponential function is possible only for one recharge cycle. However, recharge is applied intermittently for more than one cycle separated by dry periods. Manglik et al. (45), Rai et al. (46), Manglik and Rai (48), and Rai and Manglik (49) used a general scheme of recharge approximation for any number of recharge cycles. In this scheme, time-varying recharge is approximated by a number of linear elements of different lengths and slopes depending on the nature of variation of recharge rate. Later on, this scheme was modified to represent rates of recharge from any number of basins. In mathematical form, this scheme can be represented by

(15)

where h is the head at x = 0, h0 is the water level  in the river, and b and K are the thickness and hydraulic conductivity of the semipervious boundary, respectively. The purpose of solving a groundwater flow equation is to obtain the values of h(x, y, t). Generally, two types of methods, namely analytical methods and numerical methods, are used for this purpose. Numerical methods are used to solve the nonlinear groundwater flow equation to tackle the real field problems, and analytical methods are used to solve the linearized form of groundwater flow equations. Analytical methods commonly used for the solution of groundwater problems include Fourier transforms, Laplace transforms, integral balance methods, method of separation of variables, approximate analytic methods, etc. Details about these methods can be found in many books (19–25). Most of the analytical solutions developed earlier for this purpose were based on the assumption of constant recharge. Warner et al. (26) have reviewed the

Ni (t) =

rij t + cij

tj ≤ t ≤ tj+1 [j = 1, 2, · · · , k − 1]

rik t + cik

t ≥ tk

[j = k]

(16)

where rij and cij are the slope and intercept of the jth line element of the recharge rate for the ith basin and k is the number of elements. The advantage of this scheme of recharge approximation is that any type of variation in the recharge rate for any number of recharge cycles from any number of recharge basins of different dimensions located anywhere within the flow domain can be approximated with the help of the required number of linear elements of different lengths and slopes depending on the nature of variation of recharge rate. By using this recharge scheme, several analytical solutions to describe water table fluctuation in different flow systems representing different physical conditions have been developed (50–53). The following analytical solution given by Manglik and Rai (50) is considered as an example to demonstrate the application of these solutions in prediction of water table fluctuation in the presence of time-varying recharge and

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

pumping,

(a)

8a Kπ 2

m=1 n=1

 mπ x  nπ y 1 sin sin mn A B

N  mπ xi2 mπ xi1  cos − cos (17) A A i=1   k−1

nπ yi1   nπ yi2 Rij + Rik  − sin cos B B j=1

This solution is obtained by solving Eq. (8) with recharge/pumping rates defined by Eq. (16) and subjected to the horizontal water table as an initial condition and Dirichlet boundary condition. In order to demonstrate the application of Eq. (17) in the prediction of water table fluctuation, we consider an example in which an unconfined aquifer of 10 × 10 km2 dimension is having two recharge basins of dimension 60 × 40 m2 and 50 × 50 m2 centered at (4470 m, 4500 m) and (5875 m, 5530 m), respectively, and two wells each of 10 × 10 cm2 dimension centered at (5000 m, 4500 m) and (5000 m, 5500 m), respectively. The pattern of time-varying recharge rate and pumping rate are shown in Fig. 2. The recharge operation for both the basins consists of two wet periods and one dry period, each of 20 days duration. During the first wet period, the rate of recharge decreases from its initial value of 0.8 m d−1 to a lower value of 0.7 m d−1 after 2 days. It again increases and attains maximum value of 0.9 m d−1 on the fourth day. After that, it starts decreasing and reduces to zero on twentieth day. The second cycle of recharge operation begins on the fortieth day and continues until the sixtieth day. The nature of variation of recharge rate for the second cycle is considered similar to the first cycle. Pumping of groundwater at a rate of 105 m d−1 from each well is considered for two periods. The first period is from the tenth to the twentieth day, and the second period is from the fortieth to the fiftieth day after a gap of 20 days. Numerical values of other controlling parameters are h0 = 20 m, K = 8 m d−1 , and Sy = 0.20. Two water table profiles computed for t = 45 days along a line parallel to the x-axis at y = 4500 m are shown in Fig. 3. These profiles pass through the center of one recharge basin and one well centered at (4470 m, 4500m) and (5000 m, 4500 m). The profile represented by the dotted curve is in response to recharge only. Hence, it shows only growth of groundwater mound. The profile represented by continuous curve shows growth as well

(b) Pumping rate [m/d] (× 105)

rij [tj+1 exp{−λ(t − tj+1 )} − tj exp{−λ(t − tj )}] Rij = λ r cij ij − 2 − [exp{−λ(t − tj+1 )} − exp{−λ(t − tj )}] λ λ rik [t − tk exp{−λ(t − tk )}] Rik = λ r cik ik − [1 − exp{−λ(t − tk )}] − 2 λ λ

0.6 0.3 0.0 −0.3 0

15

30 Time [d]

45

60

0

15

30 Time [d]

45

60

1.2 0.9 0.6 0.3 0.0 −0.3

Figure 2. Nature of time-varying (a) recharge, and (b) pumping rates.

as depression of the water table at the respective site of recharging and pumping. This example demonstrates the capabilities of prediction of water table variations in response to time-varying recharge and withdrawal. Accurate estimation of the varying recharge rate is a major problem in groundwater resources management. If the time history of water table variation at a site of an observation well is known, then analytical

5

t = 45 d, y = 4500 m

Only recharge

4 3 h - h0 [m]

in which A and B are the length and width of the aquifer, representing number a = Kh/Sy , and m and n are integers  a π 2 m2 n2 of Fourier coefficients. λ = + 2 4 A2 B

1.2 0.9

Recharge rate [m/d]

h2 (x, y, t) = h20 +

∞ ∞  

15

Both recharge and pumping

2 1 0 −1 −2 −3 4200

4500

4800

5100

5400

5700

6000

X - distance [m]

Figure 3. Water table profiles in the presence of only recharge (dashed curve), and both recharge and pumping (solid curve).

16

ARTIFICIAL RECHARGE OF UNCONFINED AQUIFER

solutions can be used for the estimation of varying recharge rate by making a judicious selection of recharge rate using trial and error method, such that the computed water table variation matches well with the observed one. Although the application of analytical solutions is restricted to the relatively homogeneous isotropic aquifer system having boundaries of simple geometrical shapes, their application is fast and simple in comparison with that of the numerical methods. Analytical solutions are also useful for other purposes, such as analysis of the effects of various controlling parameters, such as aquifers properties, initial and boundary conditions, intensity and duration of recharge rate, shape, size, and location of a recharge basin, etc., on the response of the aquifer system. Such information is very essential for the judicious selection of a suitable recharge scheme out of many proposed schemes to achieve the preset objectives of groundwater resource management. Acknowledgment We wish to thank Dr. S. Thiagarajan for his help in preparation of this work. SNR wish to thank Dr. V.P. Dimri, Director, NGRI for according permission to publish the work.

BIBLIOGRAPHY 1. Bear, J. (1979). Hydraulics of Groundwater. McGraw-Hill, New York, p. 567. 2. Asano, T. (1985). Overview: Artificial recharge of groundwater. In: Artificial Recharge of Groundwater. T. Asano (Ed.). Butterworth Publ., New York, pp. 3–19. 3. Oaksford, E.T. (1985). Artificial recharge: methods, hydraulics and monitoring. In: Artificial Recharge of Groundwater. T. Asano (Ed.). Butterworth Publ., New York, pp. 69–127. 4. Singh, R. (2002). Building community leaders for groundwater resources management—working on people’s priorities. In: Proceedings of Intl. Conference on Sustainable Development and Management of Groundwater Resources in Semi-Arid Region with Special Reference to Hard Rocks. M. Thanagarajan, S. N. Rai, and V. S. Singh (Eds.). Oxford & IBH Publ. Co., New Delhi, India, pp. 526–533. 5. Katyal, J.C., Singh, R.P., Sharma, S., Das, S.K., Padmanabhan, M.V., and Mishra, P.K. (1995). Field Manual on Watershed Management. Central Research Institute for Dryland Agricultural, Hyderabad, India, pp. 1–87. 6. Wilderer, P.A., Forstner, U., and Kuntschik, O.R. (1985). The role of river bank filtration along the Rhine river for municipal and industrial water supply. In: Artificial Recharge of Groundwater. T. Asano (Ed.). Butterworth Publ., New York, pp. 509–528. 7. Piet, G.J. and Zoedeman, B.C.J. (1985). Bank and dune infiltration of surface water in the Netherlands. In: Artificial Recharge of Groundwater. T. Asano (Ed.). Butterworth Publ., New York, pp. 529–540. 8. Rai, S.N. (2002). Groundwater flow modeling. In: Dynamics of Earth’s Fluid Systems. S.N. Rai, D.V. Ramana, and A. Manglik (Eds.). A. A. Balkema Publ., Holland, pp. 49–62. 9. Pinder, G.F. and Gray, W.G. (1977). Finite Element Simulation in Surface and Subsurface Hydrology. Academic Press, New York, p. 295. 10. Cabral, J.J.S.P., Wrobel, L.C., and Brebbia, C. (1990). Boundary element analysis of unconfined flow in porous media

using B-splines. In: Proceedings of the 8th Intl. Conference on ‘Computational methods in water resources, Part B. Venice, pp. 405–411. 11. Rao, N.H. and Sarma, P.B.S. (1980). Growth of groundwater mound in response to recharge. Groundwater. 18: 587–595. 12. Marino, M.A. (1967). Hele-Shaw model study of the growth and decay of groundwater ridges. J. Geophys. Res. 72: 1195–1205. 13. Baumann, P. (1965). Technical development in groundwater recharge. In: Advances in Hydroscience. Vol. 2. V.T. Chow (Ed.). Academic Press, New York, pp. 209–279. 14. Ramana, D.V., Rai, S.N., and Singh, R.N. (1995). Water table fluctuation due to transient recharge in a 2-D aquifer system with inclined base. Water. Resour. Manag. 9: 127–138. 15. Mercer, J.W. and Faust, C.R. (1980). Groundwater modeling; Mathematical models. Groundwater. 16: 212–227. 16. Rai, S.N., Ramana, D.V., and Singh, R.N. (1998). On the prediction of ground water mound formation in response to transient recharge from a circular basin. Water. Resour. Manag. 12: 271–284. 17. Todd, D.K. (1980). Groundwater Hydrology, 2nd Edn. John Wiley & Sons, New York, p. 535. 18. Rushtun, K.R. (2003). Groundwater Hydrology: Conceptual and Computational Models. John Wiley & Sons, New York, p. 407. 19. Carslaw, M.S. and Jaeger, J.C. (1959). Conduction of Heat in Solids. Oxford Univ. Press, Oxford, UK, p. 510. 20. Polubarinova-Kochina, P.Ya. (1962). Theory of Groundwater Movement. Princeton University Press, Princeton, NJ, p. 613. 21. Abramowidz, M. and Stegun, I.A. (1970). Handbook of Mathematical Functions. Dover, New York, p. 1046. 22. Ozisik, M.N. (1980). Heat Conduction. John Wiley & Sons, New York, p. 687. 23. Sneddon, I.N. (1974). The Use of Integral Transforms. Tata McGraw Hill, New Delhi, India. 24. Lee, T-C. (1999). Applied Mathematics in Hydrogeology. Lewis, Boca Raton, FL, p. 382. 25. Bruggeman, G.A. (1999). Analytical Solution of Geohydrological Problems. Elsevier, New York, p. 959. 26. Warner, J.W., Molden, D., Chehata, M., and Sunada, D.K. (1989). Mathematical analysis of artificial recharge from basins. Water Resour. Bull. 25(2): 401–411. 27. Baumann, P. (1952). Groundwater movement controlled through spreading. Amer. Soc. Civ. Eng. Trans. 117: 1024–1074. 28. Hantush, M.S. (1967). Growth and decay of groundwater mounds in response to uniform percolation. Water Resour. Res. 3(1): 227–234. 29. Glover, R.E. (1961). Mathematical Derivations as Pertain to Groundwater Recharge. Agricultural Research Service, USDA, Ft. Collins, CO, p. 81. 30. Hunt, B.W. (1971). Vertical recharge of unconfined aquifers. J. Hydraul. Div. ASCE. 97(HY7): 1017–1030. 31. Rao, N.H. and Sarma, P.B.S. (1981). Groundwater recharge from rectangular areas. Groundwater. 19: 271–274. 32. Singh, V.P. (1989). Hydrologic Systems—Watershed Modelling. Vol. 2. Prentice Hall, Englewood Cliffs, NJ, p. 320. 33. Detay, M. (1995). Rational groundwater reservoir management, the role of artificial recharge. In: Artificial Recharge of Groundwater II. A.I. Johnson and R.D.G. Pyne (Eds.). ASCE, New York, pp. 231–240.

GROUNDWATER AND ARSENIC: CHEMICAL BEHAVIOR AND TREATMENT 34. Mousavi, S.F. and Rezai, V. (1999). Evaluation of scraping treatment to restore infiltration capacity of three artificial recharge projects in central Iran. Hydrogeol. J. 7: 490–500. 35. Zomorodi, K. (1991). Evaluation of response of a water table to a variable recharge rate. Hydrol. Sci. J. 36: 67–78. 36. Dagan, G. (1966). Linearized Solutions of Free-Surface Groundwater Flow with Uniform Recharge. Technion Publication No. 84, Technion, Israel Institute of Technology, Tel Aviv, Israel. 37. Rai, S.N. and Singh, R.N. (1979). Variation of water table induced by time varying recharge. Geophys. Res. Bull. 17(2): 97–109.

In: New Approaches Characterizing Groundwater Flow. K.P. Seiler and S. Wohnlich (Eds.). Balkema Pub., The Hague, The Netherlands, pp. 775–778. 53. Manglik, A., Rai, S.N., and Singh, V.S. (2004). Modeling of aquifer response to time varying recharge and pumping from multiple basins and wells. J. Hydrol. 292: 23–29.

GROUNDWATER AND ARSENIC: CHEMICAL BEHAVIOR AND TREATMENT DAVID B. VANCE

38. Rai, S.N. and Singh, R.N. (1980). Dynamic Response of an unconfined aquifer system subjected to transient recharge. Geophy. Res. Bull. 18(2): 49–56.

ARCADIS G&M, Inc. Midland, Texas

39. Rai, S.N. and Singh, R.N. (1981). A mathematical model of water table fluctuations in a semi-infinite aquifer induced by localised transient recharge. Water Resour. Res. 17(4): 1028–1032.

JAMES A. JACOBS

40. Abdulrazzak, M.J. and Morel-Seytoux, H.J. (1983). Recharge from an ephemeral stream following wetting front arrival to water table. Water Resour. Res. 19: 194–200. 41. Rai, S.N., Manglik, A., and Singh, R.N. (1994). Water table fluctuation in response to transient recharge from a rectangular basin. Water Resour. Manag. 8(1): 1–10. 42. Rai, S.N. and Singh, R.N. (1995). An analytical solution for water table fluctuation in a finite aquifer due to transient recharge from a strip basin. Water Resour. Manag. 9(1): 27–37. 43. Rai, S.N. and Singh, R.N. (1996). On the prediction of ground water mound formation due to transient recharge from a rectangular area. Water. Resour. Manag. 10: 189–198. 44. Rai, S.N. and Singh, R.N. (1998). Evolution of the water table in a finite aquifer due to transient recharge from two parallel strip basins. Water. Resour. Manag. 12: 199–208. 45. Manglik, A., Rai, S.N., and Singh, R.N. (1997). Response of an unconfined aquifer induced by time varying recharge from a rectangular basin. Water Resour. Manag. 11(3): 185–196. 46. Rai, S.N., Ramana, D.V., and Manglik, A. (1997). Modeling of water table fluctuation in finite aquifer system in response to transient recharge. In: Proceeding of Intl. Symposium of Emerging Trend in Hydrology. D. C. Singhal et al. (Eds.). pp. 243–250. 47. Rai, S.N., Ramana, D.V., Thiagarajan, S., and Manglik, A. (2001). Modelling of groundwater mound formation due to transient recharge. Hydrological Processes 15(8): 1507–1514. 48. Manglik, A. and Rai, S.N. (1998). Two-dimensional modelling of water table fluctuations due to time-varying recharge from rectangular basin. Water Resour. Manag. 12: 467–475. 49. Rai, S.N. and Manglik, A. (1999). Modelling of water table variation in response to time varying recharge from multiple basins using the linearized Boussinesq equation. J. Hydrol. 220: 141–148. 50. Manglik, A. and Rai, S.N. (2000). Modelling of water table fluctuation in response to time varying recharge and withdrawal. Water Resour. Manag. 14(5): 339–347. 51. Rai, S.N. and Manglik, A. (2000). Water table variation due to time varying recharge and withdrawal. In: Groundwater: Past Achievement and Future Challenges. Oliver Silio et al. (Eds.). Proc. of XXX IAH Congress, Minneapolis, MN, pp. 259–262. 52. Rai, S.N. and Manglik, A. (2001). Modelling of water table fluctuations due to time-varying recharge from canal seepage.

17

Environmental Bio-Systems, Inc. Mill Valley, California

INTRODUCTION Arsenic is not an abundant element in the earth’s crust; the average crustal concentration of arsenic is 1.8 mg/Kg (1) to 5 mg/Kg (2). Arsenic ranks as the 52nd element in abundance, between tin and molybdenum (1). However, through geogenic processing of crustal materials, arsenic can be concentrated in soils to a typical range of 2 to 20 mg/Kg (3) to 1 to 50 mg/Kg (2), with concentrations as high as 70 mg/Kg being unremarkable (3). Human activity generates anthropogenic arsenic, which makes it the third most common regulated inorganic contaminant found at U.S. Superfund sites. Arsenical copper was in use by 4000 BC, and the toxic effects of arsenic were documented by early Greek writers. More recently, arsenic has been linked to skin, bladder, and other cancers (4). The U.S. Environmental Protection Agency (USEPA) lowered the arsenic standard in drinking water from 50 µg/L to 10 µg/L, effective January 23, 2006. Modern usage of arsenic includes formulation of pesticides and herbicides, decolorization of glass, paint manufacturing, the production of semiconductors, and the treatment/preservation of wood. Pressure treated lumber was commonly treated for decades using copperchromium-arsenate (CCA). This product, also called ‘‘green wood,’’ has been used for foundation lumber and more recently as wood for outdoor children’s play structures and picnic tables. The CCA wood is being phased out for toxicity concerns and environmental reasons. Many of the pressure treatment lumber facilities have significant soil and groundwater contaminated with arsenic as well as chromium. CHEMICAL CHARACTER Although arsenic occurs in more than 20 minerals, only a few are commonly found in ore deposits (5). Arsenic may occur as a semimetallic element (As0 ), arsenate (As5+ ), arsenite (As3+ ), or arsine (As3− ). The biogeochemistry of arsenic involves adsorption, biotransformation, REDOX reactions, and precipitation-dissolution processes (6,7).

GROUNDWATER AND ARSENIC: CHEMICAL BEHAVIOR AND TREATMENT

The chemical character of arsenic is labile and readily changes oxidation state or chemical form through chemical or biological reactions that are common in the environment. Therefore, rather than solubility equilibrium controlling the mobility of arsenic, it is usually controlled by REDOX conditions, pH, biological activity, and adsorption/desorption reactions. Arsenic in groundwater most often occurs from geogenic sources, although anthropogenic arsenic pollution does occur. Geogenic arsenic is almost exclusively an arsenite or arsenate. The most oxidized pentavalent form, arsenic, forms oxyanions (H3 ASO4 − , H2 ASO4 − , H2ASO4-, HASO4 2− , ASO3 − ). These arsenic oxyanions are isomorphous with oxyanions of phosphorous, substituting for phosphate in both marine organisms and phosphate deposits (4). Arsenite is the trivalent form that also forms a series of oxyanions that change specific configuration and charge with pH. Of critical importance with regard to the controls of the mobility of arsenite is the fact that at a pH of 9.5 or lower, the arsenite oxyanion is not charged. This result obviates all ionic interactions of the species. Common arsenic minerals are arsenopyrite (FeAsS), enargite (Cu,AsS), proustite (Ag,AsS), and lollingite (FeAs2 ). Late-stage magmatic crystallization (pyrometasomatic and hydrothermal stages) contributes to arsenic-rich sulfides. In sedimentary rocks, arsenic is commonly found adsorbed onto fine-grained sedimentary rocks, such as iron and manganese oxides (4). According to the U.S. Geological Survey, arsenic concentrations in sedimentary iron-ores range from 65 to 650 mg/Kg (8). Arsenic is also associated with sedimentary pyrite at concentrations of 100 to 77,000 mg/Kg (6). Anthropogenic arsenic may have any form including organic arsine species. Groundwater in acidic to intermediate volcanic rocks, or in sediments derived from those rocks, will often have arsenic concentrations exceeding 50 µg/L. Figure 1 illustrates the difference in molecular structure between arsenate and arsenite. The double bond oxygen in the arsenate molecule influences its ability to become ionized through the of hydrogen ions. The process is termed dissociation. A negative charge develops on the arsenate molecule when dissociation occurs. The double bond oxygen increases the capacity to delocalize that charge, which cases the loss of hydrogen ions. The propensity for ionization is expressed by the constant of dissociation, pKa. The pKa value, which is a negative log, shows a greater degree of dissociation with a smaller value. For arsenate and arsenite, pKa values are as follows: Arsenate- H3 ASO4

pK1 = 2.19

pK2 = 6.94

pK1 = 9.20

pK2 = 14.22∗

pK3 = 11.5 Arsenite- H3 ASO3 pK3 = 19.22∗ The pH at which these ionization steps occur is significantly different between arsenate and arsenite, as *These pKa values are extrapolated from the strength of oxygen acid rules (9).

Common species of Arsenic in ground water

Arsenate

Arsenite

O

H

H O As O

O As O O

H

H

O H

H

Figure 1. Difference in molecular configuration of arsenate and arsenite.

Eh/pH conditions and arsenic speciation 1.2 Water oxidized

1 H3AsO4

0.8

H2AsO4−

0.6 Eh in volts

18

0.4 HAsO42−

0.2

AsO43−

H3AsO3

0 −0.2 −0.4

H2AsO3−

−0.6

Water reduced

−0.8 0

2

4

AsO33−

HAsO32−

6

8

10

12

14

pH

Figure 2. Control of arsenic speciation by Eh and pH conditions.

illustrated in Fig. 2 (10,11). Figure 2 also shows the control of REDOX potential (Eh) on the arsenate/arsenite transition. This Eh/pH relationship is key in understanding arsenic mobility in groundwater and the effectiveness of arsenic water treatment systems. Arsenic Immobilization The previous section described the conditions under which arsenic can become an ionized species. The most commonly recognized adsorption reactions are based on ion exchange between charged adsorption sites and charged soluble ions. However, London Van der Waals bonding is another mechanism that is also responsible for adsorption. This type of bonding is the result of complex interactions among the electron clouds of molecules, molecular polarity, and attractive forces of an atomic nucleus for electrons beyond its own electron cloud. Consequently, some degree of immobilization can occur with soluble species that are not ionized. Arsenic immobilization through ionic adsorption can be controlled within normal oxidizing Eh/pH conditions. London Van der Waals bonding is complex to the point of unpredictability except for arsenic

GROUNDWATER AND ARSENIC: CHEMICAL BEHAVIOR AND TREATMENT

ARSENIC IN GROUNDWATER

Eh/pH conditions and iron speciation

1.2

Water oxidized

1

Fe3+

2+

FeOH

0.8

Fe(OH)2+

Eh in volts

0.6 0.4 0.2

Fe2+

Fe(OH)3

0 −0.2 FeOH+

−0.4 −0.6

Fe(OH)2

Water reduced

−0.8 0

2

4

6

8

19

10

12

14

pH

Figure 3. Iron speciation as controlled by Eh/pH conditions.

Arsenic concentrations up to 12,000 µg/L have been reported for the St. Peter aquifer in eastern Wisconsin (4). In this case, the oxidation of arsenic sulfides in a sulfide cement horizon (SCH) within the aquifer is a source of the high arsenic concentrations. Figure 4 superimposes the Eh/pH relationship for the arsenic and iron systems; it illustrates the conditions under which arsenic will be immobilized in a groundwater system. Of equal importance, it illustrates how arsenic adsorbed to ferric hydroxides in sediment can be released at exposure to groundwater that is chemically reducing. Two effects would be at work: Arsenate is reduced to arsenite that will not remain ionically bound to the geologic substrate, and ferric iron is reduced to ferrous, which is soluble under normal pH conditions. Outside the immobilized zone, arsenic mobility is variable. London Van der Waals bonding of arsenite is in effect, but it is not sufficient to assure complete immobilization.

Eh/pH vs. arsenic immobilization in groundwater 1.2

0.8

Water oxidized

0.6 Eh in volts

WATER TREATMENT SYSTEMS

Zone of arsenic immobility

1

Arsenate

0.4

Following is a brief review of various technologies used for the removal of arsenic from drinking water and industrial wastewater. Table 1 summarizes the effectiveness of each and gives the source for the information.

Arsenite

0.2

Ferrous iron

0

Introduction

Ferric iron

−0.2

Arsenite Oxidation

−0.4 −0.6

Water reduced

−0.8 0

2

4

6

8

10

12

14

pH

Figure 4. Arsenic mobility in groundwater as controlled by the effect of Eh/pH conditions on the speciation of arsenic and iron.

mobility at extreme Eh/pH conditions that can be obtained in industrial settings, but not in groundwater. Components of soil that participates in both types of adsorptive reactions include clays, carbonaceous material, and oxides of iron, aluminum, and manganese. In the most shallow soils, the organic fractions typically dominate, whereas at greater depths, iron oxyhydroxides play the principal adsorptive role. The typical iron content of soil ranges from 0.5% and 5%. Not only is iron common, but as with arsenic, it is also labile and readily reflects changes in surrounding Eh/pH conditions. This relationship for iron is illustrated in Fig. 3 (12). Ferric hydroxide acts as an amphoteric ion exchanger. Depending on pH conditions, the ferric hydroxide has the capacity for cation or anion exchange. Given the average iron concentration in soil and soluble arsenic concentrations in groundwater at 50 µg/L, ferric hydroxides in sediment can potentially adsorb 0.5 to 5 pounds of arsenic per cubic yard of aquifer matrix, which may then act as a significant potential reservoir for arsenic release under changing Eh/pH conditions.

As previously described, the ionization chemistry of arsenic in groundwater precludes the removal of arsenite by ion exchange within normal pH ranges. Other technologies including coprecipitation, electrodialysis, and reverse osmosis are also affected by arsenite’s dissociation profile. One solution to this problem is an oxidation step to form arsenate. Figure 5 illustrates the Eh/pH range required for this process. Oxidation, particularly of drinking water, may be problematic. Chemical residues of the oxidant, byproducts from oxidation of other organic or inorganic species, reagent costs, and operational issues are all factors. Oxygen would be ideal, as it is thermodynamically capable of this oxidizing step. The kinetics for oxidizing arsenic compounds in groundwater are exceedingly slow (22). It is possible to use gas diffusion technologies that slowly release dissolved oxygen into aquifers and that have demonstrated the capacity to convert anaerobic groundwater systems into aerobic systems within 3 to 6 months (23). This process will convert soluble ferrous iron to insoluble ferric iron oxides capable of attracting arsenate to their surfaces. The oxidation of arsenite is complex and may take additional time or the presence of other abiotic or biological (24) stimulants. In Bangladesh, in situ concentrations of arsenic less than 0.1 mg/L were readily removed by the oxygenation of groundwater; concentrations greater than that had only 50% removal (25). Other chemicals can affect arsenite oxidation including free chlorine, hypochlorite, ozone, permanganate, and hydrogen peroxide with ferrous iron.

20

GROUNDWATER AND ARSENIC: CHEMICAL BEHAVIOR AND TREATMENT Table 1. Effectiveness of Arsenic Water Treatment Methods

Treatmenttechnology Iron Coprecipitation

Alum Coprecipitation

Lime Precipitation

Activated Alumina

Ion Exchange Reverse Osmosis Electrodialysis Sulfide Precipitation Activated Carbon ∗

Initial Arsenic Concentration

Final Arsenic Concentration∗

Reference

Oxidized Arsenic 56 mg/L Arsenate 350 µg/L Arsenite 350 µg/L Arsenate 560 µg/L Arsenate 300 µg/L Arsenite 300 µg/L Arsenate 350 µg/L Arsenite 350 µg/L Arsenate 300 µg/L Arsenite 300 µg/L Arsenate 500 mg/L Arsenite 500 mg/L Arsenate 2 mg/L Arsenite 2 mg/L Arsenate 100 µg/L Arsenite 100 µg/L Arsenate 57 µg/L Arsenite 31 µg/L Arsenate 100 mg/L Arsenate 68 mg/L Arsenite 37 µg/L Arsenate 51 µg/L Arsenite 188 µg/L Arsenate 132 mg/L Arsenite 500 µg/L

10 µg/L 6 µg/L 140 µg/L 10 µg/L 6 µg/L 138 µg/L 74 µg/L 263 µg/L 30 µg/L 249 µg/L 4 mg/L 2 mg/L 20 µg/L 160 µg/L 4 µg/L Ineffective ND at pH 6.0 Ineffective 1:1) are typical of pumped water samples from biofouled water wells in microbially active alluvial aquifers. • Elevated iron and manganese concentrations in pumped groundwater are also typically the result of bacterial activities in the aquifer, including respiratory Fe(III) and Mn(IV) reduction in the presence of abundant organic carbon and also corrosion of metal equipment. If unfiltered samples containing biofilm colloids are digested, very high Fe and Mn levels may be recorded (13). • Biofilm influences have significant impacts on oxidation-reduction potential (3,8,13) and explain discrepancies between Eh values calculated from different couples and between calculations and measured values (8).

With regard to well hydraulic performance,

• Corrosion is enhanced by biofilm action and may occur where corrosion-incrustation index calculations predict that it would not occur. Such microbially influenced corrosion (MIC; 14) can also affect materials not expected to be subject to corrosion, accelerates well aging, and contributes corrosion-product constituents to water samples (15).

• Reductions in specific capacity due to biofouling vary locally, but annual reductions are on the order of 1 to 3.4 on Long Island, NY, ranging up to 47% annually (21), for example. • Considerable biofouling buildup can occur before well specific capacity is impacted. • Wells accessing aquifers with low hydraulic conductivity (K) show faster and more enhanced performance decline compared to wells in formations with higher K. • Higher Fe and Mn levels, particularly when oxidizing conditions prevail, can accelerate performance decline, although rapid biofouling clogging can occur at low levels. Higher total P, sulfate and organic carbon are also associated with enhanced biofouling development and clogging (6,8,22).

• Intermittent total coliform positive test results, mostly caused by bacteria not necessarily part of the total coliform group, but possessing the galactosidase enzyme. It is important to note that corroding, encrusting, and nonencrusting biofouling effects, as well as test results that pose health concerns, often occur simultaneously.

BIOFOULING EFFECTS ON SYSTEM HYDRAULIC PERFORMANCE In a well, biofouling phenomena may encrust or loosely plug well borehole intake areas and screens, pumps, and other downstream equipment. The initial process is the formation of a biofilm on surfaces in the well (casing, screen, pump) and in the aquifer in the vicinity of the well. The time course of this process to result in water quality or pumping problems may vary considerably, depending on site-specific conditions: • Clogging (both formation/well and pump/discharge systems). Well clogging is usually expressed as reduced specific capacity (yield in volume/time unit versus drawdown during pumping).

37

See related items, EVALUATION OF MICROBIAL COMPONENTS BIOFOULING, and WELL MAINTENANCE, for diagnostic methods and their applications where biofouling occurs in wells. OF

BIBLIOGRAPHY 1. Borch, M.A., Smith, S.A., and Noble, L.N. (1993). Evaluation and Restoration of Water Supply Wells. AWWA Research Foundation, Denver, CO. 2. Chapelle, F.H. (1993). Ground-Water Microbiology and Geochemistry. John Wiley & Sons, New York. 3. Smith, S.A. (1992). Methods for Monitoring Iron and Manganese Biofouling in Water Supply Wells. AWWA Research Foundation, Denver, CO. 4. Alcalde, R.E. and Gariboglio, M.A. (1990). Biofouling in Sierra Colorado water supply: A case study. Microbiol. Civ.

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IN SITU BIOREMEDIATION OF CONTAMINATED GROUNDWATER Eng., FEMS Symposium No. 59. P. Howsam (Ed.). E.&F.N. Spon, London, pp. 183–191. Barbic, F., Krajcic, O., and Savic, I. (1990). Complexity of causes of well decrease. Microbiol. Civ. Eng., FEMS Symposium No. 59. P. Howsam (Ed.). E.&F.N. Spon, London, pp. 198–208. Houben, G.J. (2003). Iron oxide incrustations in wells. Part 1: Genesis, mineralogy, and geochemistry. Appl. Geochem. 18: 927–939. Harder, E.C. (1919). Iron-Depositing Bacteria and Their Geologic Relations. U.S. Geological Survey Water Supply Paper 113, U.S. Geological Survey, Washington, DC. Walter, D.A. (1997). Geochemistry and Microbiology of IronRelated Well-Screen Encrustation and Aquifer Biofouling in Suffolk County, Long Island, New York. Water Resources Investigation Report 97-4032. U.S Geological Survey, Coram, NY. McLaughlan, R.G., Knight, M.J., and Steutz, R.M. (1993). Fouling and Corrosion of Groundwater Wells: A Research Study. National Centre for Groundwater Management, University of Technology, Sydney, Australia. Vuorinen, A. and Carlson, L. (1985). Special of heavy metals in Finnish lake ores; selective extraction analysis. Int. J. Environ. Anal. Chem. 20: 179–186. Mouchet, P. (1982). From conventional to biological iron removal in France. JAWWA 84: 158–167. Van Riemsdijk, W.H. and Hiemstra, T. (1993). Adsorption to heterogeneous surfaces. Metals in Groundwater. H.E. Allen, E.M. Perdue, and D.S. Brown (Eds.). CRC Lewis, Boca Raton, FL, pp. 1–36. Smith, S. and Hosler, D.M. (2001). Report of Investigations with Recommendations, Biological Fouling of the Pressure Relief Drainage System, Pablo Canyon Dam, Montana. Bureau of Reclamations, Denver, CO. Videla, H.A. (1996). Manual of Biocorrosion. CRC Press Lewis, Boca Raton, FL. Oakley, D. and Korte, N.E. (1996). Nickel and chromium in ground water samples as influenced by well construction and sampling methods. Ground Water Monitoring and Remediation 16 (1): 93–99. Howsam, P. and Tyrrel, S., (1989). Diagnosis and monitoring of biofouling in enclosed flow systems—experience in groundwater systems. Biofouling 1: 343–351. Smith, S.A. (1990). Well maintenance and rehabilitation in North America: An overview. In: Water Wells Monitoring, Maintenance, and Rehabilitation. P. Howsam (Ed.). E.&F.N. Spon, London, pp. 8–16. Howsam, P., Misstears, B., and Jones, C. (1995). Monitoring, Maintenance and Rehabilitation of Water Supply Boreholes. Report 137, Construction Industry Research and Information Association, London, UK. Smith, S.A. (1995). Monitoring and Remediation Wells: Problems Prevention and Cures. CRC Lewis, Boca Raton, FL. Alford, G. and Cullimore, D.R. (1999). The Application of Heat and Chemicals in the Control of Biofouling Events in Wells. CRC Press Lewis, Boca Raton, FL. Walter, D.A. (1997). Effects and Distribution of IronRelated Well-Screen Encrustation and Aquifer Biofouling in Suffolk County, Long Island, New York. Water Resources Investigation Report 96-4217. U.S Geological Survey, Coram, NY. Cullimore, D.R. (1993). Practical Manual of Microbiology. CRC Press Lewis, Boca Raton, FL.

IN SITU BIOREMEDIATION OF CONTAMINATED GROUNDWATER JIM C. PHILP Napier University Edinburgh, Scotland, United Kingdom

COLIN C. CUNNINGHAM The University of Edinburgh Edinburgh, Scotland, United Kingdom

MARIA S. KUYUKINA IRENA B. IVSHINA Institute of Ecology and Genetics of Microorganisms of the RAS Perm, Russia

INTRODUCTION Globally, the use of groundwater for various purposes is enormous. It is used as potable, industrial (especially cooling), and irrigation water on a huge scale. At least twelve megacities (population over 10 million) could not function without groundwater, and typically at least 25% of the water for these cities comes from aquifers (1). China alone has over 500 cities, and two-thirds of the water for them comes from aquifers. Despite this importance, the number of instances of groundwater contamination due to accidental spills or unsatisfactory disposal is beyond counting. Up to a certain point, contamination can be attenuated by natural processes, especially biodegradation. In this regard, the biologically active zone is the vadose (unsaturated) zone where attenuation rates are highest. Contaminant removal continues in the saturated zone but usually at much lower rates, and migration of contaminants to the saturated zone can disperse the contaminants. Although it brings about dilution, this latter process often cannot be relied upon for complete decontamination. Beyond a threshold, this natural attenuation cannot continue, and a decision has to be made whether or not to intervene with cleanup technology. This decision is now based on risk assessment. Risk assessment usually uses source–pathway–receptor analysis, and its outcome also determines, which treatment technology is to be used if treatment is necessary. RISK ASSESSMENT The objective of source–pathway–receptor analysis is the identification of the linkage(s) between them. Risk-based remedial design has as an objective the selection of the strategy to break the linkages and thus remove the risk. If, for example, there is a source of pollution and potential receptors (usually human) but no pathway to link them, or the pathway(s) can be blocked, then the risk is removed. If the receptor is human, then it is more likely that the risk will be removed by treating the source.

IN SITU BIOREMEDIATION OF CONTAMINATED GROUNDWATER

WHERE IS THE BOUNDARY BETWEEN GROUNDWATER AND SOIL TREATMENT? Necessarily, if groundwater is to be treated by insitu technologies, then the boundary between soil and water treatment becomes blurred, especially for vadose zone treatment. SOME STATISTICS The U.S. EPA has a recorded history of the treatment technologies applied to contaminated soil and groundwater at Superfund and other priority sites, which is updated through annual status reports (ASRs). As this history spans several decades, it is instructive to examine the record to identify the main technologies and trends in usage. Figure 1 shows that ex situ pump and treat methods have dominated groundwater treatment. This involves pumping the groundwater to the surface and selecting a treatment. Monitored natural attenuation (MNA) is not regarded as an engineered technology (2), but it is briefly described. Specifically at Superfund sites (Fig. 2), in situ groundwater treatment remedies have been chosen 169 times at 135 sites (3). Air sparging has been chosen most frequently, followed by bioremediation. However, a comparison of those projects at the predesign, design, or installation stages (preoperational or future) and the operational and complete (current and historical) shows that

Groundwater remedy type

the popularity of bioremediation has increased and that of air sparging has dropped dramatically.

IN SITU BIOREMEDIATION TECHNOLOGIES The treatment of groundwater in-situ has several advantages and disadvantages compared to pump and treat. The advantages include reduced site disturbance, which is crucial when a site is still being used; remediation around or under buildings without disturbance; and reduced worker exposure to volatile compounds. The quoted disadvantages include that site assessment must fully describe the hydrogeology and contaminant distribution, leading to increased cost; difficulty in control of reaction conditions; difficulty in accurately predicting end points; and therefore, the need for careful monitoring of the process. As it is difficult to take samples during in situ treatment, then it is difficult to assess if contamination hot spots have been remediated. Additionally, in bioremediation, several other factors must be considered. Regularly quoted advantages of bioremediation are

Number of sites

Pump and treat In situ treatment Monitored natural attenuation Other

713 135 201 822

Figure 1. Total number of sites with a groundwater remedy (3). Sites may be included in more than one category.

Other 11.9%

Phytoremediation 3.4% Multiphase extraction 3.4%

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• It is often less expensive than other techniques. • Because the contaminants are mineralized, they are permanently eliminated. • Long-term liability is thus eliminated. • It has ‘‘green’’ credentials, representing a sustainable option. Likewise, there are quoted disadvantages: • Bioremediation cannot treat all contaminant types; most notably heavy metals cannot be biodegraded. • Low permeability soil makes in-situ bioremediation of groundwater untenable. • It may be too slow in certain circumstances.

Air sparging 20.3%

PRB 11.9%

Chemical 18.6%

Bioremediation 30.5%

Preoperational

Phytoremediation 3.6% Multiphase extraction 10.9%

Other 18%

Air sparging 41.8%

PRB 16.9%

Chemical 16.9%

Bioremediation 23.6% Operational and complete

Figure 2. In situ groundwater treatment projects by technology, fiscal years 1982–2002 (adapted from Reference 3).

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IN SITU BIOREMEDIATION OF CONTAMINATED GROUNDWATER

• It is often regarded as ‘‘incomplete’’ in that the contaminants may not be removed completely. • Physicochemical conditions that are not ideal for microbial growth can slow the process down considerably. • There is a worry that metabolites may be toxic. As a result, the suitability of bioremediation has to be assessed case by case, and small-scale treatability studies are required (4), which adds to the time and expense of a project.

Monitored natural attenuation may be useful where natural processes within the polluted area are effective in stabilizing or reducing the size of a contaminated groundwater plume, but there cannot be unacceptable impacts or risks to receptors. Monitoring and modeling are used to enable predicting the rate of attenuation and the rate of migration of the pollutant with some degree of confidence. The overall cost-effectiveness of MNA must take into consideration the long-term expenses involved in monitoring compared to the cost of a more intrusive approach, which is likely to be more costly during operational remediation, but requires no longterm monitoring.

Aerobic or Anaerobic? More than any other factor, it is likely that oxygen limitation occurs in groundwaters, and the most rapid biodegradation processes normally involve oxygenase enzymes (5). The supply of oxygen for in situ bioremediation adds considerably to the expense and the technical difficulty. This supply involves the use of either blowers or vacuum pumps to draw air through the system. A measure of the pivotal role of oxygen is in the use of hydrogen peroxide to enhance aeration. That its use has been considered at all is remarkable: concentrations of H2 O2 above 100 to 200 mg/L are toxic to microorganisms; it can be consumed very quickly, limiting treatment to the regions near the injection well; a groundwater circulation system must be created; and good soil permeability is essential. Recent advances in the knowledge of anaerobic biodegradation microbiology have opened up the prospect of greater acceptance of anaerobic bioremediation where maintaining aerobic conditions is not feasible. The existence of microorganisms capable of coupling the anaerobic reduction of Fe3+ to the oxidation of organic compounds shows promise (6). Virtually our entire knowledge of anaerobic metabolism of hydrocarbons has been gained since around 1990 (for reviews, see References 7 and 8). Several alkylbenzenes, alkanes or alkenes, are anaerobically biodegraded by denitrifying, ferric iron-reducing or sulfate-reducing bacteria. Another group of anaerobic hydrocarbon-degrading bacteria are ‘‘proton reducers’’ that rely on syntrophic associations with methanogens (8). As some of the most significant groundwater pollutants are chlorinated solvents, then a role for microbial reductive dehalogenation is feasible. Although slow, even the reductive dechlorination of dioxins is possible (9). A recent discovery has shown that even benzene can be oxidized completely under anaerobic conditions by pure cultures, using nitrate as the electron acceptor (10). Benzene is a particular problem in groundwater because it is relatively soluble and mobile.

Bioventing This method uses indigenous microbes to biodegrade organic contaminants in the unsaturated zone above the water table. Technically, then, it is an in situ source treatment for contaminated soil (13), but it can be considered here as it inevitably results in soil water remediation. Bioventing is, after all, designed to remove contamination from the vadose zone to prevent future contamination of groundwater. It combines supplying extra oxygen with vapor extraction to induce forced airflow through the contaminated area and enhance natural biodegradation (Fig. 3). Air is blown into the center of the area of contaminated soil above the water table and sucked out through peripheral boreholes to off-gas treatment prior to emission to the atmosphere. In its engineering manifestations, it closely resembles soil vapor extraction (SVE), but the two technologies have fundamentally different goals. SVE endeavors to maximize volatilization of low molecular weight contaminants, with some incidental bioremediation. Bioventing, however, endeavors to maximize biodegradation of the contaminant(s) regardless of molecular weight, with some incidental volatilization (14). To enhance the biological process, nutrients can be supplied to the contaminated area. The injection of oxygen stimulates the microbes in the contaminated soil to

Vacuum Blower

Air extraction

Air flow

Monitored Natural Attenuation This technique involves monitoring the natural physical, chemical, and biological processes in soil and groundwater that are used to destroy a pollutant or limit its spread or migration (11). In nearly all situations, however, microbial reactions are the dominant processes driving natural attenuation (12), so it can be considered a long-term in situ bioremediation process.

Vacuum

Water table

Figure 3. Bioventing.

IN SITU BIOREMEDIATION OF CONTAMINATED GROUNDWATER

degrade the organic contaminants to CO2 and water. The process also mobilizes volatile compounds (either present in the soil or produced during biodegradation) to move toward boreholes, making extraction simpler. Bioventing has now been used successfully at over 1000 sites and seems best suited to the bioremediation of middle distillate fuels such as diesel and jet fuel (15), but also nonchlorinated solvents, some pesticides, wood preservatives, and other organic chemicals have been removed. It is not successful in limited permeability soils, and application to the saturated zone relies on the water level being reduced (16). However, it can be successfully combined with biosparging.

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Nutrient tanks Treatment

Recovery well

Injection well

Oxygen & nutrients Groundwater flow

Contaminated zone

Figure 5. Pump and treat.

Biosparging Biosparging is similar to bioventing, in that air is injected, but in this case it is introduced below the water table (in the saturated zone) (Fig. 4) to increase the dissolved oxygen concentration in the groundwater and thus stimulate the activity of the indigenous microorganisms, thereby stimulating aerobic bioremediation. It can also have a similar stimulatory effect in the unsaturated zone. Like bioventing, nutrients may be added to enhance the biological process. The success of the technique depends on adequate diffusion of the injected air away from the boreholes into the surrounding groundwater and soil. The location and number of boreholes depend primarily on the subsurface soil structure and permeability, and like bioventing, biosparging is best suited to permeable soils (16). Like bioventing, biosparging uses equipment that is readily available and easy to install.

show that the remedial objectives have been achieved. Technically, then, when pump and treat is operated as a bioremediation technology, it combines features of ex situ (water is pumped to the surface) and in situ (biological treatment continues underground). Pump and treat is a technology of great flexibility because the surface treatment need not be biological. In fact, a treatment train employing a variety of different technologies is possible at the surface. A much higher degree of process control is possible at the surface, for example, temperature, pH. For these reasons, pump and treat has historically been a very popular treatment for contaminated groundwater. However, there have been frequent questions over the efficacy of pump and treat, probably because, as the technology has become accepted, expectations have become too high (17). Permeable Reactive Barriers

Pump and Treat The contaminated aqueous phase from the saturated zone is pumped, via a recovery well, to a treatment tank on the surface. In the treatment tank, nutrients, oxygen, and other electron acceptors (e.g., sulfate and nitrate) are added before the groundwater is pumped back into the ground via an injection well and recirculated through the contaminated zone (Fig. 5). The oxygen and nutrients in the injected groundwater stimulate the microbes in the contaminated zone to biodegrade contaminants dissolved in the groundwater and present in the soil. Groundwater extraction and injection continue until monitoring data

Pressure relief valve

A PRBs consists of a reactive material that is placed in the path of flowing groundwater (Fig. 6), and due to the permeability of the chosen reactive material, it removes contaminants from the flow as the groundwater passes through. These barriers allow the passage of water while prohibiting the movement of contaminants. Agents within the barrier are materials such as zero-valence metals, for example, Fe0 , chelators, sorbents, and microbes (18). The contaminants are either degraded or retained in concentrated form by the barrier material, which may need to be replaced periodically. PRBs have considerable flexibility in that combined chemical and biological treatments are possible within

Pressure gauge Flow meter

Prefilter

Cap

Manifold

Pollutants enter aquifer Porous treatment medium

Sparge control system Groundwater flow

Water table

Aerated zone

Decreased contaminant concentration

Lower confining aquiclude Treatment wall

Figure 4. Biosparging.

Figure 6. Permeable reactive barrier.

PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER

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the same barrier, or barriers can be sequenced. For example, a contaminated aquifer beneath a petrochemical site may typically contain both chlorinated solvents and petroleum hydrocarbons. The chlorinated compounds may be removed by zero-valence iron in the permeable reactive barrier. With time, the permeable barrier will become colonized by bacteria, and bioremediation of petroleum hydrocarbons will occur, given the correct conditions. Alternatively, hydrocarbon-oxidizing bacteria may be deliberately added (the practice of bioaugmentation). Another aspect of the flexibility of the PRB is that in many cases it can be installed at working sites with minimal disruption to normal operations. COSTS The costs of all remediation technologies are constantly shifting and vary greatly from country to country. So quoted figures per tonne or per cubic meter of contaminated groundwater or soil would soon be out of date. Rather, it should be possible to identify economies of scale if confidence is gained in a technology and it is used increasingly at full scale. This confidence in bioremediation has been lacking in the past, and in many parts of the world, it is still not accepted at full scale. If a technology is following a typical economies of scale pattern, then the more it is used, the more costs will decrease. The US EPA (19) examined the correlation between unit costs and quantity treated for six different soil and groundwater remediation technologies: bioremediation, thermal desorption, soil vapor extraction, on-site incineration, pump and treat, and permeable reactive barriers. Some important findings were reported: 1. Four of the six technologies (bioventing, thermal desorption, soil vapor extraction and pump and treat) evidenced a correlation between unit cost and quantity treated, thus exhibiting an economies of scale pattern. 2. Bioventing had the best correlation of these four technologies. 3. No other bioremediation technology exhibited any such correlation. 4. Pump and treat groundwater remediation systems showed a correlation for both unit capital costs and unit average operating costs. The lack of correlation for the other bioremediation technologies may simply result from lack of data from a sufficiently large number of full-scale projects, but the signs are encouraging that bioventing has proven to be a market success. BIBLIOGRAPHY 1. Morris, B.L., Lawrence, A.R.L., Chilton, P.J.C., Adams, B., Calow, R.C., and Klinck, B.A. (2003). Groundwater and Its Susceptibility to Degradation: A Global Assessment of the Problem and Options for Management. Early Warning and Assessment Report Series RS.03-3. United Nations Environment Programme, Nairobi, Kenya.

2. U.S. EPA. (1999). Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action and Underground Storage Tank Sites. OSWER Directive 9200.4-17P. 3. U.S. EPA. (2004). Treatment Technologies for Site Cleanup: Annual Status Report. 11th Edn. EPA-542-R-03-009. 4. Boopathy, R. (2000). Factors limiting bioremediation technologies. Bioresource Technol. 74: 63–67. 5. Dagley, S. (1975). A biochemical approach to some problems of environmental pollution. Essays Biochem. 11: 81–138. 6. Crawford, R.L. (1991). Bioremediation of groundwater pollution. Curr. Opin. Biotechnol. 2: 436–439. 7. Holliger, C. and Zehnder, A.J.B. (1996). Anaerobic biodegradation of hydrocarbons. Curr. Opin. Biotechnol. 7: 326–330. 8. Heider, J., Spormann, A.M., Beller, H.R., and Widdel, F. (1999). Anaerobic bacterial metabolism of hydrocarbons. FEMS Microbiol. Rev. 22: 459–473. 9. Adriaens, P., Fu, Q., and Grbic-Galic, D. (1995). Bioavailability and transformation of highly chlorinated dibenzo-pdioxins and dibenzofurans in anaerobic soils and sediments. Environ. Sci. Technol. 29: 2252–2260. 10. Coates, J.D., Chackraborty, R., Lack, J.G., O’Connor, S.M., Cole, K.A., Bender, K.S., and Achenbach, L.A. (2001). Anaerobic benzene oxidation coupled to nitrate reduction in pure culture of two strains of Dechloromonas. Nature 411: 1039–1043. 11. BioWise. (2000). Contaminated Land Remediation: A Review of Biological Technology. DTI/BW/18/5000/12/2000/NP. 12. Smets, B.F. and Pritchard, P.H. (2003). Elucidating the microbial component of natural attenuation. Curr. Opin. Biotechnol. 14: 283–288. 13. U.S. EPA. (2001). Use of Bioremediation at Superfund Sites. EPA 542-R-01-019. 14. U.S. EPA. (1995). Bioventing Principles and Practice. Volume 1: Bioventing principles. EPA/540/R-95/534a. 15. U.S. EPA. (1995). Soil Vapor Extraction (SVE) Enhancement Technology Resource Guide. EPA/542-B-95-003. 16. Martin, I. and Bardos, P. (1995). A Review of Full Scale Treatment Technologies for the Remediation of Contaminated Soil. EPP, Richmond, VA. 17. U.S. EPA. (1996). Pump-and-Treat Ground-Water Remediation. A Guide for Decision Makers and Practitioners. EPA/625/R-95/005. 18. U.S. EPA. (2001). Treatment Technologies for Site Cleanup: Annual Status Report (Tenth Edition). EPA-542-R01-004. 19. U.S. EPA. (2001). Remediation Technology Cost Compendium—Year 2000. EPA-542-R-01-009.

PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER DAVID B. VANCE ARCADIS G&M, Inc. Midland, Texas

JAMES A. JACOBS Environmental Bio-Systems, Inc. Mill Valley, California

Based on the origin and susceptibility to microbial interaction, hydrocarbons may be divided into two broad

PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER

classes: petroleum hydrocarbons and complex chlorinated hydrocarbons. Petroleum hydrocarbons are largely associated with the production, storage, or use of fuels, lubricants, and chemical feedstocks. Petroleum hydrocarbons have been demonstrated to be biodegradable by numerous species of bacteria. Over a period of 3.5 billion years, bacteria have been able to evolve genetic resources that allow some of them to potentially use petroleum hydrocarbons as a source of food. Complex industrial hydrocarbons include chlorinated aliphatic and aromatic hydrocarbons, MTBE, pesticides and herbicides, and polymers. These new synthetic compounds have been manufactured for about 100 years. Bacteria have not had time to evolve the genetic information required to utilize them as a source of food. Due to recalcitrance to microbial attack, these complex industrial chemicals are termed xenobiotic. Successful in situ bioremediation of groundwater has been demonstrated at sites impacted with petroleum hydrocarbons. The most successful method of application has been through stimulation of indigenous microbial species. Microbial stimulation is the process of ensuring that environmental conditions, nutrient availability, and requirements for an electron acceptor are adequate in the contaminated portions of the aquifer. The most common cause for failure of saturated zone in situ bioremediation is the lack of adequate mass transport of the electron acceptor (usually oxygen). In this regard, the physical setting of the site is critical. Overall permeability and the scale and degree of heterogeneity are the factors governing the advective and diffusional transport rates of contaminants and remediation reagents in the subsurface. If mass transport rates are too low, saturated zone in situ bioremediation is not a viable option. Given adequate mass transport properties, site-specific microbiological conditions can also impact the process. Unfortunately, the presence of indigenous microbes and efficient mass transport may still prove insufficient for effective bioremediation. Specific reasons for the poor performance of in situ bioremediation systems relate to unoptimized subsurface conditions. There is uncertainty with regard to the effect of hydrocarbon availability on the effectiveness of biodegradation. Can bacteria degrade hydrocarbons adsorbed to surfaces or degrade hydrocarbons with low levels of solubility? Or must the hydrocarbon be solubilized before it can be biodegraded? Contradictory laboratory evidence and field evidence have been published for both scenarios. With the predominance of evidence indicating that solubilization must take place, degradation reactions with extracellular bacterial exudates are much less likely. The answer is likely consortia specific and dependent on the ability of the bacteria to synthesize appropriate biosurfactants. This ability may be absent in some instances. Although petroleum hydrocarbons are amenable to primarily aerobic biodegradation, for it to occur the indigenous bacteria must have the appropriate genetic information. This genetic information is specific and precise. The presence of a specific hydrocarbon will

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stimulate the synthesis of an oxygenase enzyme that is expressly configured to react with that stimulating hydrocarbon. For remediation, indigenous microbes generally possess the genetic information required for appropriate enzyme production and the contaminant will stimulate the production of those enzymes. General microbial stimulation has the potential to produce a large amount of biomass that may not take part in the biodegradation process and actually cause harm through biofouling and plugging of injection wells, galleries, or surrounding formations. There is potential to lose critical subsurface mass transport capabilities. BACTERIAL TRANSPORT There are practical limits to the degree of cleanup obtainable using bioremediation. Hydrocarbons at the low parts per million (ppm) level may not be capable of supporting significant levels of microbial activity even under stimulation. Sites with relatively high levels of hydrocarbon impact may actually be better candidates for bioremediation than those lightly impacted at levels slightly above regulatory action levels. Toxicity related to the presence of heavy metals, such as chromium, arsenic, or lead, or low temperature of the groundwater have been observed to inhibit bacterial growth in a variety of settings. Stimulating electron acceptors must also be available at sufficient concentrations; native sulfate concentrations less than about 20 mg/L do not stimulate sulfate reducing bacteria even in the presence of usable carbon substrates. Xenobiotic industrial compounds are often recalcitrant to direct aerobic microbial attack. However, over the last 20 years a biodegradation process termed cooxidation (or cometabolism) has been successfully demonstrated by researchers. For example, the aerobic degradation of trichloroethylene (TCE) has been accomplished using monooxygenase and dioxygenase enzymes produced through the use of petroleum hydrocarbons as a metabolizable substrate (food source) and stimulus for enzyme production. This general process is termed co-oxidation and the hydrocarbon substrate used as a food source is the cometabolite. Many different hydrocarbon substrates have been observed to stimulate the generation of cooxidation enzymes. The currently known cometabolic substrates fall into two broad classes: 1. Analog substrates, which are hydrocarbons that have a geometry similar to the targeted xenobiotic compound. 2. Methanotrophic (which is different than methanogenic) microbial systems have proved particularly effective at generating xenobiotic active enzymes. Enzymes with co-oxidizing potential have a strong natural affinity for the hydrocarbon that originally stimulated its generation. The enzyme is genetically tailored to the compound used as a food source. Over 300 mol of methane are required to biodegrade 1 mol of TCE via co-oxidation. The efficiency of the co-oxidation process is extremely poor. Under field conditions where

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PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER

mass transport is a critical success factor, a 300-fold decrease in the effectiveness of the reactants in the contaminated zone often can be impractical. Accurate assessment of potential limiting factors such aerobic terminal electron acceptor (dissolved oxygen), geochemical conditions (pH, temperature, conductivity), and macronutrients (orthophosphate and ammonia as nitrogen) should be documented as part of the bioremediation evaluation process. Bioremediation is a dynamic process requiring monitoring of the hydrological, geochemical, and biological conditions over the life of a project. NATURAL ATTENUATION: GROUNDWATER REMEDIATION BY NO ACTION In a time of reappraisal for the allocation of financial resources to environmental action, a question of ever increasing importance is the consequence of no action concerning the release of petroleum or chlorinated hydrocarbons into groundwater. An important portion of that answer comes from the application of site-specific health based risk assessments. However, in instances where human consumption or exposure is not an issue, no action may be a reasonable alternative, even at elevated dissolved contaminant concentrations. The issue then becomes the determination of the consequence of no action under conditions where the sole process for remediation is natural attenuation. The physical, chemical, geological, and biological processes that take place in a contaminated aquifer are complex. In most instances, a ‘‘native’’ aquifer is in a long-standing state of chemical equilibrium between the groundwater and the geologic matrix through which it flows. The release of anthropogenic hydrocarbons into an aquifer upsets that equilibrium. The dissolved concentration of the contaminant as it migrates through the aquifer is controlled by adsorption, dispersion, volatilization, and degradation. Adsorption affects the overall residence time of the release and dispersion affects the downgradient shape and dissolved concentration in the plume. Only volatilization and degradation contribute to the removal of contaminant from the aquifer, and at low concentrations degradation is the dominant mechanism for attenuation. The mechanisms for attenuation through degradation can be broadly divided into two categories, biological and abiotic chemical action. This discussion is predicated on relatively ‘‘normal’’ groundwater conditions under which biological action proceeds at a rate orders of magnitude greater than abiotic processes. Extremes of pH, redox conditions, ionic strength, or temperature may make an exception to that generalization. Transformation can be chemically complex, dependent on the environmental conditions described above and affected by aquifer heterogeneity related to granular or fracture variability. The factor controlling the rate of aerobic degradation is the availability of oxygen and the rate at which it can be introduced into the groundwater (through the groundwater–table interface) or the rate at which oxygen-rich groundwater can pass through the zones of adsorbed contamination. Each pound of petroleum

hydrocarbons requires about 3.08 pounds of oxygen for complete degradation (1). Typical in situ aerobic decay rates for groundwater are in the range of 35 µg/L·d (equivalent to about 0.5 oz/d per cubic yard of aquifer matrix). Natural attenuation occurs both in the source zone and in the dissolved phase plume. In the source zone, oxygen will be rapidly consumed and portions of the aquifer will then host anaerobic degradation. Anaerobic degradation is limited by the availability of appropriate anaerobic electron acceptors such as nitrate, sulfate, or iron. When their availability is limited, degradation will stop after the production of aliphatic and aromatic organic acids; similarly, at low levels of dissolved oxygen (DO), aerobic degradation may also stop with the production of organic acids. The intrinsic biodegradation process and the alternative terminal electron acceptors are shown in Fig. 1 (2). Optimum aerobic biodegradation occurs with the dissolved oxygen above 2 mg/L. Below that, the aerobic degradation rate of aromatic hydrocarbons will decrease dramatically. Conversely, under complete anaerobic conditions, nitrate reducing, iron reducing, and suflate reducing bacteria can effectively degrade hydrocarbons. However, at DO concentrations as low as 0.1–0.4 mg/L, anaerobic degradation rates will be reduced to just a few percent of optimum. Because of all the mechanisms described above, hydrocarbon plumes tend to achieve a stable shape and size even when there is a continuous source of free phase hydrocarbon release. Steady state is achieved when the area of the plume edge is great enough to provide for a natural degradation rate equivalent to the rate of hydrocarbon infiltration. The edges of the dissolved plume do not have enough DO to support optimum rates of aerobic degradation but have too much DO to allow for optimum anaerobic degradation. The interior of a plume will support anaerobic natural attenuation, which is typically limited by the availability of iron in the mineral matrix and sulfate in the native groundwater and to a lesser extent the mineral matrix as well. However, once the source of hydrocarbon has been removed, a dissolved plume will narrow and dissipate from the edges inward, due to the availability of DO from groundwater along those edges. The selection of a no action natural attenuation option should be based on an appropriate analysis of data gathered during the assessment of the site. Firstorder decay rates are appropriate for the evaluation of degradation kinetics at low concentrations, less than 1 ppm (an appropriate level to assume at the periphery of a plume). Given first-order decay rates, the analysis has a focus that is twofold—the effect of attenuation over time and the effect over distance. Attenuation over time is measured at the edges of a plume using concentration measurements gathered repeatedly from specific monitor wells. The minimum recommended time is one year, with quarterly sampling from the selected monitor wells. The data for each well is then semilog plotted as log concentration against time. The slope of the line is the first-order decay constant in percent per day.

PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER

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Intrinsic biodegradation processes

Organics Concentration

Fe(II)

CH4

H2S

N2 Aerobic respiration

pE (mV)

Metabolic products

CO2

Denitrification

Iron (III) reduction

+100 0

O2

NO3−

Fe(III)

Sulfate Methanogenesis reduction Dominant electron acceptors SO42−

CO2

−100

Source Plume migration/ ground water flow

Attenuation with distance more accurately incorporates the effects of aquifer heterogeneity. Data for this analysis is obtained from a minimum of three monitoring wells, preferably along the long axis of the plume. This data is semilog plotted as log concentration versus distance. The slope of this line is equal to the decay constant divided by the groundwater velocity. With this data, decisions can be made based on the sitespecific contaminant dynamics under no action natural attenuation. This, in conjunction with a health based risk assessment, can allow for sound decision-making by the business and regulatory community. In summary, the adoption of a no action alternative is most applicable to the dissolved phase plume only. Except for volumetrically small releases, it will still be necessary to remove or remediate the source zone of an impacted aquifer, after which natural attenuation may be a reasonable approach to the residual dissolved phases. Also implicit in this approach is that no action does not preclude the performance of requisite assessment activity, which can represent a significant long-term liability in some cases. Nonetheless, after proper source abatement, assessment, and analysis, the reliance on natural attenuation mechanisms for the final stages of cleanup is a cost effective and, if properly managed, environmentally sound resolution to aerially extensive dissolved phase hydrocarbon contamination. NATURAL ATTENUATION: THE EFFECT OF PUMP AND TREAT REMEDIATION In situ groundwater remediation has matured over the past 25 years, particularly with regard to understanding the dynamics of the interactions between contaminants, the impacted saturated soil matrix, and microbiological activity. Recent interest in the phenomenon of natural

Figure 1. Intrinsic biodegration process (2). Petroleum hydrocarbons are referred to as organics in the diagram.

attenuation has served to illustrate the variety of microbial ecosystems that are present in a contaminant plume, each system determined by redox conditions and availability of electron acceptors. While natural attenuation is an attractive alternative to those responsible for groundwater contamination, the regulating communities are more skeptical. The need for proactive groundwater pumping remediation and the efficacy of natural attenuation pose a potentially complex balance that is governed by the subsurface conditions of each individual site. There is no universal applicable rule for the resolution of that balance. It is the responsibility of remediation designers to make those site-specific determinations and to provide the regulating community with information sufficient to support the proactive and the natural attenuation portions of each individual cleanup. Our purpose here is to point out some of the most significant factors impacting that balance. The first and most dominant control is the nature of the saturated soil matrix. Several factors must be evaluated for remedial design: 1. The degree and scale of sediment heterogeneity, which determines how much and what portion of a contaminated aquifer can be affected with advective groundwater flow. Low permeability regions must rely on diffusional transport, which will dominate the overall remediation rate in the treatment zone. 2. The time of exposure to the contaminant is a direct function of the impact of heterogeneity described above. The contaminant will diffuse into the nonadvective portions of the aquifer soil matrix. At a minimum, remediation will take as long as the initial exposure. Due to the adsorptive retardation reactions, remediation is likely to take longer than the exposure time.

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PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER

3. The geochemical composition of the soil matrix is also a factor in remedial design. Carbonaceous material and clays have a much higher propensity for the adsorption of organic contaminants. Iron oxides, in turn, have high adsorptive capacity for metal contaminants. Iron and sulfur minerals may be sources of electron acceptors as redox conditions are modified through the interaction of indigenous microbial populations and the contaminant. 4. The background geochemical makeup of the groundwater as well as that in the contaminant plume is an important remedial design factor. Dissolved oxygen, sulfate, nitrate, and iron can all potentially serve as alternate terminal electron acceptors to aid in the degradation of organic contaminants. Ferrous iron, hydrogen sulfide, and carbon dioxide are indicative end products of those reactions. 5. The distribution of the organic contaminant is another factor affecting remedial design. Free phase hydrocarbons should be recovered proactively with an extractive technology (pump and treat) or in the case of CVOCs possibly an in situ chemical oxidation injection technology. In cases where impact is shallow, excavation and disposal is still an extremely viable option. The treatment of dissolved and adsorbed hydrocarbons is the point at which the balance between proactive remediation and natural attenuation must be determined. One of the most important contributions that a pump and treat system makes to the in situ remediation of contaminated groundwater is plume capture and hydraulic control in the source zone and the core of the dissolved and adsorbed plumes. Background groundwater that is drawn through the plume perpendicular to the natural groundwater flow direction must also be evaluated. In the past, the focus of pump and treat remediation has been on how it acts to flush and remove the contaminant. The contributions made by recent developments on the mechanisms of natural attenuation reside in the role of electron acceptors present at background concentrations within the aquifer. From the exterior to the interior of a plume, the specific electron acceptor zones are aerobic, denitrification, sulfate/iron reducing, and methanogenic. The boundary between each specific redox zone and the active electron acceptor is controlled by the kinetics of the degradation process in each zone and the advective transport rates of groundwater through that zone. In most instances, the dominant effect is the groundwater transport rate. The natural concentrations of these electron acceptors cover a wide range. Natural oxygen levels commonly range from 2 to 8 mg/L. Groundwater sulfate concentrations in soils derived from sedimentary rocks are typically in the 25-mg/L range, with higher values of several hundred mg/L not uncommon. Ferric oxides are commonly present in soils in the range of 0.5–5%; the ability of indigenous iron reducing bacteria to access that iron will vary from location to location. Given adequate permeability and the presence of appropriate electron acceptors, natural

enhancement of pump and treat systems is possible and worth the relatively inexpensive analyses (some of which can be done with field kits) required to evaluate.

NATURAL ATTENUATION: TRANSVERSE DISPERSION AS THE NATURAL DRIVING FORCE Dispersion is the process by which the interface of contaminated groundwater with native groundwater does not remain abrupt. The leading edge of a contaminant plume will arrive at a given point more rapidly than it would if advection alone were the acting driving force. The mean transport velocity of the contaminant mass remains the same, but concentration gradients are setup. This occurs simultaneously with the phenomenon of the contaminant occupying, with time, an increasing volume of groundwater. There are two fundamental types of dispersion—longitudinal and transverse. Both are accentuated by the inhomogeneous and anisotropic physical configuration of the permeable matrix within a groundwater system. There is also a contribution to both from diffusional transport as well. Longtitudinal dispersion is caused by differences in groundwater velocity through pore spaces that vary in width or tortuosity. The result is dispersion that occurs along the direction of groundwater flow. Transverse dispersion is driven by groundwater flowing around individual particles in the aquifer matrix; the effect occurs perpendicular to the groundwater flow direction. Transverse dispersion is effective in mixing contaminated groundwater with native groundwater at the edges of an elongated plume and occurs only when there is a point source of contamination. In the context of natural attenuation, longitudinal dispersion is purely a dilution phenomena; transverse dispersion provides an influx of electron acceptors. Both types of dispersion are dependent on horizontal and vertical variations in permeability. Increasing anisotropy and heterogeneity increase the magnitude of dispersion. Groundwater velocity also plays a role: At low velocities the effects of diffusion may equal those of dispersion. The ratio of longitudinal to transverse dispersivity ranges from 1 to 24; most commonly, horizontal transverse dispersivity is 20% to 10% of the longitudinal dispersivity, and vertical transverse dispersivity is 2% to 1% of the longitudinal dispersivity. Instead of ‘‘football’’ shaped plumes, this difference in vertical and horizontal dispersivity tends to generate plumes that in three dimensions are ‘‘surfboard’’ shaped. The determination of the specifics of the effect of dispersivity on a contaminant concentration at a given time and location in an aquifer is an extremely complex process that requires detailed knowledge of the physical configuration of the aquifer matrix and solution to partial differential equations for final values. In many instances, a purely empirical approach is the only practical means of assessment rather than modeling. The physical scale that is examined also has an impact: values of dispersivity change as one examines an aquifer on the scale of inches, feet, or thousands of feet.

PROCESS LIMITATIONS OF IN SITU BIOREMEDIATION OF GROUNDWATER

Transverse dispersion is what serves to physically mix groundwater containing contaminants with adjacent groundwater that is unimpacted and contains natural electron acceptors. The phenomenon is probably best understood in terms of angle of divergence, that is, the angle between the two edges of the plume as it migrates from a point source. In granular materials that angle can be as low as 2◦ and in groundwater flowing through fracture systems as high as 20◦ . The most accurate determination of the angle of divergence is from information gathered as close as possible to the point representing the source of contamination. The picture we normally have of a contaminant plume as it migrates away from a point source is one that is (in the two horizontal dimensions) tear drop shaped. The plume initially spreads downgradient and cross gradient until some point is reached at which the distal edges of the plume first travel parallel to the advective groundwater flow direction, and then begin to turn inward to the plume axis and close. The distal portion of the plume is the dominant area where longitudinal dispersion (as well as transverse dispersion) is in effect. Adsorption and other attenuation reactions also take place, but for the purposes of this discussion those effects are ignored. Under pure hydrodynamic effects, a plume should dilute itself taking the shape of a cone at a constant angle of dispersion. This assumes that the nature of the geologic matrix and groundwater velocity stay constant, which is unlikely, especially in the vertical component. The most common result is that the downgradient shape of the plume is a consequence of attenuation caused by the migration of electron acceptors into the margins of the plume through the force of transverse dispersion—not through purely hydrodynamic flow effects. NATURAL ATTENUATION: METHANOGENIC SYSTEMS The ultimate no action alternative may be the use of methanogenic microbial systems. In this case, the electron acceptor is the contaminating hydrocarbon that may contribute to the degradation of petroleum and xenobiotic hydrocarbons There is a difference between methanotrophic bacteria and methanogenic bacteria. Methanotrophs use oxygen to oxidize methane into carbon dioxide (CO2 ). Methanotrophic bacterial systems have received a great deal of attention over the last 15 years since it has been found that methane monooxygenase (the enzyme generated by methanotrophs to react with methane) can degrade a wide variety of chlorinated hydrocarbons. The process is known as cometabolism and is definitely an aerobic process. Methanogenesis is the process of degrading hydrocarbons with the end product being methane (CH4 ) gas and carbon dioxide. The general reaction is as follows: 2C (organic contaminant) + 2H2 O → CO2 + CH4 This is a strictly anaerobic process; methanogenic bacteria are poisoned by the presence of oxygen at levels as low as 0.18 mg/L of soluble oxygen (as O2 ).

47

The redox conditions under which these two different microbial systems operate are literally at opposite ends of the spectrum: methanotrophic reactions occur at the Eh range of +250 mV while methanogenic reactions occur at the Eh range of −200 mV. As an aside, methanogenic bacteria are one of the three classes of bacteria termed Archaebacteria, which are representative of organisms that first appeared on Earth some 3.5 billion years ago. Although their activity is inhibited by oxygen, these bacteria are robust enough to appear in a wide variety of natural locations such as the intestinal tracts of ruminant mammals like cows, sewage digesters, groundwater, and soil. The concentrations of methane and carbon dioxide are expressed as mole percent. This data is somewhat remarkable since this should be a difficult reaction to initiate. Oxygen, nitrate, and sulfate are all toxic or inhibitory to methanogenic activity. That means that at any site at which evidence of methanogenesis is present (methane gas), there has been a series of biodegradation reactions that have consumed the alternative electron acceptors. Migrating from the exterior to the interior of a plume, the type of redox conditions (Eh) will change from + 250 mV to − 200 mV, with oxygen, nitrate, ferric iron, and sulfate progressively being consumed. The degradation of chlorinated xenobiotic compounds under methanogenic conditions is particularly enhanced via reductive dehalogenation reactions that involve substitution of hydrogen in the CVOC carbon chain. This can take place from the direct actions of bacteria and from the presence of molecular hydrogen that is produced from water during the methanogenic process. Conceptually, methanogenesis might be considered ‘‘the ultimate of no action alternatives.’’ The physical/chemical requirement for the removal of all other potential terminal electron acceptors infers that the hydrodynamics of such a system are relatively quiescent. The transport of alternate electron acceptors into the core of the contamination plume must be at a rate slow enough to allow for the consumption of all alternate electron acceptors before the methanogenic core zone is reached. If this situation occurs, it can be a positive argument for absolutely no action involving pump and treat systems. That would increase groundwater velocities, introducing inhibitory electron acceptors into the active methanogenic zone. Groundwater systems in fine grained soils, where transport properties are poor, would be ideal for the exploitation of methanogenic degradation. Of course, one problem with methanogenic degradation is the kinetics of the process. Methanogenic degradation occurs at rates that are orders of magnitude slower than the rates seen with other electron acceptors. However, at some sites with poor transport conditions or inaccessibility (i.e., a deep groundwater table), this may still be a reasonable alternative. The number and types of hydrocarbons degraded under methanogenic conditions are very limited. Some laboratory studies have demonstrated the methanogenic degradation of toluene and o-xylene with no degradation of m-xylene, p-xylene, ethylbenzene, or benzene. Others have found

48

BLACK MESA MONITORING PROGRAM

evidence of degradation of benzene with recalcitrance toward other compounds. The methanogenic process appears to be extremely selective and not capable of the complete degradation of all contaminants present in a typical hydrocarbon plume on the order of years. In addition, there is likely a great deal of microbial heterogeneity with specific degradation capacity varying from location to location. In instances where a specific compound has been released (such as toluene used as a solvent), methanogenesis may be a viable natural attenuation process. It also has value in the natural dehalogenation of chlorinated compounds. In instances where an aquifer has good transport qualities and a supply of natural electron acceptors, natural attenuation alone can be a viable process and acceptable to the regulatory community. However, responsible parties must be prepared for the longterm monitoring that will often be required in support of a natural attenuation program—more accurately termed monitored natural attenuation by the regulating community. In other cases, supplementation of electron acceptors or improvement of groundwater dynamics through pump and treat may provide an adequate minimal approach. But, in most instances, methanogenesis as the ‘‘ultimate of no action alternatives’’ is not going to be practically applicable.

2. Sloan, R. (2004). Slide: intrinsic biodegradation processes. In: Aerobic and Anaerobic Bioremediation and Monitored Natural Attenuation of VOCs Short Course. National Ground Water Association (NGWA) Conference on MTBE and Perchlorate, Costa Mesa, CA, June 3–4, p. 102.

BLACK MESA MONITORING PROGRAM Bureau of Indian Affairs and Arizona Department of Water Resources—U.S. Geological Survey

PROBLEM The N aquifer is the major source of water for industrial and municipal users in the 5,400 square-mile Black Mesa area of northeastern Arizona. The aquifer consists of three rock formations-the Navajo Sandstone, Kayenta Formation, and Wingate Sandstone, which are hydraulically connected and function as a single aquifer. Annual withdrawals from the N aquifer for industrial and municipal use have increased from about 70 acre-ft in 1965, to 4,300 acre-ft in 1972, to 7,700 acre-ft in 2000. The Navajo Nation and Hopi Tribe live in the Black Mesa area, and they depend on ground water from the N aquifer to

BIBLIOGRAPHY 1. Kun, J. (2004). Practical Design Calculations for Groundwater and Soil Remediation. CRC Press, Boca Raton, FL, p. 219.

This article is a US Government work and, as such, is in the public domain in the United States of America.

BLACK MESA MONITORING PROGRAM

meet municipal, domestic, livestock, and irrigation needs. In addition, the springs and streams fed by groundwater discharge are an important part of their culture. Peabody Western Coal Company began operating a strip mine in the northern part of Black Mesa in 1968. The company withdraws water from the N aquifer for a slurry pipeline to transport coal to a powerplant in Laughlin,

49

Nevada. Withdrawals by the company accounted for about 70–80 percent of the total withdrawals in the early and mid 1970’s and have been about 60 percent of the total withdrawals from 1978 to 2000. The Navajo Nation and Hopi Tribe have been concerned about the effect of this increasing pumpage on longterm water supply, discharge in streams and springs,

50

BRINE DEPOSITS

and quality of ground water. This concern led to the establishment of a long-term program in 1971 to monitor ground-water levels, ground-water discharge, groundwater quality, and surface-water discharge. From the 1960’s to 2001, water levels in the N aquifer declined by over 50 feet in 9 of 19 long-term monitoring wells in confined areas. No large changes have been detected in ground-water levels in unconfined areas, ground-water discharge, ground-water quality, or surfacewater discharge. OBJECTIVE To collect hydrologic data in a monitoring network that is designed to determine the long-term effects of industrial and municipal ground-water withdrawals on the N aquifer. APPROACH A long-term monitoring program has been established to collect hydrologic data in the Black Mesa area. Data are collected that describe the ground-water system, surface-water flow, and ground-water quality. Continuous measurements of ground-water levels have been made in six wells since 1972. Continuous data describing streamflow have been collected for 23 years in Moenkopi Wash, and for 3 to 6 years in three other streams. Once a year, ground-water levels are measured in about 26 wells, discharge is measured from 4 springs, and waterquality data are collected from 12 wells and 4 springs. Annual ground-water withdrawal data are collected from about 35 municipal well systems. Peabody Western Coal Company provides annual ground-water withdrawal data for industrial use. These hydrologic data are entered into a computer data base. A report is prepared each year of the program. The report contains the data collected each year, and it shows comparisons of annual and long-term changes in groundwater levels, ground-water discharge, surface-water flow, and water quality. Long-term water-level changes in the six continuous-observation wells show that water levels in those confined areas have declined 80 to 140 feet and water levels in those unconfined areas have not changed.

51

Mining of the Interior Department uses the data from this program to facilitate their oversight and regulation of the coal mining operations of Peabody Western Coal Company. The long-term ground-water, surface-water, and waterquality data collected for this program provides an important opportunity to investigate and gain a better understanding of a hydrologic system of bedrock geology in an arid climate in which there are many competing water-use interests.

BRINE DEPOSITS STEPHEN M. TESTA Mokelumne Hill, California

INTRODUCTION Brines are warm to hot, saturated to nearly saturated, highly saline, ocean and lake waters, containing Ca, Na, K, and Cl and minor amounts of other elements, typical of pore fluids in restricted basins. From an oceanographic perspective, brine is seawater that, due to high evaporation rates or freezing, contains more than the usual amount of dissolved salts, typically about 35%. A brine lake is essentially a salt lake. Brine lakes, like the Dead Sea, the Great Salt Lake in Utah, the Mono Lake region in northern California, and the Salton Sea region in southern California, develop as a result of high evaporation rates in an arid desert that lacks an outlet to the ocean (Fig. 1). The salt is derived from either minerals washed out of the surrounding watershed or from a geological deposit or area previously connected to the ocean. A brine cell or pocket is a small inclusion usually the shape of an elongated tube about 0.05 mm in diameter that contains residual liquid more saline than seawater. A brine pit is a salt well or opening at the mouth of a salt

RELEVANCE AND BENEFITS The long-term available supply of water in the Black Mesa area is critical to many parties. The Hopi Tribe and Navajo Nation use water to meet their needs for public supply, irrigation, and livestock. In addition, sustained springflow and streamflow are important to their culture. Peabody Western Coal Company uses water to transport coal in a slurry pipeline to a powerplant. The hydrologic data collected in this monitoring program are needed to understand the available water supply and the effects of industrial and municipal ground-water withdrawals. The U.S. Geological Survey has a commitment to provide data and information to Indian Tribes and other Interior Department Agencies. The Office of Surface

Figure 1. Mono Lake, a terminal brine lake situated in northeastern California. Tufa occurs at saline lakes worldwide, but not in the abundance and variety in shape as at Mono Lake. Mono brine shrimp are unique to this lake.

52

BRINE DEPOSITS

spring, from which water is taken to be evaporated for making salt. Brines also occur in the subsurface notably as subsurface oil-field waters and geothermal mineralizing fluids. Subsurface brines in sandstone and other porous rocks are largely regarded as connate or buried seawater. Some brines also form locally by solution of rock salt beds. SEAWATER BRINES Seawater is relatively complex, but also relatively concentrated (1). Its principal ions are Na+ , Mg2+ , Ca2+ , K+ , Cl− and SO4 2− . Minor constituents include the elements Br, Sr, F, Ar, Li, Rb, I, Al, Fe, Zn, and Mo. Variable minor constituents include C, N, O, Si, P, and Ba. Seawater brines become more concentrated by evaporation, and in this process, deviation from ideal chemical behavior is greater (2). Microscopic cavities in such minerals as halite contain trapped brines representative of ancient seawater. Due to the protection of the brine provided by the halite crystal, such water has not undergone any significant chemical changes. The ionic ratios in these trapped brines are similar in composition to evaporating modern seawater. Three properties of seawater and its brines predominant in evaporate production are concentration, density, and vapor pressure. Total salts are commonly stated in salinity, reflecting the resulting concentration from evaporation. The evaporation rate, however, indicates how much brine has undergone concentration via evaporation. Brine density is most important in its control of the circulation pattern in evaporate basins. The density of brines derived from seawater rises sharply from the 1.03 g/cm3 of normal seawater to 1.29 g/cm3 at the beginning of halite deposition. Vapor pressure is also important in its control of the rate of evaporation. Other properties include heat capacity, light transparency, and viscosity. Brines have substantially lower heat capacity, slightly lower thermal conductivity, and are less transparent to light than seawater. BRINE OCCURRENCES Geologically, brines are generated in a variety of environments and under varying sedimentary processes (2). Two main types of evaporate deposits are recognized: capillary evaporites and open-water evaporites (3). Capillary evaporites precipitate from interstitial brines that have a preexisting sedimentary layer. They contain remains of the host sediment, so capillary evaporites are always more or less dirty. Open-water evaporites precipitate from exposed water bodies. The detrital fraction, which is often eolian derived, makes up a small portion of these salt layers, so open-water evaporites are generally very clean. The primary depositional settings are (1) coastal intertidal and supratidal zones called sabkhas, (2) small lagoons on coasts and atolls, (3) large deep-water marine basins, (4) sub-sea-level basins with inflow of marine water, and (5) nonmarine interior basins. Tectonically, these deposits occur along continental margins and

shelves, interior shallow and deep cratonic basins, and rifted continental margins. Brine deposits are typically extremely localized and reflect their unique conditions of formation. Brine deposits including salt are derived from four primary sources: seawater, sedimentary bedded deposits, surface playa deposits, and salt domes. Such areas and conditions include deserts of the southwestern United States, High Andean salares, Middle Eastern sabkhas, the East African Rift Valley, and subsurface Permo-Triassic deposits. Recent studies have discovered brine pools on the seafloor within the Gulf of Mexico. These deposits are characterized by a distinct surface and shoreline. The pools are derived from the Luann Salt Layer and form in craterlike depressions that contain very concentrated brines and methane. These pools also support dense peripheral mussel beds. Extensive evidence also exists that brines commonly form deep along midocean ridge hydrothermal systems. Such occurrences result from heating of sea-water-derived hydrothermal fluids at supercritical conditions, although it is not clearly understood how these brines are stored in the crust. One widely accepted model is the development of dense brines formed by formation of a two-layer system, whereas a recirculating brine layer underlies a singlephase seawater cell. If the temperature of the seawater cell remains high enough for substantial supercritical two-phase separation, the brine layer will grow. Should temperatures drop into the single-range region, the brine layer erodes. Brines have also been recently discovered on other planets within our solar system. A growing body of evidence indicates that an ocean exists but is hidden beneath the icy crust of Jupiter’s moon, Europa. Hydrated salts are indicated by spectral evidence and thermal evolution models of Europa’s interior and laboratory studies of meteorites. The hydrated mineral deposits may reflect exposure of salty ocean water to the surface. On Mars, evaporate deposits may represent significant sinks of mobile anions and cations among the materials composing the Martian surface and upper crust. The nature of evaporate-precursor brines formed under Martian conditions is poorly understood at this time. Salts depress the freezing point of water significantly, so the presence of salts in the Martian soil could explain why water might be flowing in very high latitudes. EVAPORITE MINERALOGY AND THE CONCEPT OF FACIES Brines fall into three main types based on the predominating acid radical: chloride brines, sulfate brines, and alkali or volcanic brines. They include bitterns (natural brines) with other chlorides, bromides, iodides, and sulfates. The metallic ions in greatest abundance are sodium, magnesium, calcium, and potassium. Chloride type brines include many terrestrial brines as well as seawater. Terrestrial brines that contain chloride in excess of sodium are comparatively rare. Chloride type brines contain a higher proportion of sulfate than seawater. Seawater is an impure solution of sodium chloride and contains enough chloride to combine with all the sodium (the most abundant metallic

BRINE DEPOSITS

ion) and part of the magnesium (the second most abundant metallic ion). The transition from chloride to sulfate type brines is gradational, whereby sulfate is the predominate acid radical. Alkali brines contain carbonate, sulfate, and borate; chloride is present in subordinate amounts. Based on provenance, brines can be divided into marine brines that are concentrated from ocean water, continental brines concentrated from groundwater, and formation waters circulating in deeper bedrock horizons. Evaporites are sedimentary deposits composed of minerals that have precipitated from brines concentrated during evaporation. These deposits are products of waters of inland desert basins, as well as interstitial waters of sediments located along ocean margins. Evaporites also occur from replacement of rocks that are not evaporites by evaporate minerals (4). Sedimentary processes vary significantly. Source fluids for evaporate precipitation can be classified by the degree of solute concentration of undersaturated brines. The three types distinguished include waters of low salinity or hypohaline waters, intermediate salinity or mesohaline waters, and high salinity or hypersaline brines. Evaporites have a distinct and characteristic mineralogy. In the study of their geochemistry, origin, and geologic setting, mineralogy is the most important characteristic (2). Critical minerals derived from the evaporation of seawater are summarized in Table 1, along with the general facies sequence for marine evaporites. Facies in geological terms characterizes the critical variations or stages in the mineralogy of evaporite rocks during brine evaporation. Facies in this context essentially represent subdivisions of the evaporate process, recognizable through the appearance of new albeit critical minerals, regardless of proportion. For example, the mere presence of halite would characterize the assemblage as within the halite facies. The appearance of a potash-magnesia, halite, calcium sulfate, or calcium carbonate mineral is part of natural crystallization which occurs as water is removed

53

during evaporation. This process is dependent on a balance between the proportions of various components in seawater and the corresponding mineral solubility in the brine. An extensive listing of the most common evaporate minerals is presented by Braitsch (5) and Sonnenfeld and Perthuisot (3). Structurally, evaporates occur in two widely different structural forms: beds or lenses, and as salt structures which includes such features as bosses, plugs, ridges, and domes. Lenses and bed type deposits, such as the great Permian Basin, extend throughout parts of Kansas, Colorado, Oklahoma, Texas, and New Mexico. Nonbedded structures are common along or near the Gulf of Mexico. Bedded salt deposits vary in thickness but are typically of the order of 15 to 45 feet. The thickness of salt domes or diapirs remains uncertain. BRINE RESOURCES Known resources of evaporate and brine resources in the United States range from 100 years for potassium compound and iodine production, to unlimited, reflecting seawater as a source. The birthplace of subsurface brine production was in the Michigan Basin, an intracratonic basin that hosts thick successions of carbonates and evaporate-bearing strata. Natural subsurface brines are the feedstock for a wide variety of industrial minerals and chemicals. For example, bromide is produced in Arkansas from the Upper Jurassic Smackover Formation, magnesium chloride and bromide are produced from the Devonian Detroit River Group in Michigan; lithium carbonate is produced in Chile and California from aquifers beneath playas and salares; iodide is produced from the Lower Pennsylvanian Morrow Formation in Oklahoma and coproduced in Japanese natural gas wells; and calcium chloride is produced near Slave Lake, Drumheller, and Brooks, Alberta.

Table 1. Summary of Evaporite Faciesa

Facies Potash-magnesia (supersaline)

Subfacies Bittern subfacies Potash subfacies

MgSO4 subfacies

Halite facies CaSO4 facies Dolomite facies CaCO3 facies (normal marine) (brackish) (terrestrial) a

Modified after Reference 2.

Mineralogy Bischoffite, tachyhydrite Carnallite, sylvite, kainite, kieserite, halite, anhydrite or polyhalite Epsomite, bloedite, halite, polyhalite or anhydrite Halite, anhydrite or gypsum Gypsum or anhydrite, dolomite Dolomite, calcite Calcite, aragonite

Salinity, wt. %

Density, g/cm3

Fraction Evaporated, wt. % H2 O

380

1.31

99.2 98.7

120 75

78.0

375

1.29

98.4

65

68.0

300

1.20

91.0

11.5

12.2

150

1.10

72.0

3.5

3.6

35

1.02

0

1.0

1.0

10 1 0

1.01 1.00 1.00

— — —

Concentration × Seawater, wt H2 O

Concentration × Seawater, Vol. of Brines

54

CONNATE WATER

BRINE PRODUCTION

Table 1. Geochemical Composition of Connate Water Compared to Aquifer Watera

Brines are extracted via several means, including underground mining, solar evaporation, and solution mining and mechanical evaporation. Underground halite deposits are conventionally mined by drilling and blasting. Vertical shafts are about 20 feet in diameter and extend to depths of 500 to more than 2000 feet below ground into the salt deposit. In the United States, 70% of the salt produced is extracted from natural or synthetic brines or seawater; the remaining is mined as a solid. Products developed from brines include specialized industrial minerals used in the chemical industry for the manufacture of glass, fertilizers, pharmaceuticals, and batteries (6). Specialized industrial minerals produced include soda ash, potash, borax and boric acid, potassium chloride and sulfate, lithium, and nitrate and iodine. Salt produced from brine, and a small amount of dry salt, is used to produce chemicals such as chlorine gas and caustic soda (sodium hydroxide), among others. Deep unused brine-bearing aquifers are also used for to reduce greenhouse gas emissions via subsurface injection. BIBLIOGRAPHY 1. Goldberg. (1965). 2. Holzer, W.T. (1979). Mineralogy of evaporites. In: Mineralogical Society of America Short Course Notes, Marine Minerals. Vol. 6, pp. 211–294. 3. Sonnenfeld, P. and Perthuisot, J.P. (1989). Brines and Evaporites. American Geophysical Union, Short Course in Geology, Vol. 3. 4. Friedman, G.M. (1978). Depositional environments of evaporite deposits. In: Marine Evaporites. Society of Economic Paleontologists and Mineralogists (SEPM) Short Course No. 4, W.E. Dean and B.C. Schreiber (Eds.). pp. 177–184. 5. Braitsch, O. (1971). Salt Deposits Their Origin and Composition. Springer-Verlag, New York. 6. Jensen, M.L. and Bateman, A.M. (1979). Economic Mineral Deposits, 3rd Edn. John Wiley & Sons, New York, pp. 553–562.

CONNATE WATER RICHARD C. BRODY UC Berkeley Berkeley, California

CHEMISTRY Because of its long contact with rock material, connate water can change in chemical composition throughout the history of the rock and become highly mineralized (see Table 1). Connate water can be dense and saline compared with seawater; connate water salinities can range from 20 to more than 300 grams per liter. For comparison, seawater salinity is approximately 35 grams per liter total dissolved solids (1). AGE Some of the oldest connate water may stay beneath the land surface for millions of years, which is in contrast to

Chemical Component

Connate, mg/1

Aquifer, mg/1

Na+ K+ Mg2+ Ca2+ Ba2+ Sr2+ Fe2+

65,000 500 28,000 165,000 0.1 1,300 0.1

2,600 64 210 430 0.1 6.5 0.1

Cl− Br− SO4 2− HCO3 Organics

138,360 10 260 100 55

5,100 10 1,900 150 5

pH

5.8

7.65

a

Ref. 2.

the age of groundwater, which ranges from a few years or less to tens of thousands of years. Because of its long underground residence, connate water may move long distances, even though its velocity may be very low (3). OIL RESERVOIR MONITORING Crude oil is always produced with connate water, and the water-to-oil ratio is often greater than 10 to 1; therefore, the oil and gas exploration industry extensively monitors connate water (4). Connate water tends to rise toward the surface during oil and gas extraction, so resistivity logs track connate water movement to recognize and mitigate coning problems associated with excessive production yields (5). Distinguishing the difference between connate water and other types of underground water (e.g., meteoric water) through water analyses can aid in both oil and gas exploration. Variations in fluid composition can help to delimit reservoir boundaries and reveal the connectivity of different strata. Saline waters are generally more favorable for locating petroleum reservoirs. If the trapped fluid is saline, then there is less chance that any associated petroleum has been degraded by contact with meteoric water or flushed from the reservoir (1). BIBLIOGRAPHY 1. Robin, R. (2001). Oilfield Water Notes. University of Saskatchewan. Available: http://www.usask.ca/geology/classes/ geol463/46304.pdf. 2. Kane, R.D. (2000). InterCorr International, Inc. Use of Portable Monitoring Units to Assess Microbial Activity, Corrosion and Souring in Water Handling and Injection Systems. Available: http://www.corrosionsource.com/CS2000/ session01/paper0107/paper0107.htm. 3. Bouwer, H. (1978). Groundwater Hydrology. McGraw-Hill, New York, p. 7. 4. IPEC. (2001). Integrated Petroleum Environmental Consortium. Locating Oil-Water Interfaces In Process Vessels. Available: http://ipec.utulsa.edu/Ipec/lopresti abs.html.

CONSOLIDATED WATER BEARING ROCKS 5. JPT. (2002). Journal of Petroleum Hydrology. Technology Update. Available: http://www.spe.org/spe/cda/views/jpt/ jptMaster/0,1513,1648 2300 5376 3,00.html.

CONSOLIDATED WATER BEARING ROCKS NITISH PRIYADARSHI Ranchi University Ranchi, Jharkhand, India

INTRODUCTION Consolidated rocks (otherwise known as bedrock) consist of rock and mineral particles of different sizes and shapes that have been welded together by heat and pressure or chemical reaction into a rock mass. Aquifers of this type are commonly composed of one or more of the following rocks: sandstone, limestone, granite, or lava. Water flows through these rocks through fractures, gas pores, and other openings in the rock. CONSOLIDATED ROCKS Consolidated rocks are usually classified according to their origin. There are three types: sedimentary, igneous, and metamorphic. SEDIMENTARY ROCKS Sedimentary rock is formed by the deposition of material suspended in water; the material may have been material weathered from older rocks, plant or animal remains, or precipitated chemicals. Important types of sedimentary rock aquifers include sandstone, carbonate rock, and conglomerate. Sandstone Sandstone is a cemented form of sand and gravel; the sand grains are cemented together when dissolved silica and calcium contained in the pore fluid precipitate. Sandstone shows great variation in water yielding capacity which is controlled chiefly by its texture and the nature of the cementing materials. Thus, whereas coarse-grained sandstone with rather imperfect cement may prove excellent aquifers, fine-grained varieties, especially those that are thoroughly cemented, may be the worst types from which no yield of water is possible (1). Carbonate Rocks Carbonate formations include limestone (CaCO3 ) and dolomite (a mixture of CaCO3 and MgCO3 ). These deposits exhibit mostly secondary porosity due to fracturing and dissolution openings because CaCO3 is soluble in rainwater. Limestones vary widely in density, porosity, and permeability, depending on the degree of consolidation and development of permeable zones after deposition. Openings in limestone may range from microscopic original pores to large solution caverns forming subterranean

55

channels sufficiently large to carry the entire flow of a stream (2). Large springs are frequently found in limestone areas. The dissolution of calcium carbonate by water causes prevailingly hard groundwater in limestone aquifers; by dissolving the rock, water, also tends to increase the pore space and permeability with time. Solution development of limestone forms a karst terrane, characterized by solution channels, closed depressions, subterranean drainage through sinkholes, and caves. Such regions normally contain large quantities of groundwater (3). Major limestone aquifers occur in the southeastern United States and in the Mediterranean area (4,5). Conglomerates Conglomerates like sandstone are cemented forms of sand gravel. As such, their porosity and yield have been reduced by the cement. Conglomerates have limited distribution and are unimportant as aquifers (6). IGNEOUS ROCKS Igneous rocks are formed from the cooling and solidification of molten magma originating in the earth’s core. Important consolidated rock aquifers include basalts. The important points in this context are as follows: 1. Basalts rocks are often vesicular. Vesicles, in them are of considerable size and number, and if they are interconnected (by fracture), they can serve as aquifers. 2. Contraction joints (like columnar joints) and other fractures, if present, also contribute to the porosity and permeability of basalts. 3. Basalts originate as lava flows, so sometimes, they may overlie buried valleys that offer good groundwater potential. The Columbia River Plateau covering eastern Washington and Oregon, and Idaho, averages about 500 m in thickness and is one of the largest basalt deposits in the world. Basalt aquifers are critically important water sources for the Hawaiian Islands. Most of the largest springs in the United States are associated with basalt deposits. Rhyolites are less permeable than basalt. Granites have very low porosity and permeability; what little exists is primarily due to fracturing. Although not important as aquifers, these materials are candidates for the host rock for high-level radioactive waste. METAMORPHIC ROCKS These are sedimentary or igneous rocks that have been altered by heat or pressure. They generally form poor aquifers. BIBLIOGRAPHY 1. Singh, P. (1978). Engineering and General Geology. Katson, Ludhiana.

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SENSITIVITY OF GROUNDWATER TO CONTAMINATION

2. Brucker, R.W. et al. (1972). Role of vertical shafts in the movement of groundwater in carbonate aquifers. Ground Water 10(6): 5–13. 3. LeBrand, H.E. and Stringfield, V.T. (1973). Karst hydrology—a review. J. Hydrology 20: 97–120. 4. Burdon, D.J. and Papakis, N. (1963). Handbook of Karst Hydrogeology with Special Reference to the Carbonate Aquifers of the Mediterranean Region. Inst. Geology and Subsfc. Research, United Nations Spec. Fund, Karst Groundwater Inv., Athens. 5. LaMoreaux, P.E. et al. (1970). Hydrology of Limestone Terranes, Annotated Bibliography of Carbonate Rocks. Geological Survey of Alabama Bull. 94(A), Alabama. 6. Todd, D.K. (1995). Ground Water Hydrology. Wiley, Toronto, Canada, p. 51.

SENSITIVITY OF GROUNDWATER TO CONTAMINATION MICHAEL D. TROJAN Minnesota Pollution Control Agency St. Paul, Minnesota

For this discussion, groundwater sensitivity is defined as the likelihood for contaminants to reach a specified position in the groundwater system after introduction at some location above the specified position (1,2). Sensitivity is a widely applied concept that has several uses. Sensitivity analyses often lead to the development of maps showing the relative sensitivity of groundwater across a geographic area. These maps are used as a screening tool for land-use decisions and have increasingly been used in source water protection. Sensitivity analyses are also used in program development, such as identifying areas where groundwater monitoring will occur. The concept of groundwater sensitivity also has educational value because hydrogeologic concepts are easily expressed and the audience can visualize relationships between land use and aquifer protection. WHAT MAKES AN AQUIFER SENSITIVE TO CONTAMINATION? Groundwater sensitivity applies to a variety of situations; each requires a different level of analysis. Sensitivity, vulnerability, and susceptibility, which are often defined differently, are used interchangeably in this discussion. The simplest scenario involves transport of a conservative contaminant in water from the land surface to groundwater. This scenario can be modified to consider land use, account for attenuation or retardation of contaminants, consider a point below the top of the saturated zone, or consider risk to human or ecological receptors at some point within an aquifer. Under the scenario where a conservative contaminant is transported from the land surface to groundwater, factors that affect sensitivity include the permeability of geologic materials, the thickness of the unsaturated zone, and recharge. Highly permeable geologic materials

transmit water quickly and thus increase sensitivity (see HYDRAULIC CONDUCTIVITY/TRANSMISSIBILITY). In heterogeneous deposits, the geologic material that has the lowest permeability typically has the greatest effect on the rate of water movement, provided the geologic material is sufficiently thick, laterally continuous, and does not have extensive macroporosity. As the thickness of the unsaturated zone increases, sensitivity decreases because water can be stored and has a farther distance to travel. Sensitivity increases as the quantity of recharge increases because contaminants are transported by recharge water. Recharge is often computed from a water budget by subtracting evaporation from precipitation and assuming no runoff or soil storage. Even in sensitive hydrogeologic settings, groundwater will not become contaminated unless there is a source of contamination. We can generalize about how groundwater quality will be affected by land use (see LAND USE IMPACTS ON GROUNDWATER QUALITY). Agriculture can contribute significant quantities of nitrates and pesticides to an aquifer, nonsewered residential areas can contribute large quantities of nitrates and chlorides (3), and industrial and older sewered residential areas can contribute large quantities of volatile organic chemicals (VOCs), chlorides, and some metals such as manganese. Within each of these land uses, specific management practices further affect contaminant volumes. Several studies show direct, although not necessarily linear relationships, between the quantity of chemical applied at or near the land surface and concentrations in sensitive aquifers. Thus, for example, groundwater is more sensitive in areas that use greater fertilizer application rates. Nearly all contaminants attenuate to some extent within the unsaturated zone. The extent and rate of attenuation depend on the properties of the contaminant and on the chemical, physical, and biological properties of the geologic materials (Table 1). These properties are too numerous to discuss individually, but adsorption and degradation are the two primary attenuation processes that affect contaminants. Contaminants that are not readily adsorbed percolate to groundwater and result in increased sensitivity. Except in highly weathered, acidic soils, these are chemicals that have a negative charge, such as chloride, nitrate, and acid herbicides (2,4-D, Dicamba). Contaminants that are quickly degraded biologically or chemically represent a lower risk to groundwater than more persistent contaminants. In addition to the properties of a chemical, degradation is affected by factors such as temperature, the presence of oxygen, and pH. For example, most halogenated contaminants are persistent except under reducing conditions, nitrate is persistent in the presence of oxygen, and benzene is persistent in the absence of oxygen. We are often interested in the sensitivity at some point below the top of the saturated zone, such as at a well. In addition to having a greater distance to travel, a contaminant may be affected by geochemical changes within the saturated zone. Trojan et al. (4), for example, showed that in the glacial aquifers of Minnesota, nitrate was quickly attenuated in the upper few meters of groundwater, even though these aquifers were mapped

SENSITIVITY OF GROUNDWATER TO CONTAMINATION

57

Table 1. Characteristics of Different Classes of Contaminants. This Information is Generalized and Varies within Each Contaminant Class and Between Different Geologic Deposits and Hydrologic Environments Contaminant Class

Examples

Adsorption

Persistence

Toxicity

Halogenated organics with one or more benzene rings Polyhalogenated aliphatics

PCBs, dioxins, many pesticides, chlorobenzenes Industrial solvents (TCE, PCE), some pesticides

Nonhalogenated polynuclear aromatics

Pyrene, benzo(a)pyrene

High, increases with greater extent of halogenation Low, increases with greater extent of halogenation Moderate but increases rapidly with increasing molecular weight

High, increases with greater extent of halogenation Moderate, increases with greater extent of halogenation Moderate but increases rapidly with increasing molecular weight

Other nonhalogenated aromatics

Benzene, toluene

Low

Nonhalogenated aliphatics

Oil, alkanes,

Moderate to high

Metals Nonmetals

Lead, copper Arsenic, boron

Moderate to high Moderate to high

Anions Radionuclides

Chloride, nitrate Cesium-137, radon-222

Moderate to high Moderate but varies widely Low High

Low to high, depending on geochemical environment Low to high, depending on geochemical environment High High

High, increases with greater extent of halogenation Moderate, increases with greater extent of halogenation Moderate but increases rapidly with increasing molecular weight Low except for some chemicals, such as benzene Low

High High

Low High

as sensitive to contamination. In this case, denitrification was the most likely cause for disappearance of the nitrate. In other scenarios, contaminants may be adsorbed within an aquifer. Few sensitivity methods consider the risk to receptors. Risk analysis does not consider whether a chemical will reach groundwater but whether it will pose a risk to receptors. Risk is a function of dose and toxicity. Thus, consumption of 1,1,2-trichloroethene (TCE) at a concentration of 5 µg/L poses a greater risk than consumption at 1 µg/L; consumption of TCE at 5 µg/L [maximum contaminant level (MCL) = 5 µg/L] poses a greater risk than consumption of xylene at 5000 µg/L (MCL = 10,000 µg/L). Because we have to consider specific exposure points (e.g., a well), the quantity of contaminant being transported, and the contaminant toxicity, estimates of sensitivity based on risk can be very complicated. Table 2 provides a summary of factors that affect groundwater sensitivity to contamination. Figure 1 provides a schematic showing how different factors affect sensitivity. Figure 2 shows how sensitivity varies when different receptor points or contaminants are considered. METHODS FOR ASSESSING SENSITIVITY The Commission on Geosciences, Environment, and Resources (2) provides an excellent discussion of methods for assessing sensitivity. The most commonly employed methods are overlay and index methods. These involve combining various physical properties of the hydrogeologic system. Each property is assigned a score or other sensitivity value based on perceived sensitivity. Overlay and index methods use many of the factors in Table 2. DRASTIC (5) is a widely employed index and overlay

method that uses seven factors in the sensitivity assessment. Overlay and index methods can be relatively simple and may require small amounts of information. Sensitivity analyses using these methods often result in plan view maps depicting relative groundwater sensitivity, usually through a color-coded scheme. These maps are useful interpretive and screening tools. The scale of these maps is usually not appropriate for site-specific decisions, although they have been used for this purpose. Analytic or numeric methods predict the time it takes contaminants to reach groundwater. These methods typically consist of mathematical models. Attenuation of chemicals is often considered. Examples include PRZM (6), GLEAMS (7), and LEACHM (8). An advantage of analytic and numeric methods is that they quantify the processes that affect the movement of water and contaminant. The accuracy of these methods depends on the quality of data used in the model. Statistical methods use statistical techniques, such as regression analysis, to predict the likelihood of contamination. These methods require data on contaminant concentrations and may require additional information to derive sensitivity estimates. For example, concentrations of nitrate may be correlated with depth to water and sand content in the vadose zone. Because they use actual data, statistical methods can provide accurate estimates of sensitivity, information on variability in a sensitivity analysis, and estimates of certainty. They are, however, data intensive. LIMITATIONS OF THE SENSITIVITY CONCEPT In 1993, the Commission on Geosciences, Environment, and Resources prepared a report on ground water vulnerability assessments (2). The Commission, ‘‘in struggling with the manifold technical and practical difficulties

58

SENSITIVITY OF GROUNDWATER TO CONTAMINATION Table 2. Summary of Factors that Affect Groundwater Sensitivity Property

Effect on Sensitivity

Low Sensitivity

High Sensitivity

Hydrologic Factors Aquifer material

Depth to bedrock

Depth to water

Recharge Soil material

Thickness of confining layers Topography Type of bedrock

Rate at which water moves within an aquifer Distance that water must travel to reach groundwater Distance that water must travel to reach groundwater Amount of water reaching groundwater Rate at which water moves to or within groundwater Rate at which water moves to or within groundwater Amount of water reaching groundwater Rate at which water moves to or within groundwater

Shale, most hard rocks, clay, silt

Limestone, sandstone, sand, gravel

Large distances (>50 feet)

Short distances (50 feet)

Short distances (20 feet)

Small thickness ( pulses> fish> meat> fruits. In addition to foodstuffs, fluoride has also been reported in cosmetics and drugs. The use of drugs containing sodium fluoride for osteoporosis, otosteosclerosis, and dental caries is very common. Different brands of toothpaste contain excessive amounts of fluoride. Fluoride enters into the circulation directly from the oral cavity through the fine blood vessels of the mouth. Fluoride is a persistent bioaccumulator; even small amounts that enter through fluoride toothpaste one guaranteed entry in children as well as adults, and the cumulative effects of fluoride are a matter of serious concern.

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CONTROLLING FACTORS AND MECHANISM Controlling Factors Fluoride contamination in ground water is controlled by a number of parameters. The most important are • distribution of easily weathered fluoride-bearing minerals • the accessibility of circulating water to these minerals • pH of the percolating water • calcium content of the leaching water • temperature of the percolating water and the soil • exchangeable ions in the percolating water • extent of fresh water exchange in an aquifer • evaporation and evapotranspiration • complexing of fluoride ions with other ions • presence of CO2 and other chemicals in draining water • residence time of the percolating water in soil The pH of circulating water is an important factor that controls leaching of fluoride from fluoride-bearing minerals. Wodeyar and Sreenivasan (20) have indicated that higher alkalinity of waters promotes leaching of fluoride and thus affects the concentration of fluoride in groundwater. Alkaline water dissolves fluoride-bearing minerals under simultaneous precipitation of calcium carbonate (50). It has been reported that the degree of weathering and easy accessibility of circulating waters from weathered rocks, due to intensive and long term irrigation, are responsible for leaching fluoride from the parent minerals in soil and rocks. A high content of fluoride has been reported in black cotton soil due to excessive canal irrigation (20). Further, concentration of fluoride has been brought about by the arid climate of the region and the long residence time of groundwater in the aquifer (20,50). Very low fresh water exchange due to the arid climate of the region is also responsible for higher concentration of fluoride in groundwater (20). The fluoride concentration in groundwater, it was found, is positively correlated with calcium content but with a very low degree of validity (14,15). This observation is similar to that reported by Somani et al. (51), but there is no agreement to that report by Gupta et al. (31,32). This departure from the normal trend may be due to an irregular distribution of fluoride-bearing minerals in the soil, their solubilization characteristics, the nature of the product with soil, and other environmental conditions (51). Generally, waters of high calcium contents are low in fluoride content (52). The presence of carbon dioxide also affects the fluoride dissolution process in rocks. Fluoride in soil and ground water is also concentrated by evaporation and evapotranspiration due to arid and semiarid atmospheric conditions following scanty rainfall. The overwithdrawal of groundwater may also have favorable effects on fluoride concentration in groundwater. The combined effect of evapotranspiration and long-term contact of the water in the aquifer (due to low hydraulic conductivity of the weathered zone) activates the process

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FLUORIDE CONTAMINATION IN GROUND WATER

of dissolution (50). Ion exchange of the different elements in the soil and circulating water during the weathering process help in dissolving of fluoride from rocks and minerals (17,50). Besides, the formation of ion pairs such as CaSO4 , CaHCO3 and complexing of fluoride with aluminium, beryllium, ferric ions, and series of mixed fluoride hydroxide complexes with boron affect fluoride contamination (35). The temperature also plays a crucial role in the fluoride content of groundwater because the amount of fluoride ingested by living beings is influenced primarily by air temperature (20). Besides, the dissolution of fluoride from minerals/rocks is a physicochemical process that is also controlled by the temperature of the water and the soil itself. Chand (36) presented a correlation of ambient temperature with fluoride content.

concentration. Handa (55), Das (56), and Gupta et al. (57) summed up the characteristics of ground water whose fluoride content is from dissolved minerals, as follows: • • • •

negative correlation of calcium and fluoride ions positive correlation of bicarbonate and fluoride ions close of saturation with respect to calcium fluoride saturated with calcium carbonate

Thus, the dissolution mechanism of fluoride from its minerals and rocks can be explained on the basis of the solubility product. The dissolution mechanism is physicochemical and, therefore, is governed by thermodynamic parameters, too. The important parameters are temperature, pH, ionic strength, and pressure.

Mechanism During the process of chemical weathering, dissolution of fluoride species in natural water is controlled by calcium ions and governed by thermodynamic principles. The CaCO3 equilibrium in groundwater plays an important role in this process. The equilibrium constant of calcite can be evaluated from the following reactions (53). CaCO3 + H− ⇐ ⇒ Ca2+ + HCO3 − KCaCO3 = [Ca2+ ][HCO3 − ]/[H+ ] = 97at 25 ◦ C

(1)

(CaCO3 is constant) The fluoride equilibrium is given by (54): CaF2 ⇐ ⇒ Ca2+ + 2F− KCaF2 = [Ca2+ ][F− ]2 = 10−10.58 at 25 ◦ C

(2)

(CaF is constant) Dividing the first equation by the second, the solubility of calcite and fluorite can be represented by a third constant K: K= or or

[HCO3 − ] [H+ ][F− ]2

[F− ]2 = K

[HCO3 − ] (where K = 1/K) [H+ ]

[F− ] ∝ [HCO3 − ]/[H− ]

It is evident that the activity of fluoride is directly proportional to the bicarbonate ion at constant pH. Thus, according to the principle of ionic product, if the concentration of calcium and fluoride in water exceeds the solubility product of fluorite (10−10.58 at 25 ◦ C), CaF2 precipitates. Before reaching saturation the calcium ion has a positive correlation with fluoride ion, and after this stage, there will be a negative correlation between calcium and fluoride ions. In fact, the total concentration of fluoride in a solution will be somewhat greater due to the presence of other electrolytes (ionic strength and complexing effects). But it appears that a high fluoride concentration is more likely to occur in water of low calcium

REMEDIAL MEASURES There are two types of remedial measures to control fluoride contamination. The first is to control fluoride contamination in groundwater, and the second includes removal of fluoride from fluoride-containing water. The control of fluoride contamination in groundwater is very difficult because the contamination of fluoride in groundwater is controlled by a number of hydrogeologic and physicochemical parameters. However, various artificial recharge techniques, including the aquifer storage recovery (ASR) technique may be applied to improve the quality of water by dilution. The ASR technique is being followed in many parts of the world. In this technique, water is stored underground in wells when it is available, and this water is recovered from the same wells when needed to meet peak, longterm, and emergency water needs. The technique is being applied in United States, Canada, England, Australia, Israel, and other countries. The technique has proved to be a viable, cost-effective option for storing large volumes of fresh water not only in fresh, but also in brackish and other nonpotable aquifers at depths as low as 900 m. Most ASR sites store drinking water in confined aquifers containing water that is brackish or contains constituents such as nitrates, fluorides, iron, manganese, and hydrogen sulfide, all unsuitable for drinking purposes except following treatment. Mixing between the drinking water and the native water in the aquifer can be controlled in most situations by the proper design and operation of ASR wells, so that recovered water has acceptable quality. The operation includes development of a buffer zone surrounding the ASR well to contain the stored water and development of a target storage volume for each well so that recovered water will meet flow, volume, and water quality criteria with acceptability. This technique, however, still remains to be tried in India. Excess withdrawal of groundwater should be avoided to the extent possible. Aquifers should be recharged periodically so that air cannot enter the aquifer. In addition, only those types of raw materials should be used in industries, which do not release fluoride into the environment. In addition to this, contamination through nonpoint sources should also be minimized by checking

FLUORIDE CONTAMINATION IN GROUND WATER

man-made activities and the use of fluoride-containing fertilizers and pesticides. The second type of remedial measure includes removal of fluoride from fluoride-containing water (28,58). There are several methods that have been advocated for defluoridation of drinking water. These methods can be broadly divided into two categories, those based upon the addition of some chemical to the water during the softening or coagulation processes and those based upon ion exchange or adsorption processes. Adsorption or ion exchange processes are recommended for treating low concentrations. These processes are performed by using lime and alum, bone char and synthetic bone, activated carbon and bauxite, ion exchange, activated alumina, and reverse osmosis.

10.

11. 12. 13. 14.

15.

16.

CONCLUSION The problem of high fluoride concentration in groundwater resources has now become one of the most important health-related geoenvironmental issues in India influenced by the regional and local geological and hydrological conditions of the region. It is high time that an affordable solution is found to minimize fluoride contamination to maintain the health of the large population of the country. There is an immediate need to defluoride the water system either by community or by domestic defluoridation techniques. Demonstration-cum-awareness camps for the purpose should be arranged in fluorosis endemic areas. There is a need to carry out detailed fluoride mapping, hydrological studies for existing water sources to show flow lines, and hydrogeochemical surveys where fluorosis is endemic. In the affected areas, the government should apply firm guidelines for using groundwater, so that tube wells and/or hand pumps in high fluoride zones can be discouraged. Short-term solutions to minimize the fluoride level in drinking water could be using domestic defluoridation equipment or filters.

17.

18.

19.

20.

21.

22.

23.

24.

BIBLIOGRAPHY 1. Shortt, H.E. (1937). Endemic fluorosis in the Madras presidency. Ind. J. Med. Res. 25: 553–561. 2. Sudarshan, V. and Reddy, B.R. (1991). Pollution of fluoride in groundwater and its impact on environment and socioeconomic status of the people—a case study in Sivannagudem area. Indian J. Environ. Prot. 11(3): 185–192. 3. Rajiv Gandhi National Drinking Water Mission. (1994). New Delhi, Vol. 1, p. 20 4. Datta, P.S. (1996). Fluoride in ground water in Delhi area. J. Contam. Hydrol. 24(1): 85–96. 5. Sengupta, A. (1999). Proc. Workshop on Ground Water Pollution and Protection with Special Reference to Arsenic Contamination. Central Ground Water Board, Science City, Calcutta, pp. 69–74. 6. Chakraborti, D. and Bhatta, A. (1999). The Statesman. 22 August 1999, pp. 8–9. 7. Susheela, A.K. (1999). Fluorosis management programme in India. Curr. Sci. 77(10): 1250–1256. 8. Chakraborti, D. et al. (2000). Fluorosis in Assam, India. Current Science 78(12): 1421–1423. 9. Susheela, A.K., Kumar, A., Bhatnagar, M., and Bahadur, R. (1993). Prevalence of endemic fluorosis with gastro-intestinal

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31. Gupta, M.K., Singh, V., Rajwanshi, P., Srivastava, S., and Dass, S. (1994). Ground water fluoride levels in a rural area of district Agra. Indian J. Environ. Prot. 14(5): 370–372. 32. Gupta, M.K., Singh, V., Rajwanshi, P., Srivastava, S., and Dass, S. (1994). Fluoride in ground water at Agra. Indian J. Environ. Hlth. 36(1): 43–46. 33. Shrivastav, R. and Choudhary, B. (1997). Drinking water quality in an average Indian city: A case study of Agra (U.P.). Poll. Res. 16(1): 63–65. 34. Das, S., Mehta, B.C., Das, P.K., Srivastava, S.K., and Samanta, S.K. (1998). Source of high fluoride in ground water around Angul, Dhenkenal district, Orissa. Poll. Res. 17(4): 385–392. 35. Das, S., Mehta, B.C., Das, P.K., Srivastava, S.K., and Samanta, S.K. (1999). Sources of high fluoride in ground water around Angul, Dhenkenal district, Orissa. Poll. Res. 18(1): 21–28. 36. Chand, D. (1999). Fluoride and human health—cause for concern. Indian J. Environ. Prot. 19(2): 81–89. 37. Samal, U.N. (1988). Dental fluorosis in school children in the vicinity of aluminium factory in India. Fluoride 21(3): 137–141. 38. Susheela, A.K. (1991). Prevention and Control of Fluorosis. Technical Information for Training cum Awareness Camp for Doctors, Public Health Engineers and other Officers. Published by National Technology Mission of Drinking Water, New Delhi. 39. Sun, Z., Cheng, Y., Zhou, J., and Wei, R. (1998). Research on the effect of fluoride pollution in atmosphere near an aluminium electrolysis plant on regional fall wheat growth. Proc. Annu. Meet., Air Waste Manag., Assoc. 91st TPE 09/P1TPE 09P/7. 40. Clarke, M.L., Harvey, D.G., and Humphrevs, D.J. (1981). Veterinary Toxicology, 2nd Edn. Bailliere Tindal & Cassle Ltd., London, pp. 48–54. 41. Suresh, T. (1996). Fluoride in ground water of Chiknaya Kanahalli Taluk, Tumkur district, Karnataka. Proc. Workshop on Challenges in Groundwater Development, Madras, pp. 185–187. 42. Handa, B.K. (1979). Effect of return irrigation flows from irrigated land on the chemical composition of ground water from shallow unconfined aquifers. Prog. Water Tech. 11: 337–349. 43. EPA Report. (1976). National Interim Primary Drinking Water Regulation. EPA Publication No. EPA-570/9-76-003. 44. Piekos, R. and Paslawska, S. (1998). Leaching characteristics of fluoride from coal fly ash. Fluoride 31(4): 188–192. 45. Chary, V., Rao, R.J., and Naidu, M.G.C. (1975). Fluoride content of some raw vegetable foods available at Podile, Prakasam district, Andhra Pradesh. Proc. Symp. Fluorosis, Hyderabad, India, pp. 144–150. 46. Lakdawala, D.R. and Punekar, B.D. (1973). Fluoride content of water and commonly consumed foods in Bombay and supply of dietary intake. Ind. J. Med. Res. 16: 1679–1687. 47. Nanda, R.S. and Kapoor, K (1971). Fluoride content of pine and betel and its constituents. Ind. J. Med. Res. 59: 1966–1968. 48. Sengupta, S.R. and Pal, B. (1971). Iodine and fluoride contents of foodstuffs. Ind. J. Nutr. Dicter. 8: 66–71. 49. Gradien, P. (1992). Cancer incident and mortality in workers exposed to fluoride. J. Mat. Cancer Inst. 184: 1903–1909. 50. Rao, N.S., Rao, J.P., Rao, B.N., Babu, P.N., Reddy, P.M., and Devadas, D.J. (1998). A preliminary report on fluoride content in ground water of Guntur area, Andhra Pradesh, India. Curr. Sci. 75(9): 887–888. 51. Somani, I.C., Gandhi, A.P., and Palwal, K.V. (1972). Note on toxicity of fluorine in well waters of Nagpur and Jaipur

52.

53.

54. 55.

56.

57.

58.

districts of Western Rajasthan. Ind. J. Agric. Sci. 48(2): 752–754. Maithani, P.B., Gurjar, R., Banerjee, R., Balaji, K.B., Ramachandran, S., and Singh, R. (1998). Anomalous fluoride in ground water from western part of Sirohi district, Rajasthan and its crippling effects on human health. Curr. Sci. 74(9): 773–776. Hem, J.D. (1970). Study and Interpretation of the Chemical Characteristics of Natural Waters. USGS Water Supply, Paper 1475, pp. 363–369. Brown, D.W. and Roberson, C.E. (1977). Solubility of natural fluorite at 25 ◦ C. USGS J. Res. 5(4): 506–517. Handa, B.K. (1975). Geochemistry and genesis of fluoride containing ground water in India. Ground Water 13(3): 275–281. Das, D.K. (1985). Incident of high fluoride in deep ground water in Betwa basin, Madhya Pradesh, Central India. G.S.I. Res. 116(2): 23–30. Gupta, S.C., Doshi, C.S., and Paliwal, B.L. (1986). Occurrence and chemistry of high fluoride ground water in Jalore district of Western Rajasthan. Annal of Arid Zone 25(4): 255–265. Jain, A.K. (1997). Akalanka’s Water Supply and Treatment: Central Public Health and Environmental Engineering Organizations: Manual on Water Supply and Treatment. Akalank Publications, New Delhi, India, pp. 203–208.

ROCK FRACTURE NITISH PRIYADARSHI Ranchi University Ranchi, Jharkhand, India

INTRODUCTION From their origin, rocks are more or less disturbed by forces acting within the lithosphere. When a mass of rock is not strong enough to resist forces that are tending either to compress it or to stretch it, the rock is deformed. The change of form is brought about by flow in the deeper parts of the lithosphere and fractures in the upper parts (1). Fractures in rocks are either joints or faults. JOINTS The term joint is most commonly used in reference to relatively continuous and through-going fractures that are reasonably planar and along which there has been imperceptible movement. Joints may form as a result of either diastrophism or contraction. Classification of Joints Theoretically, joints may be classified according to whether they have been formed by compression or tension. Joints due to compression are (1) diagonal joints, (2) irregular cracks induced by the expansion of rocks consequent upon chemical alteration, and (3) probably a majority of the regular joint system in stratified rocks. Tension joints include (1) irregular cracks formed in the shrinkage accompanying certain kinds of rock alteration; (2) crossjoints in igneous rocks; (3) hexagonal columnar structures and associated fractures due to cooling; (4) small local

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fractures; (5) fractures clearly associated with tension faulting; and (6) cracks due to drying of muds, clays, and argillaceous limestones. Description of Joints 1. A series of parallel joints is called a ‘‘joint set.’’ 2. Two or more joints intersecting each other produce a ‘‘joint system.’’ 3. A persistent joint or set that may be horizontal or vertical is called a ‘‘master joint.’’ Probably all consolidated rocks and a good share of unconsolidated deposits contain joints. Although not well recognized by most individuals involved in groundwater problems, joints exert a major control on water movement and chemical quantity. Characteristically, joints are open and serve as major conduits or pipes. Water can move through them quickly, perhaps carrying contaminants, and, being open, the filtration effect is lost. It is a good possibility that the outbreak of many waterborne diseases that can be traced to groundwater supplies result from the transmission of infectious agents through fractures to wells and springs. Most joints, at least initially, are tight fractures, but because of weathering, the joint may be enlarged into an open fissure; this is especially common in limestone regions. Effect of Joints Knowledge of joints is important in many kinds of geologic studies. Quarry operations, especially those involved in obtaining blocks of certain dimensions and sizes, are obviously greatly influenced by joints. Closely spaced horizontal joints are obviously of great concern in tunneling. A large joint dipping into a highway cut is the site of a potential landslide. Wells drilled in granites for water supply are more productive in highly jointed rocks than in less jointed rocks (2). If joints are too numerous (i.e., more sets), closely spaced, and of great magnitude, then such a fractured site will be physically too weak to withstand the stresses of dams and bridges. Saturation with water along with the accompanying decay of rocks will make the site more unsuitable for foundations. Value of Joints Although joints are often difficult to interpret, they are nonetheless very important structures. For ages, quarry workers have taken advantage of joint-controlled planes of weakness in removing building blocks of granite and limestones from bedrock. These fracture weaknesses exert profound control on weathering and erosion and, thus, on fashioning landscape. Many scenic attractions owe much of their uniqueness to weathering and erosion of horizontal layers of systematically broken up, steeply dipping joints. Beyond their scenic value, joints constitute structures, of indisputable geologic and economic significance. Joints invite circulation of fluids, including rain and groundwater, hydrothermal mineralizing solutions, and oil and

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gas. As cracks in rocks, joints can be thought of as structures that significantly contribute to the bulk porosity and permeability of rocks. Explorers appreciate the benefits of the circulation of fluids through jointed rocks. Petroleum geologists evaluate the nature and degree of development of joints as one guide to the reservoir quality of sedimentary formations. To increase the yield of reservoir rocks in oil and gas fields where production is waning, it is common practice to ‘‘crack’’ the rocks artificially, either by explosives or by high-pressure pumping of fluids into wells. Joints can serve as sites of deposition of metallic and nonmetallic minerals. In almost all hydrothermal deposits, a part of the mineralization is localized in and around joints. The minerals are deposited either through openspace filling of joints or through selective replacement of chemically favorable rocks adjacent to the joint surfaces along which hydrothermal fluids once circulated. FAULTS A fault may be defined as a fracture along which there has been slipping of contiguous masses against one another. Points formerly in contact have been dislocated or displaced along the fracture. Solid rocks or unconsolidated sands, gravel, etc. may be dislocated in this way. Faulting may result from compression, tension, or torsion. Some faults in loose or weakly consolidated clays, sands, and gravels are produced by the removal of a support. Faults are most common in the deformed rocks of mountain ranges, suggesting either lengthening or shortening of the crust. Movement along a fault may be horizontal, vertical, or a combination. The most common types of faults are called normal, reverse, and lateral (Fig.1). A normal fault, which indicates stretching of the crust, is one in which the upper or hanging wall has moved down relative to the lower or foot wall. The Red Sea, Dead Sea, and the large lake basins in the East African highlands, among many others, lie in a graben, which is a block bounded by normal faults. A reverse or thrust fault implies compression and shortening of the crust. It is distinguished by the fact that the hanging wall has moved up relative to the foot wall. A lateral fault is one in which the movement has been largely horizontal. The San Andreas Fault, extending some 600 miles from San Francisco Bay to the Gulf of California, is the most notable lateral fault in the United States. Movement along this fault produced the 1906 San Francisco earthquakes. Recognition of Faults in the Field To recognize faults in the field, a number of criteria are used. The faults may be directly seen in the field, particularly in artificial exposures such as river cuttings and road cuttings. In most of cases, faults are recognized by stratigraphic and physiographic evidence such as (2) 1. 2. 3. 4.

repetition or omission of strata, discontinuity of structures, features characteristic of fault planes, silicification and mineralization,

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Foot wall Hanging wall

Foot wall

Hanging wall Fault

Cross section of reverse fault

Cross section of normal fault

Figure 1. Cross sections reverse, and lateral faults.

of

Plan view of lateral fault

normal,

5. sudden changes in sedimentary facies, and 6. physiographic data. Effects of Faulting Faulted areas are neither safe nor stable for the foundation of civil engineering works because of the various harmful effects produced by faults. Some important effects of faulting follow (3): 1. Faults cause considerable fracturing and shattering of rocks along fault zones, which means that they are not compact, massive, or strong. Such places are reduced to physically very weak grounds and hence are unfit as foundation sites to withstand heavy loads of structures such as dams. 2. When such porous and fractured zones are saturated with water, their strength is reduced further. 3. The same fractures act as channels for movement of groundwater, which may cause severe groundwater problems in tunnels and leakage problems in reservoirs. 4. The most dangerous features of faulting are its possible recurrence at the same place, which means that the faulted ground is unstable as long as faulting remains active there. Vertical, lateral, or rotational movements are likely to take place at the time of renewed faulting. Naturally, under such conditions, any civil engineering structure cannot have a safe or stable foundation. And if constructions are made, they are likely to collapse when renewed ground movements occur. 5. The fault plane itself is a very prominent fracture plane in the fault zone and therefore may act as a severe source of water leakage. When such percolated water reaches underground, it decomposes the

shear zone or fault zones. Such weathering further reduces the competence of rocks. BIBLIOGRAPHY 1. Lahee, F.H. (1987). Field Geology. CBS, New Delhi, p. 222. 2. Billings, M.P. (1984). Structural Geology. Prentice-Hall of India, New Delhi, pp. 140–200. 3. Kesavulu, N.C. (1997). Textbook of Engineering Geology. Macmillan India, New Delhi, pp. 241–242.

GEOCHEMICAL MODELS ´ MARIO ABEL GONCALVES ¸

Faculdade de Ciˆencias da Universidade de Lisoba Lisbon, Portugal

INTRODUCTION A geochemical model, like any other model, is an abstract representation of a given reality, normally reduced to a set of master variables and described by mathematical equations, which aim at representing natural processes that occur in a system. The output data of these models are a quantitative representation of an outcome that can be observable in the natural system or subject to experimental validation. This operational definition of a geochemical model suggests that a model is nothing more than a set of mathematical equations, which is not strictly true. Any set of equations representing a process must be bounded within limits imposed by nature; otherwise the outcome of these models may be totally unrealistic. Thus, the modeler is called to

GEOCHEMICAL MODELS

set proper initial and boundary conditions such that the natural system is correctly represented, and to feed the model with the correct parameters, most of them previously determined experimentally. The quality and self-consistency of the thermodynamic data used, as well as other parameters, such as kinetic ones, are of fundamental importance for the outcome of geochemical models. This issue is considered one of the most critical in any geochemical model. When using any of the available computer programs for geochemical modeling, the choice of the thermodynamic database is left to the modeler. Several compiled thermodynamic databases are available, but this does not mean that all data is internally self-consistent. It is, however, possible to find some databases that are self-consistent relative to some set of chemical species. Geochemical models have been extensively reviewed in the literature, such as Yeh and Tripathi (1), Mangold and Tsang (2), Appelo and Postma (3), Nordstrom and Munoz (4), and Nordstrom (5). Some textbooks on aqueous geochemistry or geochemical modeling also discuss and include several examples and case studies where specific geochemical computer models have been used (6–8). Of relevance is also the book of Albar`ede (9), which provides a wealth of mathematical methods and extensive examples on how to build and develop geochemical models in a truly wide range of applications. COMPONENTS OF GEOCHEMICAL MODELS Geochemical models may have several components that can be combined in different configurations. These components may be coupled within the model or may imply certain feedback loops. An essential component of these models is chemical reactions, and these determine, for example, the chemical speciation in solution or the saturation states relative to solid phases. Within these reactions, biological processes may be involved, which take active part in certain reactions, boosting their kinetics (catalysts), hindering the formation of certain compounds (inhibitors), or just transforming chemical compounds (such as the biodegradation of organic pollutants). Chemical species and compounds in solution are carried away with the water by advection and disperse through the medium by molecular diffusion. These are the components of transport of chemical elements in solution, which determine their spreading in the system. Transport of chemical elements are thus a function of water velocity and, consequently, of fluid flow in the system. Physical parameters of the system, such as porosity and permeability, determine the patterns of fluid flow and velocity variations in space and time. In the coupled reactive transport models, precipitation/dissolution of mineral phases can reduce/increase the porosity and permeability of the medium and change fluid flow patterns. In certain systems, such as large-scale sedimentary basins, it is necessary to consider heat transport. Heat can increase the kinetics of several chemical reactions and induce fluid flow along thermal gradients.

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TYPES OF GEOCHEMICAL MODELS The description of the different types of geochemical models is not extensive and outlines only their main characteristics, as presented by Zhu and Anderson (8). Geochemical models are generally grouped according to their level of complexity. The simplest ones are the speciation-solubility models. These models are meant to compute the thermodynamic equilibrium of species in a system at a given temperature and pressure. Therefore, the output comprises the concentration and activity of the various ionic and molecular species in a solution. It also includes the saturation state of the solution relative to several minerals and the distribution of stable species on surfaces or ion-exchange sites in equilibrium with the aqueous solution. Reaction-path models calculate the sequence of equilibrium states of a system in response to incremental additions (or subtractions) of mass to the system, change in temperature and/or pressure, and mass transfer between phases in the system. The configuration of these models can be diverse and includes the addition of a reactant (such as a titration), fixation of the activity of a chemical species modeling a buffered system, incremental feeding of a reactant solution (as in a continuous stirred tank reactor), and kinetic controls of heterogeneous reactions. Another group of models corresponds to inverse mass balance models. These specialized models derive the initial composition of a water solution from its actual final composition, which takes into account the reactions and mass transfer between water and solid and/or gas phases, in agreement with the available data of a system. Thus, the initial composition of the water is determined by subtracting the amount of dissolved species caused by reaction with minerals and other phases in the system from its final composition. Inverse mass balance calculations may also involve the determination of the fractions of different waters that have, at some given time, mixed completely. Finally, coupled reaction-transport models are the most complex. In these models, both the partial differential equations describing the advection-dispersion transport and the set of algebraic equations describing the chemical equilibrium are solved. These models can also include heat transfer and fluid flow, thus increasing the number of equations to be solved. The level of complexity depends also on the details of chemical reactions considered. These details can include multicomponent reactive transport, which accounts for the kinetics of mineral dissolution and precipitation; adsorption onto mineral surfaces; and radioactive decay, to name but a few. FINAL REMARKS: MODEL VALIDATION AND USEFULNESS The outcome of geochemical models can be either observable in nature or subject to experimental testing. Both processes are fundamental for model validation, and they are surely the ultimate test that a model must face. However, the process is not as simple as it may seem. Usually, geochemical models may adequately describe several processes and mechanisms in nature, but nature’s

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inherent complexity puts a limit to model precision and accuracy, which limits considerably its proper validation. Normally, the number of variables assumed within a model is limited and corresponds to a fraction of the ones found in nature. In such complex models, slight variations in parameters may induce diverse outcomes, such as in climate modeling. Thus, models are not only helpful tools to gain insight into the workings of nature, but they also must have some sort of predictive power. A model outcome may not be accurate enough to make a prediction relative to contaminant dispersion in a groundwater system, for example. However, it may give enough confidence to help make decisions on regulatory issues.

2. Mangold, D.C. and Tsang, C.-F. (1991). A summary of subsurface hydrological and hydrochemical models. Rev. Geophys. 29: 51–79. 3. Appelo, C.A.J. and Postma, D. (1993). Geochemistry, Groundwater and Pollution. A.A. Balkema, Rotterdam.

markedly important, their regular maintenance should not be dismissed. Programming languages evolve, as well as operating systems and computer hardware, which means that codes without regular revision become outdated and eventually useless as their compiled versions may stop working properly under new operating systems. All geochemical models rely very much on the availability of good quality, self-consistent thermodynamic data. This data is stored in database files that are accessed by the program while it is executed, making it one of its core elements. In the absence of specific data in the database for the problem to be modeled, some programs allow the incorporation of the data in simulations or, alternatively, the database can be modified by incorporation of new data. As new and improved experimental thermodynamic measurements are continuously being made, thermodynamic databases should also be regularly revised and updated. The description that follows is meant to address mainly those computer programs most readily accessible and does not pretend to be an exhaustive list of all available programs. Although presenting the address of websites where these programs and codes are stored and may be obtained, one should be aware that this information will potentially become out of date rather quickly.

4. Nordstrom, D.K. and Munoz, J.L. (1994). Geochemical Thermodynamics, 2nd Edn. Blackwell, Boston, MA.

USGS CODES

BIBLIOGRAPHY 1. Yeh, G.T. and Tripathi, V.S. (1989). A critical evaluation of recent developments in hydrogeochemical transport models of reactive multichemical components. Water Resour. Res. 25: 93–108.

5. Nordstrom, D.K. (2004). Modeling low-temperature geochemical processes. In: Treatise on Geochemistry—Surface and Ground Water, Weathering, and Soils. J.I. Drever (Ed.). Vol. 5, pp. 37–72. 6. Langmuir, D. (1996). Aqueous Environmental Geochemistry. Prentice-Hall, Upper Saddle River, NJ. 7. Drever, J.I. (1997). The Geochemistry of Natural Waters, 3rd Edn. Prentice-Hall, Upper Saddle River, NJ. 8. Zhu, C. and Anderson, G. (2002). Environmental Applications of Geochemical Modeling. Cambridge University Press, Cambridge, UK. 9. Albar`ede, F. (1996). Introduction to Geochemical Modeling. Cambridge University Press, Cambridge, UK.

GEOCHEMICAL MODELING-COMPUTER CODES ´ MARIO ABEL GONCALVES ¸

Faculdade de Ciˆencias da Universidade de Lisoba Lisoba, Portugal

An important share of geochemical studies increasingly relies on the use of computer programs1 to model diverse geochemical systems. Most of these programs are freely available to the general public, or at a symbolic cost for educational and research purposes. Although developing open source and/or precompiled codes is 1

We will use the term ‘‘computer code’’ as a set of written instructions aiming at solving a set of specific problems, and computer program to a compiled code to be executed as a standalone application under a given operating system.

The USGS supports various projects for developing software, including aqueous geochemistry computer programs, which include the chemical speciation program WATEQ4F (1), well suited for processing large numbers of water analyses. The most recent upgrades include the revision of the thermodynamic data on uranium and arsenic species. The most complete computer programs available are the ones from the PHREEQC (2) family (http://wwwbrr.cr.usgs.gov/projects/GWC coupled/). PHREEQC is a program that performs chemical speciation calculations, reaction-path modeling, one-dimensional transport, and inverse geochemical calculations. Currently, it contains a basic interpreter allowing a very flexible use of the program, meeting each one’s needs, especially for modeling kinetic data. The latest revisions (February–April 2003) include isotope fractionation modeling (3). PHREEQC uses its own thermodynamic database, and also the LLNL and WATEQ4F databases, which are still updated and corrected regularly. Two graphical user interfaces (GUI) were developed: PHREEQCI by USGS and PHREEQC for Windows by Vincent Post from the Vrije Universiteit Amsterdam (http://www.geo.vu.nl/users/posv/phreeqc). The latter allows the graphical display of the output, which is unavailable in the original program, which is rather achieved by using the PHRQCGRF program. PHAST is a three-dimensional multicomponent reaction-transport model that simulates transient groundwater flow, that may or may not include geochemical reactions. PHAST combines the HST3D simulator (4) for the transport calculations with PHREEQC for geochemical calculations. PHRQPITZ is specially designed to use with brines, as it implements Pitzer’s equation for the calculation of activity coefficients.

GEOCHEMICAL MODELING-COMPUTER CODES

Other computer programs include OTIS (5), used for the geochemical modeling solute transport in streams and rivers. Recently, Bowser and Jones (6) presented a Microsoft Excel spreadsheet for a mineral-solute mass-balance model in order to study and understand the mineralogical controls on water composition in surface and groundwater systems dominated by silicate lithologies. All of these programs, and others, are available from USGS webpages at http://water.usgs.gov/software/geochemical.html. USEPA CODES The USEPA has a series of supported computer codes, the most popular of which is the MINTEQA2/PRODEFA2 (last release in 1999 is version 4.0), widely used in environmental geochemistry problems (7), which is a chemical equilibrium computer model that is able to calculate chemical speciation, solubility equilibrium, titration, and surface complexation modeling. It also includes the Gaussian model for the interaction of dissolved organic matter (DOM) with cations. However, it lacks database maintenance. Gustafsson (8) has been developing VisualMINTEQ, a GUI version of this program that also presents other improvements, such as the NIST database, adsorption with five surface complexation models, ion-exchange, and inclusion of both the Stockholm Humic Model and the NICA-Donnan model for metal-DOM complexation to name only a few. The program BIOPLUME III (9) is a 2-D finite difference model that accounts for advection, diffusion, adsorption, and biodegradation in groundwater systems to model natural attenuation of organic contaminants. BIOCHLOR (10) and BIOSCREEN (11) are both Microsoft Excel spreadsheet-based codes that model natural attenuation of chlorinated solvents and petroleumderived hydrocarbons in water systems, respectively. CHEMFLO-2000 (12) is a model that simulates water flow and chemical transport and fate in the vadose zone. CHEMFLO-2000 is a program that is written in Java, which makes it platformindependent. All of these programs can be obtained from http://www.epa.gov/ada/csmos/models. OTHER CODES The set of computer codes known as EQ3/6 (13) supported by Lawrence Livermore National Laboratory (LLNL) was originally developed to model water-rock interactions in hydrothermal systems. It is currently one of the most complete programs applied to several problems, including municipal and industrial waste situations, and has been used to assess natural and engineered remediation processes. Unlike the programs presented until now, it must be purchased from LLNL. Closely related but mostly used for a range of high temperature and pressure is SUPCRT92 (14). This program has been discontinued, but still available on request to the authors. The Geochemist Workbench (15) is a commercial software with a range of capabilities similar to EQ3/6 and PHREEQC. It is available for Windows only, but taking advantage of this environment makes it user friendly, with

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graphical capabilities included that are normally absent from most of these programs. The Windermere Humic Aqueous Model (WHAM) version 6 (16) models the ion–humic substances interaction in surface waters using surface complexation. It also incorporates cation exchange on clays. However, precipitation and dissolution of solids as well as oxidation-reduction reactions cannot be simulated. This program must also be purchased for use. Alternatively, WinHumicV is a freely available GUI version of WHAM with model V implemented (17). Steefel and Yabusaki (18) developed the GIMRT/OS3D codes for 2-D and 3-D multicomponent coupled reactivetransport modeling for flow in porous media. Both of these programs were superseded by the program CRUNCH (http://www.csteefel.com/CrunchPublic/WebCrunch.html), which can be obtained from the developer (C. I. Steefel) on request. ORCHESTRA (19) (http://www.meeussen.nl/orchestra/) represents a new class of computer programs for use in geochemical reactive-transport modeling. This program is actually a framework where chemical speciation models can be implemented by the user and combine them with kinetic and transport processes. It is written in Java and takes advantage of objectoriented programming. In the same class of programs is MEDIA (http://www.nioo.knaw.nl/homepages/meysman/), to simulate the biogeochemistry of marine and estuarine sediments. BIBLIOGRAPHY 1. Ball, J.W. and Nordstrom, D.K. (1991). User’s Manual for WATEQ4F, with Revised Thermodynamic Data Base and Test Cases for Calculating Speciation of Major, Trace, and Redox Elements in Natural Waters. U.S. Geological Survey, Open-File Report 91-183, Washington, DC. 2. Parkhurst, D.L. and Appelo, C.A.J. (1999). User’s Guide to PHREEQC (Version 2)—A Computer Program for Speciation, Batch-Reaction, One-Dimensional Transport, and Inverse Geochemical Calculations. U.S. Geological Survey WaterResources Investigations Report 99-4259. Washington, DC, p. 312. 3. Thorstenson, D.C. and Parkhurst, D.L. (2002). Calculation of Individual Isotope Equilibrium Constants for Implementation in Geochemical Models. U.S. Geological Survey WaterResources Investigation Report 02-4172. Washington, DC, p. 129. 4. Kipp, K.L. (1987). HST3D—A Computer Code for Simulation of Heat and Solute Transport in Three-dimensional Groundwater Flow Systems. U.S. Geological Survey Water-Resources Investigations Report 86-4095. Washington, DC, p. 517. 5. Runkel, R.L. (1998). One Dimensional Transport with Inflow and Storage (OTIS): A Solute Transport Model for Streams and Rivers. U.S. Geological Survey Water-Resources Investigation Report 98-4018. Washington, DC, p. 73. 6. Bowser, C.J. and Jones, B.F. (2002). Mineralogic controls on the composition of natural waters dominated by silicate hydrolysis. Amer. J. Sci. 302: 582–662. 7. Allison, J.D., Brown, D.S., and Novo-Gradac, K.J. (1991). MINTEQA2/PRODEFA2, A Geochemical Assessment Model for Environmental Systems: Version 3.0 User’s Manual. U. S. Environmental Protection Agency, Washington, DC, p. 107.

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8. Gustafsson, J.P. (2004). Visual MINTEQ ver. 2.30. Available: http://www.lwr.kth.se/english/OurSoftware/Vminteq/index. htm. 9. Rafai, H.S., Newel, C.J., Gonzales, J.R., Dendrou, S., Kennedy, L., and Wilson, J.T. (1998). BIOPLUME III—Natural Attenuation Decision Support System, User’s Manual Version 1.0. EPA/600/R-98/010. USEPA, Washington, DC, p. 282. 10. Aziz, C.E., Newel, C.J., Gonzales, J.R., Haas, P., Clement, T.P., and Sun, Y. (2000). BIOCHLOR—Natural Attenuation Decision Support System, User’s Manual, Version 1.0. EPA/600/R-00/008. USEPA, Washington, DC, p. 46. 11. Newel, C.J., McLoed, R.K., and Gonzales, J.R. (1996). BIOSCREEN—Natural Attenuation Decision Support System, User’s Manual, Version 1.3. EPA/600/R-96/087. USEPA, Washington, DC, p. 65. 12. Nofziger, D.L. and Wu, J. (2003). CHEMFLO-2000—Interactive Software for Simulating Water and Chemical Movement in Unsatured Soils. EPA/600/R-03/008. USEPA, Washington, DC, p. 69. 13. Wolery, T.J. (1992). EQ3/6, A Software Package for Geochemical Modeling of Aqueous Systems: Package Overview and Installation Guide (Version 7.0). (UCRL-MA-110662 PT I). 14. Johnson, J.W., Oelkers, E.H., and Helgeson, H.C. (1992). SUPCRT92: A software package for calculating the standard molal thermodynamic properties of minerals, gases, aqueous species, and reactions from 1 to 5000 bar and 0 to 1000 ◦ C. Comput. Geosci. 18: 899–947. 15. Bethke, C.M. (1996). Geochemical Reaction Modeling, Concepts and Applications. Oxford University Press, New York. 16. Tipping, E. (1998). Humic ion-binding model VI: an improved description of the interactions of protons and metal ions with humic substances. Aqua. Geochem. 4: 3–48. 17. Gustafsson, J.P. (1999). WinHumicV. http://www.lwr.kth. se/english/OurSoftWare/WinHumicV/. 18. Steefel, C.I. and Yabusaki, S.B. (1996). OS3D/GIMRT, Software for Multicomponent—Multidimensional Reactive Transport. User manual and programmer’s guide. PNL-11166. Battelle, Richland, WA. 19. Meeussen, J.C.L. (2003). ORCHESTRA: An object-oriented framework for implementing chemical equilibrium models. Environment. Sci. Technol. 37: 1175–1182.

READING LIST Plummer, L.N., Parkhurst, D.L., Fleming, G.W., and Dunkle, S.A. (1988). A Computer Program Incorporating Pitzer’s Equations for Calculation of Geochemical Reactions in Brines: U.S. Geological Survey Water-Resources Investigations Report 88-4153. Washington, DC, p. 310.

GEOCHEMICAL MODELING—COMPUTER CODE CONCEPTS GEOFFREY THYNE Colorado School of Mines Golden, Colorado

INTRODUCTION This article is focused on geochemical models in aqueous systems. The applications for these models have grown

over the last 20 years as the capabilities and flexibility of the codes has increased in conjunction with the increased speed of personal computers. Many earlier computer models were designed for specific questions related to aqueous speciation (1,2). Many earlier models have been discontinued or superseded by newer programs that incorporate new features and capabilities, increased flexibility, and improved input and output options. This change has been driven by the wider use of geochemical modeling and the increase in modeling as a component of environmental studies. Some more important recent applications include modeling high-level radioactive waste disposal, environmental issues associated with mining, landfill leachate, injection of hazardous wastes into deep wells, water resources issues, and artificial recharge to aquifers (3). All available models use the same basic approach, that of calculating the thermodynamic equilibrium state of a specified system that can include water, solutes, surfaces, and solid and gas phases. These models comprise four major components. They are as follows: 1. Input: specific information that defines the system of interest such as concentrations of solutes, temperature, partial pressure of gases, and composition of solid phases. 2. Equations that are solved by the model. 3. Equilibrium and kinetic formulations between solutes of interest. 4. Output: in tabular or graphic form. CAPABILITIES AND METHODS The computer codes require initial input constraints that generally consist of water chemistry analyses, units of the measurement, temperature, dissolved gas content, pH, and redox potential (Eh). The models work by converting the chemical concentrations, usually reported in wt./wt. or wt./volume terms such as mg/kg or mg/L, to moles, and then solving a series of simultaneous nonlinear algebraic equations (chemical reaction, charge balance, and mass balance equations) to determine the activityconcentration relationship for all chemical species in the specified system. The models usually require electrical balance and will force charge balance with one of the components (can be designated), as they solve the matrix of nonlinear equations. The capabilities of modern codes include calculation of pH and Eh, speciation of aqueous species, equilibrium with gases and minerals, oxidation and reduction reactions (redox), kinetic reactions, and reactions with surfaces. The nonlinear algebraic equations are solved with an iterative approach by the Newton–Raphson method (4). The equations to be solved are drawn from a database that contains equations in the standard chemical mass action form. In theory, any reaction such as sorption of solute to surface that can be represented in this form can be incorporated into the model. These equations represent chemical interactions with reactants on the right and products on the left. Reactions are assumed to reach

GEOCHEMICAL MODELING—COMPUTER CODE CONCEPTS

equilibrium (the point of lowest free energy in the system) when there is no change in concentration on either side. CaCO3 ↔ Ca2+ + CO3 2− Note that the arrow in the calcite dissolution example above goes in both ways; that is, the reaction as written is reversible. Once a mineral reaches equilibrium with a solution, adding more mineral will not increase the dissolved concentration because we have already saturated the solution. But removing ions from the right side (e.g., lowering the concentration by dilution with distilled water) will cause more solid to dissolve. We express this in a mathematical form where, at equilibrium, the ratio of the concentration of reactants (on the bottom) and products (on the top) is equal to K, known as the proportionality constant or distribution coefficient or equilibrium constant. 2+

Kcalcite =

2−

[Ca ][CO3 ] [CaCO3 ]

Kinetic reactions, those involving time, are included by assuming that the chemical reaction will proceed to equilibrium, but at a specified rate. The available kinetic reactions include mineral dissolution and precipitation, redox reactions and microbial growth, and metabolism of solutes. The rate laws used in the codes vary, but all codes with kinetic capabilities include simple firstorder rate laws, and they may include more complex rate formulations such as cross-affinity, Michaelis–Menten, and Monod formulations (5). THERMODYNAMIC BACKGROUND The formalism that allows us to relate mass actions equations (balanced chemical reactions) to actual solutions is called chemical thermodynamics or, more precisely, equilibrium thermodynamics. The basic idea is that elements, molecules, and compounds all contain some internal energy and that all systems try to reach a state where that energy is minimized (equilibrium). Natural systems, particularly low-temperature systems, do not always reach equilibrium, but they do move in that direction. The internal energy of an element, molecule, or compound is expressed as its enthalpy or internal heat. The free energy of an element is an element, molecule or compound is the sum of its internal heat and its internal tendency toward disorder (entropy). G = H − TS Assuming constant T and P, we get G = H − TS The total free energy of a component in the system is dependent on this inherent energy of an element, molecule, or compound and the amount (concentration). When two or more elements, molecules, and compounds are combined, the result is a reaction that minimizes the energy of the new system, lowering the G. The free energy of a reaction is calculated by   G0 products − G0 reactants G0 rxn. =

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The G0 values (standard free energy) for many compounds can be found in the back of textbooks such as Drever (6). This way we can calculate a G value for any reaction that for which we can write a balanced chemical equation. The minimum energy state (equilibrium) between the reactants and products is related to the G value by log Krxn. =

−G0 rxn 2.303 RT

where R is the universal gas constant in kJ/mole; T is the temperature in Kelvin; and K is the equilibrium constant for the reaction. COMPUTER MODELS The computer models are divided into two basic types, speciation models and reaction-path models. In both cases, the models are fundamentally static; that is, no explicit transport function exists; however, some forms of transport can be simulated by manipulation of the models. More complex reaction-transport models that explicitly incorporate transport are briefly described below. The equilibrium models are speciation models in that they can calculate the speciation (distribution) of aqueous species for any element or compound included in the database. Speciation models calculate activities (chemically reactive concentration), species distribution for elements in the database, saturation indices, and ion ratios at the specified conditions of pH and redox potential (ORP or Eh). Most models allow selection of method of activity calculation (Davies, Debye-Huckel, extended Debye-Huckel, Pitzer). Some models incorporate adsorption, solid phase solutions, and kinetics. Only one model, PHREEQC, has the inverse modeling option. This features uses mass balance constraints to calculate the mass transfer of minerals and gases that would produce an ending water composition given a specified starting water composition (7). This method does not model mass transport; it only calculates and provides statistical measures of fit for possible solutions to the mass balance between starting and ending water compositions. The next step in complexity is the reaction path (mass transfer) models. The reaction path models use speciation calculation as a starting point and then make forward predictions of changes along the specified reaction path (specified change in T, P, pH, addition of new reactants such as another fluid or solid). The program makes small incremental steps with stepwise addition or removal of mass (dissolution or precipitation), and it can include changes in temperature or pressure along the reaction path. Typical questions posed by modelers include: • If I change a variable (pH, pe, PCO2 ), how does system change? • What happens if I mix water A with water B? • What is concentration of A+ in water saturated with mineral AB?

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• What happens to the water chemistry A if I react the water with mineral B? Limitations exist with any model. The field input data may be corrupt with bad analysis, missing parameter, or electrical imbalance. Speciation models assume equilibrium conditions, which may not be the case. The databases are also a source of uncertainty. They do not always contain all elements or species of interest; the data have some uncertainty, and some data may be inaccurate data (6). Some available codes try and minimize this problem by including several of the most popular databases such as the MINTEQ database (EPA-approved database specializing in metals), WATEQ (USGS database specializing in minerals), and the LLNL database (the most complete database available that is compiled and maintained by Lawrence Livermore National Laboratory). For environmental applications, the limited data for organic compounds remain a concern. Other limitations include the redox reactions that are of particular importance in metal transport. These reactions are difficult to model correctly because redox reactions may have different rates producing natural systems that are not in redox equilibrium (8). This problem can be addressed by modeling redox reactions as rate-limited (kinetic) formulations if the data are available. The most complex models explicitly incorporate transport and reaction. The codes couple and solve both the partial differential equations of flow using the advectivedispersion equation and the nonlinear algebraic equations of chemical equilibrium. The general approach is to solve for the reaction term in each cell with the chemical module of the code, and then separately solve for the effects of transport (split-operator approach). The effects of adsorption are solved in the transport module with the retardation portion of the equation (3,8,10). These models are much more complex than are the reaction path models. Presently, only three commonly used models have transport capabilities, HYDROGEOCHEM2, PHREEQC (1D) and the related PHAST (3D) code, and Geochemist’s Workbench. Table 1 lists the most common programs, their sources, and some of the most useful capabilities. The list is not meant to be exhaustive; rather, it offers an overview

to serve as a starting point for further investigation. Details of the capabilities of each program can be found on the listed websites or in the manuals. More detailed comparison of these and other models are available in related publications (1,11,12). BIBLIOGRAPHY 1. Nordstrom, D.K. et al. (1979). Comparison of computerized chemical models for equilibrium calculations in aqueous systems; chemical modeling in aqueous systems; Speciation, sorption, solubility, and kinetics. ACS Symposium Series 93: 857–892. 2. Allison, J.D., Brown, D.S., and Novac-Gradac, K.J. (1991). MINTEQA2/PRODEFA2, A Geochemical Assessment Model for Environmental Systems: Version 3.0 User’s Manual. U.S. EPA, Athens, GA. 3. Zhu, C. and Anderson, G. (2002). Environmental Applications of Geochemical Modeling. Cambridge University Press, Cambridge, UK. 4. Bethke. (1996). 5. Bethke. (2002). 6. Drever, J.I. (1997). The Geochemistry of Natural Waters, 3rd Edn. Prentice-Hall, Englewood Cliffs, NJ. 7. Garrels, R.M. and Mackenzie, F.T. (1967). Origin of the Chemical Compositions of Some Springs and Lakes in Equilibrium Concepts in Natural Waters. American Cancer Society, Washington, DC, pp. 222–242. 8. Lindberg, R.D. and Runnels, D.D. (1984). Ground water redox reactions: An analysis of equilibrium state applied to Eh measurements and geochemical modeling. Science 225: 925–927. 9. Appelo, C.A.J. and Postma, D. (1993) Geochemistry, Groundwater and Pollution. A. A. Balkema, Rotterdam. 10. Parkhurst, D. and Appelo, C.A.J. (1999). USER’S Guide to PHREEQC (VERSION 2)—A Computer Program for Speciation, Batch-Reaction, One-dimensional Transport and Inverse Geochemical Calculations. Water-Resources Investigations Report 99-4259. 11. Engesgaard, P. and Christensen, T.H. (1988). A review of chemical solute transport models. Nordic Hydrol. 19(3): 183–216. 12. Mangold, D.C. and Tsang, C.-F. (1991). A summary of subsurface hydrological and hydrochemical models. Rev. Geophys. 29(1): 51–79.

Table 1. Comparison of Selected Codes’ Capabilities and Features Program EQ3/6 GWB HYDROGEOCHEM2 MINTEQ MINEQL+ PHREEQC

Source

Speciation

Reaction Path

Tabular Output

Graphic Output

Surface Rxns.

Kinetics

Inverse

Transport

Multiple Databases

LLNL Rockware∗ SSG∗ EPA ERS∗ USGS

yes yes yes yes yes yes

yes yes yes no no yes

Yes Yes Yes yes yes yes

no yes no no some no

no yes yes yes no yes

yes yes yes no no yes

no no no no no yes

no yes yes no no yes

no yes no no no yes

∗ Commercial programs; others are freeware. EQ3/6—http://www.earthsci.unibe.ch/tutorial/eq36.htm GWB—Rockware—http://www.rockware.com HYDROGEOCHEM—http://www.scisoftware.com/environmental software/software.php MINEQL+—http://www.mineql.com/ MINTEQ—http://soils.stanford.edu/classes/GES166 266items%5Cminteq.htm PHREEQC—http://wwwbrr.cr.usgs.gov/projects/GWC coupled/phreeqc/

GEOPHYSICS AND REMOTE SENSING

READING LIST Glynn, P.D., Engesgaard, P., Kipp, K.L., Mallard, G.E., and Aronson, D.A. (1991). Two Geochemical Mass Transport Codes; PHREEQM and MST1D, Their Use and Limitations at the Pinal Creek Toxic-Waste Site. U.S. geological survey toxic substances hydrology program. Abstracts of the technical meeting, Monterey, CA, March 11–15, 1991. Open-File Report—U.S. Geological Survey OF 91-0088. Grove, D.B. and Stollenwerk, K.G. (1987). Chemical reactions simulated by ground-water-quality models. Water Resour. Bull. 23(4): 601–615.

GEOLOGICAL OCCURRENCE OF GROUNDWATER JOHN E. MOORE USGS (Retired) Denver, Colorado

The occurrence, movement, and storage of groundwater are controlled by geology. The geologic factors that control groundwater are petrography, stratigraphy, structure, geomorphology, lithology, and thickness (1). The petrography of a given rock type controls the porosity and permeability. Porosity defines the storage capacity of an aquifer. There are two types of porosity: primary and secondary. Primary porosity, such as pores between sand grains, is created when rocks are formed. The shape, sorting, and packing of grains determine primary porosity. Sedimentary rocks are poorly sorted when the grains are not the same size creating spaces between the larger grains that are filled by smaller grains. Secondary porosity such as joints, fractures, and solution opening, is formed after the rock has been deposited. The number and arrangement of fracture openings and the degree to which they are filled with fine-grained material control secondary porosity. Aquifers (1) are classified as unconfined and confined. An unconfined aquifer has the water table as its upper boundary (2). Recharge to unconfined aquifers is primarily by downward seepage through the unsaturated zone. The water table in an unconfined aquifer rises or declines in response to rainfall and changes in stream stage. When a well that tap an unconfined aquifer is pumped, the water level is lowered, gravity causes water to flow to the well, and sediments near the well are dewatered. Unconfined aquifers are usually the uppermost aquifers and, therefore, are more susceptible to contamination from activities at the land surface. A confined (artesian aquifer) contains water under pressure greater than that of atmospheric. Rocks of permeability lower than the aquifer overlie a confined aquifer. The low-permeability layer that adjoins a confined aquifer is called a confining bed. A confining bed has very low permeability that restricts the movement of groundwater either into or out of the aquifer. Confining beds are thus poorly transmissive to groundwater flow. Because the water is under pressure, water levels in wells rise above the base of the confining bed. If the water level in a well that taps a confined aquifer stands above the land surface, the well is called a flowing artesian well. In some

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cases, a fault (fractures along which rocks have moved) will allow the passage of water from a confined aquifer to the surface, resulting in a spring. Sand and gravel aquifers are the source of most of the groundwater pumped in many parts of the world, including North America, The Netherlands, France, Spain, and China. Sand and gravel aquifers are common near large to moderately sized streams. Rivers or the meltwater from glaciers formed these aquifers. Limestone aquifers are the sources of some of the largest well and spring yields because limestone is soluble in water. Openings that existed when the rocks were formed are frequently enlarged by solution (dissolved by water), providing highly permeable flow paths for groundwater (3). Basalt and other volcanic rocks are some of the most productive aquifers. Basalt aquifers contain water-bearing spaces in the form of shrinkage cracks, joints, and lava caves. Lava tubes are formed when tunneling lava ceases to flow and drains out, leaving a long, cavernous formation. Fractured igneous and metamorphic rock aquifers are the principal sources of groundwater for people who live in mountainous areas. Where fractures are numerous and interconnected, these rocks can supply water to wells and can be classified as aquifers. Wells are commonly 50 to 100 feet deep (15 to 30 meters). Granite and metamorphic rocks have not been extensively developed as aquifers. Groundwater movement in these rocks is irregular, which makes exploration for a water supply difficult (4). Sandstone aquifers are formed by the cementation of sand. Their porosity ranges from 5–30%. Their permeability is largely a function of the amount of cement (clay, calcite, and quartz). Sandstone is an important source of groundwater in Libya, Egypt (Nubian), Britain (the Permo-Triassic sandstones), the north central United States (St. Peter-Mount Simon Sandstone), and in the west central United States (the Dakota Sandstone). BIBLIOGRAPHY 1. Meinzer, O.E. (1923). The occurrence of ground water in the United States, with a discussion of principles. U.S. Geol. Survey Water Supply Paper 489. 2. Lohman, S.W. et al. (1972). Definitions of selected groundwater terms-revisions and conceptual refinements. U.S. Geol. Survey Water Supply Paper 1988. 3. Heath, R.C. (1983). Ground water regions of the United States. U.S. Geol. Survey Water Supply Paper 2220. 4. Moore, J. E. et al. (1995). Groundwater, a Primer. American Geological Institute, Alexandria, VA, p. 53.

GEOPHYSICS AND REMOTE SENSING JOHN R. JANSEN Aquifer Science & Technology Waukesha, Wisconsin

When data cannot be gathered by direct observation, such as observing an outcrop or taking a physical sample,

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GEOPHYSICS AND REMOTE SENSING

information can often be obtained by using remote sensing techniques. Remote sensing methods measure physical properties of materials by measuring changes in the flux of various forms of energy using natural or transmitted fields. The science of applying these methods to earth materials is called geophysics Geophysics is commonly used to obtain subsurface information for a variety of resource development and engineering applications. A partial list of applications for which geophysical methods can be used includes mapping aquifers, mapping water quality, mapping geologic structure, measuring flowing fluids, finding buried objects for engineering and archeological purposes, and measuring in situ soil and rock properties. The major energy fields used include magnetic fields (magnetometry), electrical fields (electrical resistivity and spontaneous potential), electromagnetic fields (electromagnetic induction and ground-penetrating radar), propagation of seismic waves (seismic reflection, seismic refraction, passive acoustic emission monitoring, and spectral analysis of surface waves), the gravitational field (gravimetry), gamma-ray radiation (gamma-ray spectroscopy), and heat transfer (geothermal). Other physical properties can be measured by bombarding the material with gamma rays (electron density) or high-energy neutrons (hydrogen content). Less commonly, properties such as thermal conductivity, electrical chargeability, or magnetic resonance are also measured. The methods can be broadly classified in several ways. Most commonly, methods are classified by their mode of deployment. Surface methods are commonly applied from the ground surface. Borehole methods are used within boreholes or wells. Several techniques can also be applied using airborne or marine systems. Methods can also be classified as active or passive measurement. Methods that measure variations in natural fields, such as gravimetry or magnetometry, are called passive methods because they use natural fields propagating through the earth. Other methods are called active methods because the measurement is based on the response of earth material to some form of transmitted energy. Active methods include most seismic methods and most electrical methods. There are many types of geophysical methods; each has specific advantages and limitations. The choice of the proper method requires understanding the target body, the matrix material, and the environment in which the survey is to be conducted. The following is a brief outline of the most common geophysical methods with a summary of the characteristics of each method. The list is by no means complete. SURFACE METHODS Electrical Methods Electrical Resistivity. The electrical resistivity method uses two electrodes planted at the surface to pass an electric current through the ground. A second pair of electrodes is used to measure the potential difference between two points. The measured potential and the applied current are used to calculate the electrical

resistance per unit length (resistivity) of the subsurface. The depth of measurement is changed by changing the relative position and spacing of the electrodes. Several newer systems use cables with multiple electrodes connected to a switching system to select the electrode pairs. The field data produce an apparent resistivity value, which is a function of all materials penetrated by the current. The data must be modeled to obtain the unique resistivity and thickness of each layer for a nonuniform subsurface.

Applications Can distinguish saturated from unsaturated materials. Can distinguish sandy material from clay-rich material. Can distinguish high-conductivity groundwater (inorganic contaminants) from low-conductivity groundwater. Less susceptible to cultural interference than electromagnetic (EM) methods. Better vertical resolution for resistive targets than EM methods.

Limitations Requires electrical coupling with the surface, which can be difficult in dry soils. Generally requires three to five times the surface array length to the depth of investigation. Field acquisition generally slower than EM methods. Poorer lateral resolution than EM methods at depths of more than a few meters. Relatively insensitive to changes in the resistivity of highly conductive bodies. Highly resistive near-surface material can mask conductive bodies at depth. Can be affected by strong EM fields such as from highvoltage power lines. Interpretation of field data produces nonunique solutions (i.e., many possible solutions can fit data nearly as well as actual conditions). Spontaneous Potential. The SP method uses two nonpolarizing electrodes and a high-impedance voltmeter to measure natural voltages between two points generated by several natural processes. Anomalies can be due to variations in the geochemical properties of soils, fluids, or moving groundwater. Although its origins are in the mining industry for finding sulfide ore bodies, the method is principally a profiling technique to find seepage pathways such as for a dam investigation.

Applications Relatively fast and simple data acquisition and interpretation. Can detect zones of high groundwater flow. Can detect oxidizing metal bodies.

Limitations Requires special nonpolarizing electrodes for accurate voltage measurements.

GEOPHYSICS AND REMOTE SENSING

Susceptible to interference from near-surface soil conditions. Interpretation is qualitative and can be ambiguous. Induced Polarization. The IP method measures a property known as chargeability, which is the ability of subsurface materials to hold a charge after an electrical current is shut off. It is basically a modification of the electrical resistivity method that measures induced electrical polarization between two electrodes caused by an applied low-frequency alternating current. Measurements are typically made by measuring the decay of voltage with time after the applied current is shut off, called the time domain method, or by measuring the suppression of voltage relative to the frequency of the applied external current, called the complex resistivity method. IP surveys have been used to determine the clay content of the subsurface or the geochemical properties of several minerals in the subsurface. The surveys can be conducted as profiles or as depth soundings.

Applications Can be used to estimate the clay content of the subsurface and thereby infer hydraulic conductivity. Can be acquired simultaneously with electrical resistivity data. Can be used to detect organic contamination of soils under certain conditions, primarily sites that have 6% to 12% clay content. Can be used to measure the in situ oxidation state of some inorganic compounds.

Limitations Requires surface electrodes and better electrode coupling than standard resistivity surveys. Requires higher currents than electrical resistivity—potentially dangerous. Ability to convert IP soundings to depth sections is limited. IP response is relatively insensitive to increases in clay content above 7% to 10%. Interpretation is qualitative and can be ambiguous. Azimuthal Resistivity. The azimuthal resistivity method is a modification of the standard electrical resistivity method that measures electrical resistivity versus azimuthal orientation at one location. The method is used to identify areas of fractured till or bedrock and determine the orientation of the principal fracture set. Several specialized electrode arrays have been developed to reduce the number of readings required, including the tripotential array and the square array.

Applications Can determine the fracture orientation and fracture porosity from the surface. Uses standard resistivity equipment.

Limitations Data acquisition can be relatively slow and require a relatively large area of open land.

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Limited ability to determine location of individual fractures. Steeply dipping or irregular bedrock surfaces can skew data. Cannot distinguish clay-filled from fluid-filled fractures. Interpretation is quantitative for uniformly fractured materials with no overburden; quantitative interpretation is more limited for more common cases such as discrete fracture zones with overburden. Seismic Methods Seismic Refraction. The seismic refraction method uses the principle of refraction (bending) of sound energy across a boundary between two materials that have different sound velocities. The travel time of sound energy from a seismic source, such as a hammer blow or explosive, to a line of ground motion detectors, called geophones, is measured by a device called a seismograph. The travel time versus distance between the source and the geophones is used to calculate the thickness and depth of the geologic layers present. The method can resolve only layered systems in which each layer has a higher sound velocity than the layer above it. A layer that is too thin or whose velocity is lower than the layer above it cannot be detected and will introduce an error in the depth calculation of deeper layers. The shape of dipping or undulating surfaces can be resolved if several source locations are used in different positions relative to the geophone string.

Applications Relatively simple and fast. Requires less interpretation than other seismic methods. Directly measures the seismic velocity of the materials present. Measured velocities can be used to determine the type of bedrock or soil. Can provide information on depth to bedrock to within about 10%. Can determine the slope and shape of a bedrock surface. Can map dense till sheets beneath softer sediments. Can identify larger faults under the right conditions.

Limitations Typically requires three to five times the surface array length to the depth of penetration. Practical depth limit approximately 150 to 200 feet without a strong seismic source such as explosives. Cannot detect low velocity layers beneath high velocity layers (i.e., sand beneath clay). Limited ability to measure changes in velocity along a layer. Cannot resolve steep dips. Water table mapping only possible in coarse, sandy soils. Sensitive to vibrations from cultural features such as highways. Seismic Reflection. The reflection methods uses a string of geophones connected to a seismograph to measure

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GEOPHYSICS AND REMOTE SENSING

direct reflections of sound energy from geologic boundaries between layers with contrasting acoustic impedance values (density times sound velocity). After significant processing, the data are typically plotted as a time section that shows the two-way travel time to major reflectors that is analogous to a geologic cross section, except that the depth axis is presented in travel time, which is not linear with depth. The method is typically used for deeper, more detailed studies than the refraction method, and the limitation of increasing velocity with depth does not apply. Faults, geologic structure, and dipping or truncating beds are commonly mapped by reflection.

Applications The depth of penetration can be several times the surface array length. More sophisticated interpretation and processing techniques are available. Produces a continuous image of the subsurface that provides a time cross section that can be converted into a depth cross section. Lateral resolution is much better than refraction. Can tolerate steeper dips than refraction. Can accurately map small faults. Can be used onshore or offshore. Can handle low velocity layers beneath high velocity layers.

Limitations Requires more sophisticated field equipment, processing, and field procedures. Data acquisition and processing are more intensive than refraction. Requires sophisticated analysis and accurate seismic velocity data from an independent source to convert from time to depth cross section. Difficult to apply to shallow exploration targets (i.e., above approximately 50 to 100 feet). Interpretation of data subjective and requires experienced interpreter. Sensitive to vibrations from cultural sources such as highways. Gravitational Methods Gravimeter Surveys. The gravitational method uses a sensitive balance, called a gravimeter, to measure variations in the force of gravity at the surface caused by variations in the density of the subsurface. A common modification of the method is called the microgravity method. Microgravity uses a more sensitive gravimeter to detect small near-surface features, such as voids or cavities.

Applications Can be used to detect buried bodies of contrasting density. Can be used to detect bedrock valleys, faults, cavities, and other geologic structures. Data acquisition relatively simple.

Forward modeling can easily be used to predict the success of the method for a given target.

Limitations Requires a series of complex data corrections requiring precise locations and elevations at points of measurement and surrounding topography. Materials of similar density do not produce a measurable anomaly (e.g., granite and quartzite) Large anomalies at depth can mask shallow, smaller anomalies. Interpretation of data produces a nonunique solution. Can be difficult to collect accurate data around buildings or in areas of highly irregular topography. Magnetic Methods The magnetic method measures variations in the intensity of the earth’s magnetic field caused by material of high magnetic susceptibility such as ferrous iron. Measurements can be made with single sensor units, so-called total field magnetometers, that measure the total magnitude of the magnetic field at a given point. Other types of magnetometers make simultaneous measurements at two elevations and compute the vertical gradient of the magnetic field, which increases the sensitivity to shallow targets and decreases the interference from adjacent objects. Some magnetometers measure only the vertical component of the magnetic field, which also reduces interference from adjacent objects. Three types of magnetometers are currently available: (1) Proton precession instruments measure the total field and commonly use two sensors to measure the vertical gradient. Proton precession magnetometers are commonly available and relatively inexpensive, but they are sensitive to steep magnetic gradients from cultural sources, such as power lines and large metal bodies, and do not provide reliable data in many highly developed sites. (2) Flux gate magnetometers measure only a single (usually vertical) component of the magnetic field. The readings are slower, and the sensor must be properly oriented, making data acquisition slower; but the sensor functions well in the presence of steep vertical gradients and is less sensitive to horizontal fields from adjacent objects. The units are less common and more expensive, but they work on sites where proton procession units do not. (3) Cesium magnetometers make fast, very accurate total field measurements, and two sensors are commonly used to make vertical gradient measurements. The sensors can handle steep magnetic gradients and operate at a higher data acquisition rate than other magnetometers. The sensor has an axial blind spot of about 30◦ that may produce erroneously low readings if the sensor is not properly aligned with the target bodies. The units are more common than flux gate systems but more expensive than proton systems.

Applications Field acquisition is fast and relatively easy. Can readily detect buried ferrous objects (e.g., buried drums, tanks, and pipelines).

GEOPHYSICS AND REMOTE SENSING

Can be used to detect geologic structures where materials of contrasting magnetic susceptibility are present (e.g., igneous vs. sedimentary rock).

Limitations Large masses of ferrous objects can have coalescing anomalies that prevent precise location of individual targets. Susceptible to interference from surface metallic objects such as fences or strong electromagnetic fields from power lines. Does not detect material of low magnetic susceptibility (e.g., nonferrous objects such as aluminum or fiber barrels). Interpretation of data produces nonunique results. Electromagnetic Methods Frequency Domain Electromagnetic Induction. Frequency domain EM systems use a two-coil system to transmit an electromagnetic field to induce ‘‘eddy’’ currents in the subsurface and measure the resultant magnetic field, which is a product of the primary field and the induced fields. In some systems, the receiver coil is coincident with the transmitter coil. Systems typically operate at a single frequency that can be selected by the operator. Some systems use a sweep across a range of frequencies. The depth of investigation is controlled by changing the separation between the coils or the frequency of the transmitted field. Two components of the magnetic field are measured, the in-phase or quadrature component and the out of phase component. Some units are designed to operate in low-conductivity soils and use the quadrature term to set the coil spacing. These units are called low induction number systems and generally provide a direct reading of subsurface conductivity. They are easier to use in the field but do not make accurate measurements in soils above a few hundred millimhos per meter. Other systems require measuring the coil spacing independently. These systems make an accurate measurement of subsurface conductivity over a broader range of conditions, but the data must be processed to provide a conductivity measurement. The in-phase component of these units can be used as a direct indicator of high induction number material, such as metal, in the subsurface. Both types of systems produce an apparent conductivity measurement that must be modeled to obtain individual layer conductivity and thickness for a nonuniform subsurface.

Applications No electrodes required; can be used on surfaces where electrode plants would be impossible. Data acquisition is fast and efficient. Can be conducted through freshwater. Different coil orientations and intercoil spacing can be used to accommodate different depths of investigation. Lateral resolution much better than electrical resistivity techniques. Can be used to detect buried metal.

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Detects nonferrous metal (e.g., copper, aluminum, brass, bronze). Sensitive to inclined conductive sheets such as faults or fracture zones. Can be used for azimuthal surveys to map fracture density and orientation. Simple profiling instruments can be used by relatively inexperienced operator (e.g., Geonics EM34 or EM31). Multiple-frequency systems can make vertical depth soundings.

Limitations Depth interpretation of sounding data requires an experienced interpreter. Vertical resolution generally poorer than with electrical resistivity. More sensitive to cultural interference (pipelines, metal fences, power lines, etc.) than electrical resistivity. Relatively insensitive to changes in the conductivity of highly resistive targets. Highly conductive surface material limits depth of penetration. Cannot be used through reinforced concrete. Problem of nonunique interpretation greater than with resistivity method. Very Low Frequency Induction. The VLF method uses low-frequency radio waves from one of several military transmitters as a plane wave source. The instrument measures the magnetic field generated by induced eddy currents. Some instruments use a pair of electrodes to measure the phase shift of the electric field. The depth of penetration of the system is limited by the frequency of the military transmitter and the conductivity of the subsurface. Penetration may be limited in areas of clayrich soils, and the method is generally incapable of detecting features that are aligned perpendicularly to the direction of propagation of the plane wave. The method is generally used for profiling to detect fractures in bedrock, although it does have some limited depth sounding capabilities.

Applications Simple instrumentation and operation. Very rapid and efficient. Sensitive to inclined conductive sheets such as faults or fracture zones. Can map lateral variations in conductivity such as conductive plumes, changes in soil type, or landfill boundaries.

Limitations Limited range in transmitted frequencies available that limits choices of the depth of investigation. Certain transmitters are out of service periodically. Relatively insensitive to changes in conductivity in highly resistive targets. Highly conductive surface materials limit the depth of penetration.

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Location of transmitters makes it difficult to detect linear conductive bodies oriented approximately perpendicularly to the direction of propagation of the plane wave. Long, narrow, near-surface conductors (such as pipelines, fences, or groundwater-filled bedrock depressions) produce strong anomalies that can mimic or mask deeper anomalies. Depth of investigation is totally dependent on subsurface conductivity for a given transmitter frequency. Interpretation is largely qualitative and nonunique. Magnetotellurics. The magnetotelluric (MT) method uses plane waves from natural electromagnetic fields generated by oscillations in the earth’s ionosphere as transmitter signals or from transmitters positioned remote from the area of investigation. The magnetic and electric fields are measured in two horizontal components using two orthogonal electrode dipoles and two orthogonal magnetometers. The intensity and phase of electric and magnetic fields are used to determine the conductivity of the subsurface. The method is commonly used to map conductive targets at depths of hundreds or thousands of meters, where other EM methods are impractical. The magnitude and frequency range of the natural EM signals are variable, and data collection often is limited to the later afternoon or evenings. Deep studies require very low source frequencies that can require hours of data collection to record a single sounding. Shallower studies frequently require a remote transmitter to provide enough EM energy in the audio-frequency range, known as a controlled source audio-frequency magnetotelluric (CSAMT) survey. Interference from power lines and other utilities is common in developed areas. Modern MT systems generally can collect both natural source MT and CSAMT data in a single sounding to provide a broader spectrum of frequencies.

Applications Can find conductive anomalies at depths of a few hundred to a few thousand meters. A large area can be surveyed with a single transmitter setup (CSAMT) or with no transmitter (MT).

Limitations The natural source is often biased toward low frequency, which seriously limits the sensitivity of the method to shallow features without a CSAMT transmitter. Relatively insensitive to small, near-surface targets. Relatively insensitive to highly resistive targets. Sensitive to interference from cultural features such as power lines and buried utilities. Interpretation is nonunique Time Domain Electromagnetic Induction. The TEM (or TDEM) method uses a loop of wire to pass a current of several amps. The current is shut off almost instantaneously to create a broad-frequency-pulsed EM source. Typical transmitter loops are square or rectangular

and a few meters to a few hundred meters on a side. The receiver uses a small magnetic coil to measure the magnetic field over a series of time windows following the shutoff of the current in the transmitter coil. The magnetic field measured at the receiver is a function of the induced eddy currents, which are a function of the conductivity of the subsurface. The data produce a sounding that can be modeled to determine the change in electrical conductivity versus depth. Several systems are available that cover a range in exploration depths of a few tens of meters to a few hundred meters.

Applications Well suited to map conductive bodies at depths of a few tens of meters to a few hundred meters. Often used to map zones of saline groundwater. Vertical resolution significantly better than most multiple-frequency FDEM methods. Can be used to find resistive targets under favorable circumstances.

Limitations Practical upper limit of method is approximately 10 to 20 meters, depending on near-surface conductivity. Interpretation produces nonunique results. Sensitive to interference from linear conductors such as power lines, fences, or pipelines within distances of one or two times the depth of investigation from the transmitter loop. Metal Detectors. Metal detectors are essentially specialized FDEM or TEM devices designed to find buried metal objects. The FDEM instruments use the high response of metal objects to distinguish them from natural materials. TEM instruments take advantage of the relatively long duration of the eddy currents induced in metal bodies. In most instruments, the depth of penetration is related to the coil size, transmitter frequency, and the surface area and shape of the metal target. The depth of penetration is generally unaffected by soil type over the conductivity range of typical soils. The instruments have high lateral resolution, but the vertical resolution of the depth of the target is generally limited.

Applications Ease of use. Very portable. High lateral resolution.

Limitations Limited depth of investigation. Response of target related to the surface area of the target rather than the volume of metal. Ground Penetrating Radar. GPR uses a high-frequency pulse of EM energy to probe the subsurface. The EM energy is reflected off boundaries of contrasting dielectric constant (the ability of a material to separate an electric charge), so the instrument can find nonmetallic or metallic

GEOPHYSICS AND REMOTE SENSING

targets. The output is a time versus amplitude plot that is generally displayed as a series of plots along a transect line. The field plots are analogous to cross sections, except that the vertical access is in two-way travel time, not depth. The data can be processed similarly to seismic reflection data to produce depth sections. The instruments are configured with single transmitter/receiver antennas or with separate transmitter and receiver antennas that can be offset to collect data at different separations to increase the data processing that can be conducted. Under ideal conditions, the GPR method provides very highresolution images of buried bodies, stratigraphic structure, or areas of disrupted soils. The penetration of the method is determined by the conductivity of the subsurface and the degree of scattering from rubble or other small targets. Penetration is severely limited in highly conductive soils or in areas with lots of buried rubble or other discontinuous reflectors. The depth of penetration can be increased to some degree by using a lower frequency antenna but at the cost of resolution of the target.

Applications Resolution can be of the order of a few centimeters. Provides a cross section of subsurface in the field. Can detect buried, ferrous or nonferrous targets. Can detect disturbed soil zones. Sophisticated data processing techniques available (similar to seismic reflection processing).

Limitations Equipment is somewhat cumbersome; usually requires a relatively flat surface. Depth of penetration is seriously limited by conductive material such as clay or water. Penetration through moist clay can be less than 1 foot. Decreasing transmitter frequency to increase penetration decreases resolution. Penetration and resolution limited by scattering effects at sites with buried cobbles or rubble. Geothermal Techniques The geothermal method uses soil temperature measurements to detect zones of anomalous flow at depth. Soil temperature is measured by using a shallow probe driven a few feet below the surface or a dedicated monitoring probe installed in a shallow borehole. Surface temperature variations diminish with depth below the surface. Daily temperature variations penetrate less than 1 or 2 meters, whereas seasonal temperature variations generally penetrate only about 10 meters. Below a depth of a few meters, soil temperature variations are largely a function of heat flow from within the earth. On a local scale, variations in heat flux often indicate zones of anomalous groundwater flow. Geothermal methods have been used to find permeable zones at depths of a few meters to over 100 meters using probes at depths of less than 1 meter to approximately 15 meters. A principal advantage of the method is that it detects anomalous flow instead of some other less directly related property, such as electrical

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resistivity. The method is susceptible to thermal interference by infiltration from shallow sources or from small features near the probes.

Applications Identifies areas of high groundwater flow such as bedrock fractures or clean sand and gravel deposits. One of the few methods that can directly detect moving groundwater from the surface. Can pinpoint location of maximum permeability. Can detect permeable features at depths of more than 100 feet beneath probes.

Limitations Requires relatively uniform soils above target interval. Small sand lenses near probes can create false anomalies or mask deeper anomalies. Interpretation is qualitative. Probe installation is relatively labor-intensive. GROUNDWATER GEOPHYSICS—BOREHOLE METHODS Most of the geophysical measurements made at the surface can also be made in a borehole using sensors lowered down the hole on a cable connected to up-hole recording equipment. Borehole measurements tend to be more accurate and precise because the borehole environment tends to be more predictable and less prone to noise. The field of measurement is typically only a few inches to a few feet around the borehole. Due to the small scale of the measurement, borehole methods tend to produce excellent vertical resolution. The combination of high resolution and greater accuracy make most borehole methods suitable for quantitative measurements and small-scale correlation of stratigraphic properties. Borehole measurements are commonly used to determine the physical properties of the formation and formation fluids. The measurements can be used to measure directly or interpret several properties such as lithology, porosity, water quality, borehole diameter or alignment, borehole flow, formation or fluid temperature, mineralogy, and other properties of interest. A variety of logging tools are available. Each tool has specific requirements for the borehole environment. Some tools can measure through casing; most cannot. Some tools require a fluid-filled hole; others operate in air, water, or drilling mud. Electrical Logs Spontaneous Potential Log. The SP log measures the electrical potential between a surface reference electrode and an exposed electrode that is raised up or down the borehole. The SP response is caused by a difference in resistivity between the borehole fluid and the formation fluid.

Applications Quick and inexpensive. Can be a good sand versus shale indicator. Can be used to estimate total dissolved solids (TDS) of formation fluid.

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Limitations If borehole fluid and formation fluid have similar resistivity (typical for a water well), there is no SP response. Works only in uncased fluid-filled boreholes. Single Point Resistance Log. The SPR log measures the resistance between a grounded surface electrode and an electrode that is moved up or down the borehole. The tool measures the resistance of the total thickness of the formation between two electrodes. As a result, the measurement is in units of resistance, which is not a material property. The log is used primarily to detect changes in the formation but is not appropriate for quantitative analysis of the physical properties of the formation or the formation fluid.

Applications Quick and inexpensive. Good indicator of formation contacts.

Limitations Measures resistance, not resistivity. Works only in uncased fluid-filled boreholes. Resistivity Log. Resistivity logs use two or more potential electrodes in the borehole to measure the change in potential difference along the borehole across a fixed electrode separation. Both current electrodes can be located on the down-hole sonde, or more commonly, one current electrode is on the sonde, and the other is planted at the surface. The depth of penetration of the current can be controlled by using different electrode spacings. Typically, a short electrode spacing is used to provide a high-resolution measurement of formation changes, and one or more longer electrode spacings are used for more accurate measurements of the bulk formation resistivity. Some sondes use additional current electrodes to focus the electrical field into the formation. The data can be corrected for borehole effects to provide a quantitative measurement of the bulk formation resistivity which is a function of the resistivity of the formation and formation fluid.

Can be run in open hole or through PVC casing. Can be run with shallow or deep depths of investigation (dual induction) to determine the invasive effects of borehole fluids.

Limitations More expensive than electrode logs. Response of highly resistive materials (freshwater aquifers) is small, so the sensitivity of the tool is reduced. Cannot be run in highly conductive borehole fluids. Passive Radioactivity Log Natural Gamma Log. Natural gamma logs use a sensor that measures the gamma-ray energy emitted by the formation. In most formations, gamma-ray emissions are almost entirely produced by shale minerals. As a result, the gamma-ray response is a good indicator of the concentration of clay in the formation.

Applications Good clay indicator Can be run in cased or uncased, fluid-filled or airfilled boreholes.

Limitations Gamma-ray units (API units or counts per second) are arbitrary units and do not relate directly to clay mineral content. Raw data must be corrected for borehole diameter effects before making direct comparisons. Spectral Gamma Log. A spectral gamma log separates gamma-ray energy by frequency to measure the spectrum of gamma rays emitted by the formation. The frequency spectrum of gamma rays emitted by the formation is diagnostic of the radioactive isotopes in the formation. The gamma-ray spectrum is commonly used to determine the concentrations of potassium, uranium, and thorium in the formation (KUT logs), but full spectrum tools are available that allow identifying additional elements.

Applications

Applications

Measures bulk resistivity of formation. Can be used to determine the formation lithology, porosity, and fluid resistivity.

Can be used to perform ‘‘complex formation analysis,’’ where the mineral content and groundwater flow history of formation are estimated. Can be used to detect uranium or radium-enriched zones of sandstone aquifers. May have application to radioactive contamination problems.

Limitations Works only in uncased fluid-filled boreholes. Induction Log. Induction logs use a frequency domain electromagnetic (EM) sonde to measure the conductivity of the formation around the borehole. Some tools include a short spacing electrode tool to measure the conductivity of the borehole fluid independently.

Limitations Equipment is significantly more expensive than natural gamma tools. Porosity Log

Applications Can be run in fluid-filled or air-filled borehole.

Several logging tools have been developed to measure the porosity of a formation. Porosity is a useful property

GEOPHYSICS AND REMOTE SENSING

to measure, but it does not always relate directly to permeability, which is generally the property of greatest interest. If more than one porosity tool is used, the data can be cross-plotted to identify formation type directly and correct the porosity measurement for lithologic effects. Most porosity tools require the use of a radioactive source in the tool. The use of radioactive sources requires special licensing and is prohibited in water wells in some states. The following sections describe the most common types of porosity tools. Neutron Density Log. Neutron density logs use a radioactive source to bombard a formation with highenergy neutrons and records gamma-ray energy ‘‘bounced back’’ from the formation. The tool measures the hydrogen content of a formation, which is equivalent to the water content of the formation if no hydrocarbons are present. The water content of the formation is equivalent to the porosity below the water table and the soil moisture content above the water table. Some tools do not compensate for changes in borehole diameter and produce erroneously high estimates of porosity in washouts in the borehole. Compensated tools can accommodate some degree of variation in borehole diameter, but erroneous data can be recorded in zones with rough borehole walls. Bound water in shale zones produces erroneously high values of measured porosity.

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Limitations Sensitive to variations in borehole diameter. Presence of hydrocarbons causes erroneously high porosity values. Stringent licensing requirements to make a log due to the radioactive source. Sonic Log. Sonic logs use a source of high-frequency sound and one or more detectors on the down-hole sonde to measure the transit time of sound across a known distance of the formation. By comparing the measured velocity of sound in the formation to the known sound velocity for the rock type, the percentage of void space (porosity) can be calculated. Sonic tools do not require a radioactive source. The measurement is sensitive to irregular hole conditions. Specialized versions of the tool record the full waveform of the sound energy transmitted and can be used to detect poor bonds between the casing and grout or find voids behind the casing.

Applications Porosity can be calculated for a given matrix (sandstone or limestone). Modified sonic log can be used to inspect the bond of the cement grout to the casing (cement bond log). Full waveform sonic logs can be run to measure the engineering properties of the formation.

Applications

Limitations

Directly measures water content of formation. Can be related to porosity of nonshale units. Measures porosity below water table and moisture content above water table. Can be run in a cased or uncased hole.

Sensitive to borehole diameter and roughness of borehole wall. Works only in a fluid-filled uncased hole (except the cement bond log). Borehole Condition Log

Limitations Sensitive to variations in borehole diameter. Will see ‘‘bound water’’ in shale zones as porosity and give unrealistically high values. Presence of hydrocarbons (pure product) erroneously suppresses calculated porosity. Stringent licensing requirements to make a log due to the radioactive source. Application to cased holes is limited. Gamma Density Log. Gamma density tools use a radioactive source to bombard a formation with highenergy gamma rays. The tool has a sensor that records gamma rays ‘‘scattered’’ back to the detector. The magnitude of the backscattered gamma rays is a measurement of the electron density of formation that is related directly to bulk density. Given a bulk density measurement, the formation porosity is calculated for an assumed rock matrix (sandstone or limestone).

Applications Will record accurate porosity values in shale zones. Can be run in cased or uncased boreholes.

Several types of logs have been developed to measure the physical condition of the borehole, rather than measuring the properties of the surrounding formation. These tools can be used to determine the integrity of a borehole, determine if the hole is straight, or determine if flow is occurring between different zones. The following section describes several of the more common logs. Down-Hole Televising Log. Down-hole televising logs use a down-hole video camera to inspect a borehole visually. Most cameras provide a fish-eye, forward-looking view with a depth reading on the image. Some cameras offer a side-looking view that looks directly at the borehole wall and can be rotated 360◦ .

Applications Produces a visual record of a well bore on VCR tape or DVD. Can be used to inspect borehole irregularities or obstructions. Can be used to inspect the casing. Can be used to detect fractures or solution cavities. Shows small-scale stratigraphic layering with more resolution than other logs.

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Valuable tool in assessing options to recover lost logging tools or drilling equipment.

Limitations Requires clear borehole fluid. May have to add clear water to flush the hole if the water is cloudy. Produces a qualitative record. Does not measure any physical property in calibrated units. Most tools do not provide an oriented image. Caliper Log. Caliper tools use spring-loaded arms that scrape along the borehole wall to record the variation in diameter. Simple tools that use two arms tend to turn so that the tool always measures the maximum hole diameter. Three-arm tools average the diameter at three points and provide better measurements of the average hole diameter. Four-arm caliper tools are available that provide oriented independent measurements in two coordinates in the horizontal plane that provide additional information on the shape of the hole.

Applications Can identify ledges or washouts that can identify incompetent zones or fractures. Can be used to calculate volumes of grout or backfill. Can be measured in x and y planes (four-arm caliper) to estimate the shape of the borehole.

Limitations Two-arm caliper measures only maximum hole diameter. Can be misleading in elliptical boreholes. Does not measure alignment of borehole. Alignment Log. Alignment tools measure the deviation of the center of the borehole from vertical. Simple alignment tools use a weighted plumb bob suspended from a tripod. The deflection of the plumb bob is measured by crosshairs at the surface. More sophisticated alignment tools use cameras with targets mounted on a system of gimbals or gyroscopes and magnetometers to measure the deviation of the borehole.

Applications Can identify deviations in a borehole that can limit the size of the casing or pump that can be installed. Can be used to calculate actual well depth and actual lateral position of a borehole.

Limitations Must be combined with a caliper log to determine the shape of the hole accurately. Large washouts can cause erroneous alignment readings unless the diameter of the hole is taken into account. Sharp doglegs or extreme deviations can cause a surface line to ‘‘hold up’’ on the side of a borehole for simple tripod and crosshair systems and make measurements from below impossible.

Alignment tools that use magnetometers do not work inside a metal casing. Acoustic Televiewer. Produces a high-resolution image of a borehole wall using a high-frequency sonic tool that is rotated to scan the hole. The log produces an image that shows relief on the borehole wall as shading or color patterns. Standard log formats show the borehole wall from 0 to 360◦ on the horizontal axis and depth on the vertical axis. The tools are generally oriented with respect to true directions. Televiewer logs are often used to map fractures and plot their strike and dip. The tool requires a fluid-filled hole, but water can be added to raise the water level artificially for logging.

Applications Can measure the aperture, strike, and dip of fractures. Can operate in either mud- or water-filled boreholes.

Limitations Does not operate in air-filled holes. Measures only the surface of the borehole; does not penetrate scale or casing. The resolution of the tool is limited to several millimeters by the wavelength of the sonic source. Scrapes or washouts from the drilling process can obscure true fractures. Optical Televiewer. Produces a high-resolution image of the borehole wall using an optical scanning tool that is rotated to scan the hole. The log produces an image of the borehole wall in true color. Standard log formats show the borehole wall from 0 to 360◦ on the horizontal axis and depth on the vertical axis. The tools are generally oriented with respect to true directions. Optical televiewer logs are often used to map fractures, plot their strike and dip, and image small-scale stratigraphic features. The tool requires a clear borehole, either air or clear water.

Applications Can measure the aperture, strike, and dip of fractures. Can image individual grains and other fine stratigraphic features.

Limitations Does not operate in mud-filled holes, cloudy water, or dusty air-filled holes. Measures only the surface of the borehole; does not penetrate scale or casing. Scrapes or washouts from the drilling process can obscure true fractures. Flowmeter Log. Flowmeter tools are used to measure the flow within a borehole that develops from head differences between different zones or while pumping. Spinner tools are simple flowmeters that use small impellers to measure the flow past the tube. The tools are usually trolled up or down the borehole to overcome the frictional resistance of the impeller and its bearings.

GEOPHYSICS AND REMOTE SENSING

For stronger flows, the tool can be held stationary to measure the absolute flow velocity past the tool without correcting for the speed of the tool in the borehole. Heat pulse tools can be used to measure lower flow rates. Heat pulse tools use a heating element to create a small pulse of heated water, and temperature sensors are placed at known distances above and below the heating element. The tool measures the time required for the pulse of heated water to move upward or downward to the sensor and calculates the flow velocity and direction. A modification of the heat pulse flowmeter has been used to measure horizontal flow in boreholes. The diameter of the hole must be known before the flow velocity can be converted into a volumetric flow rate. Specialized tools have been developed to detect flow behind a casing.

Applications Can measure the flow up or down a borehole. Can be used to estimate the head difference and transmissivity of different zones of a borehole.

Limitations The friction of the bearings and inertial mass of the impeller impose a lower limit of flow that can be measured by a spinner tool. Heat pulse tools have an upper limit of flow rate that can be accurately measured. Temperature Logs Down-Hole Temperature Log. Temperature logs use down-hole thermistors or thermocouples to measure the variation in the temperature of the borehole fluid with depth. Temperature variations in the hole can be used as indicators of flow. The data can be plotted as absolute temperature or as a differential temperature that removes the effect of the geothermal gradient from the plot.

Applications Provides direct measurement of fluid temperature. Can be used to detect zones of increased fluid flow within a borehole or between two boreholes. Directly measures the geothermal gradient.

Limitations Geothermal gradient causes a continuous drift in borehole temperatures that can mask thermal variations within the borehole. Requires temperature and head variations between units to create a thermal anomaly to detect permeable zones within a borehole. Flow within a borehole during or after drilling can mask geothermal gradient measurements. GEOPHYSICAL SURVEY DESIGN When designing a geophysical survey, several factors must be considered, including the physical properties of the

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target, the desired resolution, the physical properties of the matrix around the target, the ambient noise in the survey area, and the physical access at the site. To be detected, the feature you are trying to find must have some physical property that has sufficient contrast with the surrounding matrix. A variety of physical properties can be remotely sensed by geophysical methods. Some targets, such as buried pipes in sand, present obvious contrasts of a number of properties. Some targets, such as excavations backfilled with native soils, present less obvious contrasts or can be identified only by contrasts of related features, such as disruption of macropore drainage structures causing perched soil moisture over old graves. Geophysical methods each have characteristic resolution limits that are a function of the wavelength of the fields used for measurement and the limitations of the equipment. The resolution of the various geophysical methods varies from several tens of feet for deep studies using electromagnetic induction methods to fractions of an inch using ground penetrating radar with high-frequency antennas. The resolution of the geophysical method used must match the physical dimensions of the target body if the target is to be detected. The magnitude and size of the anomaly must be predicted with reasonable accuracy, typically by forward modeling, to determine the minimum grid size that can be used for a reasonable probability of detection. Site conditions also play an important role in determining the ultimate success of the survey. Surface and subsurface conditions can create a noisy environment that can easily obscure the signal from the target body. Noise sources are different for each method. For instance, ground vibrations from traffic or machinery can be a major source of noise for seismic or gravity methods but do not affect other methods such as electromagnetic induction or magnetics. Physical conditions at a site also affect the utility of geophysical methods. Major obstructions such as buildings or developed land often limit the area where data can be gathered and create inherent limitations on the geophysical survey. Buildings, parked cars, dumpsters, overhead power lines, buried utilities, fences, buried rubble, and other cultural features are not often considered when designing a geophysical survey. In addition, subsurface conditions such as soil type, uniformity, and other factors affect the propagation of energy and can limit the performance of most methods. Surface conditions, such as paved surfaces, heavy vegetation, frozen ground, rough topography, and surface debris, must be considered when designing a survey. Assuming that the site conditions are acceptable, it is still necessary to select a method that can detect the target body. The following is a partial list of typical survey objectives with some suggestions of methods to consider. 1. Locating optimal well locations (bedrock fractures or sand and gravel deposits) a. EM b. Electrical resistivity

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2.

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c. Azimuthal EM d. Azimuthal resistivity e. Geothermal Determining depth to bedrock and mapping bedrock surface a. Seismic refraction b. Gravity c. Seismic reflection d. GPR Locating preferred pathways for groundwater flow (bedrock fractures, voids, seepage paths, or sand and gravel lenses) a. EM b. Electrical resistivity c. Azimuthal EM d. Azimuthal resistivity e. Geothermal f. SP g. GPR Finding buried metal bodies (tanks, drums, Buicks, etc.) a. Magnetometer b. Electromagnetic induction (EM) c. Ground penetrating radar (GPR) d. Metal detectors Mapping conductive ground water plumes (inorganics in the range of hundreds of ppm) a. EM b. Electrical resistivity c. GPR Mapping continuity of till sheets or clay layers a. Seismic refraction b. EM c. Electrical resistivity d. Seismic reflection Determining the lateral extent or thickness of refuse a. EM b. Electrical resistivity c. Magnetometer Detecting fractures in tills a. Azimuthal EM b. Azimuthal resistivity Finding abandoned well casings a. Magnetometer b. EM Detecting voids a. Gravity b. GPR c. Resistivity Direct detection of hydrocarbons a. GPR (LNAPLS, DNAPLS, or biodegrading plumes) b. Electrical resistivity (biodegrading plumes) c. Induced polarization (bound only to clay)

GEOTHERMAL WATER NITISH PRIYADARSHI Ranchi University Ranchi, Jharkhand, India

Geothermal water has a temperature appreciably higher than that of the local average annual air temperature. However, in general, a spring is considered hot when its temperature is about 12.2 ◦ C higher than mean annual ambient temperature (1). The relative terms geothermal water, warm springs, and hot springs are common. From earliest times, hot geothermal water or geothermal springs have been used for bathing and cooking. The first technological success in using geothermal energy was in Italy in 1904 where the worlds first geothermally driven electrical plant was opened and operated. Today, applications of low- and moderate-temperature (100 to 300 ◦ F) geothermal waters have expanded enormously to include heating large tracts of homes and buildings (district heating), heating greenhouses for growing vegetables and flowers, fish farming (aquaculture), drying foods and lumber, and many other uses. In the United States alone, there are 17 district heating systems, 38 greenhouse complexes, 28 fish farms, 12 industrial plants, and 218 spas that use geothermal waters to provide heat. The district heating system in Boise, Idaho, has been operating since the 1890s and continues to provide heat today. In Iceland, most of the homes and other buildings are connected to geothermal district heating systems, and in the Paris basin in France, many homes are heated by bringing geothermal water to the surface. Geothermal greenhouses are prominent in Italy and in the western United States. Worldwide, there are about 12,000 thermal megawatts of installed direct uses of geothermal fluids in nearly 30 countries, replacing the combustion of fossil fuels equivalent to burning 830 million gallons of oil or 4.4 million tons of coal per year. This illustrates the trend of countries switching to alternative energy sources such as geothermal energy. Geothermal water discharges from numerous springs located mostly in mountainous or plateau areas. The springs are connected by faults to deeply buried reservoirs that contain geothermal water, which moves upward along the fault zones to discharge at the land surface. Much geothermal water discharges as hot springs that flow steadily instead of erupting at intervals. Hydrothermal phenomena involving the release of water and steam are nearly always associated with volcanic rocks and tend to be concentrated in regions where large geothermal gradients occur. By implication, aquifers must also be present that permit water to percolate to great depths—often 1500 to 3000 m (2). One theory used to explain how geothermal water becomes heated in areas that are underlain by complex geologic structures is that when precipitation falls in highland areas, it recharges the aquifer system. Some of the water moves downward along faults and fracture zones to great depths. As the water descends, it becomes heated because of the geothermal gradient. At some depth, the heated water becomes lighter

GEOTHERMAL WATER

than the overlying water and then moves upward along faults to discharge as spring flow. Deeply circulating groundwater can also become heated by cooling magma (molten igneous rock) at great depths in the crust of the earth. The water warms as it descends, possibly along fault zones that overlie the magma chamber, until it absorbs enough heat to become lighter than overlying water. The warm water then rises to the surface. The mechanism for circulating the water is the same, regardless of whether the water becomes heated by the geothermal gradient or by the buried cooling magma. It can also be heated through mantle decay of radioactive elements such as U, Th, and Ra, tectonic activity, metamorphic processes, and exothermic reactions of minerals. More than 10,000 individual thermal features, including geysers, hot springs, mud pots, and fumaroles (steam vents), have been identified. Geysers are the best known and certainly the most spectacular features. They are a type of hot spring that periodically emits sudden, violent eruptions of stream and hot water. Water from surface sources and/or shallow aquifers drains downward into a deep vertical tube where it is heated above the boiling point. As the pressure increases, the steam pushes upward; this releases some water at the surface, which reduces the hydrostatic pressure and causes the deeper superheated water to accelerate upward and to flash into stream. The geyser then surges into full eruption for a short interval until the pressure is dissipated; thereafter, the filling begins, and the cycle is repeated (2). The explosive release of pressure can cause a column of steam and hot water to rise 200 feet or more into air. The period between eruptions depends on several factors, including the volume of steam and water that is ejected and how rapidly ground water refills the tubes and chambers. A mud pot, results when only a limited supply of water is available. Here, water mixes with clay and undissolved particles brought to the surface, forming a muddy suspension from the small amount of water and steam that continues to bubble to the surface. A fumarole, meaning smoke, is an opening through which only steam and other gases such as hydrogen sulfide and carbon dioxide discharge. These features are normally found on hillsides above the level of flowing thermal springs; water can often be heard boiling underground (3). Excessive concentration of certain dissolved minerals in geothermal water poses water quality problems. The most common of these minerals are dissolved fluoride, arsenic, and iron. Concentration of dissolved fluoride in excess of 4 milligrams per liter can cause mottling of teeth, especially children’s, and can cause bones to become brittle. Dissolved iron is not detrimental to human health, but concentrations in excess of 200 micrograms per liter can cause staining of kitchen and bathroom fixtures and can cause clogging of well-screen openings and pumps. Concentrations of dissolved arsenic in excess of 30 micrograms per liter are toxic to humans. Sodium, calcium, magnesium, and their salts are nontoxic to human beings. Chloride salts of these metals in higher concentrations may be toxic to plants because of their chloride ions rather than the metals. A chloride ion concentration of >300 mg/L is in general toxic to vegetation and results in defoliation,

157

chlorosis, bronzing, and burning of plants (4). Boron in trace quantities is necessary for plants, but it is toxic to certain plants in concentrations as low as 1 to 2 mg/L. Natural steam that always contains some noncondensable gases, such as hydrogen sulphide, and ammonia, presents the greatest potential hazard. Ammonia affects primarily the upper respiratory tract and causes coughing, vomiting, and redness of the mucus membranes of the mouth, nose, lips, and pharynx (4). In geothermal operations, ammonia is not likely to present a direct toxic hazard except possibly in the immediate vicinity of the power plant. Hydrogen sulphide, which is a noticeable geothermal effluent, is much more toxic than commonly realized. A concentration of more than 600 ppm hydrogen sulfide can cause death within 1 hour. It can also paralyze the respiratory center. High values of total dissolved solids in geothermal water sometimes interfere with the usefulness of water as a source of energy. Rain, the prime source of meteoric water, which feeds the geothermal reservoir, is practically devoid of chemical ingredients. But after entering the atmosphere, it picks up gases such as CO2 , O2 , SO2 , and dust particles. As it falls to the ground, its concentration of chemical constituents may be up to 40 mg/L (5). Other important factors are climatic conditions and the state of weathering of rocks and their mineralogical compositions. Thus, even before entering the geothermal reservoir, water is considerably rich, chemically. Groundwaters whose salt content is around 1000 mg/L are not uncommon. The major cations in groundwater are Ca, Mg, Na, and the anions are HCO3 , SO4 and Cl. Geothermal water commonly contains large concentrations of silica (if the water has moved through limestone or other calcite-rich rocks). Along with these radicals, gases dissolved in meteoric water such as CO2 , O2 , He, and Ar also enter the reservoir. The chemistry of meteoric water is modified in the geothermal reservoir due to rock–water interaction under the influence of temperature, contributions from magmatic sources, mixing of different types of waters, and decay of U, Th, and Ra. Closer to the surface, within the zone of oxidation, air can also enter thermal waters. In that case, the nitrogen and oxygen content will be high, and their mutual ratio may also match that of the atmosphere. Some geothermal waters may be brines, for example, the Salton Sea area, California, whose salinity is 30% (6); others may be as pure as distilled water, for example, the Rajgir thermal spring of Bihar, India (7). Microorganisms can thrive in geothermal waters, even at boiling temperature (8). Extremely thermophilic bacteria that survive at 350 ◦ C have also been found (9). Fortunately, these microorganisms do not survive in animal systems. Pathogenic organisms can enter thermal waters through soil and air and also through contamination by animal/man in thermal water pools. Thermal springs of various kinds are found throughout the world. Notable areas exist in the United States, Iceland, New Zealand, the Kamchatka Peninsula of the former Soviet Union, Brazil, Argentina, Ethiopia, Zambia, China, Tibet, India, Thailand, Taiwan, and Japan. Yellowstone National Park in Wyoming, United States, that contains literally thousands of hydrothermal features possesses the greatest

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concentration of thermal springs in the world (10). This area marks the site of an enormous volcanic eruption 600,000 years ago. Today, a temperature of 240 ◦ C exists only 300 m below the ground surface. As economies expand and population grows, energy demand worldwide is increasing rapidly. If we are to increase our energy consumption while simultaneously reducing environmental pollution, we must change our fuel mix, which today relies heavily on fossil and nuclear fuels. The use of geothermal energy has enormous environmental advantages over the use of fossil or nuclear fuels. Among these advantages are far fewer and more easily controlled atmospheric emissions, maintenance of groundwater quality, and much smaller dedicated land requirements. The small quantities of gases emitted from geothermal power plants are naturally occurring and result from geologic processes. Because the earth is porous and permeable, these gases would eventually find their way to the surface, even in the absence of geothermal power development. BIBLIOGRAPHY 1. Alfro, C. and Wallace, M. (1994). Origin and classification of springs and historical review with current applications. Environ. Geol. 24(2): 112–124. 2. Todd, D.K. (1995). Ground Water Hydrology. Wiley, Toronto, Canada, p. 51. 3. Poland, J.F. (1972). Glossary of Selected Terms Useful in Studies of the Mechanics of Aquifer Systems and Land Subsidence due to Fluid Withdrawl. U.S. Geological Survey Water-Supply Paper 2025, p. 9. 4. Pandey, S.N. and Srivastava, G.C. (1996). In: Environmental Hazards of Indian Geothermal Fields. U.L. Pitale and R.N. Padhi (Eds.). Geol. Surv. India, Spl.Vol. 45, pp. 375–378. 5. Hawkes, H.E. and Webb, J.S. (1962). Geochemistry in Mineral Exploration. Harper and Row, New York. 6. Werner, H.H. (1975). Contribution to mineral extraction from superheated geothermal brines, Salton Sea Area, California. Geothermics 2: 1651–1657. 7. Geological Survey of India Special Publication. (1991). Geothermal Atlas of India, p. 106. 8. Yun, Z. (1986). Thermophilic microorganisms in the hot springs of Tengencheng geothermal area, West Yunan, China. Geothermics 15: 347–358. 9. Baross, J.A. and Dening, J.W. (1983). Growth of ‘‘black smoker’’ bacteria at a temperature of at least 250 ◦ C. Nature 303: 423–426. 10. Keefer, W.R. (1971). The Geologic Story of Yellowstone National Park. U.S. Geological Survey Bull. Publ. No. 1347, p. 92.

GHIJBEN–HERZBERG EQUILIBRIUM EKKEHARD HOLZBECHER ¨ Berlin Humboldt Universitat Berlin, Germany

THE ORIGINAL PAPERS In 1889 a note was published in The Hague (The Netherlands), which dealt with a suitable location for

a new borehole for the pumping of drinking water for the city of Amsterdam. The authors, Drabbe and Badon Ghijben (1), outline in length the different geological formations, their extension and their connection with each other, and especially their connection with various saline surface water bodies in that region. In the final part, salinity becomes the subject of their consideration. The authors describe that saltwater is found in the subsurface, when there is a hydraulic gradient from the surface water into the aquifer and contrast this to the situation in the dunes at the North Sea coast. There, a freshwater lens can be observed on top of the salt water. In one sentence, the topic of this contribution can be summarized: If the height of the groundwater table above the mean sea level is denoted by a, the equilibrium between saltwater and freshwater is reached at a depth of a/0.0238 = 42a, when the seawater weight exceeds the freshwater weight by a factor of 1.0238. A simple calculation example follows using the given formula and a reference to an even earlier communication of Conrad in 1881, who is cited to have used factors between 40 and 50 in analogy to the proposed 42. Twelve years later, the other note (2) appeared, in which the Ghijben–Herzberg equilibrium (GHE) is named. Herzberg, an engineer from Berlin (Germany), in a talk in Vienna on behalf of the annual meeting of the German Association of Gas- and Water-Specialists, was concerned with the water supply of villages on the German North Sea coast. More specifically, he was interested in the water supply on the North Sea islands, which, due to a rise in tourism, had an increasing demand for drinking water in the summer season. Compared with his Dutch counterpart, Herzberg provided a much more detailed derivation of the equilibrium rule. In order to illustrate his arguments, Herzberg presented a figure, which is reproduced in Fig. 1. In Herzberg’s version, t denotes the elevation of the groundwater table above mean sea level and is thus identical to Badon Ghijben’s a; h is the depth of the freshwater boundary below mean sea level. A saltwater column of length h has the same weight as a freshwater column of length H = t + h, when the condition h · 1.027 = h + t

(1)

is fulfilled, and when it is taken into account that the specific weight of saltwater exceeds that of freshwater by a factor of 1.027. Solving Eq. 1 for h delivers h = 37t

(2)

In both studies it is shown that a wide freshwater lens is to be expected in coastal regions. The extension of this length is roughly 40 times the elevation of the groundwater table above the seawater level. For practical purposes—the authors’ main concern was water supply—one conclusion is that quite a large amount of fresh groundwater is available, as the lens extends much deeper into the subsurface than above the seawater horizon. Another conclusion is that the occurrence of freshwater at greater depth does not rely on the existence of an additional freshwater source in deeper horizons.

GHIJBEN–HERZBERG EQUILIBRIUM

Groundwater level

159

t H h Figure 1. Illustration concerning the freshwater–saltwater equilibrium. Adapted from Reference 2.

Seawater

Following the Ghijben note, a strong dispute took place in the beginning of the twentieth century in The Netherlands with respect to the maximum amount of pumping by the water works, as reported by de Vries (3). Proponents of the GHE argued in favor of such a regulated maximum. In both publications the density difference is recognized as the major factor concerning the conditions of subsurface water near the sea coast. Freshwater is lighter than saltwater, and in order to reach an equilibrium (i.e., equal pressure), the freshwater column must extend further into the subsurface. The general situation with freshwater salinity ρf and water of elevated salinity ρs in the deeper subsurface, in current formulation, would be written as hf = −

ρs − ρf z ρf

(3)

where hf denotes the elevation of freshwater above sea level, while z measures the depth of the interface below that level. As both extend in different directions from the zero sea level, they have different signs. The equilibrium formulas given above can be taken as rule of a thumb. In most textbooks the formula is used with a factor of 40. The deviation of the factor in the formulation of Badon Ghijben (42) from the one given by Herzberg (37) stems from different assumptions concerning the relative weight of saltwater in relation to the freshwater. According to current knowledge, the density of seawater ρs lies between 1.02 and 1.029 g/m3 , depending on salinity and temperature. For a mean salinity of 35 ppt and a temperature of 10 ◦ C (typical for subsurface water), the density is 1.027 g/m3 (4), which is the value used by Herzberg. In real situations, it has to be taken into account that, due to mixing, dissolution, or precipitation, the subsurface saltwater may have a different salinity than seawater and that freshwater may have a slightly elevated density too. CONDITIONS The Ghijben–Herzberg equilibrium (GHE) has proved to be valid in several coastal systems all around the world. In both original publications, correspondence with observations is stressed. Herzberg describes that he already used the equilibrium formula for predictions concerning the depth of the saltwater in the sandy subsurface of several North Sea islands and that he obtained a good confirmation of the rule by measurements in boreholes. De Vries (3) lists field measurements and laboratory experiments, performed by Pennink in The

Netherlands at the beginning of the twentieth century. However, the formula is linked to several conditions. In the derivation it is assumed that the transition from the fresh to saline water appears quite rapidly. When the scale of the transition zone is very small in relation to the scale of the depth of the aquifer, one speaks of a sharp interface, which is a common term in the current literature; see, for example, Reference 5. One condition for a sharp interface is small dispersion, more specifically small transverse dispersion, as the flow direction coincides more or less with the interface (see below). But the tides and seasonal fluctuations in groundwater recharge and/or in the amount of groundwater being withdrawn also have an influence on the width of the transition zone. Extended descriptions have been derived nowadays, which are valid for situations with transition zones (see below). Herzberg demanded caution concerning the transfer of his results obtained for the islands to cities near the coast, pinpointing the different scale, especially concerning the water supply. In a remark to Herzberg in Vienna, Halbertsma, another Dutchman, stated that the situation along the Dutch coast is quite different from one place to the other, but in some parts it is very similar to the situation described by Herzberg. As a counterexample he reports a situation where a confining impermeable clay layer can be found near the surface, which separates the freshwater on top from the saline water. Custodio (6) illustrates several different situations, concerning the position of one or more clay layers, which disturb the ideal situation assumed for the GHE, showing that the GHE may over- or underestimate the position of the interface depending on local circumstances. In atoll islands a permeable layer below the upper strata of low hydraulic conductivity also results in a different situation, and the GHE can surely not be applied on coastlines of volcanic rock. In the very vicinity of the coastline the GHE is mostly not valid. The GHE predicts that both fresh and saltwater heads (see below) should approach zero. But at many locations submarine groundwater discharge of freshwater has been observed (see Fig. 2). Kooi and Groen (7) provide an overview of the phenomenon and interpretations using different sharp-interface approaches. The original formulation is given for phreatic aquifers and for the hydrostatic situation, that is, where flow in the freshwater zone as well as in the saltwater zone is neglected. Various different extensions have been made since the first formulations. Thus it is possible to use the generalizations of the Ghijben–Herzberg relation in situations for which the simple formula was not stated originally. In that sense, it is possible to use GHE as a

160

GHIJBEN–HERZBERG EQUILIBRIUM

N Phreatic surface

le h y x

H

Sea level

Fresh Salt

Figure 2. Schematic view of flow in an unconfined aquifer near the coast, describing submarine groundwater discharge. Adapted from Reference 12.

Interface

Potential lines

principle for confined aquifers, for dynamic situations, and for a dispersed transition zone, as described in the following. When the amount of water withdrawal becomes a significant part of the entire water budget of a coastal system, the Ghijben–Herzberg formula should be questioned. Hydrostatics then has to be replaced by a quantification of the hydrological balance of the system. GENERALIZATIONS The original derivation assumes hydrostatic conditions, while the real situation is hydrodynamic in all cases. The authors were aware of that: in the report Badon Ghijben describes that under the North Sea dunes there is a relatively strong freshwater flow, recharged by infiltrating rainwater in the upper part, directed toward the sea, while there is a very weak flow of saltwater in the opposite direction in the lower part. In-between there is a transition from inflow to outflow. However, the hydrostatic state can be taken as a limiting case from which the dynamic situation deviates only slightly as long as the groundwater velocities are small. The GHE idea of an equilibrium between freshwater and saltwater can then be transferred to dynamic situations. Hubbert (8) introduced the variable of freshwater head, defined as hf =

ρ − ρf ρ h− z ρf ρf

(4)

where ρ and h denote density and head, measured at depth z. As outlined in the overview papers by Reilly and Goodman (9) and Cheng and Quazar (10), the freshwater head was used also by Muskat (11) in a description of the interface between oil and water in the subsurface. For h = 0 (stagnant saltwater) and ρ = ρs , the Ghijben–Herzberg formula (Equation 3) results for the interface position at depth z.

With the help of the freshwater head hf , descriptions of situations with an interface can be simplified. A solution for freshwater head distribution is determined first; the location of the interface is calculated on the basis of the GHE afterward. This procedure is followed in many publications concerning different situations with fluids of different densities. van der Veer (12) presents a simple formula for the one-dimensional flow in a coastal aquifer and a complex solution for flow in a vertical cross section, both based on the GHE. It is shown that distant from the shore both solutions coincide; that is, the use of the simpler solution is justified. But when the coastline is approached, the differences between both solutions increase. With the two-dimensional solution, it is possible to describe the submarine groundwater discharge into the sea. Strack (13,14) uses the GHE and freshwater head to construct analytical solutions for shallow interface flow in confined and unconfined situations. In the case of the confined situation, the piezometric head h is the relevant variable, not the water table in the confined layer. Using these solutions the regional flow in coastal aquifers toward wells can be calculated in a two-dimensional horizontal domain. The cross-sectional view, given in Fig. 3, depicting schematically the flow in the confined situation, is adapted from Strack (14). Due to pumping of freshwater in the upper part of the aquifer, the interface between saltwater and freshwater rises. Such an upconing may lead to severe problems, when the interface comes close to the well filter, as the salinity of the pumped water may increase above limits, which makes the water unsuitable for the drinking water supply, irrigation, or other purposes. According to the static GHE, the interface rise exceeds the water table decline by a factor of 40 in coastal aquifers. Recent experience in Sweden shows that the GHE overestimates the rise of the interface (15). Dynamic approaches using well-known formulas for single wells were combined with the GHE by Wang (16) and by Motz (17). In regard to saltwater upconing, the GHE today

GHIJBEN–HERZBERG EQUILIBRIUM

161

Piezometric head in lower aquifer

Sea coast

Withdrawal

z hf

Salt water

h

hs

impermeable

H

Hs

Base flow

x

is often applied in combination with finite differences or finite elements (18,19). The mentioned analytical solutions are valid for horizontal flow in an aquifer with a sharp interface, but vertical fluxes cannot be treated correctly by freshwater heads. As a further generalization of the GHE, Lusczynski (20) showed that for the general three-dimensional case in variable density flow, the gradients of freshwater head are responsible for horizontal groundwater flow, while for flow in the vertical direction the so-called environmental water head he is responsible. he is a generalization of Eq. 3: he =

ρ ρ − ρa h− z ρf ρf

(5)

where ρa denotes the average density in the column between position z and sea level. Note that for the latter extensions of the GHE the condition of a sharp interface has been dropped. But with ρa another parameter is introduced, which cannot be estimated easily without any knowledge of the mean position of the interface: remember that the principal purpose of the Ghijben–Herzberg formula was to deliver an estimate for that position. In fact, saltwater intrusion is a problem of densitydriven flow (21), in which flow and transport processes are coupled and cannot be calculated independently. Naturally, the flow in the aquifer determines the salt

Figure 3. Schematic view of flow in a confined coastal aquifer with groundwater withdrawal. Adapted from Reference 14.

distribution, but on the other hand the salinity influences the flow via density and possibly via viscosity. In order to treat such complex interactions, computer models are often used today. An overview on numerical approaches, codes, and applications, concerning saltwater intrusion, is given by Sorek and Pinder (22). Henry (23) presented a widely discussed solution for the miscible displacement in a confined aquifer, which has been used by numerous modelers as a test case. Figure 4 represents the solution in a vertical cross section, calculated with the new FEMLAB code (24) for multiphysics simulations. Depicted are streamlines extending from the left freshwater boundary to the seawater boundary on the right-hand side. On the seaside, streamlines enter in the lower part and return to the seaside after a u-turn. Along the paths salinity changes due to dispersion and diffusion. The grey pattern visualizes salinity, which is high in the lower right corner. The Henry solution and with it the penetration of the saltwater wedge and its dispersion depends on various flow and transport parameters: the ambient freshwater flow and the dispersivities—dependencies that cannot be predicted by the GHE, which is based on water densities only. CONCLUSION The history of the GHE offers some more peculiarities. According to Carlston (25), the static equilibrium condition

Figure 4. The solution for saltwater intrusion presented by Henry (23).

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for saltwater and freshwater was formulated long before the referenced authors. Du Commun (26) already envisaged a U-tube that is filled with fluids of different density on both sides. The equilibrium in that tube was characterized by different fluid levels in both ends of the tube, and the height difference could be calculated from the density difference between the fluids. Du Commun did not provide a formula (like Badon Ghijben) but illustrated the calculation by an example. The background for his derivations was a practical problem: on the ground of a distillery in New Brunswick (New Jersey) a water well had been drilled. At that time, where not much was known about groundwater, the water level in the well and its fluctuations with the tide became the subject of a newspaper dispute (25). Acknowledgment The author is grateful B. Kobisch for reprinting original figures.

BIBLIOGRAPHY 1. Drabbe, J. and Badon Ghijben, W. (1889). Nota in verband met de voorgenomen putboring nabij Amsterdam. Tijdschrift van het Koninklijk Institut voor Ingenieurs, The Hague, 8–22. 2. Herzberg, A. (1901). Die Wasserversorgung einiger Nord¨ seebader. J. Gasbeleuchtung Wasser. 44: 815–819, 842–844. 3. De Vries, J.J. (2004). From speculation to science: the founding of groundwater hydrology in the Netherlands. In: Dutch Pioneers of the Earth Sciences. R.P.W. Visser and J.L.R. Touret (Eds.). Koninklijke Nederlandse Akademie van Wetenschappen, Amsterdam, pp. 139–164. 4. Fofonoff, P. and Millard, R.C., Jr. (1983). Algorithms for computation of fundamental properties of seawater. Unesco Tech. Pap. Mar. Sci. 44: 53. 5. Bear, J., Cheng, A.H.-D., Sorek, S., Quazar, D., and Herrera, I. (Eds.). (1999). Seawater Intrusion in Coastal Aquifers—Concepts, Methods and Practices. Kluwer Academic Publ., Dordrecht, the Netherlands, pp. 163–191. 6. Custodio, E. (2002). Coastal aquifers as important hydrogeological structures. In: Groundwater and Human Development. E. Bocanegra, D. Martines, and H. Massone (Eds.). 1905–1918. 7. Kooi, H. and Groen, J. (2001). Offshore continuation of coastal groundwater systems; predictions using sharp-interface approximations and variable-density flow modelling. J. Hydrol. 246: 19–35. 8. Hubbert, M.K. (1940). The theory of ground-water motion. J. Geol. XLVIII(8): 785–938. 9. Reilly, T.E. and Goodman, A.S. (1985). Quantitative analysis of saltwater–freshwater relationships in groundwater systems—a historical perspective. J. Hydrol. 80: 125–160. 10. Cheng, A.H.-D. and Quazar, D. (1999). Analytical solutions. In: Seawater Intrusion in Coastal Aquifers—Concepts, Methods and Practices. J. Bear, A.H.-D. Cheng, S. Sorek, D. Quazar, and I. Herrera (Eds.). Kluwer Academic Publ., Dordrecht, the Netherlands, pp. 163–191. 11. Muskat, M. (1937). The Flow of Homogeneous Fluids Through Porous Media. McGraw-Hill, New York. 12. van der Veer, P. (1977). Analytical solution for steady interface flow in a coastal aquifer involving a phreatic surface with precipitation. J. Hydrol. 34: 1–11.

13. Strack, O.D.L. (1976). A single-potential solution for regional interface problems in coastal aquifers. Water Resour. Res. 12(6): 1165–1174. 14. Strack, O.D.L. (1989). Groundwater Mechanics. PrenticeHall, Englewood Cliffs, NJ. ¨ 15. Backblom, G. (2002). Experience on Grouting to Limit Inflow to Tunnels. Research and development and case studies from Sweden, Working Report 2002-18, Posiva Oy, Helsinki, Finland. 16. Wang, F.C. (1965). Approximate theory for skimming well formation in the Indus plain of West Pakistan. J. Geophys. Res. 70(20): 5055–5063. 17. Motz, L.H. (1992). Salt-water upconing in an aquifer overlain by a leaky confining bed. Groundwater 30(2): 192–198. 18. Wirojanagud, P. and Charbeneau, R.J. (1985). Saltwater upconing in unconfined aquifers. J. Hydrol. Eng. 111(3): 417–434. 19. Reilly, T.E., Frimpter, M.H., LeBlanc, D.R., and Goodman, A.S. (1987). Analysis of steady-state salt-water upconing with application at Truro well field. Groundwater 25(2). 20. Lusczynski, N.J. (1961). Head and flow of ground water of variable density. J. Geophys. Res. 66(12): 4247–4256. 21. Holzbecher, E. (1998). Modeling Density-Driven Flow in Porous Media. Springer, Heidelberg, Germany. 22. Sorek, S. and Pinder, G.F. (1999). Numerical modelling issues. In: Seawater Intrusion in Coastal Aquifers—Concepts, Methods and Practices. J. Bear, A.H.-D. Cheng, S. Sorek, D. Quazar, and I. Herrera (Eds.). Kluwer Academic Publ., Dordrecht, the Netherlands, pp. 399–461. 23. Henry, H.R. (1964). Effects of dispersion on salt water enroachment in coastal aquifers. Geol. Surv. Water-Supply Paper 1613-C. 24. FEMLAB. (2004). Version 3.1, COMSOL AB, Tegn´ergatan 23, SE-111 40, Stockholm, Sweden. 25. Carlston, C.W. (1963). An early American statement of the Badon Ghyben–Herzberg principle of static fresh-water–saltwater balance. Am. J. Sci. 261: 88–91. 26. DuCommun, J. (1828). On the causes of fresh water springs, fountains, etc. Am. J. Sci. 14: 174–176.

GROUNDWATER BALANCE C.P. KUMAR National Institute of Hydrology Roorkee, India

INTRODUCTION Rapid industrial development, urbanization, and increase in agricultural production have led to freshwater shortages in many parts of the world. In view of increasing demand of water for various purposes like agricultural, domestic, and industrial, a greater emphasis is being placed on a planned and optimal utilization of water resources. The water resources of the basins remain almost constant while the demand for water continues to increase. As a result of uneven distribution of rainfall both in time and space, the surface water resources are unevenly distributed. Also, increasing intensities of irrigation from surface water alone may result in an alarming rise

GROUNDWATER BALANCE

of the water table, creating problems of waterlogging and salinization, affecting crop growth adversely and rendering large areas unproductive, which has resulted in increased emphasis on development of groundwater resources. The simultaneous development of groundwater, especially through dug wells and shallow tubewells, will lower the water table and provide vertical drainage, which can prevent waterlogging and salinization. Areas that are already waterlogged can also be reclaimed. On the other hand, continually increased withdrawals from a groundwater reservoir in excess of replenishable recharge may result in regular lowering of the water table. In such a situation, a serious problem is created resulting in drying of shallow wells and increasing pumping head for deeper wells and tubewells, which has led to an emphasis on planned and optimal development of water resources. An appropriate strategy will be to develop water resources with planning based on conjunctive use of surface water and groundwater. For a sustainable development of water resources, it is imperative to make a quantitative estimation of the available water resources. For this, the first task would be to make a realistic assessment of the surface water and groundwater resources and then plan their use in such a way that full crop water requirements are met and neither waterlogging nor excessive lowering of the groundwater table occurs. It is necessary to maintain the groundwater reservoir in a state of dynamic equilibrium over a period of a time, and the water level fluctuations have to be kept within a particular range over the monsoon and nonmonsoon seasons. A complexity of factors, hydrogeological, hydrological, and climatological, control the groundwater occurrence and movement. The precise assessment of recharge and discharge is rather difficult, as no techniques are currently available for their direct measurements. Hence, the methods employed for groundwater resource estimation are all indirect. Groundwater, being a dynamic and replenishable resource, is generally estimated based on the component of annual recharge, which could be subjected to development by means of suitable groundwater structures. For quantification of groundwater resources, proper understanding of the behavior and characteristics of the water-bearing rock formation, known as an aquifer, is essential. An aquifer has two main functions: (a) to transit water (conduit function) and (b) to store water (storage function). The groundwater resources in unconfined aquifers can be classified as static and dynamic. The static resources can be defined as the amount of groundwater available in the permeable portion of the aquifer below the zone of water level fluctuation. The dynamic resources can be defined as the amount of groundwater available in the zone of water level fluctuation. The replenishable groundwater resource is essentially a dynamic resource that is replenished annually or periodically by precipitation, irrigation return flow, canal seepage, tank seepage, influent seepage, etc. The methodologies adopted for computing groundwater resources are generally based on the hydrologic budget techniques. The hydrologic equation for groundwater

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regime is a specialized form of water balance equation that requires quantification of the components of inflow to and outflow from a groundwater reservoir, as well as changes in storage therein. Some of these are directly measurable, few may be determined by differences between measured volumes or rates of flow of surface water, and some require indirect methods of estimation. Water balance techniques have been extensively used to make quantitative estimates of water resources and the impact of man’s activities on the hydrological cycle. The study of water balance requires the systematic presentation of data on the water supply and its use within a given study area for a specific period. The water balance of an area is defined by the hydrologic equation, which is basically a statement of the law of conservation of mass as applied to the hydrological cycle. With the water balance approach, it is possible to quantitatively evaluate individual contribution of sources of water in the system over different time periods and to establish the degree of variation in water regime because of changes in components of the system. A basinwise approach yields the best results where the groundwater basin can be characterized by prominent drainages. A thorough study of the topography, geology, and aquifer conditions should be employed. The limit of the groundwater basin is controlled not only by topography but also by the disposition, structure and permeability of rocks and the configuration of the water table. Generally, in igneous and metamorphic rocks, the surface water and groundwater basins are coincident for all practical purposes, but marked differences may be encountered in stratified sedimentary formations. Therefore, the study area for groundwater balance study is preferably taken as a doab, which is bounded on two sides by two streams and on the other two sides by other aquifers or extension of the same aquifer. Once the study area is identified, comprehensive studies can be undertaken to estimate for selected period of time, the input and output of water, and change in storage to draw up the water balance of the basin. The estimation of groundwater balance of a region requires quantification of all individual inflows to or outflows from a groundwater system and change in groundwater storage over a given time period. The basic concept of water balance is: Input to the system − outflow from the system = change in storage of the system (over a period of time) The general methodology of computing groundwater balance consists of the following: • Identification of significant components, • Evaluating and quantifying individual components, and • Presentation in the form of water balance equation. The groundwater balance study of an area may serve the following purposes: • As a check on whether all flow components involved in the system have been quantitatively accounted for,

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and what components have the greatest bearing on the problem under study. • To calculate one unknown component of the groundwater balance equation, provided all other components are quantitatively known with sufficient accuracy. • As a model of the hydrological processes under study, which can be used to predict the effect that changes imposed on certain components will have on the other components of groundwater system. GROUNDWATER BALANCE EQUATION Considering the various inflow and outflow components in a given study area, the groundwater balance equation can be written as: Rr + Rc + Ri + Rt + Si + Ig = Et + Tp + Se + Og + S (1) where Rr = recharge from rainfall Rc = recharge from canal seepage Ri = recharge from field irrigation Rt = recharge from tanks Si = influent seepage from rivers Ig = inflow from other basins Et = evapotranspiration from groundwater Tp = draft from groundwater Se = effluent seepage to rivers Og = outflow to other basins S = change in groundwater storage Preferably, all elements of the groundwater balance equation should be computed using independent methods. However, it is not always possible to compute all individual components of the groundwater balance equation separately. Sometimes, depending on the problem, some components can be lumped and account only for their net value in the equation. Computations of various components usually involve errors, because of shortcomings in the estimation techniques. The groundwater balance equation, therefore, generally does not balance, even if all its components are computed by independent methods. The resultant discrepancy in groundwater balance is defined as a residual term in the balance equation, which includes errors in the quantitative determination of various components as well as values of the components that have not been accounted in the equation. The water balance may be computed for any time interval. The complexity of the computation of the water balance tends to increase with increase in area. This is because of a related increase in the technical difficulty of accurately computing the numerous important water balance components. DATA REQUIREMENTS FOR A GROUNDWATER BALANCE STUDY For carrying out a groundwater balance study, the following data may be required over a given time period:

Rainfall Data Monthly rainfall data of sufficient number of rainguage stations lying within or around the study area, along with their locations, should be available. Land Use Data and Cropping Patterns Land use data are required for estimating the evapotranspiration losses from the water table through forested area. Cropping pattern data are necessary for estimating the spatial and temporal distributions of groundwater withdrawals, if required. Monthly pan evaporation rates should also be available at a few locations for estimation of consumptive use requirements of different crops. River Data Monthly river stage and discharge data along with river cross sections are required at a few locations for estimating the river-aquifer interflows. Canal Data Monthly water releases into the canal and its distributories along with running days during each month are required. To account for the seepage losses through the canal system, the seepage loss test data are required in different canal reaches and distributories. Tank Data Monthly tank gauges and water releases should be available. In addition, depth vs. area and depth vs. capacity curves should also be available for computing the evaporation and seepage losses from tanks. Field test data are required for computing infiltration capacity to be used to evaluate the recharge from depression storage. Water Table Data Monthly water table data (or at least premonsoon and postmonsoon data) from a sufficient number of welldistributed observation wells along with their locations are required. The available data should comprise reduced level (RL) of water table and depth to water table. Groundwater Draft For estimating groundwater withdrawals, the number of each type of well operating in the area, their corresponding running hours each month, and discharge are required. If a complete inventory of wells is not available, then this can be obtained by carrying out sample surveys. Aquifer Parameters Data regarding the storage coefficient and transmissivity are required at a sufficient number of locations in the study area. ESTIMATION OF GROUNDWATER BALANCE COMPONENTS The various inflow/outflow components of the groundwater balance equation may be estimated through appropriate field experiments or other methods, as discussed below.

GROUNDWATER BALANCE

Recharge from Rainfall (Rr ) Rainfall is the major source of recharge to groundwater. Part of the rain water that falls on the ground is infiltrated into the soil. A part of this infiltrated water is used in filling the soil moisture deficiency, whereas the remaining portion percolates down to reach the water table, which is termed as rainfall recharge to the aquifer. The amount of rainfall recharge depends on various hydrometeorological and topographic factors, soil characteristics, and depth to water table. The methods for estimation of rainfall recharge involve the empirical relationships established between recharge and rainfall developed for different regions, groundwater balance approach, and soil moisture data-based methods. Empirical Relationships. Several empirical formulae have been worked out for various regions on the basis of detailed studies. For example, Kumar and Seethapathi (1) conducted a detailed seasonal groundwater balance study in Upper Ganga Canal command area (India) for the period 1972–1973 to 1983–1984 to determine groundwater recharge from rainfall. It was observed that as the rainfall increases, the quantity of recharge also increases, but the increase is not linearly proportional. The recharge coefficient (based on the rainfall in monsoon season) was found to vary between 0.05 to 0.19 for the study area. The following empirical relationship was derived by fitting the estimated values of rainfall recharge and the corresponding values of rainfall in the monsoon season through the nonlinear regression technique. Rr = 0.63 (P − 15.28)0.76

(2)

where Rr = Groundwater recharge from rainfall in monsoon season (inch) P = Mean rainfall in monsoon season (inch) The relative errors (%) in the estimation of rainfall recharge computed from the proposed empirical relationship was compared with groundwater balance study. In almost every years, the relative error was found to be less than 8%. Therefore, Eq. (2) can conveniently be used for better and quick assessment of natural groundwater recharge in Upper Ganga Canal command area. It is to be noted that the relationships, tentatively proposed for specific hydrogeological conditions, have to be examined and established or suitably altered for application to other areas. If adequate data of groundwater levels are not available, rainfall recharge may be estimated using the rainfall infiltration method. The same recharge factor may be used for both monsoon and nonmonsoon rainfall, with the condition that the recharge because of nonmonsoon rainfall may be taken as zero, if the rainfall during nonmonsoon season is less than 10% of annual rainfall. An additional 2% of rainfall recharge factor may be used in areas where watershed development with associated soil conservation measures are implemented. This additional factor is separate from contribution because of water conservation structures, such as check dams, nalla bunds, percolation tanks, etc., for which the norms are defined separately.

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Groundwater Balance Approach. In this method, all components of the groundwater balance Eq. (1), except the rainfall recharge, are estimated individually. The algebraic sum of all input and output components is equated to the change in groundwater storage, as reflected by the water table fluctuation, which in turn yields the single unknown in the equation, namely, the rainfall recharge. A prerequisite for successful application of this technique is the availability of very extensive and accurate hydrological and meteorological data. The groundwater balance approach is valid for the areas where the year can be divided into monsoon and nonmonsoon seasons, with the bulk of rainfall occurring in former. Groundwater balance study for monsoon and nonmonsoon periods is carried out separately. The former yields an estimate of recharge coefficient and the latter determines the degree of accuracy with which the components of water balance equation have been estimated. Alternatively, the average specific yield in the zone of fluctuation can be determined from a groundwater balance study for the nonmonsoon period, and using this specific yield, the recharge because of rainfall can be determined using the groundwater balance components for the monsoon period. Soil Moisture Data-Based Methods. Soil moisture databased methods are the lumped and distributed model and the nuclear methods. In the lumped model, the variation of soil moisture content in the vertical direction is ignored and any effective input into the soil is assumed to increase the soil moisture content uniformly. Recharge is calculated as the remainder when losses, identified in the form of runoff and evapotranspiration, have been deducted from the precipitation with proper accounting of soil moisture deficit. In the distributed model, variation of soil moisture content in the vertical direction is accounted for, and the method involves the numerical solution of partial differential equation (Richards equation) governing onedimensional flow through unsaturated medium, with appropriate initial and boundary conditions.

Soil Water Balance Method. Water balance models were developed in the 1940s by Thornthwaite (2) and revised by Thornthwaite and Mather (3). The method is essentially a bookkeeping procedure that estimates the balance between the inflow and outflow of water. When applying this method to estimate the recharge for a catchment area, the calculation should be repeated for areas with different precipitation, evapotranspiration, crop type, and soil type. The soil water balance method is of limited practical value, because evapotranspiration is not directly measurable. Moreover, storage of moisture in the unsaturated zone and the rates of infiltration along the various possible routes to the aquifer form important and uncertain factors. Another aspect that deserves attention is the depth of the root zone, which may vary in semiarid regions between 1 and 30 meters. Results from this model are of very limited value without calibration and validation, because of the substantial uncertainty in input data. Nuclear Methods. Nuclear techniques can be used for the determination of recharge by measuring the travel of

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moisture through a soil column. The technique is based on the existence of a linear relation between neutron count rate and moisture content (% by volume) for the range of moisture contents generally occurring in the unsaturated soil zone. The mixture of Beryllium (Be) and Radium (Ra) is taken as the source of neutrons. Another method is the gamma ray transmission method based on the attenuation of gamma rays in a medium through which it passes. The extent of attenuation is closely linked with moisture content of the soil medium.

Seepage refers to the process of water movement from a canal into and through the bed and wall material. Seepage losses from irrigation canals often constitute a significant part of the total recharge to groundwater system. Hence, it is important to properly estimate these losses for recharge assessment to groundwater system. Recharge by seepage from canals depends on the size and cross section of the canal, depth of flow, characteristics of soils in the bed and sides, and location as well as level of drains on either side of the canal. A number of empirical formulae and formulae based on theoretical considerations have been proposed to estimate the seepage losses from canals. Recharge from canals that are in direct hydraulic connection with a phreatic aquifer, underlaid by a horizontal impermeable layer at shallow depth, can be determined by Darcy’s equation, provided the flow satisfies Dupuit assumptions. hs − hl A L

(3)

where, hs and hl are water-level elevations above the impermeable base, at the canal, respectively, and at distance L from it. For calculating the area of flow cross section, the average of the saturated thickness (hs + hl )/2 is taken. The crux of computation of seepage depends on correct assessment of the hydraulic conductivity, K. Knowing the percentage of sand, silt, and clay, the hydraulic conductivity of undisturbed soil can be approximately determined using the soil classification triangle showing relation of hydraulic conductivity to texture for undisturbed sample (4). A number of investigations have been carried out to study the seepage losses from canals. United States Bureau of Reclamation (USBR) recommended the channel losses based on the channel bed material as given below:

Material Clay and clay loam: Sandy loam: Sandy and gravely soil: Concrete lining:

Rc− max = K(B + 2D) (in case of deeper water table)

Recharge from Canal Seepage (Rc )

Rc = K

Specific results from case studies may be used, if available. The above norms take into consideration the type of soil in which the canal runs while computing seepage. However, the actual seepage will also be controlled by the width of canal (B), depth of flow (D), hydraulic conductivity of the bed material (K), and depth to water table. Knowing the values of B and D, the range of seepage losses (Rc− max and Rc− min ) from the canal may be obtained as

Seepage Losses (Cumec Per Million Square Meter of Wetted Area) 1.50 2.40 8.03 1.20

These values are valid if the water table is relatively deep. In shallow water table and waterlogged areas, the recharge from canal seepage may be suitably reduced.

(4a)

Rc− min = K(B − 2D) (in case of water table at the level of channel bed)

(4b)

However, the various guidelines for estimating losses in the canal system are only approximate. The seepage losses may best be estimated by conducting actual tests in the field. The methods most commonly adopted are: Inflow-Outflow Method. In this method, the water that flows into and out of the section of canal under study is measured using current meter or Parshall flume method. The difference between the quantities of water flowing into and out of the canal reach is attributed to seepage. This method is advantageous when seepage losses are to be measured in long canal reaches with few diversions. Ponding Method. In this method, bunds are constructed in the canal at two locations, one upstream and the other downstream of the reach of canal with water filled in it. The total change in storage in the reach is measured over a period of time by measuring the rate of drop of water surface elevation in the canal reach. Alternatively, water may be added to maintain a constant water surface elevation. In this case, the volume of water added is measured along with the elapsed time to compute the rate of seepage loss. The ponding method provides an accurate means of measuring seepage losses and is especially suitable when they are small (e.g., in lined canals). Seepage Meter Method. The seepage meter is a modified version of permeameter developed for use under water. Various types of seepage meters have been developed. The two most important are seepage meter with submerged flexible water bag and falling head seepage meter. Seepage meters are suitable for measuring local seepage rates in canals or ponds and used only in unlined or earth-lined canals. They are quickly and easily installed and give reasonably satisfactory results for the conditions at the test site, but it is difficult to obtain accurate results when seepage losses are low. The total losses from the canal system generally consist of the evaporation losses (Ec ) and the seepage losses (Rc ). The evaporation losses are generally 10–15% of the total losses. Thus, the Rc value is 85–90% of the losses from the canal system. Recharge from Field Irrigation (Ri ) Water requirements of crops are met, in parts, by rainfall, contribution of moisture from the soil profile, and applied

GROUNDWATER BALANCE

irrigation water. A part of the water applied to irrigated field crops is lost in consumptive use and the balance infiltrates to recharge the groundwater. The process of re-entry of a part of the water used for irrigation is called return flow. Percolation from applied irrigation water, derived both from surface water and groundwater sources, constitutes one of the major components of groundwater recharge. The irrigation return flow depends on the soil type, irrigation practice, and type of crop. Therefore, irrigation return flows are site specific and will vary from one region to another. The recharge because of irrigation return flow may be estimated, based on the source of irrigation (groundwater or surface water), the type of crop (paddy, nonpaddy), and the depth of water table below ground surface. For surface water, the recharge is to be estimated based on water released at the outlet from the canal/distribution system. For groundwater, the recharge is to be estimated based on gross draft. Where continuous supply is used instead of rotational supply, an additional recharge of 5% of application may be used. Specific results from case studies may be used, if available. For a correct assessment of the quantum of recharge by applied irrigation, studies are required to be carried out on experimental plots under different crops in different seasonal conditions. The method of estimation comprises application of the water balance equation involving input and output of water in experimental fields. Recharge from Tanks (Rt ) Studies have indicated that seepage from tanks varies from 9–20% of their live storage capacity. However, as data on live storage capacity of large number of tanks may not be available, seepage from the tanks may be taken as 44–60 cm per year over the total water spread, taking into account the agroclimatic conditions in the area. The seepage from percolation tanks is higher and may be taken as 50% of its gross storage. In the case of seepage from ponds and lakes, the norms as applied to tanks may be taken. Groundwater Resource Estimation Committee (5) has recommended that, based on the average area of water spread, the recharge from storage tanks and ponds may be taken as 1.4 mm/day for the period in which the tank has water. If data on the average area of water spread is not available, 60% of the maximum water spread area may be used instead of the average area of water spread. In the case of percolation tanks, recharge may be taken as 50% of gross storage, considering the number of fillings, with half of this recharge occurring in monsoon season and the balance in nonmonsoon season. Recharge because of check dams and nala bunds may be taken as 50% of gross storage (assuming annual desilting maintenance exists), with half of this recharge occurring in the monsoon season and the balance in the nonmonsoon season. Influent and Effluent Seepage (Si & Se ) The river-aquifer interaction depends on the transmissivity of the aquifer system and the gradient of the water table in respect to the river stage. Depending on the water level in the river and in the aquifer (in the vicinity of

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river), the river may recharge the aquifer (influent) or the aquifer may contribute to the river flow (effluent). The effluent or influent character of the river may vary from season to season and from reach to reach. The seepage from/to the river can be determined by dividing the river reach into small sub-reaches and observing the discharges at the two ends of the sub-reach along with the discharges of its tributaries and diversions, if any. The discharge at the downstream end is expressed as: Qd .t = Qu .t + Qg .t + Qt .t − Qo .t − E.t ± Srb (5) where Qd = discharge at the downstream section Qu = discharge at the upstream section Qg = groundwater contribution (unknown quantity; −ve computed value indicates influent conditions) Qt = discharge of tributaries Qo = discharge diverted from the river E = rate of evaporation from river water surface and flood plain (for extensive bodies of surface water and for long time periods, evaporation from open water surfaces cannot be neglected) Srb = change in bank storage (+ for decrease and − for increase) t = time period The change in bank storage can be determined by monitoring the water table along the cross section normal to the river. Thus, using the above equation, seepage from/to the river over a certain period of time, t, can be computed. However, this would be the contribution from aquifers on both sides of the stream. The contribution from each side can be separated by the following method: Contribution from left bank =

IL TL · Qg (6a) IL TL + IR TR

Contribution from right bank =

IR TR · Qg (6b) IL TL + IR TR

where IL and TL are gradient and transmissivity, respectively, on the left side and IR and TR are those on the right. Inflow from and Outflow to Other Basins (Ig and Og ) For the estimation of groundwater inflow/outflow from/to other basins, regional water table contour maps are drawn based on the observed water level data from wells located within and outside the study area. The flows into and out of a region are governed mainly by the hydraulic gradient and transmissivity of the aquifer. The gradient can be determined by taking the slope of the water table normal to water table contour. The length of the section, across which groundwater inflow/outflow occurs, is determined from contour maps, the length being measured parallel to the contour. The inflow/outflow is determined as follows: Ig or Og =

L 

T I L

(7)

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where T is the transmissivity and I is the hydraulic gradient averaged over a length, L, of contour line.

distribution point and multiplying the average draft value with the number of units of electricity consumed, the draft at each point can be computed for every month.

Evapotranspiration from Groundwater (Et ) Evapotranspiration is the combined process of transpiration from vegetation and evaporation from both soil and free water surfaces. Potential evapotranspiration is the maximum loss of water through evapotranspiration. Evapotranspiration from groundwater occurs in waterlogged areas or in forested areas with roots extending to the water table. From the land use data, area under forests is available whereas the waterlogged areas may be demarcated from depth to water table maps. The potential evapotranspiration from such areas can be computed using standard methods. Depth to water table maps may be prepared based on well inventory data to bring into focus the extensiveness of shallow water table areas. During well inventory, investigation should be specifically oriented toward accurately delineating water table depth for depths less than 2 meters. The evapotranspiration can be estimated based on the following equations: Et = PEt ∗ A

if h > hs

(8a)

Et = 0

if h < (hs − d)

(8b)

Et = PEt ∗ A(h − (hs − d))/d

if (hs − d) ≤ h ≤ hs (8c)

where Et = evapotranspiration in volume of water per unit time[L3 T −1 ] PEt = maximum rate of evapotranspiration in volume of water per unit area per unit time [L3 L−2 T −1 ] A = surface area [L2 ] h = water table elevation [L] hs = water table elevation at which the evapotranspiration loss reaches the maximum value d = extinction depth; when the distance between hs and h exceeds d, evapotranspiration from groundwater ceases [L] Draft from Groundwater (Tp ) Draft is the amount of water lifted from the aquifer by means of various lifting devices. To estimate groundwater draft, an inventory of wells and a sample survey of groundwater draft from various types of wells (state tubewells, private tubewells, and open wells) are required. For state tubewells, information about their number, running hours per day, discharge, and number of days of operation in a season is generally available in the concerned departments. To compute the draft from private tubewells, pumping sets, rahats, etc., sample surveys have to be conducted regarding their number, discharge, and withdrawals over the season. In areas where wells are energized, the draft may be computed using power consumption data. By conducting tests on wells, the average draft per unit of electricity consumed can be determined for different ranges in depth to water levels. By noting the depth to water level at each

Change in Groundwater Storage (
S ) To estimate the change in groundwater storage, the water levels are observed through a network of observation wells spread over the area. The water levels are highest immediately after monsoon, in the month of October or November, and lowest just before rainfall, in the month of May or June. During the monsoon season, the recharge is more than the extraction; therefore, the change in groundwater storage between the beginning and end of monsoon season indicates the total volume of water added to the groundwater reservoir, whereas the change in groundwater storage between the beginning and end of the nonmonsoon season indicates the total quantity of water withdrawn from groundwater storage. The change in storage (S) is computed as follows: S =



h A Sy

(9)

where h = change in water table elevation during the given time period A = area influenced by the well Sy = specific yield Groundwater Resource Estimation Committee (5) recommended that the size of the watershed as a hydrological unit could be of about 100 to 300 sq. km area, and there should be at least three spatially well-distributed observation wells in the unit, or one observation well per 100 sq. km, whichever is more. However, as per IILRI (6), the following specification may serve as a rough guide:

Size of the Area (ha) 100 1,000 10,000 1,00,000

Number of Observation Points

Number of Observation Points per 100 Hectares

20 40 100 300

20 4 1 0.3

The specific yield may be computed from pumping tests. The values of specific yield in the zone of fluctuation of water table in different parts of the basin can also be approximately determined from the soil classification triangle showing relation between particle size and specific yield (7). ESTABLISHMENT OF RECHARGE COEFFICIENT Groundwater balance study is a convenient way of establishing the rainfall recharge coefficient, as well as to cross check the accuracy of the various prevalent methods for the estimation of groundwater losses and recharge from other sources. The steps to be followed are:

HYDRAULIC HEAD

1. Divide the year into monsoon and nonmonsoon periods. 2. Estimate all the components of the water balance equation other than rainfall recharge for monsoon period using the available hydrological and meteorological information and employing the prevalent methods for estimation. 3. Substitute these estimates in the water balance equation and thus calculate the rainfall recharge and, hence, recharge coefficient (recharge/rainfall ratio). Compare this estimate with those given by various empirical relations valid for the area of study. 4. For nonmonsoon season, estimate all the components of water balance equation including the rainfall recharge, which is calculated using recharge coefficient value obtained through the water balance of monsoon period. The rainfall recharge (Rr ) will be of very small order in this case. A close balance between the left and right sides of the equation will indicate that the net recharge from all the sources of recharge and discharge has been quantified with a good degree of accuracy. By quantifying all the inflow/outflow components of a groundwater system, one can determine which particular component has the most significant effect on the groundwater flow regime. Alternatively, a groundwater balance study may be used to compute one unknown component (e.g., the rainfall recharge) of the groundwater balance equation when all other components are known. The balance study may also serve as a model of the area under study, whereby the effect of change in one component can be used to predict the effect of changes in other components of the groundwater system. In this manner, the study of groundwater balance has a significant role in planning a rational groundwater development of a region. CONCLUDING REMARKS • Water balance approach, essentially a lumped model study, is a viable method of establishing the rainfall recharge coefficient and for evaluating the methods adopted for the quantification of discharge and recharge from other sources. For proper assessment of potential, present use, and additional exploitability of water resources at optimal level, a water balance study is necessary. • Groundwater exploitation should be such that protection from depletion is provided, protection from pollution is provided, negative ecological effects are reduced to a minimum, and economic efficiency of exploitation is attained. Determination of exploitable resources should be based on hydrological investigations. These investigations logically necessitate use of a mathematical model of groundwater system for analyzing and solving the problems. The study of water balance is a prerequisite for groundwater modeling.

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• A need exists for studying unsaturated and saturated flow through weathered and fractured rocks for finding the recharge components from rainfall and from percolation tanks in hard rock groundwater basins. The irrigation return flow under different soils, crops, and irrigation practices needs to be quantified. Assessment of groundwater quality in many groundwater basins is a task yet to be performed. A hydrological database for groundwater assessment should be established. Also, user-friendly software should be developed for quick assessment of regional groundwater resources. • Nonconventional methods for utilization of water, such as through interbasin transfers, artificial recharge of groundwater, and desalination of brackish or seawater as well as traditional water conservation practices like rainwater harvesting, including rooftop rainwater harvesting, need to be practiced to further increase the usable water resources. BIBLIOGRAPHY 1. Kumar, C.P. and Seethapathi, P.V. (2002). Assessment of natural groundwater recharge in Upper Ganga Canal command area. J. Appl. Hydrol. Assoc. Hydrologists of India XV(4): 13–20. 2. Thornthwaite, C.W. (1948). An approach towards a rational classification of climate. Geogr. Rev. 38(1): 55–94. 3. Thornthwaite, C.W. and Mather, J.W. (1955). The water balance. Publ. Climatol. Lab. Climatol. Drexel Inst. Technol. 8(1): 1–104. 4. Johnson, A.I. (1963). Application of Laboratory Permeability Data, Open File Report. USGS, Water Resources Division, Denver, CO, p. 34. 5. Groundwater Resource Estimation Methodology. (1997). Report of the Groundwater Resource Estimation Committee. Ministry of Water Resources, Government of India, New Delhi, India. 6. IILRI. (1974). Drainage Principles and Applications, Survey and Investigation. Publication 16, III. 7. Johnson, A.I. (1967). Specific Yield—Compilation of Specific Yields for Various Materials, Water Supply Paper, USGS. Denver, CO, p. 74.

HYDRAULIC HEAD MATTHEW M. ULIANA Texas State University—San Marcos San Marcos, Texas

Hydraulic head is defined as the fluid energy per unit weight at a given point in a fluid system (like a pipe filled with flowing water or an aquifer). In simple, everyday terms, we can conceptualize (and measure) hydraulic head as an elevation—more specifically, the elevation to which water rises in a manometer, in a pressurized water pipe or in a piezometer. Water elevations is a manifestation of the fluid energy in a groundwater system, and measuring a water elevation

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in a piezometer is really a measurement of the energy in the fluid at a certain point in a groundwater system. WHAT DO WE MEAN BY ‘‘ENERGY’’? Everything in the universe has some amount of energy associated with it, and that energy is present in various forms. Energy drives every natural process, and the key to understanding physical processes is understanding how energy is distributed in a system. There are two basic types of energy: Potential Energy: energy stored in a piece of matter or at a point in a system, generally associated with position or with the thermodynamics of the system (elevation, pressure, chemical, thermal). Kinetic: energy of motion (velocity). At every point in an aquifer, the fluid possesses some total amount of energy. That energy is the sum of all the potential and kinetic energy in the fluid associated with the velocity of the fluid, pressure of the fluid, temperature of the fluid, chemical bonds in the fluid, etc. As previously stated, the fluid energy at a point in an aquifer manifests itself as the water level in a piezometer. So we could also say that the water level, or hydraulic head, represents the total energy in the aquifer at a given point, and we can use the various energy components of the hydraulic head (elevation, pressure, velocity, etc.) to understand the driving forces behind fluid motion in the subsurface. THE BERNOULLI EQUATION As previously stated, fluid energies (and, subsequently, water levels) vary from one point in an aquifer to the next (Figs. 1 and 2). Now, if we recall the first and second laws of thermodynamics: First law of thermodynamics: Energy is conserved. Energy added − energy subtracted = change in total energy

Figure 2. Cross section of aquifer showing hydraulic heads in three wells.

We could also express the first law in terms of the difference in energy at two points in a dynamic system: Total energy(atpoint1) + energy added/lost(betweenpoint1and2) = total energy(atpoint2) Second law of thermodynamics: Closed systems move toward increasing entropy. — In a dynamic system like an aquifer, water will move from a point of higher energy (i.e., lower entropy) to a point of lower energy (higher entropy). — In other words, groundwater moves in the direction of decreasing hydraulic head. The Bernoulli equation, which describes the total energy of a fluid at all positions along a flow path in a closed system, is basically an expression of the first and second laws of thermodynamics: z1 + where

v2 v2 p1 p2 + 1 + Ha = z2 + + 2 + HL + HE ρw g 2g ρw g 2g

(1)

z = elevation (L) p = pressure (M·L−1 ·t−2 ) ρw = fluid density (M·L−3 ) g = gravitational acceleration (L·t−2 ) v = velocity (L·t−1 ) Ha = heat energy added (L) HL = mechanical energy lost (L) HE = heat energy extracted (L)

and the numerical subscripts represent two different positions along the flow path. Let’s assume that we don’t add or subtract any heat energy from the system and the only change is the loss of mechanical energy from one point to the next; then we can rewrite the equation as z1 +

Figure 1. Pipe of flowing water with manometers showing the loss of head along the flow path.

v2 v2 p1 p2 + 1 + I1 = z2 + + 2 + I2 ρw g 2g ρw g 2g

(2)

where I is the internal energy at each point (i.e., the rest of the potential energy not described by the other terms) and I2 I1 is equal to HL in Eq. 1. We can understand this equation by considering that each term in the equation represents a specific component

HYDRAULIC HEAD

of energy with units of length. The terms of the equation represent, (respectively), — — — —

elevation pressure kinetic (velocity) other internal energies (thermal, chemical)

This equation describes the change in energy from one point along a flow path to the next point. This equation expresses all the components of energy in the same units (i.e., length), so we can use it to compare the relative magnitude of the individual components. When we do that, we see that for most groundwater situations, the internal and kinetic components of the total energy are so small that we can ignore them. The result is that we can describe fluid energy in a groundwater system by only the elevation and pressure components. HYDRAULIC HEAD AND HYDRAULIC POTENTIAL If we assume that we can ignore velocity and internal energy components when dealing with groundwater, we can drop all that out of the equation and express the fluid energy as the sum of the elevation and pressure components. That sum is what we call hydraulic head; in physical terms, it is the fluid energy per unit weight, and in mathematical terms, it is h = z + p/ρw g where

(3)

From Fig. 3b, we see that the elevation head is the height of the screened interval above the datum. Keep in mind that the datum is arbitrarily chosen—if we wanted, we could choose the bottom of the well as the datum, and the elevation head would be zero. However, it is important to realize that the head is really only important to us when we are looking at multiple wells in the same aquifer, and we need to have a constant datum for all those wells make the comparison meaningful. Sea level is usually chosen, but it is not the necessary datum. We also see that the pressure head is the length of the column of water in the well above the screened interval. Keep in mind that pressure and pressure head are two different things; the pressure at the screened interval is the force per unit area of the column of fluid above that point; the pressure head is the pressure divided by the product of the density of the fluid and gravitational acceleration and is manifest as the length of the column of water above the screen. If we know the density of the fluid and the length of the column of water in the well, we can calculate the pressure at the well screen. Finally, we see that the total head is just the sum of the other two heads, or more generally, the height of the water level in the well above the datum, this brings us full circle to the concept of water levels reflecting fluid energy. DISTRIBUTION OF HEADS (I.E., FLUID ENERGY) IN A 3-D AQUIFER An aquifer is a dynamic system of flowing water. The fluid energy varies throughout the system and is different from one point to the next. The result is

h = hydraulic head (L) z = elevation (L) p = pressure (M·L−1 ·t−2 ) ρw = fluid density (M·L−3 ) g = gravitational acceleration (L·t−2 )

— different water levels throughout the aquifer

If we multiply both sides of the equation by the gravitational constant, g, we get a quantity called hydraulic potential (), which is the fluid energy per unit mass, or (4)  = gz + p/ρw such that  = gh

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(5)

The hydraulic potential is simply a way of expressing the same fluid energy, so that it is independent of gravity. PHYSICAL DESCRIPTION OF THE COMPONENTS OF HEAD From Eq. 3, we can see that hydraulic head is the sum of the elevation component (or elevation head) and the pressure component (or pressure head). The physical meaning of pressure head and elevation head are defined and described in Fig. 3a–c. These figures show an idealized cross section of a piezometer that has a screened interval at the bottom. Figure 3a shows the depths and elevations measured in the field (relative to some datum, like mean sea level). Figure 3b shows the pressure, elevation, and total heads, and Fig. 3c is a combination of the other two figures.

— hydraulic gradients — movement of water driven by those gradients Head is a scalar quantity; it is measured at a point, and it has a single magnitude that doesn’t vary with respect to direction. So, when we talk about heads, from a theoretical standpoint, we are talking about the energy at an infinitesimal point in an aquifer. However, that aquifer exists in three dimensions, contains an infinite number of points, and the head varies from point to point. The trends in that variation control the directions and magnitudes of flow. We call this variation the hydraulic gradient. The hydraulic gradient is defined as the change in hydraulic head over the change in length, and it is directly analogous to other physical gradients (topographic slope, thermal gradient, concentration gradients, etc.) Hydraulic heads exist in three dimensions, so hydraulic gradients are not necessarily horizontal. In reality, most groundwater flow is generally in a horizontal direction, and it is often a realistic assumption to ignore fluid movement in a vertical dimension. However, there are many situations where vertical gradients (and, subsequently, vertical flow) are significant.

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THE ROLE OF HEAT IN GROUNDWATER SYSTEMS (a)

(b) Piezometer

Piezometer Ground surface

Water depth

Ground surface

Pressure head Piezometer depth Total (or hydraulic) head

Surface elevation

Elevation head

Datum (mean sea level)

Datum (mean sea level) (c)

Piezometer Water depth

Piezometer depth

Surface elevation

Ground surface

Pressure head

Total (or hydraulic) head

Elevation head

Datum (mean sea level)

Figure 3. Hypothetical well showing depths, elevations, and heads.

THE ROLE OF HEAT IN GROUNDWATER SYSTEMS JAMES A. JACOBS Environmental Bio-Systems, Inc. Mill Valley, California

DAVID B. VANCE ARCADIS G&M, Inc. Midland, Texas

The role of heat in the subsurface includes the study of the overall transport of heat in saturated groundwater systems. Once site specific dynamics of heat transport are understood the effects of temperature on physical, chemical, and biological processes that govern the transport and degradation of contaminants in an aquifer can be evaluated. FUNDAMENTALS Of the many parameters that are monitored in association with the assessment and remediation of contaminated

groundwater, temperature is typically regarded as a background condition that must be accepted for a given site. In the northern tier of states, it is customary that in the winter months remediation processes will be slowed or stopped depending on the depth of groundwater. DEFINITIONS Heat is an added or external energy that causes a rise in temperature, physical expansion, evaporation or other physical change. The rise or drop of temperature in an aquifer is the result of the transport of heat into or away from the aquifer. The key measure of the relationship between heat flow and resulting temperature is heat capacity, Cp. Heat capacity is used as specific heat in English units, which is the number of BTUs required to raise one pound of aquifer matrix by one degree Fahrenheit. There are two primary materials through which heat is transported in an aquifer, the mineral matrix and the entrained groundwater. Density of the mineral matrix and water, porosity, and heat capacity combine to govern

THE ROLE OF HEAT IN GROUNDWATER SYSTEMS

the temperature response of an aquifer to heat flow. The specific heat Cp (as BTU/lb ◦ F), dry density (rho, as lbs/ft3 ) and percent porosity (phi) of some soils and minerals are as follows:

Specific Heat Water Clay Sand Granite Limestone Organic Fraction

CpDensity (rho) % Porosity lbs/ft3 (phi)

1.00 0.27 0.19 0.20 0.22 0.30

62.4 65.0 110.0 165.0 155 55

NA 75 35 1 4 65

The thermal energy capacity per unit volume per degree of temperature change is called the aquifer thermal capacity (q). With the above information, it is possible to calculate the thermal capacity (q) of an aquifer: q = (Cp × rho)rock (1 − phi) + (Cp × rho)water × phi Assuming a sandy aquifer with a specific heat of 0.19 for the mineral matrix, a density of 110 lb/ft3 , and a porosity of 0.35, the specific thermal capacity per cubic foot of aquifer is q = (0.19 × 110)rock × (1 − 0.35) + (1.0 × 62.4)water × 0.35 q = 13.59rock + 21.84water = 35.4 BTU per cubic foot per 1 degree Fahrenheit EXAMPLE Given a volume in an aquifer 100 feet by 100 feet by 20 feet in depth, it would require 7,080,000 Btu to raise the temperature 1 ◦ F. Converting Btu to kilowatt hours gives 2073 KW hours. At 6 cents per Kilowatt, that is equal to $125 in electrical cost. That estimate does not take into account the efficiency of the heat generation system, transfer of that heat into the aquifer, and effects of heat transport away from the zone undergoing heating. Three physical processes are responsible for heat transfer in groundwater systems: 1. Conduction—heat flows from hotter regions to cooler regions through the molecular transfer of kinetic energy. 2. Convection—heat is transported along with overall mass transport of groundwater. 3. Radiation—the transference of heat through space via electromagnetic radiation. The dominant mechanisms in groundwater systems are convection and conduction. Thermal conductivity (k) is the amount of heat that will flow through a unit area in a unit time with a unit temperature difference. Under saturated conditions, the value of thermal conductivity is primarily governed

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by the texture, degree of compaction, and mineralogical composition of the aquifer matrix. Thermal conductivity decreases with the reduction of particle size. The units of thermal conductivity are Btu per hour per foot per degree Fahrenheit. Typical values are as follows: Water Air Wet Sand Dry Sand Wet Clay Dry Clay

0.346 0.0145 0.95 0.157 0.87 0.138

Notice that the thermal conductivity of water is about 23 times that of air. In the past, the use of heat as an augmentation to remediation has been applied to the vadose zone through the application of steam. For heat transport efficiency, that is probably not the most efficient application of a subsurface heating process. Heat flow through a mass (qx) is calculated by the equation: qx = Ak(dT/dX) where A is the cross-sectional area normal to the heat flow, k is the thermal conductivity, T is temperature in Fahrenheit, X is the distance over which the heat must flow. The solution to the above function requires substitution of an equation describing the system geometry for the dX term followed by integration. That is beyond the scope of this article. Thermal diffusivity (in units of square feet per hour) is equal to the thermal conductivity (k) divided by the thermal capacity (q) of the aquifer; it measures the rate at which temperature changes occur in the soil mass. Higher values of thermal diffusivity result in more rapid changes in temperature and deeper penetration of heat into the soil. Ratio of Thermal Velocity to Groundwater Velocity Versus Porosity Lastly, as heated groundwater moves through an aquifer, thermal energy is transferred to the mineral matrix. As a consequence the thermal front associated with the advective flow of heated groundwater will move at a lower velocity than the groundwater. Aquifer porosity is the dominant element in this dynamic. The ratio of the velocity of the migration of the thermal front versus the groundwater velocity is affected by aquifer porosity. The smaller the ratio, the slower the thermal front velocity compared with the overall groundwater velocity. EFFECT ON PHYSICAL PROPERTIES AND REACTION KINETICS The two dominant physical characteristics of groundwater that change with temperature are viscosity and density. Their relationship is not linear. In the temperature range

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THE ROLE OF HEAT IN GROUNDWATER SYSTEMS

of 0 ◦ C to 30 ◦ C, the viscosity decreases by over 50%. This change improves the efficiency of both advective fluid flow and diffusional transport. In contrast, water density decreases significantly at temperatures over 60 ◦ C. Next, chemical reactions should be considered. Most salts increase in solubility as the temperature increases, the common exception to this general rule being carbonates in which solubility decreases as temperature increases. As temperature increases, the solubility of gases in groundwater decreases. The solubility (in milligrams/liter) of environmentally important gases over the temperature range of 5 ◦ C to 35 ◦ C is as follows: Oxygen with air as the source Carbon Dioxide Methane

12.7 to 7.1 2774 to 1105 34.1 to 17.3

The relationship between temperature and the kinetics of chemical reactions is expressed by the Arrhenius equation: ln k = ln A − Ea /RT where k A Ea R T

is the rate constant, is an integration constant, is the energy of activation, is the universal gas constant, and is the temperature in Kelvin.

Commonly, this relationship is graphed as ln k versus 1/T, which results in a straight line in which k increases exponentially as T increases. Under proper circumstances, small increases in temperature can have large effects on kinetic rates. In addition to reaction rates, temperature also affects equilibrium constants, particularly those that involve reactions with noncovalent bonds. Covalent bonds are relatively resistant to thermal perturbation. However, van der Waals interactions, hydrogen bonds, weak ionic bonds, and hydrophobic interactions are disrupted by relatively small temperature changes. Hydrogen bonds, ionic bonds, and van der Waals interactions all form with the release of heat, and hydrophobic interactions form with heat consumption. Increases in temperature will destabilize the first three, whereas it will stabilize the hydrophobic interactions. As temperature increases, the solubility of hydrocarbons increases, not uncommonly two to three times in the temperature range of 5 ◦ C to 30 ◦ C. Through this and the bonding reactions mentioned above, the forces responsible for the retardation of hydrocarbons within the soil matrix are reduced as temperature increases, improving the flushing action of groundwater flow induced by pump and treat recovery systems. Vapor pressure also increases with temperature, and increased vapor pressure results in increased mass transport rates in soil vapor extraction systems, which is where many of the prior applications of external heat to subsurface remediation have been used. Examples of the percent increase in vapor pressure as a result

of a temperature increase from 5 ◦ C to 35 ◦ C are as follows: Benzene Ethylbenzene Toluene TCE Tetrachloroethylene

417% 638% 513% 476% 545%

With regard to free product recovery, an increase in temperature will reduce the hydrocarbon viscosity and increase the recovery of free product from the surface of a water table. A temperature increase from 10 ◦ C to 50 ◦ C can increase the recoverable amount of middle distillates by 10%. Molecular diffusion is driven by the presence of a chemical concentration gradient between adjacent zones in the subsurface. The presence of a temperature gradient will also set up the diffusion of dissolved components in a fluid. The phenomena is called the Soret Effect and is known generally as thermal diffusion. Mass transport rates under thermal diffusion are described by the equation: NAT = DT (Rho)d ln (T/dz) where NAT is the mass transport rate as number of moles/cm2/second, DT is the coefficient of thermal diffusion in centimeters squared/second, T is temperature in kelvins, z is the distance in the direction of diffusion in centimeters, and Rho is the fluid density in gmoles/centimeters cubed. The magnitude of thermal diffusion is dependent on the size and chemistry of the molecules involved as well as on the temperature. At maximum, it may reach a diffusional mass transport rate of 30% that is seen for molecular diffusion driven by chemical concentration gradients, and it is typically significantly less than that. Lastly, a brief discussion on temperature and biological systems. Bacteria typically have a relatively narrow temperature range (about 10 ◦ C) in which they experience maximum growth and metabolic activity. As groundwater temperatures increase, the dominant bacterial consortia will change in response. This is true with increases of temperature up to about 35 ◦ C; temperatures of 50 ◦ C or above will cause traumatization and partially kill many bacterial species indigenous to the subsurface. It is common for the rates of physiological and biochemical processes to undergo a two-fold increase because of a temperature change of only about 10 ◦ C. This change is caused by the nature of the Arrhenius equation previously discussed as applied to biological systems. A practical note with regard to bench scale studies of soil bacteria: Incubation of bacteria in the laboratory ideally should be conducted in a temperature range that is from 5 ◦ C below the in situ groundwater temperature to 10 ◦ C above. On average, the groundwater temperature in the United States is approximately 10 ◦ C. Most bench scale laboratory testing is done at room temperature (about 25 ◦ C).

GROUNDWATER FLOW IN HETEROGENETIC SEDIMENTS AND FRACTURED ROCK SYSTEMS

In conclusion, many of the processes exploited in aid of subsurface remediation are impacted by temperature. The overall compounded physical/chemical/biological effects that can be anticipated within a relatively normal temperature range (5 ◦ C to 35 ◦ C) could on an additive basis make a three- to five-fold increase in remediation rates, which is not insignificant. The issue is then site specific, i.e., shallow versus deep contamination. Passive heating systems using modification of the ground surface to improve heat adsorption or low-intensity intrusive systems such as those associated with heat pump systems could offer remediation actions with possible economic benefits.

GROUNDWATER FLOW IN HETEROGENETIC SEDIMENTS AND FRACTURED ROCK SYSTEMS JAMES A. JACOBS Environmental Bio-Systems, Inc. Mill Valley, California

DAVID B. VANCE ARCADIS G&M, Inc. Midland, Texas

Groundwater flow in heterogenetic sediments and fractured rock systems is highly complex. Contaminants within these types of subsurface conditions are relatively common. Similarities and differences exist in flow characteristics of heterogenetic sediments and fractured rock systems, which influence contaminant fate and transport and subsequent remediation efforts at environmentally impacted properties. FLOW IN HETEROGENETIC SEDIMENTS Heterogeneity in granular aquifers is common and is caused by the large variability in water flow regimes in various depositional environments. The vertical and lateral variability in granular aquifers can occur on a scale of centimeters or less. Anisotropic advective flow, dispersion, diffusional transport, contaminant adsorption, and other physical/chemical processes taking place in these systems is complicated. Complete evaluation and understanding of these processes at any given site is impractical, and attempting to gather comprehensive data in support of that understanding is prohibitively expensive. Anisotropic advective groundwater flow dominates the heterogenetic sedimentary systems. Heterogeneity and subsequent anisotropic groundwater flow conditions are considered normal at most contaminated sites. Near-surface granular aquifers that are impacted by the release of contaminants are usually poorly indurated and can be classified according to the depositional regime as follows: • Fluvial that includes rivers, streams, and alluvial fans • Glacial such as tills • Lacustrine lake deposits • Eolian sand dune systems

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In these depositional systems (particularly fluvial and glacial), it is common to have granular deposits of high permeability contrast juxtaposed to each other, and therein lies the source of great complexity. Granular media advective flow regimes can be classified into three broad types: • Uniform, medium-to-coarse granular soils capable of supporting significant intergranular advective flow • Low permeability silts, clays, and tills in which any advective flow actually occurs in secondary permeable channels such as fractures or joints • Complex interbedded soils with zones of high permeability contrast resulting in anisotropic advective flow Sediment source and energy of transport are the primary controllers of the characteristics of waterdeposited materials. A rain storm, rising tide, or flood waters carry an initial high-energy load of coarser material (that may be mixed with finer grained silts and clays). As a given deposition event begins to subside, the denser and typically larger size grains settle first, followed by granular materials of decreasingly less dense and smaller sizes. The last stage may involve quiescent waters from which the finest clay and silt particles will settle over a significant period of time. The result of a given episode is a layer stratified by size within the overall soil mass. Scale is important when evaluating water flow characteristics. The smallest scale heterogeneity is expressed by variations in pore size and shape. Slightly larger, bench scale heterogeneity is expressed by variance in the particle size; on the larger scale of a road cut, heterogeneity is seen as the layering of individual bedding planes; and on larger scales, as changes in size of the layers occurs, pinch outs and facies change. In layered sediments of this type, horizontal flow is dominated by the most permeable units in the sequence and vertical flow by the least permeable. Mega scale depositional heterogeneity can also be created by processes such as braided stream sediments or glacial till systems, where lenticular pods of sand may be deposited in a matrix of lower permeability silts and clays. Material that is two or three orders of magnitude higher in permeability than the bulk soil matrix will totally govern the groundwater flow system. For example, 1 in. of sand can dominate the flow through tens of feet of silt or clay. The practical problem this presents is the determination of the spatial configuration of the physically small but hydraulically dominant units in the soil matrix. Because of this significant problem, continuous soil coring is recommended during subsurface investigations of soils with mixed sands and clay or silts, where practical. One potentially most damaging situation includes conditions in which the permeable channels or layers are preferentially exposed to the surface. These zones are susceptible to contaminant impact followed by surface recharge, which can act as a hydraulic driver of contaminants within a matrix that overall does not support high rates of advective groundwater flow, even

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in the more permeable (but unexposed) portions of the coarser soil units. The testing and analysis of these anisotropic heterogenetic flow systems is potentially extremely complex. As a matter of practicality, a significant number of assumptions must be made in the analysis of these aquifers. As readily available computing power has increased over the past decades, the number and degree of required assumptions have declined in the more sophisticated applications. Irrespective, two prime assumptions are that at some scale the heterogenetic soil can be treated as a homogenous block and that the spatial configuration of the heterogeneity has been defined. Increasing computational power typically allows for a finer mesh of blocks, but it has not addressed the issue of spatial configuration. A significant new tool that is beginning to be put to use is the employment of fractal concepts to study the hydrology of heterogenetic soils. This approach points to a serious flaw in the value of using finer blocks in the modeling matrix. If the geometry of the heterogeneity is fractal in the soil matrix (which evidence increasingly indicates it often is), the assumption that the matrix can be represented by averaging is false. A fractal matrix will not become homogenous with averaging irrespective of block size. With further use of the concepts of fractal analysis, it is possible to model general features and hydraulic behavior and interpolate hydraulic dynamics from sparse data, an ability of potentially great value in these complex groundwater flow systems. With this approach, water level, flow, and soil data can be used from an operating groundwater recovery system to fine tune understanding of the site. Through an iterative process, the fractal model resolves hydrogeological attractors in the flow system. Known hydraulic properties model drawdown in the forward iteration, and observed drawdown and flow rates model hydraulic properties in the inverse iteration. This process has the potential to be a powerful tool for defining heterogenetic anisotropic groundwater systems. At this juncture, computer capacity allows for primitive two-dimensional modeling in this manner. With improvements, it will soon be possible to robustly model three-dimensional systems. This technology will spatially identify which portions of the soil matrix have the greatest impact on the groundwater flow regime. It can also subtly determine areas where data gaps of high impact exist and that require additional testing wells. FLOW IN FRACTURED ROCK SYSTEMS Usually shallow water-bearing zones underlying a site are in shallow unconsolidated sediments. However, in some cases where bedrock outcrops or lies just meters below the surface, fractured rock hydrogeologic systems occur. In the latter case, any significant groundwater flow occurs in fracture systems within the bedrock. If soils are fine grained (i.e., tills or clays), fractures may also play a dominant role in advective flow. The study of fluid flow through porous media was first established by Henri Darcy in 1857. Much of the science of hydrogeology was developed with and was designed for use in granular porous media because of

the incentive to find and produce aquifers for largescale groundwater consumption. It took almost 100 years before fractured flow was studied in great detail. The study of fluid flow through fractured rock was first developed by the petroleum industry in the 1950s (1). These petroleum studies resulted from observations that oil and gas production could be significantly increased by fracturing the oil-producing formations near the well bore (2). Fractured media in most instances will not produce groundwater on the same scale as homogeneous granular aquifers. Fracture flow systems are also more complex to analyze and hydraulically respond differently than do those in porous granular matrices. A good overview of contaminant transport in fractured media is presented in Schmelling and Ross (3). To compare groundwater flow in fractured rock systems with granular media, under identical hydraulic gradients, one square meter of granular material with a hydraulic conductivity of 8.1 × 10−2 cm/s has the equivalent waterconducting capacity of one fracture in one square meter of rock with an aperture of 1 mm. In granular media, grain size, shape, and degree of sorting are the prime microscale parameters that determine hydraulic character. Fracture density, orientation, aperture, and type of rock matrix are the major parameters affecting groundwater flow in fractured media. The typical range for fracture aperture is from 0.2 to 25 mm, and fracture spacing is from 2 mm to 3 m. Individual fractures are not infinite in extent. Therefore, where flow is supported, fracture density must be high enough to ensure connectivity through the system. The fracture density required to sustain advective flow is termed the ‘‘percolation threshold.’’ Below that threshold, fractures may be connected, but only in small localized regions. Above, localized regions become interconnected, and flow over significant distances can take place. With an increase in fracture density, the system becomes increasingly previous. An expression of the parameter that determines the percolation through fracture is N(LF )2 , where N is fracture density and LF is equal to fracture length times pi/2. The percolation threshold has been found to fall around 0.3. As an example, in an area underlain by metamorphic rocks with discrete water-producing fracture systems: • Fractures were approximately 0.5 m in length • Rocks with a fracture density of 50 to 200 per 0.5 m This process gave values of NLF 2 of 30 to 120, well above the percolation threshold of 0.3. Given adequate connectivity, the flow an individual fracture can support is proportional to the cube of the fracture aperture. This cube rule means that a few fractures with preferentially higher apertures can dominate the flow system, and those are the ones most important to delineate. Typically, fracture aperture will decrease with depth. Usually the highest flow rates occur in the upper 9 m of a fracture system, with flow decreasing to near zero below depths of 30 m. To some degree, all rocks or soils are reactive. In most igneous, metamorphic, or fine-grained sedimentary

HORIZONTAL WELLS

rocks, geochemical reactions tend to seal a given fracture over time. In carbonates, the opposite can be true, with the fracture aperture increasing with time (and the potential flow rate increasing with the cube of that aperture). The hydrodynamics of fracture flow systems will approach those in porous media in systems where fractures are randomly oriented and density is high. These systems can be analyzed with conventional granular media methodologies. To use those methods on other fracture systems is an error with potentially significant consequences. Characterization of a fracture flow system is potentially an expensive process. Ideally, data should be gathered on fracture length, orientation, aperture, and density. Additionally, information on hydraulic head, the porosity and hydraulic characteristics of the bulk matrix, the type of contaminants, and the potential interactions between contaminants and the matrix are also important. Hydrogeologic characterization of fracture systems can be accomplished through coring, complex pumping tests, tracer tests, geophysical evaluation, or bore hole flow meters. In addition, evaluation of the hydrodynamics of a fracture system that has been defined with multiple orientations is mathematically extremely complex. However, some level of useful knowledge can be inferred based on the structural setting of a site. The tectonic and depositional history of a given site is generally available in the geologic literature. The removal of overburden introduces stresses caused by reduction of overburden pressure, uplift of the region, and thermal stresses caused by cooling; the net stress is extensional. Fractures that form in the tensile stress field can be placed into two classes: • Unloading fractures, which include vertical fractures, indicating response to tensile stress in the horizontal plane and fractures horizontal or parallel to the topographic surface • Release fractures that are fabric controlled in their orientation. If a site has undergone even a mild degree of tectonic deformation, structural analysis can be a powerful predictive tool for the orientation of the dominant fracture sets at a site. Several points are important to use this concept: • The deformational pattern observed in the large scale is replicated at medium, small, and microscopic scales. For example, the NE trend of the Appalachians is generally reflected as an NE-oriented fabric at all scales. • Deformation imparts a fabric to the impacted rocks. This fabric imparts anisotropicity, which in turn will control subsequent fracture generation. For example, fractures will tend to propagate perpendicular to a strong linear fabric element. • The stronger the degree of imprinted fabric, the greater the density of fractures in the controlled orientations.

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• Depositional features such as bedding planes also impart anisotropic fabric. Important differences exist between fractured and porous media with regard to contaminant fate and transport. Rapid transport in preferred directions can occur through rocks that normally would be thought impervious, such as bedrock fractured flow systems. Because of the general lack of organics in fractured rock systems, the flow of organic chemicals traveling through fractured rock systems is generally not retarded as much as in organic-rich granular sediments. Retardation is a function of matrix and contaminant chemistry and surface area. As an estimate, a block of granular media, such as well-sorted sand or gravel, will have a surface area 1000 to 100,000 times greater than a similar block of fractured media, such as fractured limestones or granites. Contaminant retardation will be roughly proportional. Contaminants in fractures can be mobile over long distances or until they are transported into a granular media. Fractured rock systems create interesting challenges as well as opportunities. When drilling contaminant recovery wells in fractured terrain, it is suggested to orient the well bore such that it is perpendicular to the major water-bearing fracture set. The resulting well may not necessarily be vertical, but it will maximize the potential to intersect flow-bearing fractures, which allows for more complete contaminant removal. BIBLIOGRAPHY 1. Gale, J.E. (1982). Assessing the Permeability Characteristics of Fractured Rock. Geological Society of America, Special Paper 189, pp. 163–181. 2. Duguid, J.O. and Lee, P.C.Y. (1977). Flow in fractured porous media. Water Resources Res. 13: 558–566. 3. Schmelling, S.G. and Ross, R.R. (1989). Contaminant Transport in Fractured Media: Models for Decision Makers. United States Environmental Protection Agency Superfund Ground Water Issue, EPA/504/4-89/004, Washington, DC.

HORIZONTAL WELLS MILOVAN BELJIN Cincinnati, Ohio

Horizontal wells are a new technology for solving problems in the environmental industry. Due to interest in horizontal wells for oil production, a large number of technical papers have been published regarding the reservoir engineering aspects of horizontal drilling and reservoir simulation. In the groundwater supply industry, the first theoretical analysis of groundwater flow to horizontal drains (collector wells) can be traced back to the early 1960s. In recent years, however, there has been renewed interest in horizontal wells for subsurface remediation. Horizontal wells offer significant advantages over vertical wells in environmental remediation and protection

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HORIZONTAL WELLS IN GROUNDWATER REMEDIATION

in many hydrogeologic scenarios. Horizontal well-screen orientation complements common site logistics, typical aquifer geometry, and groundwater flow patterns. A single horizontal well can replace many closely spaced vertical wells. The logistical advantages of horizontal wells are obvious. Horizontal wells avoid the need for installing wellheads inside buildings or in the midst of complex manufacturing or process facilities. Landfills, spoil mounds, and landfill liners need not be penetrated to extract leachate or other underlying contaminants. In addition, contaminant concentrations are often highest directly beneath buildings, landfills, and other obstacles to remedial operations, so that treatment facilities are constructed tens or hundreds of feet away from the target zone of remediation. Soils are naturally stratified and individual aquifers or water-bearing zones are much wider than they are thick. Despite the dominance of the horizontal direction in aquifer shapes and groundwater flow, the dominant tool for extracting contamination from subsurface sources is a vertical well. However, in many environmental remediation scenarios, a horizontal well offers a better match of form and function than a vertical well. The tabular geometry of many aquifer zones renders horizontal wells more productive than vertical wells. The flow characteristics of many aquifers create elongate contaminant plumes, and extracting contaminated groundwater is often more efficient using horizontal wells. A horizontal well placed through the core of a plume can recover higher concentrations of contaminants at a given flow rate than a vertical well. Horizontal wells also offer many advantageous over vertical wells in fractured aquifers. Fractures in an aquifer are commonly vertical. Because fluid or vapor recovery from fractured zones requires penetration of numerous fractures, a horizontal well oriented normally to vertical fractures is the optimal tool for pump-and-treat or soil vapor extraction systems in vertically fractured zones. By analogy, vertical wells are efficient in highly stratified soils that have little vertical communication between strata, where fluid or vapor recovery from many thin layers through a single wellbore is required. Injection of groundwater is part of some remediation systems, either to create a water table mound or to reinject treated effluent from a manufacturing plant into the subsurface. Water table mounds can help control flow of contaminants toward recovery wells or trenches, or they can serve as hydraulic barriers. Manufacturing plants can avoid high sewer discharge costs if their treated plant effluent can be reinjected into a nondrinking water aquifer. Reinjection can cause mounding, but the mounding can be minimized by using horizontal wells. Hydraulic barriers are most efficiently created using horizontal wells oriented perpendicularly to the groundwater flow direction. Various drilling technologies are capable of installing horizontal wells for subsurface pollution control. Such wells are typically installed in unconsolidated soils, 10 to 200 feet deep. Selection of drilling technique depends on surface access, well placement and completion requirements, and subsurface hydrogeology. During drilling of a well, drilling mud can invade the aquifer and change its

permeability in the vicinity of the well and cause formation damage or a ‘‘skin effect.’’ The thickness of the skin zone depends on drilling technology and also on the permeability of the aquifer. The additional drawdown due to the change in permeability and the turbulent flow around the well is called ‘‘well losses.’’ Because of lower flow rate per unit screen length, horizontal wells show smaller well losses due to drilling mud invasion than vertical wells. The current high installation cost of a horizontal well compared to that of a vertical well is offset by operating and maintenance cost savings. New developments in horizontal drilling technology will further reduce the cost of installing horizontal wells, and subsurface pollution control using horizontal wells should become as common as using vertical wells.

HORIZONTAL WELLS IN GROUNDWATER REMEDIATION JAMES A. JACOBS Environmental Bio-Systems, Inc. Mill Valley, California

DAVID B. VANCE ARCADIS G&M, Inc. Midland, Texas

Horizontal wells have been used in an increasing number of remediation projects. In addition to chemical and biological reactive barriers, horizontal borings can be used for groundwater extraction and control, air sparging, bioremediation, groundwater injection, for vadose zone soil vapor extraction or bioventing systems, and free product recovery. Although, in areas where there is a great deal of fluctuation in the groundwater table, use for free product recovery can be problematic. Horizontal well technology was first used in the late 1920s by the petroleum industry to increase oil reservoir thickness per well and greatly increase oil production. In the 1970s, the technology was applied by utility companies to cross rivers and other natural or manmade barriers such as highways. The use of horizontal drilling technology for environmental applications began in the late 1980s and has escalated ever since. Horizontal wells offer distinct advantages over trenching or vertical wells. Trenching produces massive quantities of excavation spoils and is economically limited in depth of application. In addition, horizontal wells can be installed beneath structures and other surface obstructions that would be impossible to access using trenches or vertical wells. Horizontal wells have a unique advantage that originates in the geometry of the typical contaminated groundwater system. Horizontal permeability is on the order of ten times greater than vertical, because of the stratigraphic layering of near-surface soil horizons. The effect of increased horizontal permeability enhances the spread of contaminant plumes horizontally. Horizontal wells can be installed through contaminated zones, along their leading edges, or along a property line. All active groundwater remediation systems rely on the mass

HORIZONTAL WELLS IN GROUNDWATER REMEDIATION

transport of water, air, or other chemicals. Mass transport is induced through the interface offered by the screened sections of a well that are exposed to the contaminated strata. Compared with vertical wells, horizontal wells can increase well screen interface by an order of magnitude or more. Various drilling heads and cutting removal technologies have been applied to horizontal drilling. Fluid-cutting systems, mechanical cutting, augers, percussion drilling, and sonic methods are all available. The selection of which is dependent on the geology and competency of the subsurface (and the selected technology vendor). With appropriate techniques, bore holes may be extended through difficult conditions such as gravel formations, coral reefs, and even boulders or bedrock. The removal of cuttings is usually based on mud slurry systems for installation beneath the groundwater table and air systems for installation in the vadose zone. The formulation and engineering of a mud slurry system takes great care; historically this has been one of the most common points for the failure of an installation. The mixture must be capable of holding the hole open, removing the cuttings without erosion of the walls of the well bore, and it must be capable of decomposing with time to restore the well bore to permeable operation. Inherent with the horizontal drilling process is the ability to accurately direct the placement of the horizontal well bore. Two dominant methods of achieving directional control of the drill head are available: magnetometer/accelerometer and radio beacon. A magnetometer/accelerometer array is expensive, and the long configuration of an array makes it prohibitive for short radius bends. It is also subject to magnetic interference, which can be a serious issue around tanks or other surface or near-surface structures (i.e., buildings, piers and piles, pipelines, or utility runs). Radio beacons previously have been limited to a depth of 25 feet or less, although there have been instances where down hole wirelines have increased accessible depth. Well casings for horizontal wells must be more highly engineered than for vertical wells. They must have great tensile strength to withstand the significant forces associated with installation. Compressive strength is required to resist the overburden load as the horizontal well bore collapses with time. The screen pack is also a critical issue with regard to installation and subsequent operational efficiency. Conventional sand packs can be installed, but they require great care. Many vendors now use prepacked screens. These are nested screens, an exterior screen (typically HDPE), a packing system (sand, filter cloth, or other filter media or mixtures of media), and an internal screen typically constructed of stainless steel. These prepacked screens are stiffer and subject to larger turn radii, but they obviate the difficulties associated with installing a conventional sand pack in a horizontal boring. Horizontal drilling is a relatively costly process to use. It is driven by necessity, or the economic advantages of scale. Necessity drivers are generally the need for the installation of a remediation system underneath a structure or facility that cannot be disturbed. Economics of scale come in to play with regard to large contaminant

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plumes. A single horizontal boring can replace 10 to 30 vertical wells in a plume of large aerial extent. The breakeven point for the choice between horizontal or vertical wells occurs around the need for four to five vertical wells. Installations requiring more than five vertical wells along a linear trend can often be more cost effectively addressed with horizontal wells. Two basic methods of horizontal well installation are available. The simplest involves two ends, boring downward to near the desired depth at one end, curving the boring to the horizontal, traversing the required path and distance horizontally, and then recurving to the surface at the other end. The casing is then pulled back through the boring from the distal end. A more sophisticated approach is to use a single end, through which the boring is advanced, followed by the casing being pushed through the boring, a more difficult, risky, and time-consuming process. Horizontal drilling using the two-end approach can cost as little as $30 to $40 per linear foot; the cost for single-ended systems start at $80 to $90 per linear foot. As with many other technologies applied to the environmental industry, developers over the last decade have focused on the creation of a ‘‘just good enough’’ technology. Placement accuracy is sufficient, boring sizes are minimal, and construction materials are designed for the limited life of the project. Successful installation of horizontal wells depend on the selection of a vendor who is flexible enough to use an approach that is appropriate for the subsurface condition of your site; who is experienced in the use of the specialized drilling equipment and screens; who can engineer the application as well as the installation procedures; and who is experienced enough to be quick. Aside from failure caused by inadequate mud engineering, taking too long is another key source of failure; the well bores simply cave in before installation is complete. Horizontal wells have been adapted for use in many soil and groundwater projects. The use includes groundwater extraction, air sparging, free product recovery, in situ bioremediation and bioenhancement, soil vapor extraction, in situ soil flushing, in situ radio-frequency heating, treatment walls, hydraulic and pneumatic fracturing, and leachate containment and collection (1). The overall performance of horizontal wells used for air sparging and vacuum extraction at the Savannah River Site in South Carolina (2) shows the following: • A five-time increase in chlorinated solvent removal as compared with conventional vertical wells. • Eight tons of chlorinated solvents were removed over a 20-week period. The equivalent would have been 11 vertical wells at a pump and treat system, each extracting at a rate of 500 gallons per minute; The Savannah River Site project managers estimate that a 40% cost savings was predicted when compared with the use of conventional pump and treat technologies. In the six years between 1987 and 1993, over 100 horizontal wells were drilled in the United States as part of environmental remediation projects. One quarter of them were used for groundwater extraction, one quarter of them were used for soil vapor extraction, and one half

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HEAD

of them were used for other technologies (air injection, bioventing, and free product recovery). Eighty percent of the horizontal wells were installed at depths of 25 feet or less (2). The advantages of horizontal wells are well known: Horizontal well screens contact a larger surface area of contaminated aquifer than do conventional vertical wells; the cost of horizontal wells, although more than vertical wells, is less expensive when comparing the fewer number of wells required for a particular remediation project; and finally because horizontal transmissivity generally exceeds vertical transmissivity in most aquifers, horizontal wells can deliver and recover more fluids, gases, and groundwater than can vertical wells (1). With these advantages and the improvements in directional drilling and global positioning systems (GPS), it is likely that horizontal wells will become more common in future remediation projects. BIBLIOGRAPHY 1. Miller, R.R. (1996) Horizontal Wells, Ground-Water Remediation Technologies Analysis Center (GWRTAC). Technology Overview Report TO-96-02. 2. Hazardous Waste Remedial Actions Program (HAZWRAP). (1995). In situ Bioremediation Using Horizontal Wells. Innovative Technology Summary Report, prepared for U.S. Department of Energy. Available: http://www.gnet.org/gnet/tech/ reports/sbu.htm.

Water surface

p=0

z=H

Elevation Pressure

z=0

p=H

Figure 1. Total head is the sum of the elevation and pressure heads. For example, the elevation decreases and the pressure increases when moving downward from the water surface in a column of still water.

where p = P/γ is the fluid pressure head. Figure 1 shows how the pressure head and elevation head compensate in a column of still water. The pressure term is omitted when the water surface elevation is measured, because the fluid pressure is zero, p = 0, at this point. This simplification yields the equation, H = z, where z is the elevation of the water surface. Important assumptions required for the use of water levels to determine the head include (1) hydrostatic conditions exist with no vertical water movement; (2) the water velocity within the aquifer is sufficiently small; (3) water within the monitoring borehole or piezometer is pure water at a standard temperature and density; and (4) the air pressure on the water surface equals the mean barometric pressure.

HEAD CAPILLARY AND OSMOTIC FORCES

TODD RASMUSSEN The University of Georgia Athens, Georgia

Water moves from zones of higher energy to zones of lower energy. For example, gravity causes water to move from higher to lower elevations. Water also moves from higher to lower pressures. The total energy is the sum of the all forces acting on water. This total energy is referred to as the head, which can determine the direction and rate of water movement. Bernoulli’s equation commonly combines the dominant forces that cause water to move: H =z+

v2 P + γ 2g

(1)

where H is the total head, z is the elevation of the point where the head is measured, P is the fluid pressure at the point of measurement, γ = ρg is the fluid specific weight, ρ is the fluid density, g is the gravitational constant, and v is the fluid velocity. The velocity component can be omitted if the flow is slow, v ≈ 0, which is generally true in groundwater systems. When the velocity is neglected, the total head is H =z+p

(2)

Capillary and osmotic forces also affect the total head. Failure to account for these forces may result in incorrect predictions of water flow and transport. Capillary forces develop because of the tendency of soil surfaces to attract water. Water held on soil surfaces resists the downward force of gravity, and it does not readily drain from the soil. The head must consider the negative fluid pressure that develops because of capillary forces. A measure of the force by which water is held is the matric tension, which can be measured with tensiometers. Water may move upward above the regional water table because of capillary forces. The height of the saturated zone formed above the water table (i.e., the capillary fringe) is largely determined by the magnitude of capillary forces, which is a function of the pore surface area. Finer grained media have greater capillary forces, which result in higher capillary fringes. Capillary forces generally increase with decreasing pore size, as shown in Fig. 2. The capillary rise equation relates the pore size to the height of rise: ψ=

2σ cos α γr

(3)

where ψ is the capillary height of rise, σ is the surface tension of water, α is the solid–liquid contact angle, and r is the pore radius (1).

HEAD

r1

181

r2 Y2 g

Y2

Figure 2. Height of rise of water in capillary tubes. Note that the height of rise increases as the radius of the tube decreases. Equivalently, small pores hold water to greater tensions than do large pores.

An additional force that induces fluid movement is a change in solute concentration. The osmotic potential decreases the head of water, which causes water to flow from areas of low solute concentrations to areas where solute concentrations are higher. In arid areas with high soil solute concentrations, salts may become concentrated as water evaporates at the soil surface, which causes an increase in the salt concentration. This increase in salt concentration induces an additional force that causes water to move upward from less saline groundwater to the surface. Thus, evaporation can increase the height of the capillary fringe by augmenting the capillary force with osmotic forces. The osmotic height of rise in dilute solutions is kTC (4) φ= γ where φ is the osmotic height of rise, k is the Boltzmann constant, T is the absolute temperature, and C is the solute concentration (2). FLUID DENSITY Water levels can be affected by the density of the fluid within the borehole. For example, water levels in a well monitoring a deep aquifer may be affected by the temperature of the water in the borehole. Warmer water is less dense than is colder water (except below 4 ◦ C) and so a column of warm water will display a higher water level than will a column of colder water for the same pressure at the bottom of the borehole. Because of the geothermal gradient (the tendency of temperature to increase with depth below the surface), deeper water within the borehole is warmer and slightly less dense than is water near the surface. Other factors besides temperature affect the water density in a well. The dissolved solids concentration (salinity) causes the water density to increase. Suspended sediments also increase the weight of the fluid, whereas air bubbles rising to the surface lower the density of the water column. Small changes in density over a long water column cause an appreciable difference in observed water levels.

z

H = z0 + g (z − z 0) g0

z0

Aquifer

Figure 3. Total head, H, adjustment for conditions when the borehole specific weight, γ , is different from the freshwater weight, γo . The effect is greater for larger differences in elevation between the water level, z, and the screened zone, zo .

To account for the variation in fluid density, observed water levels should be adjusted with a standard water density, called the freshwater density. The corrected head is called the freshwater head, which accounts for the fluid density within the borehole: H = zo +

γ (z − zo ) γo

(5)

where H is the freshwater head, zo is the elevation of the screened interval, z is the observed water level elevation in the well, γ = ρg is the average fluid specific weight in the well, and γo = ρo g is the standard freshwater specific weight. Figure 3 illustrates the geometry of this problem. The freshwater head equation can also be written as H = zo +  z

(6)

where  = γ /γo is the specific gravity of the fluid within the well and z = z − zo is the height of the column of water within the monitoring well above the screened zone. These equations show that the freshwater head correction is larger for longer water columns and for water with a density that is substantially different from the freshwater density. SPATIAL AND TEMPORAL VARIATION Head is often observed to vary over both space and time. These variations are particularly important when mapping the regional potentiometric surface, because long-term averages may not accurately reflect the surface dynamics over time. Temporal variations result from barometric (atmospheric pressure) influences, tidal effects, fluid density (sediment, salinity) changes, and vertical flow (nonhydrostatic conditions) within the water column (2). Spatial variations are a function of aquifer properties as well as of the regional hydrogeologic flow environment. Heads within a specific hydrogeologic unit generally vary smoothly over space, unless some kind of intervening boundary causes a jump in the head within the formation. For example, water levels may decline smoothly, only to change abruptly when a fault displaces one side of an

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WELL HYDRAULICS AND AQUIFER TESTS

aquifer relative to the other, which causes a sharp change in the head across the boundary. Also, heads in different hydrogeologic units may be different because the recharge and discharge patterns for each aquifer are different. A unit that outcrops in one area may have a different head than one that outcrops at a higher elevation. Wells that tap different aquifers are commonly observed by scientists to have different water levels, even when the wells are situated next to each other. This variation of head by aquifer is a result of the regional hydrogeologic setting, which must be considered when trying to evaluate water head data. Spatial variations are plotted on maps with equipotentials, which are lines of constant head within a hydrogeologic unit. Each aquifer generally has a unique set of equipotentials that can also vary over time. In isotropic media (i.e., aquifers with no preferential flow direction), the direction of fluid flow is shown by placing lines, called streamlines, on the map that are perpendicular to the equipotentials. Water levels in wells can also vary over time. Water levels in wells are commonly affected by barometric pressure—they fall as barometric pressure rises, and they rise as the pressure falls—because the total head in the aquifer is the sum of the water level elevation plus the atmospheric pressure on the water surface in the borehole. Although the influence of barometric pressure is commonly neglected when monitoring water levels in wells, large changes in barometric pressure (such as when large storms pass overhead) can occasionally cause large errors in the determination of the total head. To correct for the effects of barometric pressure, observed water levels, W, can be adjusted by the variation in pressure about the mean, B = B − B: H = W + B

(7)

The mean pressure, B, can be taken as the average global sea-level barometric pressure, B = 1013.25 hPa ≈ 33.9 ft, or, alternatively, equal to the local average barometric pressure, which varies with elevation and local weather conditions. Air pressure variations can be neglected by sealing the well, or by measuring absolute pressure instead of the gauge pressure within the monitored interval. Barometric pressure changes cause many wells to fluctuate over short time periods (e.g., from day to day), whereas precipitation, evapotranspiration, and pumping patterns often cause longer term variations. Climatic variation over time also can have a large influence on observed head. In trying to establish a long-term average head for a well, one must consider all possible sources of short- and long-term variation. BIBLIOGRAPHY 1. Hillel, D. (1971). Soil and Water: Physical Principles and Processes. Academic Press, New York. 2. Spane, F.A., Jr. and Mercer, R.B. (1985). HEADCO: A program for converting observed water levels and pressure measurements to formation pressure and standard hydraulic

head. Pacific Northwest National Laboratory, Rockwell Hanford Operations, Richland, WA, Report PNL-10835.

WELL HYDRAULICS AND AQUIFER TESTS JOHN E. MOORE USGS (Retired) Denver, Colorado

The objectives of an aquifer test are to identify the performance of a well (to estimate the yield capability) and to estimate aquifer properties. Accurate estimates of the hydraulic characteristics of aquifers depend on reliable aquifer test data. The tests are done on existing wells or a well drilled specifically for that purpose. An aquifer test is a controlled, field site experiment to determine hydraulic conductivity and aquifer storage. The test consists of measuring groundwater discharge and observing water level changes in the pumped well. These are the hydrologic and geologic conditions needed for a successful aquifer test: • Hydrogeologic conditions should not change over short distances. • No discharging well or stream nearby. • Discharge water should not return to the aquifer. • The pumped well should be completed to the bottom of the aquifer and should be screened or perforated through the entire thickness of the aquifer. • Observation wells (at least three) should be screened at the middle point in the aquifer. One observation should be located outside the area of influence of the pumping well drawdown. • Location of observation wells should be based on the aquifer character. • Determination of prepumping water-level trend. The following conditions and field measurements are needed for an aquifer test: • Accurate water-level measurements during pumping and recovery • Pumped well developed prior to test (several hours of pumping and surging) • Dependable power source to provide a constant pumping rate • A flow meter that can read instantaneous and cumulative discharge • Electrical conductance, Eh, pH, DO, and temperature measurements • Water levels measured several hours before the test begins • Pumping rate maintained at 5% tolerance. An optimal rate is 50% of maximum yield • Water level is measured with an electric sounder or pressure transducer • Remove the discharge water from the site

WELL HYDRAULICS AND AQUIFER TESTS

• Observation wells should be tested by injecting a known volume of water and measuring the recovery response • Establish baseline trends of regional water level changes and barometric pressure changes • Pumping well lithology and construction data TYPES OF TESTS Specific Capacity Test The amount of water that a well will yield can be determined by a specific-capacity test, in which the pumping rate and water-level changes are monitored for a set period of time. The first step is to measure the initial water level in the well. Commonly, a well is pumped at several successively increasing rates for uniform periods (typically 1 hour) to establish a rate that can be maintained for long-term pumping. The well is then pumped at a steady rate and the water-level changes are monitored at the pumped well. Water levels should also be monitored in at least one observation well 2 to 20 meters (6 to 65 feet) from the pumped well. The water level will decline quickly at first, as water is removed from the well, then more slowly as the rate of flow into the well approaches the pumping rate. The ratio of the discharge rate (Q) to water-level change (drawdown, dd) gives the well’s specific capacity, or Sc = Q/dd. For example, if the discharge rate is 6 liters per second (L/s) (100 gallons per minute) and the drawdown is 3 meters (10 feet), the specific capacity of the well is 2 L/s per meter (10 gpm/ft) of drawdown. Once the specific capacity and the available amount of drawdown are known, the yield of the well can be determined from the formula Q = Sc × dd. The pump should be deep enough that the water level does not go below the pump intake. The pump depth should also be sufficient to allow drawdown caused by pumping and natural declines in water level during periods of drought Step-Drawdown Test

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the water level is measured. The aquifer transmissivity can be determined from the time–drawdown or recovery data. The disadvantages of the test are that a data logger is needed to measure water-level changes and water removed represents only a small volume of the aquifer. Slug test data are evaluated by the Bower and Rice (1) method for unconfined aquifers and the Cooper, Bredehoeft, and Papadopulos (2) method for confined conditions. The advantages of these tests, compared to those of full aquifer tests with observation wells, are reduced cost and time. The disadvantage is that a storage coefficient is not determined and only a small volume of the aquifer is sampled. Many factors contribute to error in slug tests as follows: entrapped air, partial penetration, leaky joints, and the radius of influence of the test. ANALYSIS OF AQUIFER TEST DATA USING THE THEIS EQUATION The Theis equation (3) is used to determine the hydraulic characteristics of an aquifer. In this test, a well is pumped, and the rate of decline of the water level in nearby observation wells (two or more) is noted. The time drawdown is then interpreted to yield the aquifer parameters. In 1935, C.V. Theis (3) developed the first equation to include pumping time as a factor (4). The following are assumed: 1. The pumping well is screened only in the aquifer being tested. 2. The transmissibility of the aquifer is constant during the test to the limits of the cone of depression. 3. The discharging well penetrates the entire thickness of the aquifer, and its diameter is small compared to the pumping rate. These assumptions are most nearly met by confined aquifers at sites far from the aquifer boundaries. However, if certain precautions are observed, this equation can also be used to analyze tests of unconfined aquifers (5).

The step-drawdown test evaluates the performance of a well. Well performance can be affected by resistance to flow in the aquifer itself; partial penetration of the well screen, incomplete removal of mud from the gravel envelope, or invasion of fines into the envelope; and blockage of part of the screen area. The well should be developed prior to the test using a surge block and/or pumping until the well discharge is clear In this test, the well is pumped at several (three or more) successively higher pumping rates, and the drawdown for each rate is recorded. The test is usually conducted for 1 day. The discharge is kept constant through each step. The test measures the change in specific capacity. The data provide a basis to choose the pump size and discharge rate for the aquifer test and for long-term production.

1. Bower, H. and Rice, R.C. (1976). A slug test for determining hydraulic conductivity of unconfined aquifers with completely or partially penetrating wells. Water Resources Research 12: 423–438.

Slug Test

4. Heath, R.C. (1989). Basic ground-water hydrology. USGS Water Supply Paper 2220, p. 84.

In this test, a small volume of water is removed from a well, or a small volume is added and the recovery of

5. Lohman, S.W. (1972). Ground-water hydraulics. USGS Professional Paper 708, p. 70.

BIBLIOGRAPHY

2. Cooper, H.H. et al. (1996). Response of a finite-diameter well to an instantaneous charge of water. Water Resources Research 3: 263–269. 3. Theis, C.V. (1935). The significance of the cone of depression in groundwater bodies. Economic Geology 33(8): 880– 902.

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HYDRAULIC PROPERTIES CHARACTERIZATION

HYDRAULIC PROPERTIES CHARACTERIZATION WALTER W. LOO Environmental & Technology Services Oakland, California

Groundwater is becoming an important issue as a source of fresh water. The groundwater that exists at the subsurface is called an aquifer. An aquifer is a geologic unit that can store and transmit water at rates fast enough to supply reasonable amounts to wells. Thus, understanding a groundwater body is an essential factor in groundwater development and protection. The geohydrologic properties of a groundwater body consist of physical, chemical, and biological parameters of the solid matrix and liquid within its pore space. These parameters determine the quantitative, qualitative, and interpretive aspects of the groundwater body. The solid matrix may consist of various sedimentary deposits or rock types. The liquid within the pore space of the groundwater body consists of water, dissolved minerals and gases, microorganisms, and colloidal material.

INTRODUCTION Many groundwater professionals define a groundwater body by its flow property, such as aquifer, aquitard, and aquiclude (1), which is only good in defining the rate of water yield from a groundwater body. When dealing with water quality and environmental issues, we need better definition of the other properties and characteristics, not just dissolved chemicals, hydraulic conductivity, and effective porosity. The following parameters characterize groundwater body properties: Geophysical parameters: electrical conductivity, hydraulic conductivity, storativity, effective porosity, temperature, media thickness, pressure or hydrostatic level, adsorption, absorption, radioactivity, electrical charge, electrokinetic flow, porous or fracture media, recharge, discharge Biochemical parameters: pH, bacteria counts, inorganic chemicals, organic chemicals, oxidation/reduction, ion exchange potential Most of these parameters are not required for site characterization by natural resources agencies and the U.S. EPA (2,3). It is very clear that most groundwater bodies are poorly characterized because there is no incentive or pressure to understand these parameters, as long as the groundwater is fit for use. Most textbooks define only some of the parameters but without standardized procedures. The following sections provide typical definitions of the parameters, analytical methods, use, interpretation, limitations, and the necessity for future improvement in characterization.

PARAMETER DEFINITIONS The definitions of the following parameters are provided in Driscoll (4), Freeze and Cherry (1), Todd (5), Lohman (6), and Batu (7), Ferris et al. (8) Loo et al. (9) unless otherwise noted. Geophysical Parameters Electrical conductivity is the ability of a medium to conduct electricity measured in µmho/cm. Resistivity is the inverse of electrical conductivity measured in ohm-meters. Hydraulic conductivity is the measurement of the ease of water flowing through a medium or water permeability. It is often expressed as cm/s or gpd/ft2 . Storativity of a saturated aquifer is the volume of water release from storage per unit surface area per unit decline of hydraulic head. Effective porosity or flow porosity is the interconnected pore space available for water to flow through (10). Media thickness is defined as the aquifer thickness or productive water-bearing zone. Potentiometric surface is expressed as the pressure in the water head in a confined aquifer. It is called the groundwater table when the aquifer is unconfined. Adsorption/Absorption is the process by which molecules of dissolved chemicals in a fluid attach to solid surfaces (11–13). Radioactivity is the radiation emitted by natural elements or radioactive wastes. Radon and tritium are commonly occurring radioisotopes of uranium or other radioactive elements (14). Electrical charge is measured as the electrical potential difference (in millivolts) between layers of different materials. Fine-grain material such as clay and silt are slightly positively charged whereas sand and gravel are slightly negatively charged. An electrokinetic gradient occurs when a flow of water is created by dc electricity flow from an anode to a cathode (1,15). It is common to create several feet of water head at the cathode when the applied direct current potential difference is 50 volts over a short distance of 40 to 50 feet between electrodes. Recharge/Discharge Boundaries can be detected during a pump test when the drawdown curve becomes flat, recharge condition and change (steepened) slope, negative or discharge condition (16). Pumping near a surface stream can often detect a recharge condition. Pumping near a fault or lithologic discontinuity can often detect a negative boundary condition. Biochemical Parameters pH is an indication of the acidity or alkalinity of soil or water. Bacteria count is the enumeration of the population of indigenous bacteria in soil or water which often is expressed in colony forming units, CFU (17).

HYDRAULIC PROPERTIES CHARACTERIZATION

Inorganic Chemicals are any chemicals not based on carbon such as any salts of major cations and anions and trace metals. Organic Chemicals are carbon based compounds and can be naturally occurring or occurring as spilled chemicals. Oxidation/Reduction Potential redox potential in short as is the state of electrochemical reaction which is measured in millivolts. Ion exchange potential is related to the electrical charge measured in millivolts of colloidal particles which carry relatively large surface area. Colloidal particles have diameters in the range of 10−3 to 10−6 millimeters. Clayey material has very high cation exchange potential. FIELD TESTING AND ANALYSIS OF PARAMETERS Geophysical Parameters Many of the following physical parameters can be defined by field pump tests: electrical conductivity hydraulic conductivity storativity effective porosity or flow porosity temperature radioactivity media thickness pressure or hydrostatic level or potentiometric surface porous or fracture media recharge/discharge boundaries Pumping Test. Before a pumping test is conducted, geological and hydrological information such as the geological characteristics of the subsurface, the type of aquifer and confining bed; the thickness and lateral extent of the aquifer and confining beds, including boundary conditions, is obtained, preferably by surface resistivity and electric logs. Also, data on the groundwater flow system, including the hydraulic gradient and regional groundwater flow and existing wells in the area, should be collected. Then, the site for the well is selected considering the area representative of hydrological conditions, not near railroads and motorways, not in the vicinity of an existing discharge well, or a low water level gradient. After the well site has been chosen, the drilling operation can begin. The pumping well should be drilled to the bottom of the aquifer. Then the pump size and pump type are planned. Besides the well diameter and well depth, the location of the well screen is to be determined. The length of the well screen will largely be decided by the depth at which coarse materials are found. A general rule is to screen the well 100% of the aquifer thickness or full penetration. Partially penetrated observation wells can be used if the distance of the observation is more than 2.5 times the thickness of the aquifer tested, which is true for isotropic and anisotropic aquifer pump tests.

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With regard to a pumping rate, too low or too high a pumping rate is not desirable. Too high a rate can produce water loss, and too low a rate may make too gentle a drawdown curve. The water levels measured in observation wells represent the average head at the screen of the observation wells. The measured drawdown responses to be taken during a pumping test are of two kinds, measurements of the water levels in the well and the observation wells and measurements of the discharge rate of the well. To be practical, each pump test should last no more than one day (1440 minutes), which is enough to cover more than three time log cycles; 10, 100 and 1000 minutes, respectively. Measurements should also be made of the atmospheric pressure, temperature, the levels of nearby surface water, if present, and any precipitation. A longer duration pump test is good only for the analysis of boundary conditions (16). Drawdown data for the time period to overcome well bore storage should not be plotted for analysis. A step drawdown test should not be conducted because it is a test for well yield and well efficiency for well drillers’ use and not for the analysis of geohydrologic parameters (18). Single Dimensional Pump Test Analysis. The data analysis for the result of the pumping test is dependent on the aquifer conditions such as isotropic or anisotropic aquifer, homogeneous or non-homogenous aquifer, nonleaky or leaky aquifer, confined or unconfined aquifer, confining aquitard, etc. Data analysis also depends on the drawdown condition such as steady state or transient conditions. Before interpreting the pumping test result, conversion of the data into appropriate units and correction of the data for external influences are necessary. The interpretation of the pumping test data is primarily a matter of identifying an unknown system. Theoretical models comprise the type of aquifer and the initial and boundary conditions, which, in a pumping test, affect the drawdown behavior of the system in their own individual ways. Old fashioned curve matching techniques and computer fitting analysis all have inherent errors such as matching inaccuracy and oversimplified statistical assumptions. Therefore, curve matched analysis results may not be very accurate. It is more practical to do a pump test analysis on semilog paper using the straight line plot interpretations established by Cooper and Jacob (19) for hydraulic conductivity, storativity, and radius of influence from drawdown versus time and drawdown versus distance semilog plots. The scatter of the data points is also an indication of whether anisotropic or heterogeneous conditions are encountered. An isotropic and homogeneous aquifer will be reflected by subparallel slopes on the drawdown versus time plots for observation wells, and all points on the drawdown versus distance plot will fall on a straight line (only isotropic). If the slopes of the drawdown versus time plots are not subparallel, the aquifer is heterogeneous. If the points on the drawdown versus distance do not fall or fit onto a straight line (scattered), the aquifer is anisotropic or shows a potential

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preferential flow direction such as channel flow, fracture media, or artificial conduits of migration (20,21). For aquifer test analysis in unconfined aquifers, the Boulton analytical method will apply for determining hydraulic conductivity and specific yield or storativity (22). New analytical solutions for evaluating the drawdown near horizontal and slanted wells with finite length screens in water table aquifers are presented in Zhan et al. (23). These fully three-dimensional solutions consider instantaneous drainage or delayed yield and aquifer anisotropy. For the analysis of the leakage property of the confining layers of the aquifer, Hantush (24) developed a semilog straight line analytical method, known as the ‘inflection point’ method. Multidimensional Pump Test and Analysis. Two-dimensional horizontal hydraulic conductivity anisotropic tests (major and minor tensor) require three or more observation wells at different distances and orientations from the fully penetrating pumping well (25,26). The Hantush and Thomas analytical method is a hand contouring method which is not very accurate and may require many observation wells. The Papadopulos method requires drawdown data from three or more observation wells at the same elapsed time and a least squares or equivalent numerical fitting technique for an elliptical fit. The fitted ellipse orientation and axes thus define the horizontal anisotropic hydraulic conductivity tensors (20,21) Loo (27). For the analysis of vertical hydraulic conductivity (not leakage from confining layers), the pump test requires nonoverlapping partially penetrated wells, one pumping well and one or more observation well, located within 2.5 times the aquifer’s thickness from the pumping well (28). The pump test procedure for multi dimensional hydraulic conductivity tensors was first conducted in an alluvial deposit at Christensen Ranch in Powder River Basin, Wyoming, at an in situ uranium mining test site. Loo et al. (29) conducted a similar test at the Equity/DOE BX in situ oil shale project. All pump test and analysis procedures described were documented in the manual written by Loo (20), Loo (30) and Loo (27). Effective Porosity Testing and Analysis. No pump test can analyze effective porosity. The specific yield value from pump test analysis provides only a partial characterization because the specific retention is not defined. Freeze and Cherry (1) described the tracer test analysis for dispersivity and effective porosity. However, the tracer test procedures have not considered retardation of the sorption properties of the solid matrix. The selection of different tracers for different media is an art more than a science because there is no perfect tracer. Therefore, a tracer test for determining effective porosity is not very practical. The analysis of effective porosity can actually be easily determined (though seldom used) from stressing or loading an aquifer. This can be done by tidal efficiency or barometric efficiency tests or simply by sucking a vacuum on the pump test well. The effective porosity can then be

calculated once the storativity is defined by the pump test analysis. Jacob (31) provides an analytical solution for a confined aquifer. Hantush (32) provides an analytical solution for an unconfined aquifer. Radioactivity can be mapped by a natural gamma ray borehole geophysical log. If an abnormal level of radiation is mapped, then it may be necessary to test for radioisotopes of uranium, radon, and tritium and the level of their radioactivity (14). Electrokinetic Parameter Testing and Analysis. A resistivity survey can be conducted in boreholes to distinguish the lithologic layering sequence. A surface resistivity survey can define lateral continuity or discontinuity or boundaries (33,34). These geophysical surveying methods provide large areal extent and vertical definition or threedimensional mapping of the groundwater body. The electrokinetic gradient was mentioned in only one paragraph in Freeze and Cherry (1). The electrokinetic gradient can be tested between monitoring wells by impressing a dc voltage across the wells (15). The response of the water level rise in the cathode well can be quite large (some times more than 10 feet) between wells 50 feet apart at a modest 50 volts potential difference and 10 amperes of dc flow. However, there is no standard test procedure at this time. This is an evolving field testing technology which may ultimately partly replace the standard pumping test. It may also provide artificially induced desorption and oxidation/reduction environments for contaminant treatment. Elemental adsorption can be tested for cation and anion exchange capacity on soil samples in the laboratory (35). The amount of adsorbed chemical can be estimated from the Freundlich and Langmuir approximations (11). There is no standardized laboratory test for adsorption/absorption parameters of on organic compounds a fine-grain solid matrix at this time. But these are important properties because for most contaminant mass migration (mostly not very soluble), more than 90% by weight is adsorbed/absorbed by the solid matrix. Biochemical Parameters pH bacteria count (17) inorganic chemicals organic chemicals oxidation/reduction ion exchange potential All these parameters can be determined by standardized U.S. EPA laboratory analytical methods for water and wastewater. The redox potential of groundwater was required only recently for evaluating natural attenuation or intrinsic bioremediation. However, there is very little emphasis on the test procedures for a solid matrix below the groundwater level. Most people treat soil as soil in the vadose zone. There is no requirement for characterizing soil or rock properties underneath the water table (2,3). It is important to understand that the weight of solid matrix underneath the groundwater table represents more than 90% of the total

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weight. The laboratory procedures can easily be changed to accommodate the characterization of these properties in a solid matrix submerged under groundwater.

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geophysical surveys, and more in-depth characterization of the solid matrix submerged under groundwater. New Test Methods

USE, INTERPRETATION, AND LIMITATIONS After the aquifer system’s properties are properly characterized, then groundwater use planning and management can be implemented. Groundwater Resources Use Groundwater is a limited resource, and it is very easy to overdraw a groundwater system. The result is water shortages, coastal saline water intrusion, and land subsidence. The remedy for the situation is to ensure that the design of the groundwater withdrawal system is within safe yield limits that is, using groundwater at less than the natural recharge rate. Groundwater recharge usually can be expressed as 1 to 10% of natural precipitation. The rate of recharge is dependent on the evapotranspiration rate of the area. The remedy for saline water intrusion or upwelling is either to pump less or design a water mound or recharge mound near the coastal area (5). It is most unfortunate that, once the aquifer is intruded by saline water, the aquifer will be difficult to clean up by natural dilution. The remedy for land subsidence is to use surface water to recharge the groundwater during the wet season. This is cyclic recharge and pumping management of a groundwater basin. Groundwater Pollution, Prevention, and Remediation Since the U.S. EPA was formed about 30 years ago, groundwater pollution has not stopped. Nitrate pollution due to overfertilization and from feed lots, dairy farms, pig farms, and chicken farms continues unabated. This has resulted in the ‘dead zone’ at the estuary of the Mississippi River in the Gulf of Mexico. Inland, the Salton Sea and the Kesterson Reservoir in California here become irrigation drainage wastewater catch basins. Farmers and agricultural business are often exempted from environmental protection regulations. Most shallow groundwater in these areas is heavily polluted. Organic solvents such as PCE and TCE exist at almost all Superfund sites, electroplating shops, dry cleaners, and electronic manufacturing sites. Fuel hydrocarbons containing benzene, MTBE, and PAHs (such as naphthalene in diesel) have reached the groundwater at many groundwater supply sources. Chromium and naturally occurring arsenic are also reaching groundwater supplies. Fortunately, there are cost-effective technologies for cleanup at the source (15). FUTURE IMPROVEMENTS As mentioned in the previous sections, there is much room for improvement in geohydrologic characterization. Characterization efforts need large-scale representation area. This will lead to large-scale geohydraulic tests,

Future geohydraulic tests will most likely replace the aquifer pump test by stressing the aquifer by seismic or electrokinetic methods, so that it is more cost-effective. The mapping of anisotropic flow can be done better by horizontal and vertical resistivity profiling to map these geohydrologic anomalies. Electrokinetic surveys may in the future characterize adsorption, ion exchange, electrical charge properties in situ and have a large areal representation. Interpretive Techniques When all is said and done, then it’s time to do some real groundwater modeling using real data. There is no doubt that the theory and modeling effort have advanced much further than parameter characterization methods. After trying for more than 20 years, leaders in geohydrology admitted in the early 1990s that groundwater modeling does not work because of the general lack of real geohydrologic data. Hopefully, groundwater modeling will work in the next decade with real data on hand. BIBLIOGRAPHY 1. Freeze, A.R. and Cherry, J.A. (1979). Groundwater. Chapter 2, 9, Prentice Hall, New York, 384–462, pp. 1579. 2. RCRA Groundwater Monitoring Technical Enforcement Guidance Document. (1986). National Water Well Association. 3. Mercer, J.W. and Spalding, C.S. (1991). Site Characterization for Subsurface Remediation, EPA/625/4-91/026. 4. Driscoll, F.G. (1986). Groundwater and Wells. Glossary, Johnson Filtration Systems Inc., St. Paul, MN, pp. 885–892. 5. Todd, D.K. (1980). Groundwater Hydrology. In: Groundwater and Well Hydraulics. John Wiley & Sons, New York, pp. 111–163, Chapter 4. 6. Lohman, S.W. (1979). Ground-water Hydraulics. USGS Prof. Paper 708, p. 72. 7. Batu, V. (1998). Aquifer hydraulics. A Comprehensive Guide to Hydrogeologic Data Analysis. John Wiley & Sons, New York. 8. Ferris, Knowles, and Stallman. (1965). Theory of Aquifer Tests. USGS Water Supply Paper 1536-E, p. 174. 9. Loo, W.W., Wang, I.S., and McSpadden, W.R. (1984). Effective Porosity, Basalt Waste Isolation Project (BWIP). Richland, WA, Report no. SD-BWI-TI-254. 10. Loo, W.W. (1984). Practical Groundwater Modelling and Effective Porosity. Proceedings of Groundwater Modelling Conference, National Water Well Association. 11. Devinny, J.S. et al. (1990). Subsurface migration of hazardous wastes. In: Chemical and Physical Alteration of Wastes and Leachates. Van Nostrand Reinhold, New York, pp. 142–168, Chapter 5. 12. Myers, D. (1991). Principles and Applications of Surfaces, Interfaces and Colloids. VCH Publishers, New York. 13. Rosen, M.J. (1989). Surfactants and Interfacial Phenomena. John Wiley & Sons, New York.

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14. Murray, R.L. (1994). Understanding Radioactive Wastes. Battelle Press, Colombus, OH, pp. 11–22. 15. Lehr, J.H. (Ed.). (2001). Standard Handbook of Environmental Health, Science and Technology. Chapters 14.4 & 14.6 on Bioremediation & Electrokinetic Treatment of Hazardous Wastes by Walter W. Loo, McGraw-Hill, New York. 16. Heath, R.C. and Trainer, F.W. (1981). Introduction to Groundwater Hydrology. Chapter 2, 12, Water Well Journal Publishing Co., pp. 13–17, 157–170. 17. Gaudy, A.F. and Gaudy, E.T. (1980). Microbiology for Environmental Scientists and Engineers. McGraw-Hill, New York. 18. Walton, W.C. (1987). Groundwater Pump Tests Design and Analysis, Step Drawdown Test Analysis. Lewis Publishers, Boca Raton, FL, pp. 77–79. 19. Cooper, C.C. and Jacob, C.E. (1946). A generalized graphical method for evaluating formation constants and summarizing well field history. Trans. Amer. Geophysical Union 27: pp. 526–534. 20. Loo, W.W. (1987). Standard Operational Procedures of Pump Tests and Analysis. Certificate Seminar Manual for SUPERFUND and HAZMAT Conferences. 21. Loo, W.W. (1989–1995). Groundwater Hydrology for Beginners Certificate Seminar. HAZMAT & SUPERFUND Conferences. 22. Boulton, N.S. (1954). Unsteady State Radial Flow to a Pumped Well Allowing for Delayed Yield from Storage. International Assoc. Sci. Hydrology, General Assembly, Rome, vol. 2, Publ. 37. 23. Zhan, H., Zlotnik, V.A., and Park, E. (2001). Hydraulics Of Horizontal And Slanted Wells. In: Water Table Aquifers, GSA Annual Meeting, November 5–8, 2001 Proceedings, Session No. 117. 24. Hantush, M.S. (1956). Analysis of Data from Pumping Test in Leaky Aquifer. Trans. Amer. Geophysical Union 37: pp. 702–714. 25. Hantush and Thomas. (1966). Analysis of Data From Pumping Tests in Anisotropic Aquifers. Journal of Geophysical Research 71: No. 2 26. Papadopulos, I.S. (1965). Nonsteady flow to a well in an infinite anisotropic aquifer. Symp. Intern. Assoc. Sci. Hydrology, Dubrovinik Symposium 1: pp. 21–31. 27. Loo, W.W. (1989–1995). Groundwater Plume Management Certificate Seminar. SUPERFUND Conferences, and HAZMAT Conferences, Washington, DC. 28. Weeks, E.P. (1969). Determining the ratio of horizontal to vertical permeability by aquifer test analysis. Water Resources Research 5: 1. 29. Loo, W.W., Markley, D.E., and Dougan, P. (1979). Three Dimensional Multiple Well Testing and Reservoir Analysis of the Leached Zone of the Green River Formation. Equity/DOE BX In-situ Oil Shale Project, Proceedings of the Ninth Annual Rocky Mountain Groundwater Conference, Reno, NV, October 22, 1979. 30. Loo, W.W. (1989–1995). Principles of Groundwater Hydrology Certificate Seminar. HAZMAT Conferences. 31. Jacob, C.E. (1950). Flow of groundwater. In: Engineering Hydraulics. John Wiley & Sons, New York, Chapter 5. 32. Hantush, M.S. (1964). Hydraulics of Well. In: Advances in Hydroscience. Academic Press, New York. 33. Milsom, J. (1989). Field Geophysics, Electrical Methods. Geological Society of London Handbook Series. Open University Press and Halsted Press, pp. 72–106. 34. Vogelsang, D. (1995). Environmental Geophysics. Geoelectrical Methods. Springer-Verlag, New York, pp. 9–31.

35. Dragun, J. (1988). The Soil Chemistry of Hazardous Materials. Hazardous Materials Control Research Institute, pp. 75–262, Chapters 3–6.

MOBILITY OF HUMIC SUBSTANCES IN GROUNDWATER GUNNAR BUCKAU ¨ Nukleare Institut fur Entsorgung Karlsruhe, Germany

INTRODUCTION Humic substances are a relatively stable part of the global carbon inventory. They are found mainly in solid sources, in both seabed and land sediments. High concentrations are found especially in lignite and peat. The most obvious source is the organic/humus inventory in soil. Humic substances are also found in natural water. The dissolved inventory is relatively small compared to that found in sediments. Nevertheless, this inventory is a key in the mobilization of numerous trace elements and pollutants, which includes the key influence in determining the stability of and interface reactions with minerals. Some key information on the origin, stability, and mobility of dissolved aquatic humic substances is presented here. Also described are the experimental methods required, the approach for derivation of conclusions from experimental results, and background information. The focus is on the hydrophilic part of the total humic substance inventory. The hydrophilic aquatic humic substances originate from the same source as the stationary less hydrophilic ones, including partial oxidation from this stationary inventory. The hydrophilic nature is given by oxygen-containing functional groups. Some of these, especially carboxylic acids, may change the degree of hydrophilicity depending on physicochemical conditions, that is protonation, metal ion complexation and localization of counterions, pH, metal ion concentration, and ionic strength. Consequently, the whole inventory and changes in physicochemical conditions must be kept in mind when discussing the behavior of the hydrophilic aquatic part. Humic acid is used both as a collective term for humic substances with a sufficiently high content of hydrophilic groups to dissolve in pH neutral range or specifically for the fraction that is dissolved in pH neutral range and flocculates in the acidic range with the complementary fraction of fulvic acid, soluble also under acidic conditions. Furthermore, a fraction of organic carbon found in soil not soluble in aqueous media (humin) but with some properties common with humic and fulvic acids is also frequently considered part of the overall term. This term may also be considered for rather hydrophobic humic material or humic acid precursor material also in sediments, including peat, lignite, and mineral-bound natural organic substances, such as clay organic matter. Below, humic and fulvic acids are used for the aquatic hydrophilic part of the overall inventory unless otherwise specified.

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First, some basic relevant properties of humic substances are discussed. In order to study the origin, stability, and mobility of aquatic humic substances they must be isolated and purified. In order to ensure the quality of results, the success of isolation and purification must be verified by adequate characterization of the sampled humic material. Therefore, isolation, purification, and verification methods of aquatic humic substances are described as a prerequisite for interpretation of experimental results. Also discussed are the origin, stability, and mobility of aquatic humic acid followed by a summary. BASIC RELEVANT PROPERTIES OF HUMIC ACID The composition, mass distribution, and functional group content vary within limits, reflecting different origins and histories. Disregarding sulfur and nitrogen (present in varying low concentrations generally on the order of up to around 1% by weight), the atomic composition is dominated by carbon, oxygen, and hydrogen. The contributions of these substances vary around CO0.5 H; that is, the formal oxidation state of carbon is around zero. With respect to aquatic humic and fulvic acids, the mass distribution centers around approximately 500 mass units. In dissolved form, the molecules are highly hydrated and possibly form associates. The size distribution is generally determined to center slightly above 1 nm diameter. Oxygen-containing functional groups and structural entities are abundant, as seen from the elemental composition. These oxygen-containing structural elements are of key importance for the hydrophilic character of humic substances. The oxygen to carbon ratio of aquatic humic and fulvic acids is somewhat higher than for the bulk sedimentary hydrophobic humic matter. The proton exchange capacity is on the order of 7 meq/g, where protonation/deprotonation takes place from about pH 10 down into the acidic range (below pH 3 a considerable number of the groups are still ionized). There is clear spectroscopic evidence for both carboxylic and phenolic types of proton exchanging groups. Carboxylic types of groups are normally quantified with about two-thirds of the total capacity. The hydrophilic character and thus the stability in aqueous solution, of humic and fulvic acids, vary with pH, ionic strength, and metal ion complexation. ISOLATION AND PURIFICATION METHODS Isolation Isolation of humic and fulvic acids from natural water, including groundwater, is based on sorption chromatography. For this purpose, the XAD-8 resin is used where humic and fulvic acids sorb at low pH. Depending on the humic substance concentrations in the water sample, a pH of typically 1 or 2 will be used, the latter especially in order to limit the total amount of acid where the sample volumes are very large. Elution of both humic and fulvic acids is achieved by desorption at high pH. Typically, HCl and NaOH are used for adjustment of pH. In order to remove salt from the original sample, the column should be washed with HCl solution (pH not higher than that of

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the acidified water sample) prior to alkaline humic and fulvic acid elution. Where the humic acid concentration is sufficiently high, it, or at least a large part of it, will flocculate at low pH and thus is separated prior to sorption chromatography treatment. The alkaline elute solution should be taken to pH neutral range within a relatively short time (ideally after not more than around 10 min) in order to prevent alkaline hydrolysis and sample oxidation. If desired, the humic and fulvic acid fractions are separated from each other from this concentrate by acidification, where the humic acid fraction flocculates and may be separated by centrifugation. The so obtained humic and fulvic acid concentrates are then further processed and purified. In many cases, however, preconcentration is desired, especially directly in the field. The application of RO to humic and fulvic acid preconcentration makes use of the retention of these substances over the RO membrane. The retentate, containing salts, particles, and the humic and fulvic acids, is recirculated as long as practical with the clean water from penetration of the membrane being discarded. With increasing retentate salt concentration, the permeation of clean water through the RO membrane is decreased and finally ceases for all practical purposes (pressure buildup). Therefore, the level of preconcentration that can be obtained by RO depends on the salt content of the original water sample. Furthermore, where Ca ions and carbonate are present, calcite will precipitate unless the sample is slightly acidified and carbonic acid released. The concentrate obtained by RO is then treated by XAD-8 chromatography as described above, resulting in further sample concentration and removal of salt. Purification Purification has a number of objectives, including removal of organic and inorganic contaminants such as salt, complexed metal ions, and inorganic mineral constituents. For the purpose of dissolving Si-based minerals, NaF is added and the samples are left for typically 24 h. With respect to humic acid, subsequent purification is relatively simple. Humic acid is flocculated in HCl (pH 1) and centrifuged, and the supernatant is discarded, followed by dissolution in weak NaOH. The cycle is repeated until the flocculate is finally washed with HCl until no Na (from NaOH) is found. The sample is then in its protonated form and is brought to the final product by freeze-drying. The fulvic acid does not flocculate at low pH and thus the procedure is more tedious. The fulvic acid is sorbed/desorbed on XAD-8 in a number of cycles. Finally, the slightly alkaline solution is acidified and protonated by cation exchange chromatography. The protonated fulvic acid in its final form is then obtained by freeze-drying of this solution. The method described has successfully been applied for isolation of about 200 mg fulvic acid from about 10 m3 of groundwater. Subsequent analysis was successfully used for 14 C dating and determination of general characteristic properties (1). Details on the isolation and purification of humic and fulvic acids, including description of the RO technique, can be found in Artinger et al. (2).

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Verification Verification of the purity of humic and fulvic acids is done by characterization of the final products. It should also be borne in mind that trust in published results on the mobility of humic and fulvic acids requires evidence for successful separation and purification from the source. Characterization of humic and fulvic acids can be done by a broad spectrum of methods. For the purpose of study concerning the present topic, the quality of the samples needs to be demonstrated by a minimum of characterization methods. The elemental composition, the concentrations of inorganic constituents, and UV/Vis and IR spectra should be compared with expectation values from the literature. Inorganic constituents will be exhibited if inorganic minerals and complexed metal ions are removed to an acceptable level. The UV/Vis absorption increases uniformly with decreasing wavelength and the logarithm of the absorption should be close to linear with the wavelength. For fulvic acids the absorption ratio of 300–400 nm is expected to fall in the range between about 5 and 8 (abs. units/g and cm), whereas the values for humic acids are lower and are expected to fall around 2.5–4. These indicators show whether the UV/Vis absorbing carbon inventory is in agreement with humic and fulvic acids. IR spectroscopy shows a number of characteristic bands. Comparison to published spectra will reveal if there are considerable amounts of non-UV/Vis absorbing organic contaminants. Presence of such contaminants may also be identified by strong deviation from an atomic ratio around CO0.5 H. For the purpose of determining the origin and age of humic and fulvic acids, the 14 C concentration is essential. In this context the 13 C concentration is a good indicator for possible contaminants. The 13 C concentration should be close to −27‰ (rel. PDB) for C-3 plant cycle origin and around −13‰ for C-4 plant cycle origin. Again, only if a sufficient purity of the isolated substances is demonstrated can the results be considered trustworthy. ORIGIN, STABILITY, AND MOBILITY OF AQUATIC HUMIC ACID Determination of the origin, stability (with respect to decomposition and sorption), and mobility of aquatic humic and fulvic acids in groundwater is based on several indicators. The origin can be determined based on 14 C content, functional entity distribution, and general information on climatic conditions, vegetation history for soil recharge, and sedimentary and groundwater composition. Another important indicator is the cogeneration with dissolved inorganic carbon (DIC) of biogenic origin, where deviations from such a correlation show that either the DIC or the humic and fulvic acid inventory deviates from ideal tracer transport and stability behavior. Origin The principal origin of humic and fulvic acids is plant material. In sea sediments, marine organisms may also

play a role. From the viewpoint of aquatic humic and fulvic acids in a groundwater, two principal sources may be distinguished. These are introduction from the soil zone with recharge groundwater and in situ generation by conversion of organic sediment material. In the former case, the source term will depend on climatic conditions and the type of vegetation, including extensive vegetation versus, for example, intense modern agriculture and land conditions such as wetland and peat deposits. In the case of in situ generation, an oxidizing agent and partial oxidation of the hydrophobic source material by microbial activity are required (3,4). The different origin can be distinguished by some characteristic properties and especially the 14 C content (5). In aquifer systems where the groundwater residence time is sufficiently low, compared to the half-life of 14 C (57,300 years), the 14 C content of recharge humic and fulvic acids is that of the soil recharge source. Numbers deduced for fulvic acid for conditions prior to nuclear atmospheric testing are around 55 pmc (percent modern carbon) (5), basically the same as for the cogenerated dissolved inorganic carbon of biogenic origin (6). Depending on local conditions, however, this value may vary. One example is modern agriculture, where the turnover of the organic soil inventory may be higher and thus the 14 C concentration is higher, reflecting the lower average residence time. The 14 C concentration of fulvic acid and cogenerated DIC from in situ generation is basically zero. The relative fractions of in situ generated and recharge originating fulvic acid can be deduced by the overall 14 C concentration (5). Another important indicator for the origin of aquatic humic and fulvic acids is the concentration varying with recharge conditions. Drainage of wetland results in a decrease in the inflow of both aquatic humic and fulvic acids and the cogenerated dissolved inorganic carbon of biogenic origin (7), which is an indicator of changes in the land use and also of changes in climatic conditions with a drastic impact on land structure and vegetation. Recharge humic and fulvic acids can only be found in water originating from areas with vegetation. For this reason, the absence of humic and fulvic acids is an indicator for vegetation free recharge conditions, also indicative of climatic changes. In one study 14 C groundwater dating on the DIC indicated ages well beyond 15,000 years for the most distant part of the groundwater flow system (1). Not only did the 14 C concentration of fulvic acid show that these age determinations were wrong, but also the pure presence of fulvic acid showed that the groundwater at the end of the flow path could not be older than 15,000 years. The reason is that, in this area, vegetation started to develop around 15,000 years ago at the decline of the Pleistocene conditions. Some characteristic properties of aquatic humic and fulvic acids reflect their origin. Both spectroscopic properties and basic functional entity distribution are sufficiently insensitive to the physicochemical environment to retain origin-related properties, which is especially true for the 14 C content, which solely depends on the concentration upon their generation and subsequent decay with a halflife of approximately 5730 years. Extraction of hydrophilic

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humic and fulvic acids from clay sediments revealed that the distribution of functional entities and photodynamic behavior vary with marine and terrestrial origin of the clay sediments (8,9). Other characteristic properties reflect chemical reactions of functional groups, especially redox reactions. It should thus be kept in mind that not all characteristic properties can be used as indicators for the origin of aquatic humic or fulvic acid. The content and redox state of sulfur functional entities partly reflect the redox conditions and presumably also the microbial sulfate reduction as part of the in situ generation process (5).

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subsequent introduction of hydrophilic aquatic humic and fulvic acids will behave differently and remain in solution rather than competing with sorbed hydrophobic/sediment bound organic material. Care must therefore be taken when analyzing experimental data that the real situation is reflected. Strong sorption may be expected where a recent increase in the inflow of humic acid/fulvic acid is given, such as a deposit leaking organic material into sediment previously in contact with very low organic carbon concentrations. Under stable groundwater conditions, no indication is given for retention or decomposition of humic and fulvic acids from recharge.

Stability and Mobility In groundwater systems where the in situ generation is negligible, the fulvic acid inventory is given solely by recharge and possible follow-up geochemical reactions. Where the residence time is considerable compared to the half-life of 14 C, the age of the fulvic acid can be determined. A linear increase in 14 C age of fulvic acid with flow distance from recharge was found, with the maximum age of fulvic acid of about 15,000 years (cf. above and Ref. 1). The agreement with groundwater flow velocity calculations and the age of fulvic acid shows that the fulvic acid has an ideal tracer transport behavior over this time period, which is direct evidence for the stability and mobility with an ideal tracer behavior of aquatic fulvic acid over 15,000 years. In the case of in situ generation of fulvic acid, the quantification of the recharge originating inventory by 14 C content and supporting spectroscopic and composition information is required. The outcome of one study is that in situ generated fulvic acid is flocculated in deeper brines, whereas the recharge originating fulvic acid remains stable in the brines. Both the conserved 14 C concentration and structural entity content by C-XANES show that the fulvic acid in the deeper brines (200–250 m depth) originates from recharge. The flocculation of in situ generated fulvic acid in the high ionic strength brines is shown to be the result of their less hydrophilic nature compared to the recharge fulvic acid. The similar concentrations in recharge and in the deep brines show the stability and mobility of these fulvic acids over at least several hundreds years and, furthermore, that the recharge conditions have not changed dramatically within this time period. In addition, the close correlation between the DIC of biogenic origin and humic and fulvic acids in the broad spectrum of groundwater samples shows that there is no considerable decomposition or retention of fulvic acid from either recharge or in situ generation. This correlation is also found for deep groundwater from a former wetland. This shows also that the elevated concentration from this previous source is not subject to sorption or decomposition but the dissolved humic and fulvic acids remain stable and mobile over long time periods (7). Studies on clay organic matter show that the organic matter sorbed on the sediments are mainly of hydrophobic character and are strongly bound to mineral surfaces. Studies where humic or fulvic acid is in contact with fresh mineral surfaces frequently show strong sorption. The natural situation, however, is that active sites on sediments are coated with organic substances and

CONCLUSION With respect to the origin of aquatic humic and fulvic acids, a large number of studies have shown that they are cogenerated with inorganic carbon of biogenic origin from microbial processes and that there is a strong variation in the concentration of humic and fulvic acids in natural groundwater—an indicator for recharge conditions, including climatic changes. With respect to the stability and mobility, there is no indication for decomposition of aquatic humic and fulvic acids under natural groundwater conditions for as long as 15,000 years and no indication for their retardation. An exception is the selective flocculation of less hydrophilic in situ generated fulvic acid in high ionic strength brines. BIBLIOGRAPHY 1. Buckau, G. et al. (2000). Development of climatic and vegetation conditions and the geochemical and isotopic composition in the Franconian Albvorland aquifer system. Appl. Geochem. 15(8): 1191–1201. 2. Artinger, R. et al. (2000). Characterization of groundwater humic substances: influence of sedimentary organic carbon. Appl. Geochem. 15(1): 97–116. 3. Buckau, G. et al. (2000). Groundwater in-situ generation of aquatic humic and fulvic acids and the mineralization of sedimentary organic carbon. Appl. Geochem. 15(6): 819–832. 4. Buckau, G. et al. (2000). Origin and mobility of humic colloids in the Gorleben aquifer system. Appl. Geochem. 15(2): 171–179. ¨ 5. Schafer, T., Buckau, G., Artinger, R., Wolf, M., Kim, J.I., Geyer, S., Bleam, W.F., Wirick, S., and Jacobsen, C. (2004). Vertical exchange of Gorleben fulvic acids of different origin, Annex 1. In: Humic Substances in Performance Assessment of Nuclear Waste Disposal: Actinide and Iodine Migration in the Far-Field (Second Technical Progress Report). G. Buckau (Ed.). Report FZKA 6969, Research Center, Karlsruhe, Germany. 6. Buckau, G. et al. (2000). 14 C Dating of Gorleben groundwater. Appl. Geochem. 15(5): 583–597. 7. Buckau, G., Wolf, M., Geyer, S., Artinger, R., and Kim, J.I. (2003). Origin and mobility of aquatic humic substances from wetland recharge in the Gorleben aquifer system. In: Humic Substances in Performance Assessment of Nuclear Waste Disposal: Actinide and Iodine Migration in the Far-Field (First Technical Progress Report). G. Buckau (Ed.). Report FZKA 6800, Research Center, Karlsruhe, Germany. ¨ 8. Claret, F., Schafer, T., Rabung, Th., Bauer, A., Wolf, M., and Buckau, G. (2004). Complexation properties of humic and

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ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

fulvic acids extracted from Callovo-Oxfordian and opalinus clay, Annex 23. In: Humic Substances in Performance Assessment of Nuclear Waste Disposal: Actinide and Iodine Migration in the Far-Field (Second Technical Progress Report). G. Buckau (Ed.). Report FZKA 6969, Research Center, Karlsruhe, Germany. 9. Claret, F. et al. (2005). Differences in characteristic properties and Cm(III) complexation between terrestrial and marine origin of humic and fulvic acids from Callovo-Oxfordian and opalinus clay. Appl. Geochem.

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA CHAKRESH K. JAIN National Institute of Hydrology Roorkee, India

The groundwater quality of District Hardwar in the state of Uttaranchal (India) has been assessed to see the suitability of groundwater for domestic use. Forty-eight groundwater samples from shallow and deep aquifers were collected each during pre- and postmonsoon seasons during the year 2002. Various water quality constituents, pH, conductance, total dissolved solids, alkalinity, hardness, sodium, potassium, calcium, magnesium, chloride, sulphate, phosphate, fluoride, total coliforms, and fecal coliforms were determined. The data were analyzed with reference to BIS and WHO standards, and hydrochemical facies were determined. The concentration of total dissolved solids exceeded the desirable limit of 500 mg/L in about 25% of the samples analyzed, but the values were well within the maximum permissible limit of 2000 mg/L. The alkalinity exceeded the desirable limit of 200 mg/L in about 50% of the samples, but these were also within the maximum permissible limit of 600 mg/L. From the hardness viewpoint, about 80–85% of the samples were within the desirable limits. One sample of the study area exceeded the maximum permissible limit of 100 mg/L for nitrate. Other constituents such as chloride, sulphate, and fluoride were within the desirable limits. The bacteriological analysis of the groundwater samples indicated bacterial contamination in about 40% of the samples analyzed. Inadequate maintenance of hand pumps, improper sanitation, and unhygienic conditions around the structure may be responsible for bacterial contamination in groundwater of the region and is a cause of concern. It is recommended that the water drawn from such sources be properly disinfected before being used for drinking and other domestic purposes. The grouping of samples according to their hydrochemical facies indicates that all the samples of the study area fall under Ca–Mg–HCO3 hydrochemical facies. INTRODUCTION Water is an essential and vital component of our life support system. In tropical regions, groundwater plays an

important role in the context of fluctuating and increasing contamination of water resources. Groundwater has unique features, which render it particularly suitable for public water supply. It has excellent natural quality; is usually free from pathogens, color and turbidity; and can be consumed directly without treatment. Groundwater is widely distributed and can be frequently developed incrementally at points near water demand, thus avoiding the need for large-scale storage, treatment, and distribution systems. It is particularly important because it accounts for about 88% of safe drinking water in rural areas, where the population is widely dispersed and the infrastructure for treating and transporting surface water does not exist. Unfortunately, the availability of groundwater is not unlimited, nor it is protected from deterioration. In most instances, extracting excessive quantities of groundwater has resulted in drying of wells, damaged ecosystems, land subsidence, saltwater intrusion, and depletion of the resource. It has been estimated that once pollution enters the subsurface environment, it may remain concealed for many years, becoming dispersed over wide areas of groundwater aquifer and rendering groundwater supplies unsuitable for consumption and other uses. The rate of depletion of groundwater levels and deterioration of groundwater quality are of immediate concern in major cities and towns of the country. The creation of the new state of Uttaranchal has posed many challenges for planners and policy makers. Problems such as drinking water, transportation, power, housing and construction, and safety against natural hazards are very serious and require immediate attention. For sustainable development of a society, it is essential that the natural resources are used judiciously for the benefit of the existing population and also to meet the needs and aspirations of future generations. Drinking water is one such precious commodity for which a planned strategy is needed for immediate demands and also for sustainability for future needs. A large part of the state of Uttaranchal lies in the hills, where distribution of drinking water supply and its quality is a major problem that needs immediate attention. About 90% of the rural population of this region depends on natural springs for their daily water. However, due to population pressure, unplanned construction, garbage disposal, and changes in land use patterns, the water of these springs is becoming contaminated, and the discharge of these springs is declining. A wide number of activities are associated with, the human introduction of foreign chemical and biological materials into the subsurface environment. In the long run, the most potentially hazardous of these may be the chemical fertilizers and pesticides used in agriculture. But it is possible that tremendous use of chemical fertilizers as plant nutrients may be a more significant problem, causing an increasing buildup of nutrients in some groundwaters. Bacteriological parameters are of great importance from the human point of view. It is essential to examine the presence of toxic substances and pathogenic organisms in potable water. Experience has established the significance of coliform group density as a criterion of the degree of pollution and thus of sanitary quality. The significance

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

of the various tests and the interpretation of results are well authenticated and have been used as a basis for standards of the chemical and bacteriological quality of water supplies. In this article, the groundwater quality of District Hardwar has been assessed to see the suitability of groundwater for drinking, which will provide a proper basis for judicial management of drinking water supplies in the state.

census, the population of District Hardwar is 14,44,213, a population density of 612 per km2 . Physiographically, the area is generally flat except for the Siwalik Hills in the north and northeast. The area is devoid of relief features of any prominence except for deep gorges cut by gullies and rivers flowing through the area. The area is bounded by River Yamuna in the west and River Ganga in the east. The climate of the area is characterized by moderate subtropical monsoons. The average annual rainfall in the region is about 1000 mm; the major part is received during the monsoon period. The major land use is agriculture, and there is no effective forest cover. The soils of the area are loam to silty loam and are free from carbonates. The most common groundwater use is achieved by hand pumps and tube wells. Based on the lithologic logs and water table fluctuation data,

STUDY AREA District Hardwar, part of the Indo-Gangetic plains, lies between latitude 29◦ 30 and 30◦ 20 N and longitude 77◦ 40 to 78◦ 25 E in the state of Uttaranchal (Fig. 1). It is the largest district (in population) of Uttaranchal State and occupies an area of about 2360 km2 . Per the 2001

N

30° 15′

ad hr

De

1

un dis

3

ur district

44 8

43

45 46

6 7

48

18

15

27

.

25 Laksar

0

5

Scale

10 km 77° 45′

38

24 20

zaf fa dist rnaga rict r

36

23

Mu

Tehsil

35

22

19

Sampling sites

t

References

37

26

16 17

c tri

13

km 5

dis

28

14

34

32

29

11

District H/Q

33

31 41

12

l wa rh

Roorkee

9 10

47 . Hardwar 40 39 30

42

Ga

Saharanp

0′

ct

30°

tri

2 4 5

193

o

jn

Bi 78° 0′

i rd

t

ric

st

21

78° 15′

Figure 1. Study area showing location of sampling sites, District Hardwar.

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ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

two types of aquifers have been delineated in the area. The upper one is the shallow unconfined aquifer which generally extends to depths of around 25 m. The deeper one consists of semiconfined aquifers located at a depth of about 25 to 150 m below ground level separated by three to four aquifers at average depths of 25 to 55, 65 to 90, and 120 to 150 m. Water table contours in the area indicate a southward trend of groundwater flow in both unconfined and confined aquifers.

colonies so as to obtain a good areal and vertical representation. The samples were preserved by adding an appropriate reagent (1,2). The hand pumps and tube wells were continuously pumped prior to sampling to ensure that the groundwater to be sampled was representative of the groundwater aquifer. The water samples for bacteriological analysis were collected in sterilized highdensity polypropylene bottles covered with aluminum foil. All samples were stored in sampling kits maintained at 4 ◦ C and brought to the laboratory for detailed chemical and bacteriological analysis. The details of sampling locations and source and depthwise distribution are given in Tables 1 and 2, respectively. The physicochemical analysis was performed following standard methods (1,2). The total coliforms and fecal coliforms were determined by the multiple tube fermentation technique using MacConkey broth and EC medium, respectively.

EXPERIMENTAL METHODOLOGY Forty-eight groundwater samples from District Hardwar were collected each during pre- (June 2002) and postmonsoon (October 2002) seasons from various abstraction sources at various depths covering extensively populated area, commercial, industrial, agricultural, and residential

Table 1. Description of Groundwater Sampling Locations in District Hardwar S. No.

Location

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24

Mohand Banjarewala Buggawala Kheri Dadapatti Bahbalpur Sikandarpur Bhagwanpur Chudiala Balswa Ganj Manakpur Iqbalpur Jharera Sherpur Narsen Manglour Libarheri Mahesari Sahipur Khanpur Chandpuri Kalan Laksar Kalsiya Niranjanpur

Source

Depth, m

S. No.

OW OW HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP

10 10 30 38 38 15 25 30 40 25 40 30 30 30 30 35 35 20 35 15 15 15 10 15

25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48

Location Sultanpur Shahpur Pathri Subashgarh Marghubpur Bahadarabad Alipur Katarpur Kankhal Shyampur Rasiya Garh Gandikhatta Laldhang Kottawali Hardwar Jwalapur Roorkee Gumanwala Manubas Bandarjud Beriwala Hazara Aurangbad Daulatpur

Source

Depth, m

HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP HP

15 25 15 20 30 20 20 10 25 30 40 10 90 40 30 30 10 30 10 15 5 10 10 35

OW: Open well. HP: Hand pump.

Table 2. Source and Depthwise Distribution of Sampling Sites in District Hardwar Depth Range Source Structure Hand pumps

Tube wells Open wells Total

40 m

Total Number

6,18,20,21,22, 23,24,25,27,28, 30,31,32,36,41, 43,44,45,46,47

3,4,5,7,8,9, 10,11,12,13, 14,15,16,17,19, 26,29,33,34,35, 38,39,40,42,48 – – 25

37

46

– – 1

– 2 48

– 1,2 22

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

RESULTS AND DISCUSSION During 1983, the Bureau of Indian Standards (BIS), earlier known as Indian Standards Institution (ISI), laid down standard specifications for drinking water which have been revised and updated from time to time. To enable the users to exercise their discretion toward water quality criteria, the maximum permissible limit was prescribed especially where no alternate source was available. The national water quality standards describe the essential and desirable characteristics that must be evaluated to assess the suitability of water for drinking (3). The hydrochemical data for the two sets of samples collected from District Hardwar during pre- and postmonsoon seasons are presented in Table 3. General Characteristics The pH of the groundwater of District Hardwar is mostly confined within the range 6.22 to 7.58 during the premonsoon season and 6.70 to 7.70 during the postmonsoon season. The pH values of all samples are well within the limits prescribed by the BIS (3) and the WHO (4) for various uses, including drinking and other domestic supplies. The measurement of electrical conductivity is directly related to the concentration of ionized substances in water and may also be related to excessive hardness and/or other mineral contamination. The conductivity values in the groundwater samples of District Hardwar vary from 233–1440 µS/cm during the premonsoon season and from 221–1442 µS/cm during the postmonsoon season; about 10% of the samples had conductivity values above 1000 µS/cm during both pre- and postmonsoon seasons. The maximum conductivity of 1440 and 1442 µS/cm was observed at village Manubas (hand pump, 10 m depth) during pre- and postmonsoon season, respectively. In natural waters, dissolved solids consists mainly of inorganic salts such as carbonates, bicarbonates, chlorides, sulfates, phosphates, and nitrates of calcium, magnesium, sodium, potassium, iron, etc. and small amounts of organic

Table 3. Hydrochemical Data for Groundwater Samples from District Hardwara Characteristics pH Conductivity, µS/cm TDS, mg/L Alkalinity, mg/L Hardness, mg/L Chloride, mg/L Sulphate, mg/L Nitrate, mg/L Phosphate, mg/L Fluoride, mg/L Sodium, mg/L Potassium, mg/L Calcium, mg/L Magnesium, mg/L Boron, mg/L a

Min

Max

Average

6.22 (6.70) 233 (221) 149 (141) 78 (71) 81 (80) 0.1 (1.2) 0.5 (0.7) 0.1 (0.1) 0.01 (0.02) 0.01 (0.01) 5.0 (4.5) 0.8 (0.2) 22 (24) 5.0 (5.0) 0.12 (0.12)

7.58 (7.70) 1440 (1442) 922 (923) 460 (482) 464 (427) 32 (35) 72 (62) 140 (130) 0.41 (1.60) 0.94 (0.88) 69 (71) 42 (32) 140 (135) 36 (38) 0.92 (0.87)

6.87 (7.10) 642 (647) 411 (414) 213 (215) 209 (212) 9.0 (9.0) 22 (21) 13 (13) 0.03 (0.25) 0.36 (0.42) 26 (25) 7.0 (7.4) 55 (56) 18 (17) 0.50 (0.49)

Values given in parenthesis represent postmonsoon data.

195

matter and dissolved gases. In the present study, the values of total dissolved solids (TDS) in the groundwater varied from 149–922 mg/L during the premonsoon season and from 141–923 mg/L during the postmonsoon season, indicating low mineralization in the area. More than 75% of the samples analyzed were within the desirable limit of 500 mg/L, and about 25% of the samples were above the desirable limit but within the maximum permissible limit of 2000 mg/L. An almost similar trend was observed during the postmonsoon season. The TDS content at deeper levels (>40 m depth) is comparatively low and lies well within the desirable limit of 500 mg/L. The TDS distribution maps for the pre- and postmonsoon seasons are shown in Fig. 2a,b. Water containing more than 500 mg/L of TDS is not considered desirable for drinking water, though more highly mineralized water is also used where better water is not available. For this reason, 500 mg/L as the desirable limit and 2000 mg/L as the maximum permissible limit have been suggested for drinking water (3). Water containing more than 500 mg/L TDS causes gastrointestinal irritation (3). No sample of District Hardwar exceeded the maximum permissible limit of 2000 mg/L. Carbonates, bicarbonates, and hydroxides are the main cause of alkalinity in natural waters. Bicarbonates represent the major form because they are formed in considerable amounts by the action of carbonates upon the basic materials in the soil. The alkalinity in the groundwater varies from 78–460 mg/L during the premonsoon season and from 71–482 mg/L during the postmonsoon season. About 50% of the samples of the study area fall within the desirable limit of 200 mg/L both during the pre- and postmonsoon seasons, and the remaining 50% of the samples exceeds the desirable limit but are within the maximum permissible limit of 600 mg/L. No sample of the study area exceeded the maximum permissible limit of 600 mg/L. The high alkalinity may be due to the action of carbonates upon the basic materials in the soil. Calcium and magnesium along with their carbonates, sulfates, and chlorides make the water hard. A limit of 300 mg/L has been recommended for potable water (3). The total hardness values in the study area range from 81–464 mg/L during the premonsoon season and from 80–427 mg/L during the postmonsoon season. About 80% of the samples of the study area fall within the desirable limit of 300 mg/L and the remaining samples exceed the desirable limit but are well within the maximum permissible limit of 600 mg/L. From the point of view of hardness all samples of District Hardwar were within the permissible limit of 600 mg/L. The desirable limits for calcium and magnesium for drinking water are 75 and 30 mg/L, respectively (3). In the groundwater of the study area, the values for calcium and magnesium range from 22–140 mg/L and 5.0–36 mg/L, respectively, during the premonsoon season. An almost similar trend was observed during the postmonsoon season. In groundwater, the calcium content generally exceeds the magnesium content in accordance with their relative abundance in rocks. The increase in magnesium is proportionate to calcium in both seasons. All the

196

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

30.2

1

0

45

45

0

2 3

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45

0

45

8 43

6

30

46 47

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450

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0

0

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0

25

23

35

20

21 29.6 77.8

77.9

78

78.1

78.2

78.3

Figure 2. (a) Distribution of TDS in groundwater from District Hardwar (premonsoon 2002). (b) Distribution of TDS in groundwater from District Hardwar (postmonsoon 2002).

samples of the study area fall within the desirable limit at most places. The concentration of sodium in the study area varied from 5.0–69 mg/L during the premonsoon season and from 4.5–71 mg/L during the postmonsoon season. The violation of BIS limits could not be ascertained for sodium as no permissible limit of sodium has been prescribed in BIS drinking water specifications. Groundwater high in sodium is not suitable for irrigation due to the sodium sensitivity of crops/plants. The concentration of potassium in the groundwater of District Hardwar varied from 0.8–42 mg/L during the premonsoon season and from 0.2–32 mg/L during the

postmonsoon season. Potassium, an essential element for humans, plants, and animals, is derived in the food chain mainly from vegetation and soil. The main sources of potassium in groundwater include rainwater, weathering of potash silicate minerals, use of potash fertilizers, and use of surface water for irrigation. It is more abundant in sedimentary rocks and commonly present in feldspar, mica, and other clay minerals. The Bureau of Indian Standards has not included potassium in drinking water standards. However, the European Economic Community has prescribed a guideline level of 10 mg/L potassium in drinking water. Per EEC criteria, about 10–15% of the samples of the study area exceeded the 10 mg/L

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

30.2

1

0

50

400

2 3

30.1 4

44

5

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300

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43

6

47

6 50 00 0

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400

48 4 00

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33

31 400 34

41 32

0

50

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11 13 29.8

28

500

16

7

27 26

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400 35

36

400

14300

25

17

38

18 200

15

22 24

19

0

0

30

23

50

20

40

0

0

60

29.7

21 29.6 77.8

77.9

78

78.1

78.2

78.3

Figure 2. (Continued)

guideline level. Though potassium is found extensively in some igneous and sedimentary rocks, its concentration in natural waters is usually quite low because potassium minerals offer resistance to weathering and dissolution. A higher potassium content in groundwater is indicative of groundwater pollution. The concentration of chloride in the study area is quite low and varies from 0.1–32 mg/L during the premonsoon season. An almost similar trend was observed during the postmonsoon season. The limits of chloride have been laid down primarily from taste considerations. A limit of 250 mg/L chloride has been recommended as a desirable limit for drinking water supplies (3,4). However, no adverse health effects on humans have been reported from intake of waters containing an even higher chloride

content. No sample in the study area exceeded the desirable limit of 250 mg/L. The sulfate content of groundwater generally occurs as soluble salts of calcium, magnesium, and sodium. The sulfate content changes significantly with time during infiltration of rainfall and groundwater recharge, which takes place mostly from stagnant water pools and surface runoff water collected in low-lying areas. The concentration of sulfate in the study area varied from 0.5–72 mg/L during the premonsoon season and from 0.7–62 mg/L during the postmonsoon season. It is clearly evident from the distribution maps that all the samples from District Hardwar fall within the desirable limit of 200 mg/L prescribed for drinking water supplies.

198

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

permissible limit of 100 mg/L during both pre- and postmonsoon seasons. The nitrate distribution maps for the pre- and postmonsoon seasons are shown in Fig. 3a,b. The higher level of nitrate at Jwalapur may be attributed to improper sanitation and unhygienic conditions around the structure. Nitrate is an effective and moderately toxic plant nutrient. A limit of 45 mg/L has been prescribed by the WHO (4) and the BIS (3) for drinking water. Its

Excess nitrate content in drinking water is considered dangerous for its adverse health effects. The occurrence of high levels of nitrate in groundwater is a prominent problem in many parts of the country. The nitrate content in District Hardwar varies from 0.1–140 mg/L during the premonsoon season and from 0.1–130 mg/L during the postmonsoon season. About 95% of the samples shows nitrate content less than the desirable limit of 45 mg/L. Only one sample from Jwalapur exceeded the maximum

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4 44 5

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17 18 15

22

0

20

24

19

0

20

20

29.7 23

21 29.6 77.8

77.9

78

78.1

78.2

Figure 3. (a) Distribution of nitrate in groundwater from District Hardwar (premonsoon 2002). (b) Distribution of nitrate in groundwater from District Hardwar (postmonsoon 2002).

78.3

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

30.2

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15

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30

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19 23

21 29.6 77.8

77.9

78

78.1

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Figure 3. (Continued)

concentration above 45 mg/L may prove detrimented to human health. In higher concentrations, nitrate may produce a disease known as methemoglobinemia (blue babies) which generally affects bottle-fed infants. Repeated heavy doses of nitrates by ingestion may also cause cancer. The concentration of phosphate in the study area is generally low at all locations. Phosphorous is an essential plant nutrient that is used extensively as a fertilizer. Phosphate is adsorbed or fixed as aluminium or iron phosphate in acidic soils or as calcium phosphate in alkaline or neutral soils; as a result, the concentration of phosphate in groundwater is usually low, but various chemical processes in soil strata may induce the mobility of phosphate in sub-soil and groundwater.

The fluoride content in the groundwater of District Hardwar varies from 0.01–0.94 mg/L during the premonsoon season and from 0.01–0.88 mg/L during the postmonsoon season; it lies well below the desirable limit of 1.0 mg/L in all samples. The fluoride distribution maps for pre- and postmonsoon seasons are shown in Fig. 4a,b. The presence of fluoride in groundwater may be attributed to the localized effects of natural sources. Fluoride present in soil strata is from geological formations, such as fluorspar, and fluorapatite, and amphiboles, such as hornblende, tremolite, and mica. Weathering of igneous and sedimentary alkali silicate rocks, especially shales, contribute a major portion of fluorides to groundwaters. In addition to natural sources, considerable amounts of fluoride may be contributed by human activities. Fluoride

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

30.2

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0 .4

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3 0.

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3 27 0.

28 13

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6 0. 0.5

7

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0.

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0.

0.

0.2

3 0.4 0.5 0.6

0.1

21 29.6 77.8

77.9

78

78.1

78.2

78.3

Figure 4. (a) Distribution of fluoride in groundwater from District Hardwar (premonsoon 2002). (b) Distribution of fluoride in groundwater from District Hardwar (postmonsoon 2002).

salts are commonly used in the steel, aluminum, brick, and tile industries. Fluoride containing insecticides and herbicides may be contributed through agricultural runoff. Phosphatic fertilizers, which are extensively used, often contain fluorides as impurities, and these may increase the levels of fluoride in soil. The accumulation of fluoride in soil eventually results in leaching it by percolating water, thus increasing the fluoride concentration in groundwater. The study has clearly indicated that the concentration of total dissolved solids exceeds the desirable limit of 500 mg/L in about 25% of the samples analyzed, but the values are well within the maximum permissible limit of 2000 mg/L in all samples. The alkalinity exceeds the

desirable limit of 200 mg/L in about 50% of the samples, but the levels are well within the maximum permissible limit of 600 mg/L. Total hardness exceeds the desirable limit of 300 mg/L in about 15% of the samples. The nitrate content exceeds the maximum permissible limit of 100 mg/L in only one sample (Jwalapur) of the study area. The fluoride content is well within the desirable limit in all samples analyzed. Bacteriological Parameters The coliform group of bacteria is the principal indicator of the suitability of water for domestic, industrial, and other uses. The density of the coliform group is the criterion for

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

30.2

0.

201

1 0. 4

3

0.4

2 3

30.1

0.4 4 44

8

42

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27 0.

4 26

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9

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22 0.4 19

24

29.7 23

20

0.5 21 29.6 77.8

77.9

78

78.1

78.2

78.3

Figure 4. (Continued)

the degree of contamination and has been the basis for the bacteriological water quality standard. In ideal conditions, all samples taken from the distribution system should be free from coliform organisms, but in practice, this is not always attainable, and therefore, the following standard for water has been recommended (4): — 95% of water samples should not contain any coliform organisms in 100 mL throughout the year. — No water sample should contain E. coli in 100 mL water. — No water sample should contain more than 10 coliform organisms per 100 mL.

— Coliform organisms should not be detected in 100 mL of any two consecutive water samples. However, from bacteriological considerations, the objectives should be to reduce the coliform count to less than 10 per 100 mL, and more importantly, the absence of fecal coliform should be ensured. The presence of coliforms in water is an indicator of contamination by human or animal excrement. The presence of fecal colifirms in groundwater indicates a potential public health problem because fecal matter is a source of pathogenic bacteria and viruses. Groundwater contamination from fecal coliform bacteria is generally caused by percolation from sources of contamination (domestic sewage and

202

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

septic tanks) into the aquifers and also from poor sanitation. Shallow wells are particularly susceptible to such contamination. Indiscriminate land disposal of domestic waste on the surface and improper disposal of solid waste, and leaching of wastewater from landfill areas further increase the chances of bacterial contamination in groundwater. The results of the bacteriological analysis of groundwater samples from District Hardwar are given in Table 4.

Bacteriological analysis of the groundwater samples collected from District Hardwar indicates bacterial contamination in about 40% of the samples analyzed. About 20% of the samples even exceed the permissible limit of 10 coliforms per 100 mL of sample. Inadequate maintenance of hand pumps, improper sanitation, and unhygienic conditions around structures may be responsible for bacterial contamination in the groundwater of the region and is a cause of concern. The water from such sources should be

Table 4. Bacteriological Contamination in Groundwater from District Hardwar Premonsoon 2002

S. No.

Location

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48

Mohand Banjarewala Buggawala Kheri Dadapatti Bahbalpur Sikandarpur Bhagwanpur Chudiala Balswa Ganj Manakpur Iqbalpur Jharera Sherpur Narsen Manglour Libarheri Mahesari Sahipur Khanpur Chandpuri Kalan Laksar Kalsiya Niranjanpur Sultanpur Shahpur Pathri Subashgarh Marghubpur Bahadarabad Alipur Katarpur Kankhal Shyampur Rasiya Garh Gandikhatta Laldhang Kottawali Hardwar Jwalapur Roorkee Gumanwala Manubas Bandarjud Beriwala Hazara Aurangbad Daulatpur

Postmonsoon 2002

Total Coliform per 100 mL

Fecal Coliform per 100 mL

Total Coliform per 100 mL

Fecal Coliform per 100 mL

Nil Nil 4 9 23 75 7 23 Nil 9 14 28 15 Nil 4 4 Nil 4 Nil Nil Nil 43 23 4 4 Nil Nil 4 Nil 4 Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil 23 Nil Nil Nil

Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil

Nil Nil Nil Nil Nil 4 Nil Nil Nil Nil Nil 4 Nil Nil Nil Nil Nil Nil Nil Nil 4 Nil 15 Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil

Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil Nil

ASSESSMENT OF GROUNDWATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

properly disinfected before being used for drinking and other domestic purposes. Classification of Ground Water The groundwater of District Hardwar has been classified per Chadha’s diagram (5). The diagram is a somewhat modified version of the Piper trilinear diagram (6). In the Piper diagram, the milliequivalent percentages of the major cations and anions are plotted in two base triangles, and the type of water is determined on the basis of the position of the data in the respective cationic and anionic triangular fields. The plottings from the triangular fields are projected further into the central diamond field,

203

which represents the overall character of the water. A Piper diagram allow comparisons among numerous analyses, but this type of diagram has a drawback, as all trilinear diagrams do, in that it does not portray actual ion concentration. The distribution of ions within the main field is unsystematic in hydrochemical process terms, so the diagram lacks a certain logic. This method is not very convenient when plotting a large volume of data. Nevertheless, this shortcoming does not lessen the usefulness of the Piper diagram in representing some geochemical processes. In contrast, in Chadha’s diagram, the difference in milliequivalent percentage between alkaline earths

100 80 60

(HCO3)−(CI + SO4)

40 20

−100

−80

−60

−40

0

−20

0

20

40

60

80

100

−20 −40 −60 −80 −100 (Ca + Mg)−(Na + K)

100 80 60

(HCO3)−(CI + SO4)

40 20

−100

−80

−60

−40

−20

0 20

40

60

80

100

−20 −40 −60 −80 −100 (Ca + Mg)−(Na + K)

Figure 5. (a) Chadha’s diagram showing the chemical character of groundwater in District Hardwar, Uttaranchal (premonsoon 2002). (b) Chadha’s diagram showing the chemical character of groundwater in District Hardwar, Uttaranchal (postmonsoon 2002).

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IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

(calcium plus magnesium) and alkali metals (sodium plus potassium), expressed as a percentage of reacting values, is plotted on the x-axis, and the difference in milliequivalent percentage between weak acidic anions (carbonate plus bicarbonate) and strong acidic anions (chloride plus sulfate) is plotted on the y-axis. The resulting field of study is a square or rectangle depending upon the size of the scales chosen for the x and y coordinates. The milliequivalent percentage differences between alkaline earth and alkali metals and between weak acidic anions and strong acidic anions would plot in one of the four possible subfields of the diagram. The main advantage of this diagram is that it can be produced easily on most spreadsheet software packages. The square or rectangular field describes the overall character of the water. The diagram has all the advantages of the diamond-shaped field of the Piper trilinear diagram and can be used to study various hydrochemical processes, such as base cation exchange, cement pollution, mixing of natural waters, sulfate reduction, saline water (end product water), and other related hydrochemical problems (5). The chemical analysis data of all samples collected from District Hardwar have been plotted on Chadha’s diagram (Fig. 5a,b). It is evident from the results that all samples of the study area fall in Group 5 (Ca–Mg–HCO3 type) during both pre- and postmonsoon seasons. Chadha’s diagram has all the advantages of the diamond-shaped field of the Piper trilinear diagram and can be conveniently used to study various hydrochemical processes. Another main advantage of this diagram is that it can be produced easily on most spreadsheet software packages.

be developed. Landfill site(s) should be identified and must be scientifically designed. A mass awareness should be generated about water quality, its effect on human health, and the responsibilities of the public to safeguard water resources. BIBLIOGRAPHY 1. Jain, C.K. and Bhatia, K.K.S. (1988). Physico-Chemical Analysis of Water and Wastewater, User’s Manual, UM-26. National Institute of Hydrology, Roorkee, India. 2. American Public Health Association. (1992). Standard Methods for the Examination of Water and Waste Waters. 18th Edn. American Public Health Association, Washington, DC. 3. Bureau of Indian Standards. (1991). Specifications for Drinking Water. IS:10500:1991. Bureau of Indian Standards, New Delhi, India. 4. World Health Organization. (1996). Guidelines for Drinking Water. Vol. 2, Recommendations. World Health Organization, Geneva, Switzerland. 5. Chadha, D.K. (1999). A proposed new diagram for geochemical classification of natural waters and interpretation of chemical data. Hydrogeol. J. 7(5): 431–439. 6. Piper, A.M. (1944). A graphical procedure in the geochemical interpretation of water analysis. Trans. Am. Geophys. Union 25: 914–923.

IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA CHAKRESH K. JAIN National Institute of Hydrology Roorkee, India

CONCLUSIONS AND RECOMMENDATIONS The groundwater quality in District Hardwar varies from place to place and with the depth of the water table. The water drawn for domestic use should be tested and analyzed to ensure the suitability of groundwater for human consumption. The groundwater abstraction sources and their surroundings should be properly maintained to ensure hygienic conditions; no sewage or polluted water should be allowed to percolate directly to a groundwater aquifer. Proper cement platforms should be constructed surrounding the ground water abstraction sources to avoid direct wellhead pollution, and the surrounding surface area should be frequently chlorinated by using bleaching powder. The hand pumps and wells, which have been identified as of suspect water quality, should be painted red to indicate and warn the public that the water drawn from the source is not fit for human consumption. In the absence of an alternate safe source of water, the water with excessive undesirable constituents must be treated by a specific treatment process before it is used for human consumption. The untreated sewage and sewerage flowing in various open drains are one of the causes of groundwater quality deterioration. A proper underground sewage system must be laid in inhabited areas, and the untreated sewage should not be allowed to flow in open drains. A proper system of collection and transportation of domestic waste should

The groundwater quality of District Hardwar in the state of Uttaranchal (India) was assessed to determine the suitability of groundwater for irrigation. Two sets of 48 groundwater samples from shallow and deep aquifers were collected during pre- and postmonsoon seasons in the year 2002, and the suitability of the groundwater for irrigation was evaluated based on salinity, sodium adsorption ration (SAR), residual sodium carbonate (RSC), and boron content. The values of SAR ranged from 0.24–1.75 during the premonsoon season and from 0.22–1.92 during the postmonsoon season, respectively. Values of the SAR indicate that the majority of samples of the study area falls under the category of low sodium hazard, indicating no risk of sodification. In general, the groundwater of District Hardwar is safe for irrigation. According to the U.S. Salinity Laboratory classification of irrigation water, the majority of the samples (70%) falls under water type C2-S1 followed by the C3-S1 type. INTRODUCTION The intensive use of natural resources and the large production of wastes in modern society often pose a threat to groundwater quality and have already resulted in many incidents of groundwater contamination. Pollutants are

Previous Page 204

IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

(calcium plus magnesium) and alkali metals (sodium plus potassium), expressed as a percentage of reacting values, is plotted on the x-axis, and the difference in milliequivalent percentage between weak acidic anions (carbonate plus bicarbonate) and strong acidic anions (chloride plus sulfate) is plotted on the y-axis. The resulting field of study is a square or rectangle depending upon the size of the scales chosen for the x and y coordinates. The milliequivalent percentage differences between alkaline earth and alkali metals and between weak acidic anions and strong acidic anions would plot in one of the four possible subfields of the diagram. The main advantage of this diagram is that it can be produced easily on most spreadsheet software packages. The square or rectangular field describes the overall character of the water. The diagram has all the advantages of the diamond-shaped field of the Piper trilinear diagram and can be used to study various hydrochemical processes, such as base cation exchange, cement pollution, mixing of natural waters, sulfate reduction, saline water (end product water), and other related hydrochemical problems (5). The chemical analysis data of all samples collected from District Hardwar have been plotted on Chadha’s diagram (Fig. 5a,b). It is evident from the results that all samples of the study area fall in Group 5 (Ca–Mg–HCO3 type) during both pre- and postmonsoon seasons. Chadha’s diagram has all the advantages of the diamond-shaped field of the Piper trilinear diagram and can be conveniently used to study various hydrochemical processes. Another main advantage of this diagram is that it can be produced easily on most spreadsheet software packages.

be developed. Landfill site(s) should be identified and must be scientifically designed. A mass awareness should be generated about water quality, its effect on human health, and the responsibilities of the public to safeguard water resources. BIBLIOGRAPHY 1. Jain, C.K. and Bhatia, K.K.S. (1988). Physico-Chemical Analysis of Water and Wastewater, User’s Manual, UM-26. National Institute of Hydrology, Roorkee, India. 2. American Public Health Association. (1992). Standard Methods for the Examination of Water and Waste Waters. 18th Edn. American Public Health Association, Washington, DC. 3. Bureau of Indian Standards. (1991). Specifications for Drinking Water. IS:10500:1991. Bureau of Indian Standards, New Delhi, India. 4. World Health Organization. (1996). Guidelines for Drinking Water. Vol. 2, Recommendations. World Health Organization, Geneva, Switzerland. 5. Chadha, D.K. (1999). A proposed new diagram for geochemical classification of natural waters and interpretation of chemical data. Hydrogeol. J. 7(5): 431–439. 6. Piper, A.M. (1944). A graphical procedure in the geochemical interpretation of water analysis. Trans. Am. Geophys. Union 25: 914–923.

IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA CHAKRESH K. JAIN National Institute of Hydrology Roorkee, India

CONCLUSIONS AND RECOMMENDATIONS The groundwater quality in District Hardwar varies from place to place and with the depth of the water table. The water drawn for domestic use should be tested and analyzed to ensure the suitability of groundwater for human consumption. The groundwater abstraction sources and their surroundings should be properly maintained to ensure hygienic conditions; no sewage or polluted water should be allowed to percolate directly to a groundwater aquifer. Proper cement platforms should be constructed surrounding the ground water abstraction sources to avoid direct wellhead pollution, and the surrounding surface area should be frequently chlorinated by using bleaching powder. The hand pumps and wells, which have been identified as of suspect water quality, should be painted red to indicate and warn the public that the water drawn from the source is not fit for human consumption. In the absence of an alternate safe source of water, the water with excessive undesirable constituents must be treated by a specific treatment process before it is used for human consumption. The untreated sewage and sewerage flowing in various open drains are one of the causes of groundwater quality deterioration. A proper underground sewage system must be laid in inhabited areas, and the untreated sewage should not be allowed to flow in open drains. A proper system of collection and transportation of domestic waste should

The groundwater quality of District Hardwar in the state of Uttaranchal (India) was assessed to determine the suitability of groundwater for irrigation. Two sets of 48 groundwater samples from shallow and deep aquifers were collected during pre- and postmonsoon seasons in the year 2002, and the suitability of the groundwater for irrigation was evaluated based on salinity, sodium adsorption ration (SAR), residual sodium carbonate (RSC), and boron content. The values of SAR ranged from 0.24–1.75 during the premonsoon season and from 0.22–1.92 during the postmonsoon season, respectively. Values of the SAR indicate that the majority of samples of the study area falls under the category of low sodium hazard, indicating no risk of sodification. In general, the groundwater of District Hardwar is safe for irrigation. According to the U.S. Salinity Laboratory classification of irrigation water, the majority of the samples (70%) falls under water type C2-S1 followed by the C3-S1 type. INTRODUCTION The intensive use of natural resources and the large production of wastes in modern society often pose a threat to groundwater quality and have already resulted in many incidents of groundwater contamination. Pollutants are

IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

205

STUDY AREA

being added to groundwater systems by human activities and natural processes. Solid waste from industrial units is being dumped near factories, is subjected to reaction with percolating rainwater, and reaches the groundwater level. The percolating water picks up a large amount of dissolved constituents, reaches the aquifer system, and contaminates the groundwater. Groundwater plays an important role in agriculture for both watering crops and for irrigating dry season crops. It is estimated that about 45% of the irrigation water requirement is met from groundwater sources. Indiscriminate use of fertilizers and pesticides in agricultural fields has become common practice in India and has already resulted in very high concentrations of these constituents in groundwater, which in the long term, may create the greatest hazard to groundwater. Many problems originate from inefficient management of water for agricultural use, especially when it carries high salts. The problem of groundwater contamination in several parts of the country has become so acute that unless urgent steps for detailed identification and abatement are taken, extensive groundwater resources may be damaged. Keeping in view the severity of the problem, the groundwater quality of District Hardwar in the newly created state of Uttaranchal (India) was studied to examine the suitability of groundwater for irrigation.

District Hardwar is part of the Indo-Gangetic plains and lies between latitude 29◦ 30 to 30◦ 20 N and longitude 77◦ 40 to 78◦ 25 E in the state of Uttaranchal (Fig. 1). It is the largest district (populationwise) of Uttaranchal State and occupies an area of about 2,360 km2 . Per the 2001 census, the population of the District Hardwar is 14,44,213, and the population density is 612 per km2 . Physiographically, the area is generally flat except for the Siwalik Hills in the north and north east. The area is devoid of relief features of any prominence except for deep gorges cut by drains and rivers flowing through the area. The area is bounded by River Yamuna in the west and River Ganga in the east. The climate of the area is characterized by a moderate type of subtropical monsoon. The average annual rainfall in the region is about 1000 mm; the major part is received during the monsoon period. The major land use is for agriculture, and there is no effective forest cover. The soils of the area are loam to silty loam and are free from carbonates. The most common groundwater use is achieved by hand pumps and tube wells. Based on lithologic logs and water table fluctuation data, two types of aquifers have been delineated in the area. The upper is the shallow unconfined aquifer which generally extends to depths around 25m. The deeper one is confined to semiconfined and located at

N

30° 15′ De un

ad hr

1

3

district

4 5

44 8

47

31

14 References

0

37 28

16 17 18

15 19

20

Mu

5 10 km

zaf f dis arnag tric ar t

26 35 25 Laksar 22 24 23

78° 0′

t

o

jn

Bi

i rd

36 38

ric

st

21

Scale 77° 45′

27

t

13

34

32

29

c tri

41

dis

12

l wa rh

Roorkee

48

11

km 5

Hardwar 40 39 33 30

42 9

Sampling sites District H/Q Tehsil

45 46

6

7

10

43

Ga

Saharanpur

30° 0′

ct

tr i

dis

2

78° 15′

Figure 1. Study area showing location of sampling sites in District Hardwar.

206

IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA

a depth about 25 to 150m below ground level, separated by three to four aquifers at average depths of 25–55, 65–90, and 120–150m. Water table contours in the area indicate a southward trend of groundwater flow in both unconfined and confined aquifers. EXPERIMENTAL METHODOLOGY Two sets of 48 groundwater samples from District Hardwar were collected during pre- and postmonsoon seasons in the year 2002 from various abstraction sources at various depths. They were preserved by adding an appropriate reagent (1,2). The hand pumps and tube wells were continuously pumped prior to sampling to ensure that groundwater sampled was representative of the groundwater aquifer. The details of sampling locations and source and depth distribution are given in Tables 1 and 2, respectively. The physicochemical analysis was performed following standard methods (1,2).

its quality for irrigation. The quality of water is an important consideration in any appraisal of salinity or alkaline conditions in an irrigated area. Good quality water has can provide maximum yield under good soil and water management practices. The most important characteristics of water which determine suitability of ground water for irrigation are 1. 2. 3. 4.

Salinity Relative proportion of sodium to other cations (SAR) Residual sodium carbonate (RSC) Boron content

The safe limits of electrical conductivity for crops of different degrees of salt tolerance under varying soil textures and drainage conditions are given in Table 4.

Table 2. Source and Depth Distribution of Sampling Sites in District Hardwar Depth Range

RESULTS AND DISCUSSION

Source Structure

The hydrochemical data for the two sets of samples collected during pre- and postmonsoon seasons are presented in Table 3. The suitability of groundwater for domestic use was discussed in an earlier report (3). The quality of water plays an important role in irrigated agriculture. Many problems originate from inefficient management of water for agricultural use, especially when it carries high salts. The concentration and composition of dissolved constituents in water determine

Hand pumps 6, 18, 20, 21, 22, 3, 4, 5, 7, 8, 9, 23, 24, 25, 27, 10, 11, 12, 13, 28, 30, 31, 32, 14, 15, 16, 17, 36, 41, 43, 44, 19, 26, 29, 33, 45, 46, 47 34, 35, 38, 39, 40, 42, 48 Tube wells — — Open wells 1,2 — Total 22 25

40 m Number 37

46

— — 1

— 2 48

IRRIGATION WATER QUALITY IN DISTRICT HARDWAR, UTTARANCHAL, INDIA Table 3. Hydrochemical Data for Groundwater Samples from District Hardwara Characteristics pH Conductivity, µS/cm TDS, mg/L Alkalinity, mg/L Hardness, mg/L Chloride, mg/L Sulfate, mg/L Nitrate, mg/L Phosphate, mg/L Fluoride, mg/L Sodium, mg/L Potassium, mg/L Calcium, mg/L Magnesium, mg/L Boron, mg/L a

Min

Max

Average

6.22 (6.70) 233 (221) 149 (141) 78 (71) 81 (80) 0.1 (1.2) 0.5 (0.7) 0.1 (0.1) 0.01 (0.02) 0.01 (0.01) 5.0 (4.5) 0.8 (0.2) 22 (24) 5.0 (5.0) 0.12 (0.12)

7.58 (7.70) 1440 (1442) 922 (923) 460 (482) 464 (427) 32 (35) 72 (62) 140 (130) 0.41 (1.60) 0.94 (0.88) 69 (71) 42 (32) 140 (135) 36 (38) 0.92 (0.87)

6.87 (7.10) 642 (647) 411 (414) 213 (215) 209 (212) 9.0 (9.0) 22 (21) 13 (13) 0.03 (0.25) 0.36 (0.42) 26 (25) 7.0 (7.4) 55 (56) 18 (17) 0.50 (0.49)

Values in parentheses represent postmonsoon data.

The quality of water is commonly expressed by classes of relative suitability for irrigation with reference to salinity levels. The recommended classification with respect to electrical conductivity, sodium content, sodium absorption ratio (SAR), and residual sodium carbonate (RSC) are given in Table 5. The values of sodium percentage (% Na), SAR, and RSC in the groundwater of District Hardwar are given in Table 6.

leaf burns at margin or top. The electrical conductivity values for water in District Hardwar are well within the prescribed limits of 1500 µS/cm and are therefore safe for irrigation. Relative Proportion of Sodium to Other Cations A high salt concentration in water leads to saline soil, and high sodium leads to development of an alkaline soil. The sodium or alkaline hazard in using water for irrigation is determined by the absolute and relative concentration of cations, expressed in terms of the sodium adsorption ratio (SAR). If the proportion of sodium is high, the alkaline hazard is high, and conversely, if calcium and magnesium predominate, the hazard is less. There is a significant relationship between SAR values of irrigation water and the extent to which sodium is absorbed by the soil. If water used for irrigation is high in sodium and low in calcium, the cation-exchange complex may become saturated with sodium. This can destroy the soil structure owing to dispersion of the clay particles. A simple method of evaluating the danger of high-sodium water is the sodium adsorption ratio (SAR) (4): Na+ SAR =  (Ca2+ + Mg2+ )/2 The sodium percentage is calculated from %Na =

Salinity Salinity is broadly related to total dissolved solids (TDS) and electrical conductivity (EC). High concentrations of TDS and electrical conductivity in irrigation water may increase the soil salinity, which affects the salt intake of a plant. The salts in the water affect the growth of plants directly and also affect the soil structure, permeability, and aeration, which indirectly affect plant growth. Soil water passes into a plant through the root zone by osmotic pressure. As the dissolved solid content of the soil water in the root zone increases, it is difficult for the plant to overcome the osmotic pressure and the plant root membranes are able to assimilate water and nutrients. Thus, the dissolved solids content of the residual water in the root zone also has to be maintained within limits by proper leaching. Negative effects are visible in plants by stunted growth, low yield, discoloration, and even

207

Na+ + K+ 2+

Ca

+ Mg2+ + Na+ + K+

× 100

where all ionic concentrations are expressed in milliequivalents per liter.

Table 5. Guidelines for Evaluation of Irrigation Water Quality

Water Class

Sodium (Na), %

Electrical Conductivity, µS/cm

SAR

RSC, meq/L

Excellent Good Medium Bad Very bad

80

4000

26 >26

3.0

Table 4. Safe Limits of Electrical Conductivity for Irrigation Water

S. No. 1. 2. 3. 4.

Nature of Soil Deep black soil and alluvial soil whose clay content is more than 30% in soils that are fairly to moderately well drained Textured soil of clay content of 20–30% in soils that are well drained internally and have a good surface drainage system Medium textured soils whose clay is 10–20% internally very well drained and has a good surface drainage system Lightly textured soils whose clay is less than 10% of soil that has an excellent internal and surface drainage system

Crop Growth Semitolerant Tolerant Semitolerant Tolerant Semitolerant Tolerant Semitolerant Tolerant

Upper Permissible Safe Limit of EC in Water, µS/cm 1500 2000 2000 4000 4000 6000 6000 8000

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Table 6. SAR, % Na, and RSC Values in the Groundwater of District Hardwar Premonsoon S.No. 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48

Location Mohand Banjarewala Buggawala Kheri Dadapatti Bahbalpur Sikandarpur Bhagwanpur Chudiala Balswa Ganj Manakpur Iqbalpur Jharera Sherpur Narsen Manglour Libarheri Mahesari Sahipur Khanpur Chandpuri Kalan Laksar Kalsiya Niranjanpur Sultanpur Shahpur Pathri Subashgarh Marghubpur Bahadarabad Alipur Katarpur Kankhal Shyampur Rasiya Garh Gandikhatta Laldhang Kottawali Hardwar Jwalapur Roorkee Gumanwala Manubas Bandarjud Beriwala Hazara Aurangbad Daulatpur

Postmonsoon

SAR % Na RSC SAR % Na RSC 1.02 0.81 0.31 0.56 0.86 0.89 0.72 0.79 0.72 0.90 0.69 1.04 0.77 0.54 0.50 0.80 0.24 1.11 0.74 1.75 1.60 1.30 0.71 1.28 1.31 0.46 0.25 0.45 1.53 0.54 0.49 0.42 0.63 0.59 0.52 0.54 0.54 0.48 0.54 1.05 0.46 0.58 1.09 0.30 0.72 1.03 0.99 0.79

27.4 0.40 22.7 −0.32 11.1 −0.18 17.5 −0.17 25.4 0.77 23.4 0.61 24.5 0.48 25.3 0.67 24.4 0.60 25.0 −0.27 23.0 0.54 23.7 0.48 21.3 −0.10 23.0 0.28 21.0 0.14 21.6 −0.38 14.9 −0.06 33.1 0.69 28.9 0.59 45.5 1.48 47.9 1.11 33.2 0.74 24.4 −0.17 33.5 0.01 35.7 1.04 19.8 0.14 9.9 −0.36 17.3 −0.37 29.7 0.12 14.9 −0.70 17.0 −0.16 17.4 −0.69 20.0 −0.91 19.8 0.07 16.1 −0.11 18.0 −0.35 15.7 −0.55 16.1 −0.07 20.3 −0.46 23.7 −1.76 16.6 −0.19 19.7 −0.15 24.5 −0.10 9.1 −0.45 25.4 0.46 28.5 0.15 32.2 0.32 24.10 0.44

0.97 0.49 0.30 0.67 0.76 0.96 0.66 0.68 0.64 0.71 0.73 0.97 0.76 0.51 0.56 0.90 0.22 1.18 0.71 1.92 1.51 1.19 0.64 1.39 1.14 0.44 0.24 0.42 1.57 0.61 0.53 0.39 0.62 0.50 0.51 0.58 0.48 0.41 0.64 1.07 0.48 0.72 1.37 0.28 0.66 0.62 0.75 0.88

27.1 16.4 10.3 18.3 25.2 23.5 23.2 22.4 21.9 22.5 23.3 23.4 20.8 21.5 23.0 22.1 13.5 33.1 27.1 47.8 42.9 30.3 23.7 33.5 30.9 18.1 10.2 17.1 30.4 20.3 17.7 16.0 19.7 17.5 15.5 19.3 14.3 13.9 23.6 22.9 18.8 24.6 28.8 9.2 23.1 21.1 27.1 26.7

−0.05 −0.62 −0.05 −0.03 0.60 0.47 0.51 0.49 0.42 −0.52 0.52 0.55 −0.33 0.07 0.26 −0.35 −0.19 1.01 0.48 1.54 0.93 0.66 −0.18 0.26 0.11 0.13 −0.45 −0.17 0.18 0.17 −0.07 −0.77 −0.40 −0.06 −0.17 −0.23 −0.65 −0.31 −0.36 −2.11 0.05 0.31 1.09 −0.40 0.25 −0.62 0.24 0.64

Calculation of SAR for a given water provides a useful index of the sodium hazard of that water for soils and crops. A low SAR (2 to 10) indicates little danger from sodium; medium hazards are between 7 and 18, high hazards between 11 and 26, and very high hazards above that. The lower the ionic strength of the solution, the greater the sodium hazards for a given SAR (4). The values of SAR in the groundwater of District Hardwar vary from 0.24–1.75 during the premonsoon season and from 0.22–1.92 during the postmonsoon

season. As evident from the SAR values, the groundwater of the study area falls under the category of low sodium hazard, which reveals that the groundwater of the study area is free from any sodium hazard. The sodium percentage in the study area varies from 9.1–47.9% during the premonsoon season and from 9.2–47.8% during the postmonsoon season. All samples are well within the permissible limit for irrigation water and are free of any sodium hazard. Residual Sodium Carbonate In addition to total dissolved solids, the relative abundance of sodium with respect to alkaline earths and boron and the quantity of bicarbonate and carbonate in excess of alkaline earths also influence the suitability of water for irrigation. This excess is denoted residual sodium carbonate (RSC) and is determined by the following formula: RSC = (HCO3 − + CO3 2− ) − (Ca2+ + Mg2+ ) where all ionic concentrations are expressed in meq/L. Groundwater containing high concentrations of carbonate and bicarbonate ions tends to precipitate calcium and magnesium as carbonate. As a result, the relative proportion of sodium increases and is fixed in the soil thereby decreasing soil permeability. If the RSC exceeds 2.5 meq/L, the water is generally unsuitable for irrigation. Excessive RSC causes the soil structure to deteriorate, as it restricts the water and air movement through soil. If the value is between 1.25 and 2.5, the water is of marginal quality; values less than 1.25 meq/L indicate that the water is safe for irrigation. During the present study, the RSC values clearly indicate that the groundwater of District Hardwar does not have any residual sodium carbonate hazard. Boron Boron is essential to the normal growth of all plants, but the concentration required is very small and if exceeded may cause injury. Plant species vary in boron requirement and in tolerance to excess boron, so that concentrations necessary for the growth of plants that have high boron requirements may be toxic to plants sensitive to boron. Although boron is an essential nutrient for plant growth, generally it becomes toxic beyond 2 mg/L in irrigation water for most field crops. It does not affect the physical and chemical properties of the soil, but at high concentrations, it affects the metabolic activities of the plant. During the present study, the boron content in the groundwater clearly indicates that the groundwater is safe for irrigation. U.S. Salinity Laboratory Classification The U.S. Salinity Laboratory classification (5) is used to study the suitability of groundwater for irrigation. In classifying irrigation waters, it is assumed that the water will be used under average conditions with respect to soil texture, infiltration rate, drainage, quantity of water used, climate, and salt tolerance of the crop. Sodium concentration is an important criterion in irrigation-water

Sodium (alkali) hazard

100

500

1000

5000 30

30

Very high S4

28 26 24

High S3

22 S

20

A

18

R

16

Medium S2

20

14 12 10

10

8 Low S1

6 4 2

M N D He S [2]11 D MSC D GM C DS MCDCD D S C e 11

100

250

W F S nR S e G

K

750

2250

Conductivity, micromhos/cm at 25°C C1

C2

C3

C4

Low

Medium

High

Very high

Figure 2. U.S. Salinity Laboratory classification, District Hardwar (premonsoon 2002).

Salinity hazard

Sodium (alkali) hazard

100

Very high S4

500

1000

5000 30

30 28 26 24

High S3

22 20

20 S A Medium S2

R

18 16 14 12 10

10

8 Low S1

6 4 2

K L O M E L CC Cr E 2D O C C0 A RDC H E FD D E FMH r M DO o 2M xD

H

100

250

U

r

750

2250

Conductivity, micromhos/cm at 25°C C1 Low

C2

C3

Medium

High

Salinity hazard

209

C4 Very high

Figure 3. U.S. Salinity Laboratory classification, District Hardwar (postmonsoon 2002).

210

INFILTRATION AND SOIL WATER PROCESSES Table 7. Summarized Results of U.S. Salinity Laboratory Classification Sample Numbers Classification/Type C1-S1 C2-S1

C3-S1

Premonsoon 2002

Postmonsoon 2002

17 3, 4, 5, 6, 7, 8, 9, 11, 14, 15, 18, 19, 20, 21, 22, 23, 26, 27, 28, 31, 32, 34, 35, 36, 37, 39, 40, 41, 42, 44, 45, 47, 48 1, 2, 10, 12, 13, 16, 24, 25, 29, 30, 33, 38, 43, 46

17 3, 4, 5, 7, 8, 9, 11, 14, 15, 18, 19, 20, 21, 22, 23, 26, 27, 28, 30, 31, 32, 34, 35, 36, 37, 38, 39, 41, 42, 44, 45, 47, 48 1, 2, 6, 10, 12, 13, 16, 24, 25, 29, 33, 40, 43, 46

classification because sodium reacts with the soil to create sodium hazards by replacing other cations. The extent of this replacement is estimated by the sodium adsorption ratio (SAR). A diagram for use in studying the suitability of groundwater for irrigation is based on the sodium adsorption ratio (SAR) and the electrical conductivity of water expressed in µS/cm. The chemical analytical data on groundwater samples from District Hardwar were processed as per the U.S. Salinity Laboratory classification for the two sets of data (Figs. 2 and 3), and the results are summarized in Table 7. It is evident from the results that majority of the samples (about 70% of total samples) falls under water type C2-S1 (medium salinity and low SAR); such water can be used if a moderate amount of leaching occurs, and plants with moderate salt tolerance can be grown in most cases without special practices for salinity control. About 30% of the samples falls under water type C3-S1 (high salinity and low SAR); such water cannot be used on soils where drainage is restricted. Even with adequate drainage, special management for salinity control may be required, and plants with good tolerance should be selected. An almost similar trend was observed during the postmonsoon season.

CONCLUSIONS Groundwater quality varies from place to place and with the depth of the water table. The SAR values indicate that the majority of samples from the study area falls under the category of low sodium hazard indicating no risk for sodification. In general, the groundwater of District Hardwar is safe for irrigation. According to the U.S. Salinity Laboratory classification of irrigation water, the majority of the samples (70%) falls under water type C2-S1 followed by the C3-S1 type.

BIBLIOGRAPHY 1. American Public Health Association. (1992). Standard Methods for the Examination of Water and Waste Waters, 18th Edn. American Public Health Association, Washington, DC. 2. Jain, C.K. and Bhatia, K.K.S. (1988). Physico-Chemical Analysis of Water and Wastewater, User’s Manual, UM-26. National Institute of Hydrology, Roorkee, India.

3. Jain, C.K. (2003). Ground Water Quality in District Hardwar, Uttaranchal, Technical Report. National Institute of Hydrology, Roorkee, India. 4. Richards, L.A. (Ed.). (1954). Diagnosis and Improvement of Saline and Alkali Soils, Agricultural Handbook 60. U.S. Dept. Agric., Washington, DC. 5. Wilcox, L.V. (1955). Classification and Use of Irrigation Water. U.S. Dept. of Agric. Circular 969, Washington, DC.

INFILTRATION AND SOIL WATER PROCESSES JOSEPH HOLDEN University of Leeds Leeds, United Kingdom

Water at the soil surface is drawn into the soil pores under the influence of both suction and gravity. Suction (or tension/capillary) forces are largest in the small pore spaces between soil particles. Therefore, small pores fill first, and large pores empty first. As a soil becomes saturated, suction forces approach zero. As a soil dries, suction forces dominate, and there is little or no gravitational flow. For a dry soil during the early stages of wetting, the effect of soil pore suction is predominant. However, as the depth of wet soil increases, the suction gradient decreases (the same difference in suction is spread over an ever increasing depth interval) and the gravitational head gradient becomes the main driving force (1). Infiltration processes cannot easily be separated from the processes involved in movement of water within soil because the rate of infiltration is partly controlled by the rate of soil water movement below the surface. Soil water movement continues after an infiltration, as the infiltrated water is redistributed. Infiltration and soil water processes play a key role in surface runoff, groundwater recharge, ecology, evapotranspiration, soil erosion, and transport of nutrients and other solutes in surface and subsurface waters. Bodman and Colman (2) suggested that for a uniform soil, there would be a series of zones in the wetting part of the soil profile during an infiltration. The zone nearest the surface is a saturated zone (typically in the upper centimeter of the soil profile). As water penetrates more deeply, a zone of uniform water content, the transmission zone, develops behind a well-defined wetting front (Fig. 1).

INFILTRATION AND SOIL WATER PROCESSES (a)

Water supply to soil surface and infiltration

211

(b)

Saturated zone 2 Transmission zone

Soil with antecedent water content

Depth, cm

Wetting front

4 6 8 10 12 14 10

20

30 40 50 Water content, %

There is a sharp change in water content at the wetting front because the water at the boundary takes up a preferred position of minimum potential energy in the narrowest pores, for which the hydraulic conductivity is very low, and does not move at an appreciable rate until the large pores begin to fill. Figure 1 shows the water content versus depth for a sandy soil 4 minutes after ponded infiltration at the surface. The sharp transition in water content can be seen at the wetting front. Note that the soil below the wetting front still has some preevent moisture. There is also a sharp transition in the moisture profile between the upper saturated zone and the transmission zone, sometimes called the transition zone (1). Gravitational and suction gradients result in continued downward movement of water within the soil profile which may occur long after infiltration at the surface has ceased. As a result, the transmission zone which existed during infiltration becomes a draining zone that releases water from the wetter upper layers of the soil to deeper drier layers. As the draining zone becomes drier, the hydraulic conductivity reduces, and so the rate of soil water redistribution slows down. The suction gradient also weakens as the soil water content becomes more uniform. While the wetting front continues to advance into deeper soil layers, its movement is reduced over time until eventually the water content changes only very slowly and the soil is said to be at field capacity (1). In some soils, this may take several days (2). Theoretically, soil water redistribution following rainfall would continue until gravity and soil suction forces were in balance (4,5). At this stage and at all depths, there would be equal soil water potential. Therefore, without a hydraulic gradient, there would be no soil water movement, which is rarely observed in the field because soils are highly heterogeneous, rainfall itself is rarely an on–off uniform event (a theoretical model examining this issue was applied to a range of soil types

60

Figure 1. Soil water zones during infiltration: (a) theoretical zonation, (b) measured water content for a sandy soil 4 minutes after ponded infiltration commenced.

by Ref. 6), and because evaporation from soil and plants can result in drying of surface soil layers. Evaporation at the soil surface encourages upward movement of soil water, which therefore reduces downward movement of water over time. There is probably a two-way effect whereby drainage reduces the amount of evaporation and evaporation reduces the amount of drainage (3). During water redistribution following infiltration, the upper part of the soil profile may be drying through drainage and evaporation while the lower part is still becoming wetter, which complicates the water content–soil moisture tension (suction) relationship within the soil because it will depend on the history of wetting and drying in any given part of the soil profile (soil water hysteresis; 4,7,8). Often, the soil water redistribution that follows an infiltration is complicated by the nonuniform nature of most soils. In addition to the effects of soil layering, many studies, for example, often assume that the water table has little impact on postinfiltration water redistribution. However, in many soils, the water table may be shallow (e.g., peats), and this may exert considerable influence on the distribution of water in a soil profile (9). Trapped soil gases may also affect water redistribution by blocking pores and preventing water entry (10). Furthermore, noncapillary macropores or voids such as decayed root channels, worm holes, and structural cracks that are open at the soil surface can capture the free water available at the surface during rainfall and overland flow. This causes the water flowing through the macropores to bypass the soil matrix. Often, water and solutes (e.g., fertilizer applications) can infiltrate into the soil via macropores but then are quickly transferred to much deeper parts of the soil profile where the water and nutrients are not readily available to plants (11,12). Bypassing flow also causes some parts of the soil profile at any given depth

212

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to wet well before other parts of the soil that are at the same depth. BIBLIOGRAPHY 1. Ward, R.C. and Robinson, M. (2000). Principles of Hydrology. McGraw-Hill, London. 2. Bodman, G.B. and Colman, E.A. (1943). Moisture and energy conditions during downward entry of water into soils. Proceedings of the Soil Science Society of America 8: 116–122. 3. Gardner, W.R., Hillel, D., and Benyamini, Y. (1970). Post irrigation movement of soil water, 2, Simultaneous redistribution and evaporation. Water Resources Research 6: 1148–1153. 4. Childs, E.C. (1969). An Introduction to the Physical Basis of Soil Water Phenomena. John Wiley & Sons, London. 5. Rawls, W.J., Ahuja, L.R., Brakensiek, D.L., and Shirmohammadi, A. (1993). Infiltration and soil water movement. In: Handbook of Hydrology. D.R. Maidment (Ed.). McGraw-Hill, New York, pp. 5.1–5.51. 6. Corradini, C., Melone, F., and Smith, R.E. (1997). A unified model for infiltration and redistribution during complex rainfall patterns. Journal of Hydrology 192: 104–124. 7. Vachaud, G. and Thony, J.L. (1971). Hysteresis during infiltration and redistribution in a soil column at different initial water contents. Water Resources Research 9: 160–173. 8. Beese, F. and van der Ploeg, R.R. (1976). Influence of hysteresis on moisture flow in an undisturbed soil monolith. Proceedings of the Soil Science Society of America 40: 480–484. 9. Evans, M.G., Burt, T.P., Holden, J., and Adamson, J. (1999). Runoff generation and water table fluctuations in blanket peat: evidence from UK data spanning the dry summer of 1995. Journal of Hydrology 221: 141–160. 10. Beckwith, C.W. and Baird, A.J. (2001). Effect of biogenic gas bubbles on water flow through poorly decomposed blanket peat. Water Resources Research 37: 551–558. 11. Beven, K.J. and Germann, P. (1982). Macropores and water flow in soils. Water Resources Research 18: 1311–1325. 12. Germann, P.F. (1990). Macropores and hydrologic hillslope processes. In: Process Studies in Hillslope Hydrology. M.G. Anderson and T.P. Burt (Eds.). John Wiley & Sons, Chichester, pp. 327–363.

INFILTRATION/CAPACITY/RATES JOSEPH HOLDEN University of Leeds Leeds, United Kingdom

Infiltration is the process of water entry into the surface of a soil. It involves surface water entry, transmission through the soil, and depletion of the soil water storage capacity. Surface water entry is influenced by surface vegetation cover, soil texture, soil porosity, soil structure (e.g., cracks, surface crusting), and compaction. Land management can influence all of these parameters and therefore influence infiltration. The infiltration rate is the volume of water passing into the soil per unit area per unit time and therefore has dimensions of velocity (e.g., m s−1 ; mm h−1 ). The

infiltration rate is often measured using an infiltrometer (INFILTROMETERS) and is one of the most easily and frequently measured soil water processes. Despite the units of velocity, the values of infiltration rate are actually volume fluxes expressed as volume flux density and therefore do not indicate the effective velocity of vertical water movement in the soil (1). The maximum rate at which water soaks into or is absorbed by the soil is the infiltration capacity. Infiltration capacity may be very important in determining the proportion of incoming rainfall that runs off as infiltrationexcess overland flow and the proportion that moves into the soil. The infiltration capacity is different from the infiltration rate. Infiltration may be limited by the rate of supply of rainfall (or other water supply) or the capacity of the soil to absorb water. In the former case, infiltration is limited by the water supply to the soil surface, and in the latter case, infiltration is restricted by the infiltration capacity of the soil. Use of the term infiltration rate therefore indicates that infiltration is proceeding at less than the infiltration capacity (2). If infiltration is occurring at less than the infiltration capacity, then all rain reaching the soil surface that is not held as surface storage will infiltrate into the soil. Thus, there will be a direct relationship between infiltration rate and rainfall intensity. However, when rainfall intensity exceeds the infiltration capacity of the soil surface, this relationship breaks down and may be inverted, as infiltration capacity declines through a storm as the soil wets (2). Excess rainwater that cannot infiltrate into the soil because the surface water supply is greater than the infiltration capacity will pond on the soil surface and eventually run over the surface as infiltration-excess overland flow. The infiltration capacity of a soil generally decreases during rainfall, rapidly at first and then more slowly, until a more or less stable value has been attained (Fig. 1). Water initially crosses the soil surface interface at an initial rate (fo ), depending on existing soil moisture content. As the rainfall or other water supply to the soil surface continues, the rate of infiltration decreases (because it is limited by decreasing infiltration capacity) as the soil becomes wetter and less able to take up water. Philip (3) put forward the theory that the infiltration capacity of a soil declines over time in a manner described by the equation, f = A + Bt−0.5 where f = infiltration capacity at time t, and A and B are constants. This equation may be solved using simultaneous equations, allowing constants A and B to be evaluated: ) B = (f1 − f0 )/(t1−0.5 − t−0.5 0 A = (f0 t−0.5 − f1 t−0.5 )/(t−0.5 − t0−0.5 ) 1 0 1 The curve of infiltration rate over time shown in Fig. 1 reduces to a constant value fc , which is the infiltration capacity at time tc . The Philip model is useful for short infiltration events, but other models (e.g., those based on the Richards equation) are more reliable for longer periods (4,5).

INFILTRATION/CAPACITY/RATES

Infiltration capacity

fo

fc to

Time

tc

Figure 1. The decline of infiltration capacity during a rainfall.

Soil surface conditions may impose an upper limit on the rate at which water can be absorbed, despite a large available capacity of the lower soil layers to receive and to store additional infiltrating water. Often, the infiltration capacity is reduced by surface sealing resulting from compaction, washing of fine particles into surface pores, and by frost (6,7). Frozen moisture will block soil pores, and the greater the frozen moisture content, the lower the infiltration rate upon melt, as the larger pores remain blocked for longest. Holden and Burt (8) noted that infiltration capacity increases with the depth of standing water on the surface because of increased water head (9,10) and unblocking of pores due to flotation of fines (11). The number of cracks and fissures at the surface and the ground slope also influences infiltration capacity. Cultivation techniques may either increase or decrease infiltration capacity (12). Field plowing can increase soil infiltration capacity, whereas raindrop compaction subsequent to field plowing can reduce infiltration capacities when splashed silt and clay soil particles clog soil pores as the aggregates break down. Organic matter also binds aggregates together. New vegetation, however, tends to increase the infiltration capacity of a soil by retarding surface water movement, stabilizing loose particles, reducing raindrop compaction, and improving soil structure. Soils with well developed humus and litter layers tend to have high infiltration capacities. Most forest floors that have a litter layer tend, for example, to have higher infiltration than that beneath grass (2). The infiltration capacity of a soil is dependent on surface properties and also on transmission through the soil below the surface and the storage capacity of the soil. Water cannot be absorbed by the soil surface unless the underlying soil profile can conduct the infiltrated water away. Transmission rates depend on soil texture, porosity, pore size distribution, soil stratification, antecedent soil moisture content, initial gradient of soil water potential, salinity, and biotic activity. Storage capacity depends on pore size distribution, porosity, and antecedent moisture conditions. Where the soil profile contains a relatively impermeable layer at some distance below the surface, the curve of infiltration capacity against time may display

213

a sudden reduction of infiltration capacity because when available storage in the surface soil horizons has been filled, infiltration will be governed by the rate at which water can pass through the underlying impeding layer. Soil properties are highly heterogeneous. Infiltration capacities and infiltration rates vary greatly even within a small area and in a short time. The effect of spatial variability on infiltration produces a difference between point infiltration values and apparent infiltration rates associated with measurements from large plots, composite areas, and estimates from watershed calculations, which may provide a reason that mean infiltration capacity across a plot increases with rainfall intensity (13,14; see also RAIN SIMULATORS). In other words, the nonuniform nature of soil across a large plot may cause in the surface of one part of the plot to have a higher infiltration capacity than the rest of the plot. Therefore, as Hawkins (13) demonstrated numerically, the mean infiltration rate across a plot will increase with rainfall intensity simply because a greater flux of water is occurring through the parts of the plot surface that have higher relative infiltration capacities. It must be acknowledged, therefore, that only a fraction of the area within a plot or hillslope need contribute to overland flow, even when the rainfall rate exceeds the mean infiltration capacity across the plot or hillslope, which also suggests that the spatial variability in infiltration parameters has more effect at lower rainfall rates, closer to the saturated hydraulic conductivity of the soil (5). BIBLIOGRAPHY 1. White, R.E. (1997). Principles and practice of soil science. The Soil as a Natural Resource. Blackwell Science Ltd, Oxford. 2. Ward, R.C. and Robinson, M. (2000). Principles of Hydrology. McGraw-Hill, London. 3. Philip, J.R. (1957). The theory of infiltration 1. The infiltration equation and its solution. Soil Science 83: 345–357. 4. Kirkby, M.J. (1985). Hillslope hydrology. In: Hydrological Forecasting. M.G. Anderson and T.P. Burt (Eds.). John Wiley, Chichester, pp. 37–75. 5. Rawls, W.J., Ahuja, L.R., Brakensiek, D.L., and Shirmohammadi, A. (1993). Infiltration and soil water movement. In: Handbook of Hydrology. D.R. Maidment (Ed.). McGraw-Hill, New York, pp. 5.1–5.51. 6. Poesen, J. (1986). Surface sealing as influenced by slope angle and position of simulated stones in the top layer of loose sediments. Earth Surface Processes and Landforms 11: 1–10. 7. Romkens, M.J.M., Prasad, S.N., and Parlange, J.Y. (1990). Surface seal development in relation to rainsform intensity. Catena Supplement 17: 1–11. 8. Holden, J. and Burt, T.P. (2002). Infiltration, runoff and sediment production in blanket peat catchments: Implications of field rainfall simulation experiments. Hydrological Processes 16: 2537–2557. 9. Schiff, L. (1953). The effect of surface head on infiltration rates based on the performance of ring infiltrometers and ponds. Transactions of the American Geophysical Union 34: 257–266. 10. Philip, J.R. (1958). The theory of infiltration 6. Effect of water depth over soil. Soil Science 85: 278–286.

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11. Bowyer-Bower, T.A.S. (1993). Effects of rainfall intensity and antecedent moisture on the steady-state infiltration rate in a semi-arid region. Soil Use and Management 9: 69–76. 12. Imeson, A.C. and Kwaad, F.J.P.M. (1990). The response of tilled soils to wetting by rainfall and the dynamic character of soil erodibility. In: Soil Erosion on Agricultural Land. J. Boardman, I.D.L. Foster, and J.A. Dearing (Eds.). John Wiley, Chichester, pp. 3–14. 13. Hawkins, R.H. (1982). Interpretations of source area variability in rainfall-runoff relations. In: Rainfall-runoff relationship. Proceedings of the International Symposium on rainfall runoff modelling held May 18–21, 1981, Mississippi State University. Water Resources Publications, Littleton, CO, pp. 303–324.

Water Soil

14. Smith, R.E. (1983). Flux infiltration theory for use in watershed hydrology. In: Advances in Infiltration. American Society of Agricultural Engineers, St. Joseph, MI, pp. 313–323.

Infiltration

INFILTROMETERS

Figure 1. Infiltration from a single ring infiltrometer.

JOSEPH HOLDEN University of Leeds Leeds, United Kingdom

Infiltrometers allow measuring of the rate of infiltration of water into a given medium. Rain simulators, a particular type of infiltrometer, are discussed elsewhere (see entry RAIN SIMULATORS). Unlike rain simulators, other infiltrometers do not simulate raindrop activity but measure the rate of infiltration from a reservoir of water at the ground surface. Their main advantage is their simplicity and ease of use. There are two main types of infiltrometer, a ponded ring infiltrometer and a disk infiltrometer.

Mariotte tube Inner ring

Clamp to support Mariotte tube

Outer ring

Soil

PONDED RING INFILTROMETER A ponded ring infiltrometer is the most common type of infiltrometer used. It is inexpensive to construct or operate and requires relatively little water for measurements. One person can set up and run several tests simultaneously. A cylindrical ring of stainless steel or plastic pipe is used, and the ring is pushed a few centimeters into the ground (Fig. 1). Care is taken to minimize disturbing the soil surface and soil structure during installation. Water is then flooded into the ring. The water inside the ring gradually infiltrates the soil, so measurements of the depth of water in the ring over time can be used to provide an infiltration rate usually in units of length per unit time (e.g., mm hr−1 ) allowing easy comparison with rainfall intensity data. If the ring empties, it can be easily refilled. There are two main problems in this method: (1) lateral flow below the ring; (2) changing pressure of water in the ring as the water level decreases in the ring during infiltration. To get around the lateral flow problem, it is advisable to use two concentric rings; the larger ring forms a buffer compartment around the inner ring (Fig. 2). While water is topped up in the outer ring, measurements are taken only of infiltration from the inner ring. It is preferable that a constant head of water is maintained

Infiltration Figure 2. Double-ring infiltrometer with constant-head supply.

within the rings to avoid changing pressures. The head of water should also be less than 5 cm deep. The constant head can be supplied by a Mariotte bottle placed above the rings, and readings of water usage can be made from the Mariotte bottle (Fig. 2). Thus the optimum design for a ring infiltrometer is a constant-head, double-ring device whose rings have large diameters, so that they can represent the soil surface more readily. A plot of water uptake against time should show a leveling off as the infiltration capacity is reached. It often takes several hours before a constant rate of infiltration is achieved from the infiltrometer, depending on soil type, texture, and antecedent soil moisture. The method of placement is serious limitation to the use of ring infiltrometers. Knocking rings into the ground can result in destruction of the soil structure or compression. If a soil is shattered, it can disturb the contact between the ring edge and the soil, resulting in leakage and high infiltration rates. Research has demonstrated that

INFILTROMETERS

ring infiltrometers give higher infiltration rates than rain simulators because of the effect of a pond on the soil surface, lateral seepage, and soil cracking. DISK INFILTROMETER Tension disk infiltrometers are a standard tool for in situ determination of saturated and near-saturated soil hydraulic properties. These infiltrometers have a porous membrane disk at their base which rests on the soil surface of interest. To assess the role of matrix and macropore flow, a tension infiltrometer allows infiltration of water into the soil matrix, while preventing flow into larger pores that may otherwise dominate the infiltration process. The infiltrometer provides a source of water at a small negative pore water pressure at the surface. The supply pressure head is controlled with a Mariotte bottle. A schematic diagram of a tension disk infiltrometer is given in Fig. 3. As for a ring infiltrometer, infiltration rates can be measured manually by observing the volume of water lost from the Mariotte bottle over time. The negative pressure prevents the larger pores that fill at greater pore water pressures from wetting and short-circuiting the flow. Hence, by subtraction, the hydrological role of larger pores during the infiltration process can be evaluated. Further details on designs for these infiltrometers are given in Ankeny et al. (1) and Zhang et al. (2). Most studies using tension disk infiltrometers have been conducted at the soil surface, although Azevedo et al. (3) looked at infiltration properties of an Iowa loamy soil at 0.15 m depth, and Logsdon et al. (4), and Messing and Jarvis (5) conducted measurements at different depths under different agricultural tillages. Careful preparation of the soil surface is required to use a disk infiltrometer, which is to ensure even and sound contact of the disk with the soil surface. At each location, vegetation must be cut back to the soil surface, and a

Valves

215

fine layer of moist fine sand of the same diameter as the circular base of the infiltrometer should be applied, which must be smoothed out to remove any irregularities at the soil surface and improve contact between the disk and soil surface. Moist sand is essential as air-dry sand may readily fall down into surface-vented macropores, forming ‘wicks’ (5). The infiltrometer is then placed on the sand. The weight of the infiltrometer must not be too great so as compress the soil surface significantly, as this will restrict the water flux. Therefore, it is usual to make the water supply reservoir narrow so that the total volume of water held in the infiltrometer is low, resulting in reduced weight and also aiding accurate measurements of discharge. If a range of supply heads is to be used, then infiltration tests are normally conducted with the lowest supply head first. Reversal of this may lead to hysteresis where drainage occurs close to the disk while wetting continues near and at the infiltration front (6). Infiltration measurements should proceed until a steady state is achieved. Users should be careful to ensure that sunlight does not heat the supply reservoir significantly: this can be reduced by shading. Methods for analyzing the data from disk infiltrometers (e.g., hydraulic conductivity values and macropore contribution to infiltration) are given in Reynolds and Elrick (6). Typically, Wooding’s (7) solution for infiltration from a shallow pond is combined with Gardner’s (8) unsaturated hydraulic conductivity function. A range of assumptions is involved in using of these techniques, including that the hydraulic conductivity before the test is much less than that imposed under the infiltration experiment and that the soil below the tension disk is homogeneous, isotropic, and uniformly saturated, which are unlikely to be satisfied fully in most soils, and so it is necessary to evaluate the potential errors involved (9). Water levels in supply bottles for both the ring and disk infiltrometers can be measured automatically by using electronic pressure sensors with data recorders, thus allowing an experiment to continue unattended for several hours if the infiltration rates are very low.

Release valve

BIBLIOGRAPHY

Air entry ports

Water level

Water level

2. Zhang, Y., Smith, R.E., Butters, G.L., and Cardon, G.E. (1999). Analysis and testing of a concentric-disk tension infiltrometer. Soil Science Society of America Journal 63: 544–553.

o Bubble tower

1. Ankeny, M.D., Kaspar, T.C., and Horton, R. (1988). Design for an automated tension infiltrometer. Journal of the Soil Science Society of America 52: 893–896.

Mariotte column o

3. Azevedo, A.S., Kanwar, R.S., and Horton, R. (1998). Effect of cultivation on hydraulic properties of an Iowa soil using tension infiltrometers. Soil Science 163: 22–29. 4. Logsdon, S.D., Jordahl, J.L., and Karlen, D.L. (1993). Tillage and crop effects on ponded and tension infiltration rates. Soil and Tillage Research 28: 179–189. 5. Messing, I. and Jarvis, N.J. (1993). Temporal variation in the hydraulic conductivity of a tilled clay soil as measured by tension infiltrometers. Journal of Soil Science 44: 11–24.

Porous disk Figure 3. Schematic diagram of a tension disk infiltrometer.

6. Reynolds, E.D. and Elrick, D.E. (1991). Determination of hydraulic conductivity using a tension infiltrometer. Journal of the Soil Science Society of America 55: 633–639.

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SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY

7. Wooding, R.A. (1968). Steady infiltration from a shallow circular pond. Water Resources Research 4: 1259–1273. 8. Gardner, W.R. (1958). Some steady state solutions of the unsaturated moisture flow equation with application to evaporation from a water table. Soil Science 85: 228–232. 9. Holden, J., Burt, T.P., and Cox, N.J. (2001). Macroporosity and infiltration in blanket peat: the implications of tension disc infiltrometer measurements. Hydrological Processes 15: 289–303.

SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY SEGUN ADELANA University of Ilorin Ilorin, Nigeria

INTRODUCTION The word contaminant is a general term to describe dissolved constituents, nonaqueous phase liquids (NAPLs), or industrial solvents which, when added to water as a result of human activities, impair groundwater quality. To an important extent, the complexity of a groundwater contamination problem can be determined by whether or not NAPLs are present (1). Obviously, contaminants related to anthropogenic activities are one of the major threats facing groundwater resources today. Groundwater contamination can have a significant impact on the present and future use of groundwater resources as a source of drinking water for an increasing population living in cities and rural areas (2). In the investigation of problems posed by contaminants in the hydrologic environment, the modern approach is to use a combination of tracers. These tracers include radioactive isotopes (such as 3 H, 13 C, 14 C,), stable isotopes (18 O, 2 H), and dissolved geochemical constituents (Cl− , SO4 2− , NO3 − and some cations). An important facet of contaminant hydrogeology is the movement of liquids that are immiscible with water through the vadose zone as well as below the water table. Such liquids are usually called nonaqueous phase liquids. There are numerous cases of groundwater contamination involving organic liquids that are either insoluble or partially soluble in water. When the densities of these liquids are greater than that of water, they are called dense nonaqueous phase liquids (DNAPLs). Those whose densities are less than that of water are light nonaqueous phase liquids (LNAPLs). They could also be partially soluble in water, such that both dissolved and nonaqueous phases may be present (3–7). The major groups of contaminants in hydrogeology include radionuclides, trace metals, and nutrients. They could also be classified as inorganic and organic contaminants. These include synthetic organic compounds, hydrocarbons, and pathogens (7). Some types of these contaminants are inorganic chemicals that are either metallic or nonmetallic (1,8). Included in the group of organic contaminants are NAPLs, which naturally subdivide into LNAPLs and DNAPLs (7).

A wide variety of materials have been identified as contaminants found in groundwater. An extensive listing of these compounds is tabulated in (Reference 7, pp. 3–10). Many of these materials will dissolve in water to different degrees. Some of the organic compounds are only slightly soluble and will exist in both dissolved form and as an insoluble phase, which can also migrate through the ground (7). According to (2), the main relevant organic compounds of the LNAPLs are benzene, toluene, ethylbenzene, and xylenes (BTEX) from spills at gas stations and industrial facilities. The most common compounds of the group found in groundwater are perchloroethene (PCE), trichloroethene (TCE), cis-1,2-dichloroethene (cis1,2-DCE), and dichloroethane (DCA) (9). In terms of environmental contamination, the major concern is focused on those organic compounds that are not readily degraded by bacteria, either in the soil zone or in sewage treatment facilities. The presence of these substances in the surface environment is becoming increasingly pervasive (8). The sources that contribute to groundwater contamination are varied. Among these are land disposal of solid wastes, inorganic substances, metal and nonmetal trace elements. Other sources of groundwater contamination are from introducing of salts on roads to combat ice in winter, the infiltration of leachate from tailings in the mining industry, seepage from industrial water lagoons and septic tanks of urban waste disposal systems, wastewater discharge into rivers and lakes, agricultural wastes and fertilizers, saline or geogenic polluted water attracted by overexploitation of groundwater, and fossil fuels (1,7,8,10,11). Inorganic cations and anions occur in nature and may come from natural as well as anthropogenic sources. Some of the radionuclides are naturally occurring; others are man-made and come from the production and testing of nuclear weapons. There are three important attributes that distinguish sources of groundwater contamination: (1) their degree of localization, (2) their loading history, and (3) the kinds of contaminants emanating from them (1). From the work of Domenico and Schwartz (1) contaminants are organized by groups according to reaction type and mode of occurrence. Six categories have been tabulated: Category 1—Sources designed to discharge substances, for example, subsurface percolation (from septic tanks and cesspools), injection wells (hazardous and nonhazardous wastes), and land application. Category 2—Sources designed to store, treat, and/or dispose of substances; discharge through unplanned release, for example, landfills, open dumps (including illegal dumping waste), residential or local disposal wastes, waste tailings, material stockpiles (nonwaste), graveyards, animal burial, surface and underground storage tanks, and containers. Category 3—Sources designed to retain substances during transport or transmission, for example, pipelines and materials transport and transfer operations. Category 4—Sources discharging substances as a consequence of other planned activities, for example,

SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY

irrigation practices, pesticide applications, fertilizer applications, animal feeding operations, deicing salt applications, urban runoff, percolation of atmospheric pollutants, and mining and mining drainage. Category 5—Sources providing conduits or inducing discharge through altered flow patterns, for example, production wells (oil/gas wells, geothermal and heat recovery wells, water supply wells), other nonwaste wells (monitoring and exploration wells), and construction excavation. Category 6—Naturally occurring sources whose discharge is created and/or exacerbated by human activity, for example, groundwater–surface water interactions, natural leaching, saltwater intrusions and brackish water upconing, or intrusion of other poor-quality natural water. For details of sources and background concentrations of contaminants in groundwater, refer to Forstner and Wittman (10). More information on inorganic contaminants is available in Reference 7 (Chapter 6), Reference 1 (Chapter 17); methods of detecting groundwater contamination are fully discussed in Domenico and Schwartz (1), and the sources and hydrochemical behavior of contaminants are described in Chapter 9 of Reference 8 as well as Chapter 1 of Reference 11. As a general introduction to groundwater contamination and related literature, the following are suggested for further reading: Baetsle (12), Fuhriman and Barton (13), Hall (14), Scalf et al. (15), Miller et al. (16), Summers and Spiegal (17), Cherry et al. (18), Back and Cherry (19), Fried (20), Todd and McNulty (21), Wilson et al. (22), Freeze and Cherry (8), IAEA (23, 24), Pearson et al. (25), Fetter (7), Pankow and Cherry (9), Domenico and Schwartz (1), and Seiler (11). ENVIRONMENTAL ISOTOPES IN CONTAMINANT HYDROGEOLOGY Introduction For many years, environmental isotopes have been used in hydrogeology to address a range of issues such as evaluation of recharge areas, determination of groundwater residence time, and evaluation of surface water–groundwater interactions. Isotope studies applied to a wide spectrum of hydrologic problems related to both surface and groundwater resources as well as environmental studies in hydro-ecological systems are presently an established scientific discipline, called ‘‘Isotope Hydrology.’’ However, it is now becoming increasingly popular to use environmental isotopes as tracers to provide information about sources of contaminants and processes that affect these compounds in the groundwater flow system (2). In essence, isotope techniques are becoming an integral part of many hydrologic investigations and sometimes a unique tool in studies related to groundwater contamination. Stable and radioactive environmental isotopes have now been used for more than four decades to study hydrologic systems, particularly in determining the hydrodynamics of aquifers, and have proved useful for

217

understanding groundwater systems. The applications of isotopes in hydrogeology are based on the general concept of ‘‘tracing,’’ in which either intentionally introduced isotopes or naturally occurring (environmental) isotopes are employed. Environmental isotopes (either radioactive or stable) have the distinct advantage over injected (artificial) tracers in that they facilitate the study of various hydrologic processes on a much larger temporal and spatial scale through their natural distribution in a hydrologic system. Thus, environmental isotope methodologies are unique in regional studies of water resources to obtain the time and space integrated characteristics of groundwater systems. The use of artificial tracers generally is effective for site-specific, local applications (26,27). Environmental isotopes can now be used to trace the pathways as well as spatial distribution and temporal changes in pollution patterns for assessing pollution migration scenarios and planning for aquifer remediation (27,28). Recent research and the development of new analytical methods have led to fast expansion of the application of isotopes in hydrogeology. In particular is the introduction of techniques that make it possible to measure isotopic ratios of individual compounds at low concentrations (29). These have opened new and exciting possibilities for using isotopes in organic contaminant studies in groundwater. For example, nitrogen and carbon isotopes can now provide information regarding such attenuation processes as the denitrification and biodegradation of halogenated solvents, respectively (28). Environmental isotopes have, therefore, become useful tools in differentiating between different sources of organic contaminants or in assessing biotic and abiotic transformation processes associated with natural attenuation or engineered groundwater remediation techniques (30–32). The isotope method used to evaluate the degradation of organic contaminants relies on the frequent occurrence of small differences in the degradation rate between molecules of the same contaminant that contain different isotopes. Usually, the rate is slightly faster for molecules that contain light isotopes (e.g., 12 C) compared to those that contain heavy isotopes (e.g., 13 C) because bonds between light isotopes are weaker. As a result, degrading contaminants should contain increasingly more of the heavier isotope (13 C) as degradation proceeds, and thus the ratio between heavy and light isotopes (13 C/12 C) is expected to increase (2). Oxygen-18 in dissolved oxygen also has a significant potential as a tracer to evaluate the fate of oxygen in contaminated plumes. The sequel to these environmental isotope studies has focused on the light elements and their isotopes: hydrogen (1 H, 2 H, 3 H), carbon (12 C, 13 C, 14 C), nitrogen (14 N, 15 N), oxygen (16 O, 18 O), and sulfur (32 S, 34 S). These are the most important elements in biological systems that occur in relatively great abundance and also participate in most geochemical reactions (33). Other less important and less abundant isotopes (17 O, 33 S, 36 S) are not often employed in environmental studies but have significance in very specific studies. Table 1 lists the isotopes of major elements used in environmental studies and their average abundances in natural compounds.

218

SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY Table 1. Isotopes of Major Elements Used in Environmental Studiesa

Element Hydrogen

Isotopes

Average Terrestrial Abundance, (%)

1

H H 3 H 12 C 13 C 14 C 16 O 17 c O 18 O 14 N 15 N 32 S 33 c S 34 S 36 c S 84 Src 86 Sr 87 Sr 88 Src 234 U 235 c U 238 U 2

Carbon

Oxygen

Nitrogen Sulphur

Strontium

Uranium

99.984 0.015 10−14 to 10−16 98.89 1.11 ∼ 10−10 99.76 0.037 0.1 99.34 0.366 95.02 0.75 4.21 0.02 0.56 9.86 ∼7.02 82.56 ∼0.0056 0.7205 99.274

Commentsb

Radioactive, t1/2 = 12.35 years Radioactive, t1/2 = 5730 years

Radioactive, t1/2 = 2.47×105 years Radioactive, t1/2 = 7.13×108 years Radioactive, t1/2 = 4.51×109 years

a

Reference 33. t1/2 = half-life. c These isotopes are not presently used in environmental studies. b

Uses of Isotopes in Contaminant Hydrogeology General uses of environmental isotopes in contaminant studies are often based on evaluating the various sources of contaminants in both surface and groundwater. The isotope approach to evaluating these sources relies mostly on the distinct isotope composition, that can characterize the different sources of contaminant (2,34). Environmental isotopes have found good use as tracers both artificially and naturally (26,35–38). For general review of the use of environmental isotopes in contaminant hydrogeology and potential future use of these tracers in vulnerability studies in groundwater, see References 2,27 and 28. Environmental isotopes have been found widely applicable, especially in identifying the origin of water that is infiltrating an aquifer (28). Obviously, there is a relationship between recharge and the vulnerability of groundwater to contamination because the transport of many contaminants to the saturated zone of the aquifer occurs in the dissolved phase as part of the recharge process (39). Generally, 18 O and 2 H, among the various environmental isotopes, are most widely used in defining groundwater recharge areas as well as their origins. Changes in the isotopic relationships between 18 O/16 O and 2 H/1 H, during the evaporation and condensation of water allow defining its origin (28). New isotope applications are being developed in organic contaminant studies of groundwater. These studies have shown that carbon isotopes allow very sensitive evaluation of the fate of precursors and by-products of the biodegradation of chlorinated compounds, which are the most common organic contaminants found in groundwater. Potentially new applications include 18 O to evaluate the fate of oxygen in contaminant plumes and

11

B in salinization and nitrate studies of groundwater (2). Table 2 shows the potential/common applications of isotopes in groundwater pollution studies. Boron-11, a stable isotope of boron, has also been used to differentiate nitrate from sewage and fertilizers as well as to trace the sources of salinity in groundwater. This approach was used to differentiate isotopic composition of injected treated wastewater and irrigation-affected water (δ 11 B > 40‰) distinctly and to trace the effect of injection into an alluvial aquifer near El Paso, Texas (40). It has also been possible to approach the tracing of various organic pollutants (chlorinated hydrocarbons) in groundwater by using a combination of the isotopes 13 C/12 C, 2 H/1 H, and 37 Cl/35 Cl. For example, the isotopic ratios 37 Cl/35 Cl and 10 B/11 B have been used recently in pollution studies of groundwater (41,42). The strontium isotopic ratio (87 Sr/86 Sr) is also a valuable tracer for studies related to the mixing and origin of mineralized groundwater. It is a tracer for the origin of salinity, groundwater movement, and water–rock interactions (43,44). In the past, a combination of stable and radioactive isotopes that has dissolved geochemical constituents has been used in tracing the migration of contaminants in groundwater (45–47). According to Hirata and Varnier (28), isotope techniques are useful in determining groundwater pollution vulnerability, to the extent that they allow for better understanding of the degree of access provided for penetration of contaminants and the attenuative capacity offered by the geologic medium (Table 3). The most frequently used environmental isotopes in contaminant hydrogeology include nitrogen isotopes (14 N/15 N), chlorine - 37, the sulphur isotope (34 S), carbon

SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY

219

Table 2. Common/Potential Isotopes Used in Groundwater Pollution Studiesa Isotope(s)

Potential/Common Applications −

13

C in HCO3 and CH4

34

S and 18 O in SO4 2−

15

N and 18 O in NO3 − , NH4 − , N2

11

B in B (OH)4 − and B (OH)3 −

Chlorine-37 (37 Cl) Tritium (3 H) Strontium isotope ratio (87 Sr/86 Sr) Uranium isotopes (234 U, 238 U) a

Identifying of methane sources Monitoring CO2 and CH4 production in biodegradation studies Identifying of sources of pollution Acidification and acid mine drainage Origin of nitrates Sources of pollution Microbial denitrification processes Identifying of pollution sources Characteristics of brines, source of salinity Identifying pollution sources Characteristics of brines, sources of salinity Indicators of recent pollution/recharge Water transport in the unsaturated zone Tracer for the origin of salinity, mixing, and origin of mineralized groundwater, groundwater movement, and water–rock interactions. Sensitive indicators of radioactive pollution in natural waters, Rock–water interaction, dating of old groundwater

Modified from Reference 27.

Table 3. The Use of the Isotope Technique for Groundwater Pollution Vulnerability Characterizationa Component of Aquifer Pollution Vulnerability Hydraulic accessibility of the saturated aquifer to penetration of pollutants

Attenuative capacity of strata overlying the saturated zone, resulting from the physicochemical retention or reaction of pollutants

Information Related to

Isotope Technique

Identification of recharge/discharge areas Degree of aquifer confinement Definition of period of recharge Identification of paleowaters Interconnection of surface and groundwater Determination of transit time of water through the saturated and unsaturated zones (advective flow)

Identification of physico chemical reactions of pollutants (degradation)

18

O in H2 O H in H2 O 85 Kr flow through fissured rock 3 H in H2 O (environmental and applied tracer) 51 Cr, 58 Co, 60 Co, 131 I (applied tracers) 2

Dating young water 3 H 3 He Dating midage water 39 Ar 14 C, correction with 13 C in HCO3 − 81 Kr Dating old water 81 Kr 234 U 15 N and 18 O in NO3 − , NH4 + , N2 , microbial denitrification processes 13 C, biodegradation of synthetic solvents and fuel 34 S, oxidation/reduction of sulfur

Pollution hazard factors

Information related to

Isotope technique

Definition of pollutant origin

Identification of source of pollution

34

a

S and 18 O in SO4 2− , acidification and acid mine drainage 37 Cl, 11 B, in B(OH)4 − and B(OH)3 − characteristics of brines; sources of salinity 15 N and 18 O in NO3 − , NH4 + , N2 18 O, 2 H in H2 O, salinization mechanisms, recycling of irrigation water 14 C and 13 C for origin of sources of organic compounds

After Reference 28.

isotopes (13 C, 14 C), and the boron isotope (11 B). Those of the water molecules (18 O, 2 H) are also sometimes used with some of these isotopes in groundwater contamination studies. Following are brief descriptions of the isotopes commonly employed in contaminant hydrogeologic studies.

Nitrogen Isotopes (14 N/15 N). The most common contaminant identified in groundwater is dissolved nitrogen in the form of nitrate (NO3 − ). Other forms of inorganic nitrogen are nitrite (NO2 − ), nitrogen gas (N2 ), ammonia (NH4 + ), and cyanide (CN− ). Nitrogen is also a major constituent

220

SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY

of organic matter as amino acids (7). These contaminants are becoming widespread due to increased agricultural activities worldwide and disposal of sewage on or beneath the land surface. Nitrogen occurs as two isotopes in nature: 14 N(≈99.6%) and 15 N(≈0.36%). 14 N is more abundant in the atmosphere than 15 N. Hence, the 15 N/14 N ratio in nitrate can be used to distinguish the origin of nitrates either from animal/human wastes or from the application of chemical fertilizers (48,49). The origin of contamination by nitrates can be determined through the relationship between 15 N/14 N and standard values (δ 15 N). Combined isotope analyses of nitrogen and oxygen in NO3 − leave fingerprints on natural and anthropogenic sources of nitrate; on microbial denitrification, nitrification and biological fixation processes; and on the nitrogen budget in groundwater (50–53). Therefore, the δ 15 N values of dissolved nitrates, ammonium, and organic nitrogen in soil water are well distinguished from one region to another. The δ 15 N of animal waste is between +5 to 15‰, and fertilizers have a δ 15 N of about 0‰. However, for nitrogenous fertilizers, these values lie between −4‰ and +4‰ (50,54). For more information and related nitrogenous inputs in groundwater, refer to Foster and Crease (55), Kreitler and Jones (56), L´etolle (57), Kreitler and Browning (58), White (59), Flipse et al. (60), Heaton (60), Mariotti (61), Flipse ¨ and Bonner (62), Tessendorff (63), Hubner (64), Mariotti (65), B¨ottcher et al. (51), Exner and Spalding (66), Aravena et al. (52), Buzek (67), R´ev´esz et al. (68), Pauwels et al. (69), Tore et al. (70), and Anwar and Ghosh (71). Chlorine-37. Naturally, chlorine exists with two common stable isotopes: 35 Cl (≈75.7%) and 37 Cl (≈24.2%). Chloride ions are generally not reactive and as such, do not participate in redox reactions, are not sorbed onto mineral or organic surfaces, and do not form insoluble precipitates. As a result, 37 Cl is sometimes used as a conservative tracer in groundwater studies. Data are expressed as δ 37 Cl with respect to the standard mean oceanic chloride (SMOC). The ratio 37 Cl/35 Cl is measured by isotope ratio mass spectrometry (IRMS). Precision must be better than ± 0.1‰ (72). δ 37 Cl values contain information on the origin of chloride ions in fresh and polluted groundwater as well as in subsurface brines (30,41,42). In geochemical studies generally, chlorine isotopes are not highly fractionated, but research has shown that measurable fractionations can be used to distinguish sources of groundwater mixing (73). The 37 Cl content of chloro contaminants in groundwater can fingerprint the sources of contaminants and salinity. It is therefore a potential tool in contaminant hydrogeology (27,43). Sulphur Isotope (34 S). Sulphur has four stable isotopes: S (95.02%), 33 S (0.75%), 34 S (4.21%), and 36 S (0.02%). The abundance ratio of 34 S and 32 S is generally given as a δ 34 S value. Iron sulfide from the troilite phase of the Diablo Canyon iron meteorite (DCT) (that has a 32 S/34 S ratio of 22.220) is conventionally used as a standard (72). 34 S is one of the distinct isotopes that can be used to characterize different sources of sulfate contamination. 32

For example, high sulfate in aquifers in areas close to mine tailings could be associated with the oxidation of sulfide minerals in the tailings. However, other sources of sulfate in aquifers such as gypsum of marine origin, input from volcanic sources, or sulfate from recirculated water from agriculture have been identified (2). Generally, sulfate produced from sulfide minerals is much more depleted in 34 S and 18 O than marine sulfate (74,75). Moreover, owing to isotope fractionation during sulfate reduction, 34 S and 18 O in sulfate can also be used to evaluate the fate of sulfate in groundwater flow systems (76). The isotopic composition of sulfate can also be an indicator of anthropogenic pollution of groundwater (75). Carbon Isotopes (13 C, 14 C). Carbon has three isotopes in nature: 12 C, 13 C, and 14 C. The heavy carbon isotope, 14 C, is unstable and decays radioactively into 14 N. The stable isotopes of carbon, 13 C and 12 C, have an average ratio of about 1:100 (72). The isotopic composition of carbon in the dissolved carbon constituents of groundwater is very variable. The sources of carbon dissolved in groundwater are soil CO2 , CO2 of geogenic origin or from magmatic CO2 (from deep crustal or mantle sources) or in fluid inclusions, living and dead organic matter in soils and rocks, methane, and carbonate minerals (53). Each of these sources has a different carbon isotopic composition and contributes to total dissolved carbon in various proportions. Therefore, the isotopic composition of dissolved inorganic carbon compounds in groundwater has a wide range of δ 13 C values. Carbon isotopes (13 C, 14 C) can be used to evaluate the origin of methane in landfills (2). The origin of methane is a common problem in landfills located in old quarries. Methane in this type of environment can be thermocatalytic in origin, associated with gases present in the bedrock or/and originating in the landfill itself (2). According to Schoell (77), gases of thermocatalytic origin are generally more enriched in carbon-13 (δ 13 C > −50‰) than methane of biogenic origin (δ 13 C < −50‰). In addition, Hackley et al. (78) proved that catalytic gases are devoid of carbon-14, whereas methane in landfills has a 14 C content higher than 100% modern carbon (pmc). Obviously, due to carbon-14 from thermonuclear explosions, any carbon in domestic waste that was deposited in a landfill after 1950 is usually characterized by a 14 C content higher than 100 pmc (2). Furthermore, carbon isotopes have also been employed in monitoring CO2 and CH4 production in biodegration studies in unsaturated areas impacted by organic contamination (79,80). Boron Isotope (11 B). Natural boron has two stable isotopes, 11 B (≈80%) and 10 B (≈20%). The variation in the ratio of these two isotopes is expressed in δ 11 B (‰) with respect to SRM-951 NBS standard (81). Generally, boron is present in groundwater mainly as trigonal undissociated boric acid, B (OH)3 0 , and its conjugate base, tetrahedral borate, B (OH)4 − . Tetrahedrally coordinated boron is more isotopically depleted (i.e., lower in δ 11 B) relative to trigonally coordinated boron; thus the δ 11 B of dissolved boron in groundwater is usually higher than that in aquifer rocks (81).

SUMMARY OF ISOTOPES IN CONTAMINANT HYDROGEOLOGY

Boron is used as tracer in groundwater because of its high solubility in aqueous solution, natural abundance in all waters, and the lack of effects from evaporation, volatilization, and oxidation–reduction reactions. The concentrations of boron in pristine groundwaters are generally low (S]tot = [>S] + [>SA], Eq. 1d can be rewritten as [>SA] = [>S]tot

Kad [A] Kad [A] + 1

≡ X(OH)0 ⇔≡ XO− + H+ app

app

Ka2

(1b)

app

where Ka1 and Ka2 are the apparent equilibrium constants of reactions 1a and 1b. The participation of hydrogen ions in the above reactions indicates that the relative concentrations of the protonated and deprotonated surface species, and hence the amount of surface charges, are controlled by the pH of the aqueous environment. The pH at which there are equal amounts of >XOH2 + and >XO− (i.e., an uncharged surface) is defined as the pH of zero net proton charge, below which the surface is positively charged for >XOH2 + is more abundant than >XO− and above which the surface is negatively charged because of the excess of >XO− . In the presence of other sorbing ions, say M (n+1)+ and (n+1)− , in solutions, competitive complexation becomes L operative as M can exchange for H and L for OH at the >X(OH)0 sites to result in the formation of surface complexes such as >XOM n+ and >XLn− . Surface complexation under this condition becomes adsorption where both chemical bonding and electrostatic interaction take place between solute particles and surface atoms (3). Empirical approaches to adsorption employ simple equations to describe the adsorption isotherm (relationship between the solution concentration or activity and the surface concentration of an adsorbate), without concerning the specific interfacial interactions between adsorbate and adsorbent. For example, the Langmuir isotherm, one of the most common empirical models, simply treats adsorption as a free surface site >S occupied by an adsorbate A to form a sorbed site >SA: >S + A ⇔ >SA

[>SA] = kf [A]n

(1f)

where both kf and n are constants. The Freundlich model can be thought of as the Langmuir isotherm integrated over many types of surface sites (4). In the surface complexation model, adsorption is treated as a chemical reaction

The adsorption constant Kad given by the mass law is Kad =

[>SA] [>S][A]

(1d)

(1g)

The equilibrium constant K app K app =

[>XOM n+ ][H+ ] [>XOH0 ][M (n+1)+ ]

= K int K coul

(1h)

now has contributions from chemical (intrinsic) as well as electrostatic (coulombic) interactions between the reactants. Quantitatively, K coul = exp(FZ), where Z is the change of charges in surface species because of adsorption. The surface potential  is a function of surface charge σ , and the charge-potential relationship can be derived from electrochemistry theories. A common one σ = (8RTεε0 I)1/2 sinh(zF/2RT)

(1i)

is given by the Gouy–Chapmen theory (diffuse doublelayer model). Here ε is the relative dielectric constant of water, ε0 is the permittivity of vacuum, I is the ionic strength of the bulk solution, and z is the valence of the symmetrical background electrolyte. The quantity of surface charge can be computed from the concentration of charged surface species. For Eq. 1g, σ = F[≡ XOM (n−1)+ ]/AS

(1j)

where A and S are the specific area (m2 /g) and solid concentration (g/l) of the adsorbent material, respectively. Overall, assuming the total amount of surface sites in Eq. 1g is approximated by [>XO]tot = [>X(OH)0 ] + [>XOM n+ ], one can rearrange Eq. 1h into  exp(nF)K int [M (n+1)+ ] exp(nF)K int [M (n+1)+ ] + [H+ ] (1k) It becomes apparent that, after inspecting Eqs. 1e, 1f, and 1h, the surface complexation approach is advantageous over the empirical isotherms in that the formal provides additional insight into the dependence of adsorption upon pH and ionic strength (through Eq. 1i). Moreover, the 

[>XOM n+ ] = [>XO]tot

(1c)

(1e)

∼ [>S]tot , High concentrations of [A] lead to [>SA] = whereas low values of [A] reduces Eq. 1e to a linear form of [>SA] ∼ = Kd [A]. Another common empirical adsorption model, the Freundlich isotherm, uses a logarithmic relationship between [>SA] and [A] so that

>X(OH)0 + M (n+1)+ ⇔ >XOM n+ + H+

and

567

568

WATER/ROCKS INTERACTION

use of K int in surface complexation models makes it possible to compare results from different experiments because, unlike empirical parameters such as Kad (Eq. 1d) and Kf (Eq. 1f), the intrinsic surface complex formation constant is not experimental condition dependent. It is noteworthy to point out that although surface complexation reactions can occur on any mineral crystals, effective adsorption requires small grain sizes. Because of the inverse relationship between particle size and specific surface area (defined as the ratio of surface area to volume of a particles), small grains (