Environmental Management Handbook, Second Edition – Six Volume Set [2 ed.] 1138342629, 9781138342620

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Managing Global Resources and Universal Processes

Environmental Management Handbook, Second Edition Edited by Brian D. Fath and Sven E. Jørgensen

Volume 1 Managing Global Resources and Universal Processes Volume 2 Managing Biological and Ecological Systems

Volume 3 Managing Soils and Terrestrial Systems

Volume 4 Managing Water Resources and Hydrological Systems

Volume 5 Managing Air Quality and Energy Systems

Volume 6 Managing Human and Social Systems

Managing Global Resources and Universal Processes Second Edition

Edited by

Brian D. Fath and Sven E. Jørgensen Assistant to Editor

Megan Cole

Cover photo: Paphos Archeological Park, Cyprus, N. Fath

Second edition published 2021 by CRC Press 6000 Broken Sound Parkway NW, Suite 300, Boca Raton, FL 33487-2742 and by CRC Press 2 Park Square, Milton Park, Abingdon, Oxon, OX14 4RN © 2021 Taylor & Francis Group, LLC First edition published by CRC Press 2013 CRC Press is an imprint of Taylor & Francis Group, LLC Reasonable efforts have been made to publish reliable data and information, but the author and publisher cannot assume responsibility for the validity of all materials or the consequences of their use. The authors and publishers have attempted to trace the copyright holders of all material reproduced in this publication and apologize to copyright holders if permission to publish in this form has not been obtained. If any copyright material has not been acknowledged please write and let us know so we may rectify in any future reprint. Except as permitted under U.S. Copyright Law, no part of this book may be reprinted, reproduced, transmitted, or utilized in any form by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying, microfilming, and recording, or in any information storage or retrieval system, without written permission from the publishers. For permission to photocopy or use material electronically from this work, access www.copyright.com or contact the Copyright Clearance Center, Inc. (CCC), 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400. For works that are not available on CCC please contact [email protected] Trademark notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation without intent to infringe. ISBN: 978-1-138-34263-7 (hbk) ISBN: 978-0-429-34613-2 (ebk) Typeset in Minion by codeMantra

Contents Preface. ...................................................................................................................... ix Editors . ...................................................................................................................... xi Contributors ......................................................................................................... ..xiii

Section i Anthropogenic chemicals: Human Manufactured and Activities

1

Acaricides ......................................................................................................... . 3

2

Endocrine Disruptors . ....................................................................................... 9

3

Herbicides .... ................................................................................................... . 27

4

Herbicides: Non-Target Species Effects . . ....................................................... . 33

5

Insecticides: Aerial Ultra-Low-Volume Application ..................................... . 49

6

Neurotoxicants: Developmental Experimental Testing ................................ . 53

7

Persistent Organic Pesticides . .......................................................................... 61

8

Pollutants: Organic and Inorganic . ................................................................ 69

9

Pollution: Genotoxicity of Agrotoxic Compounds ....................................... ..81

10

Pollution: Pesticides in Agro-Horticultural Ecosystems ............................. .. 99

11

Pollution: Pesticides in Natural Ecosystems . ................................................. 111

Doug Walsh

Vera Lucia S.S. de Castro Malcolm Devine Céline Boutin He Zhong

Vera Lucia S.S. de Castro Gamini Manuweera A. Paul Schwab

Vera Lucia S.S. de Castro and Paola Poli J.K. Dubey and Meena Thakur J.K. Dubey and Meena Thakur

v

vi

Contents

12

Polychlorinated Biphenyls (PCBs) ................................................................ . 119

13

Toxic Substances . ........................................................................................... 139

Marek Biziuk and Angelika Beyer Sven Erik Jørgensen

Section ii

natural elements and chemicals

14

Allelochemics . ................................................................................................ 149

15

Aluminum ...................................................................................................... 155

16

Boron: Soil Contaminant . .............................................................................. 175

17

Cadmium: Toxicology . ................................................................................... 181

18

Carbon: Soil Inorganic . .................................................................................. 185

19

Chromium . ..................................................................................................... 195

20

Cobalt and Iodine . ......................................................................................... 203

21

Copper . .......................................................................................................... 209

22

Globalization . ................................................................................................. 213

23

Heavy Metals . ................................................................................................ 223

24

Inorganic Carbon: Composition and Formation ........................................ . 229

25

Lead: Ecotoxicology ..................................................................................... . 235

26

Lead: Regulations . ......................................................................................... 243

27

Mercury . ........................................................................................................ 257

28

Mycotoxins ................................................................................................... ..261

John Borden

Johannes Bernhard Wehr, Frederick Paxton Cardell Blamey, Peter Martin Kopittke, and Neal William Menzies Rami Keren

Sven Erik Jørgensen Donald L. Suarez

Bruce R. James and Dominic A. Brose Ronald G. McLaren

David R. Parker and Judith F. Pedler Alexandru V. Roman Mike J. McLaughlin

Larry P. Wilding and H. Curtis Monger Sven Erik Jørgensen Lisa A. Robinson

Sven Erik Jørgensen J. David Miller

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Contents

29

Nitrogen ......................................................................................................... 269

30

Phenols . .......................................................................................................... 275

31

Phosphorus: Agricultural Nutrient .............................................................. 303

32

Potassium.. ...................................................................................................... 315

33

Radionuclides ................................................................................................ 323

34

Rare Earth Elements .. .................................................................................... 331

35

Strontium ....................................................................................................... .339

36

Sulfur . ........................................................................................................... . 345

37

Sulfur Dioxide . ............................................................................................... 355

38

Vanadium and Chromium Groups ............................................................. .. 367

Oswald Van Cleemput and Pascal Boeckx Leszek Wachowski and Robert Pietrzak

John Ryan, Hayriye Ibrikci, Rolf Sommer, and Abdul Rashid Philippe Hinsinger Philip M. Jardine

Zhengyi Hu, Gerd Sparovek, Silvia Haneklaus, and Ewald Schnug Silvia Haneklaus and Ewald Schnug

Ewald Schnug, Silvia Haneklaus, and Elke Bloem Marianna Czaplicka and Witold Kurylak Imad A.M. Ahmed

Section iii

Basic environmental Processes

39

Adsorption . .................................................................................................... 387

40

Cadmium and Lead: Contamination ............................................................ .417

41

Heavy Metals: Organic Fertilization Uptake ............................................... 427

42

Inorganic Carbon: Global Carbon Cycle ....................................................... 431

43

Inorganic Carbon: Modeling ........................................................................ 435

44

Inorganic Compounds: Eco-Toxicity ............................................................ .441

45

Leaching . ....................................................................................................... 447

Puangrat Kajitvichyanukul and Jirapat Ananpattarachai Gabriella Kakonyi and Imad A.M. Ahmed

Ewald Schnug, Alexandra Izosimova, and Renata Gaj William H. Schlesinger

Leslie D. McFadden and Ronald G. Amundson Sven Erik Jørgensen Lars Bergström

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Contents

46

Aquatic Communities: Pesticide Impacts .................................................... .451

47

Phosphorus: Riverine System Transport ...................................................... 465

48

Nitrogen: Biological Fixation ....................................................................... . 475

49

Nutrients: Best Management Practices . ......................................................... 481

50

Nutrients: Bioavailability and Plant Uptake ................................................ . 501

51

Nutrient–Water Interactions .. ........................................................................ 511

52

Pollution: Non-Point Source . ......................................................................... 515

53

Pollution: Point Sources .. ............................................................................. .. 519

54

Radioactivity . ................................................................................................. 527

55

Telecouplings . ................................................................................................ 539

David P. Kreutzweiser and Paul K. Sibley

Andrew N. Sharpley, Peter Kleinman, Tore Krogstad, and Richard McDowell Mark B. Peoples

Scott J. Sturgul and Keith A. Kelling Niels Erik Nielsen

Ardell D. Halvorson

Ravendra Naidu, Mallavarapu Megharaj, Peter Dillon, Rai Kookana, Ray Correll, and W.W. Wenzel Ravendra Naidu, Mallavarapu Megharaj, Peter Dillon, Rai Kookana, Ray Correll, and W.W. Wenzel Bogdan Skwarzec

Vilma Sandström

Index . ..................................................................................................................... 545

Preface Given the current state of the world as compiled in the massive Millennium Ecosystem Assessment Report, humans have changed ecosystems more rapidly and extensively during the past 50 years than in any other time in human history. These are unprecedented changes that need certain action. As a result, it is imperative that we have a good scientific understanding of how these systems function and good strategies on how to manage them. In a very practical way, this multivolume Environmental Management Handbook provides a comprehensive reference to demonstrate the key processes and provisions for enhancing environmental management. The experience, evidence, methods, and models relevant for studying environmental management are presented here in six stand-alone thematic volumes, as follows: VOLUME 1 – Managing Global Resources and Universal Processes VOLUME 2 – Managing Biological and Ecological Systems VOLUME 3 – Managing Soils and Terrestrial Systems VOLUME 4 – Managing Water Resources and Hydrological Systems VOLUME 5 – Managing Air Quality and Energy Systems VOLUME 6 – Managing Human and Social Systems In this manner, the handbook introduces in the first volume the general concepts and processes used in environmental management. The next four volumes deal with each of the four spheres of nature (biosphere, geosphere, hydrosphere, and atmosphere). The last volume ties the material together in its application to human and social systems. These are very important chapters for a wide spectrum of students and professionals to understand and implement environmental management. In particular, features include the following: • The first handbook that demonstrates the key processes and provisions for enhancing environmental management. • Addresses new and cutting-edge topics on ecosystem services, resilience, sustainability, food–energy–water nexus, socio-ecological systems, etc. • Provides an excellent basic knowledge on environmental systems, explains how these systems function, and gives strategies on how to manage them. • Written by an outstanding group of environmental experts. Since the handbook covers such a wide range of materials from basic processes, to tools, technologies, case studies, and legislative actions, each handbook entry is further classified into the following categories: APC: Anthropogenic chemicals: The chapters cover human-manufactured chemicals and their activities COV: Indicates that the chapters give comparative overviews of important topics for environmental management ix

x

Preface

CSS: The chapters give a case study of a particular environmental management example DIA: Means that the chapters are about diagnostic tools: monitoring, ecological modeling, ecological indicators, and ecological services ELE: Focuses on the use of legislation or policy to address environmental problems ENT: Addresses environmental management using environmental technologies NEC: Natural elements and chemicals: The chapters cover basic elements and chemicals found in nature PRO: The chapters cover basic environmental processes. Overall, these volumes will be a valuable resource for all libraries supporting programs in environmental science and studies, earth science, geography, and policy. In this volume, #1, the collection of over 50 entries provides an overview of global resources and universal processes. This serves as a good introduction to the key aspects of environmental management and includes descriptions of elements of the periodic table as well as organic and inorganic processes leading to pollution and alteration of natural conditions. A new chapter on telecoupling shows the long distance relations and interactions that mark most environmental systems. Brian D. Fath Brno, Czech Republic December 2019

Editors Brian D. Fath is Professor in the Department of Biological Sciences at Towson University (Maryland, USA) and Senior Research Scholar at the International Institute for Applied Systems Analysis (Laxenburg, Austria). He has published over 180 research papers, reports, and book chapters on environmental systems modeling, specifically in the areas of network analysis, urban metabolism, and sustainability. He has co-authored the books A New Ecology: Systems Perspective (2020), Foundations for Sustainability: A Coherent Framework of Life–Environment Relations (2019), and Flourishing within Limits to Growth: Following Nature’s Way (2015). He is also Editor-in-Chief for the journal Ecological Modelling and Co-Editor-in-Chief for Current Research in Environmental Sustainability. Dr. Fath was the 2016 recipient of the Prigogine Medal for outstanding work in systems ecology and twice a Fulbright Distinguished Chair (Parthenope University, Naples, Italy in 2012 and Masaryk University, Czech Republic in 2019). In addition, he has served as Secretary General of the International Society for Ecological Modelling, Co-Chair of the Ecosystem Dynamics Focus Research Group in the Community Surface Modeling Dynamics System, and member and past Chair of Baltimore County Commission on Environmental Quality. Sven E. Jørgensen (1934–2016) was Professor of environmental chemistry at Copenhagen University. He received a doctorate of engineering in environmental technology and a doctorate of science in ecological modeling. He was an honorable doctor of science at Coimbra University (Portugal) and at Dar es Salaam (Tanzania). He was Editor-in-Chief of Ecological Modelling from the journal inception in 1975 until 2009. He was Editor-in-Chief for the Encyclopedia of Environmental Management (2013) and Encyclopedia of Ecology (2008). In 2004, Dr. Jørgensen was awarded the Stockholm Water Prize and the Prigogine Medal. He was awarded the Einstein Professorship by the Chinese Academy of Sciences in 2005. In 2007, he received the Pascal Medal and was elected a member of the European Academy of Sciences. He had published over 350 papers, and has edited or written over 70 books. Dr. Jørgensen gave popular and well-received lectures and courses in ecological modeling, ecosystem theory, and ecological engineering worldwide.

xi

Contributors Imad A.M. Ahmed Lancaster Environment Center Lancaster University Lancaster, United Kingdom

Frederick Paxton Cardell Blamey School of Agriculture and Food Sciences The University of Queensland St. Lucia, Queensland, Australia

Ronald G. Amundson College of Natural Resources University of California—Berkeley Berkeley, California

Elke Bloem Institute for Crop and Soil Science Julius Kuhn Institute (JKI) Braunschweig, Germany

Jirapat Ananpattarachai Faculty of Engineering Center of Excellence for Environmental Research and Innovation Naresuan University Phitsanulok, Thailand Lars Bergström Department of Soil Science Swedish University of Agricultural Sciences (SLU) Uppsala, Sweden Angelika Beyer Department of Analytical Chemistry Chemical Faculty Gdansk University of Technology Gdansk, Poland Marek Biziuk Department of Analytical Chemistry, Chemical Faculty Gdansk University of Technology Gdansk, Poland

Pascal Boeckx Faculty of Agricultural and Applied Biological Sciences University of Ghent Ghent, Belgium John Borden Department of Biological Sciences Simon Fraser University Burnaby, British Columbia, Canada Céline Boutin Science and Technology Branch Environment Canada Carleton University Ottawa, Ontario, Canada Dominic A. Brose University of Maryland College Park, Maryland Ray Correll Commonwealth Scientific and Industrial Research Organization (CSIRO) Adelaide, South Australia, Australia

xiii

xiv

Marianna Czaplicka Institute of Non-Ferrous Metals and Department of Analytical Chemistry Silesian University of Technology Gliwice, Poland Vera Lucia S.S. de Castro Ecotoxicology and Biosafety Laboratory Brazilian Agricultural Research Corporation (Embrapa Environment) São Paulo, Brazil Malcolm Devine Aventis CropScience Canada Co. Saskatoon, Saskatchewan, Canada Peter Dillon Commonwealth Scientific and Industrial Research Organization (CSIRO) Adelaide, South Australia, Australia J.K. Dubey Department of Entomology Dr. Y.S. Parmar University of Horticulture and Forestry Solan, India Renata Gaj Institute of Soil Science Agricultural University Poznan, Poland

Contributors

Zhengyi Hu Institute of Soil Science Chinese Academy of Sciences Nanjing, China Hayriye Ibrikci Soil Science and Plant Nutrition Department Cukurova University Adana, Turkey Alexandra Izosimova St. Petersburg Agricultural Physical Research Institute St. Petersburg, Russia Bruce R. James University of Maryland College Park, Maryland Philip M. Jardine Oak Ridge National Laboratory Oak Ridge, Tennessee Sven Erik Jørgensen Institute A, Section of Environmental Chemistry Copenhagen University Copenhagen, Denmark

Ardell D. Halvorson U.S. Department of Agriculture (USDA) Fort Collins, Colorado

Puangrat Kajitvichyanukul Faculty of Engineering Center of Excellence for Environmental Research and Innovation Naresuan University Phitsanulok, Thailand

Silvia Haneklaus Institute for Crop and Soil Science Julius Kuhn Institute (JKI) Braunschweig, Germany

Gabriella Kakonyi Kroto Research Institute Sheffield University Sheffield, United Kingdom

Philippe Hinsinger Sun and Environment Unit National Institute for Agricultural Research (INRA) Montpellier, France

Keith A. Kelling Professor Emeritus, Department of Soil Science University of Wisconsin—Extension Madison, Wisconsin

xv

Contributors

Rami Keren Agricultural Research Organization of Israel Bet-Dagan, Israel Peter Kleinman Pasture Systems and Watershed Management Research Unit U.S. Department of Agriculture (USDA) University Park, Pennsylvania Rai Kookana Commonwealth Scientific and Industrial Research Organization (CSIRO) Adelaide, South Australia, Australia Peter Martin Kopittke School of Agriculture and Food Sciences The University of Queensland St. Lucia, Queensland, Australia David P. Kreutzweiser Canadian Forest Service Natural Resources Canada Sault Sainte Marie, Ontario, Canada Tore Krogstad Department of Plant and Environmental Sciences Norwegian University of Life Science Aas, Norway Witold Kurylak Institute of Non-Ferrous Metals Gliwice, Poland Gamini Manuweera Chemicals and Health Branch United Nations Environmental Program Geneva, Switzerland

Ronald G. McLaren Soil, Plant, and Ecological Sciences Division Lincoln University Canterbury, New Zealand Mike J. McLaughlin Land and Water, Commonwealth Scientific and Industrial Research Organization (CSIRO) Glen Osmond, South Australia, Australia Mallavarapu Megharaj Commonwealth Scientific and Industrial Research Organization (CSIRO) Adelaide, South Australia, Australia Neal William Menzies School of Agriculture and Food Sciences The University of Queensland St. Lucia, Queensland, Australia J. David Miller Department of Chemistry Carleton University Ottawa, Ontario, Canada H. Curtis Monger Department of Agronomy and Horticulture New Mexico State University Las Cruces, New Mexico Ravendra Naidu Commonwealth Scientific and Industrial Research Organization (CSIRO) Adelaide, South Australia, Australia Niels Erik Nielsen Plant Nutrition and Soil Fertility Laboratory Department of Agricultural Sciences Royal Veterinary and Agricultural University Frederiksberg, Denmark

Richard McDowell AgResearch Ltd., Invermay Agricultural Center Mosgiel, New Zealand

David R. Parker Department of Soil and Environmental Sciences University of California—Riverside Riverside, California

Leslie D. McFadden Department of Earth and Planetary Sciences University of New Mexico Albuquerque, New Mexico

Judith F. Pedler Department of Environmental Sciences University of California—Riverside Riverside, California

xvi

Contributors

Mark B. Peoples Agriculture and Food, Commonwealth Scientific and Industrial Research Organization (CSIRO) Canberra, Australian Capital Territory, Australia

A. Paul Schwab Department of Agronomy Purdue University West Lafayette, Indiana

Robert Pietrzak Department of Chemistry Adam Mickiewicz University Poznan, Poland

Andrew N. Sharpley University of Arkansas Fayetteville, Arkansas

Paola Poli Department of Genetics, Biology of Microorganisms, Anthropology, and Evolution University of Parma Parma, Italy Abdul Rashid Pakistan Atomic Energy Commission Islamabad, Pakistan Lisa A. Robinson Independent Consultant Newton, Massachusetts Alexandru V. Roman School of Public Administration Florida Atlantic University Boca Raton, Florida John Ryan International Center for Agricultural Research in the Dry Areas (ICARDA) Aleppo, Syria Vilma Sandström Sustainability Science LUT-University Lahti, Finland William H. Schlesinger Department of Geology and Botany Duke University Durham, North Carolina Ewald Schnug Institute for Crop and Soil Science Julius Kuhn Institute (JKI) Braunschweig, Germany

Paul K. Sibley School of Environmental Sciences University of Guelph Guelph, Ontario, Canada Bogdan Skwarzec Faculty of Chemistry University of Gdansk Gdansk, Poland Rolf Sommer International Center for Agricultural Research in the Dry Areas (ICARDA) Aleppo, Syria Gerd Sparovek College of Agriculture Graduate School of Agriculture Luiz de Queiroz (ESALQ) University of São Paulo São Paulo, Brazil Scott J. Sturgul Nutrient and Pest Management Program University of Wisconsin Madison, Wisconsin Donald L. Suarez U.S. Salinity Laboratory Agricultural Research Service (USDA-ARS) U.S. Department of Agriculture Riverside, California Meena Thakur Department of Entomology Dr. Y.S. Parmar University of Horticulture and Forestry Solan, India

xvii

Contributors

Oswald Van Cleemput Faculty of Agricultural and Applied Biological Sciences University of Ghent Ghent, Belgium

W.W. Wenzel Institute of Soil Research University of Natural Resources and Life Sciences Vienna, Austria

Doug Walsh Washington State University Prosser, Washington

Larry P. Wilding Department of Soil and Crop Sciences Texas A&M University College Station, Texas

Leszek Wachowski Department of Chemistry Adam Mickiewicz University Poznan, Poland Johannes Bernhard Wehr School of Agriculture and Food Sciences The University of Queensland St. Lucia, Queensland, Australia

He Zhong Pesticide Environment Impact Section Public Health Entomology Research and Education Center Florida A&M University Panama City, Florida

Anthropogenic Chemicals: Human Manufactured and Activities

I

1

1 Acaricides Introduction ...................................................................................................... 3 Spider Mite Pests .............................................................................................. 4 A Big Drain from the Feeding of Such Small Pests ......................................4 Spider Mite Outbreaks Are Promoted by Hot, Dry Weather ..................... 4 Smothering Agents ...........................................................................................5 Organochlorines ............................................................................................... 5 Organophosphates and Carbamates .............................................................. 5 Organotins.........................................................................................................6 Propargite ..........................................................................................................6 Amidines ........................................................................................................... 6 Ovicides .............................................................................................................6 Antimetabolites ................................................................................................6 Synthetic Pyrethroids

Doug Walsh

Tetronic Acids ................................................................................................... 7 Application Technology................................................................................... 7 Combating Miticide Resistance ......................................................................7 References .......................................................................................................... 7

introduction Approximately 45,000 species of mites and hundreds of species of ticks are described worldwide. Many thousands of species still remain unidentified. About half are plant-feeding species, and among these, about half are in the superfamily Eriophyoidea (gall, bud, and rust mites). Most of the other plantfeeding mites are classified into the superfamilies Tetranychoidea and Tarsonemidae. The superfamily Tetranychoidea includes the economically important spider, flat, and fowl mites, and the superfamily Tarsonemidae includes the economically important broad, cyclamen mites and Varroa mites. Over another 3000 mite species are loosely classified in the order Astigmata. Economically important species include feather and scabies mites. Ticks are placed in the superfamily Ixodoidea and all are ectoparasites (blood feeders) of vertebrate animals.[1] Most mites are small to minute and mites are universally cryptic, making them difficult to detect. Often infestations are overlooked. Mites are often colonizers of new or disturbed habitats, and once established on a new host, mites possess biological characteristics that permit rapid increases in population abundance. Factors in most mites’ lifestyle that lead to rapid population buildup include high egg production, various modes of reproduction (parthenogenesis, pedogenesis, and sexual), short life cycles, a myriad of dispersal techniques, and adaptability to diverse ecological conditions.[1] These traits combined with an exponential increase in worldwide transport of humans and plant and animal products will likely contribute to increased concerns over mite pests in the future.

3

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Managing Global Resources and Universal Processes

In plant-based agriculture, Van de Vrie et al.[2] observed that outbreaks of mite populations were uncommon historically in systems where productivity languished far below the levels achieved in modern production agriculture. Spider mite populations stayed below observable levels due to natural regulation by predators, disease, and poor nutrition from low-quality host plants. Van de Vrie et al.[2] went on to observe that mite populations often experienced outbreaks in agroecosytems where production levels were bolstered by the use of synthetic inputs including fertilizers and pesticides. When crop production is optimized (i.e., not limited by water, nutrients, competition from weeds, or predatory mites and insects), the plants in production become an excellent food source for mite pests. Under these conditions, the developmental rate, fecundity, and life span of mites are increased and contribute to population outbreaks.

Spider Mite Pests A number of mite species can achieve pest status at high population abundance. Spider mites develop through several stages: egg, six-legged larva, eight-legged protonymph, deutonymph, and adult. Males typically reach maturity before females and will position themselves near developing quiescent females. When an adult female emerges, copulation will often occur immediately. Under optimal conditions, most mite species can develop from egg to adult in 6 to 10 days. Egg laying can begin as soon as one or two days after maturing to adults. Most spider mite species overwinter as mated adult females. A notable exception is the European red mite that overwinters as eggs.[3]

A Big Drain from the Feeding of Such Small Pests At the microscopic level, significant quantities (relative to mite size) of plant juices pass through the digestive tract of spider mites as they feed on leaf tissues. McEnroe[4] estimated this volume at 1.2 × 10−2 microliters per mite per hour. This quantity represents roughly 50% of the mass of an adult female spider mite. Leisering and Beitrag[5] calculated that the number of photosynthetically active leaf cells that are punctured and emptied per mite is 100 per minute. In gut content studies of two-spotted mites, Mothes and Seitz[6] observed only thylakoid granules inside their digestive tract following feeding. The thylakoid grana on which spider mites focus their feeding are key photosynthetic engines in plant cells. The grana consist of 45%–50% protein, 50%–55% lipid, and minute amounts of RNA and DNA.[7] Water and other low-density plant cell contents are directly excreted.[4] At the macroscopic level, damage from mite feeding can cause leaf bronzing, stippling, or scorching. For most horticultural crops, economic loss is caused by a drop in yield or quality due to reduction in photosynthesis.

Spider Mite outbreaks Are Promoted by Hot, Dry Weather Water stress, wind, and dust all contribute to the potential for mite outbreaks. When mite outbreaks occur, acaricide treatments are often used for suppression. Varroa mites Varroa jacobsoni provide an ideal example of how rapidly a mite species can spread and exploit a new habitat. First recorded in honeybee colonies in Southeast Asia in 1904, Varroa mites are now pandemic. Varroa mites feed parasitically on an individual bee’s hemolymph fluid, weakening the bee and often causing premature death. Mites attach themselves to foraging workers in order to spread themselves from one hive to another. This mite can severely weaken bees, and an unchecked mite population will almost certainly lead to the premature death of a honeybee colony. Apiculturists speculate that Varroa mite has contributed substantially to the collapse of feral honeybee populations worldwide.[8] The northern fowl mite Ornithonyssus sylviarum is a common pest of domestic fowl and other wild birds commonly associated with human settlements. The nymphs and adults have piercing mouthparts

Acaricides

5

and seek blood meals. Mite populations build up rapidly and a generation can be completed in 5 to 12 days. Several generations occur each year. Northern fowl mite spends virtually its entire life on the host bird.[9] Deer and dog ticks Ixodes scapularis and Dermacentor variabilis are two common ticks to which acaricides are applied for on a consistent basis, especially since both are parasitic feeders on mammals. Deer ticks are a significant concern since they are the primary vector for Lyme disease.[10] Mange or scabies in livestock is a skin condition caused by microscopic mites in or on the skin. The  mites cause intense itching and discomfort that is associated with decreased feed intake and production. Scratching and rubbing result in extensive damage to hides and fleece. Mange mites are able to cause mange on different species of livestock but are somewhat host specific, thus infecting some species more severely than others. The three most important types of mange are as follows: sarcoptic mange, caused by Sarcoptes scabiei feeding; psoroptic mange, caused by Psoroptes ovis feeding; and chorioptic mange, caused by Chorioptes bovis feeding.[11] Infestations of these mites on their respective livestock, domestic pet, or human host will cause skin irritation and itching and leave entry points for secondary infections. Weight gain can be reduced in livestock, pets can lose hair and itch persistently, and disfigurement can occur in humans. Acaracides are often applied to suppress mite populations parasitizing humans, pets, and livestock.

Smothering Agents Solutions containing petroleum-based horticultural oils, vegetable oils, or agricultural soaps are applied to many crops and, occasionally, livestock. Application of these types of products kills spider mites through suffocation. Unfortunately, oils and soaps can prove phytotoxic to crop plants and are typically not effective on mites or ticks infesting livestock, pets, or humans. Mites on animal hosts are typically cryptic or subcutaneous, so acaricide coverage is an impediment to effective control.

organochlorines Endosulfan and dicofol are organochlorine miticides registered for use on many crops. Unlike many other organochlorine pesticides, endosulfan and dicofol are relatively non-persistent in the environment. Organochlorine acaricides interfere with the transmission of nerve impulses and disrupt the nervous system of pest mites. Organochlorine acaricides are more effective at killing mites at warmer temperatures. Overuse of organochlorine acaricides in commercial situations has resulted in the development of tolerance in many pest mite populations. Organochlorines were used substantially in the mid-20th century, but regulatory actions and public health and environmental concern have eliminated their use in most developed countries (though some continue to use them). Lindane was commonly used for mange mite in pets, livestock, and humans. Only in limited circumstances is lindane still permitted as a pharmaceutical second-line treatment. However, use of lindane continues in developing countries due to its low cost, effectiveness, and persistence.

organophosphates and carbamates Many organophosphate and carbamate pesticides have acaricidal activity. Studies have demonstrated significant mite control with applications of parathion, TEPP (tetraethyl pyrophosphate), and aldicarb. Spider mites are listed as target pests on many organophosphate and carbamate products. However, many mite populations following long-term exposure have developed resistance to the toxic effects of organophosphates.[12] Carbaryl, a common carbamate, continues to be a mainstay for mite control on livestock and poultry, but its use on domestic pets and households is no longer permitted in most developed countries. The use of carbaryl continues extensively in many developing countries in domestic settings.

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Managing Global Resources and Universal Processes

organotins Miticides in this category were synthesized in the 1960s and 1970s and registered for commercial use in the United States in the 1970s. They have been used extensively for their ability to quickly knock down spider mite populations through contact activity. Fenbutatin-oxide has been used extensively since the 1970s. Cyhexatin was used extensively in the 1970s and 1980s, but regulatory actions have now limited its use. Efficacy of the organotin acaricides is improved with warmer weather. Overuse of cyhexatin during the mid-1980s led to the development of resistance in several cropping and livestock production systems. However, populations of pest mites can regain susceptibility to organotins following a period of non-exposure.[13]

Propargite This acaricide has been used since the 1960s. It provides effective suppression of pest mites on many crops. Regulatory constraints have resulted in the cancellation of a number of uses. Identification of propargite as a dermal irritant has led to substantial increases in time required following application before re-entry is permitted into the treated site.

Amidines Amitraz is a miticide that once had significant use in plant and animal agriculture. At present, its use is restricted to only a small subset of the domestic pet care market.

ovicides Clofentazine and hexythiazox are selective carboxamide ovicidal acaricides. Spider mite eggs exposed to either compound fail to hatch. These acaricides are selective and aid in the conservation of populations of beneficial arthropods. These acaricides are typically used relatively early in the production season before mite populations reach outbreak conditions.

Antimetabolites A number of miticidal compounds have been developed within the past 30 years. These include avermectins, pyridazinones, carbazates, and pyrroles. Pest mortality results from disruption of metabolic pathways typically within the mitochondria of nerve cells of spider mites.[14] Avermectins, ivermectins, and related compounds are fermentation products derived from mycelial extracts of Streptomyces species (reviewed by Burg and Stapley).[17] Avermectins are locally systemic (translaminar) in plant tissues, [15] and ivermectins can be applied dermally, by injection, suppository, or in a bolus to livestock and domestic pets. The ivermectins are the predominant parasiticide used in livestock production today. A number of products have been commercialized in recent years. Pyridaben is a pyridazinone recently registered for use on ornamentals and some tree crops. Bifenazate is a carbazate acaricide. It has a new mode of action that is not clearly understood, but it has proven toxicologically safe in mammalian studies. Bifenazate is registered on ornamentals and food products. Chlorfenapyr is a synthetic pyrrole that has been commercially available on cotton. Other uses are pending.

Synthetic Pyrethroids Fenpropathrin and bifenthrin are two synthetic pyrethroid insecticides registered for control of spider mites in plant agriculture. Permethrin is registered for mite control on livestock and poultry. Mites have a well-documented history of rapidly developing resistance to pyrethroid insecticides in both plant and animal production systems, and resurgence of spider mite populations following pyrethroid application is typical.[18]

Acaricides

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tetronic Acids Spiromesifen and spirotetramat are acaricides in a recently introduced class of selective chemistry tetronic acids that exhibit a broad-spectrum insecticidal acaricidal activity against mites. Their mode of action is by inhibition of lipid biosynthesis that affects the egg and immature stages of mites. Foliar sprays of spiromesifen are translaminar in plants and effective against mites in many cropping systems. Spirotetramat has a relatively unique property among currently registered acaricides in that it is phloem systemic within the plant it is applied to. These two acaricides have recently entered the acaricide market and are quickly gaining in use in production agriculture for mite control.

Application technology Mite pests can prove difficult to control with acaricides due to their potential for high population abundance, small size, and propensity to live on the bottom surfaces of leaves or within the folds of plant tissues. Good acaricide spray coverage is essential for mite control, particularly for acaricides that kill on contact with the pest mite.

combating Miticide Resistance Following repeated exposure, spider mite populations have a history of rapidly developing resistance to acaricides.[16] Alternating acaricides that have different modes of action reduces the potential for development of resistance to acaricides within specific modes of activity. Other techniques to discourage resistance development include spraying only when necessary and treating only infested portions of the crop. Organophosphate, carbamate, and pyrethroid insecticide applications can induce spider mite outbreaks. If possible, avoid early-season insecticide application or apply insecticides that are less disruptive to beneficial arthropods. Careful selection and use of insecticides can potentially reduce the number of miticide applications required later in the season.

References 1. Krantz, G.W.; Walter, D.E., Eds. A Manual of Acarology, 3rd Ed.; Texas Tech University Press: Lubbock, Texas, 2009; 807 pp. 2. Van de Vrie, M.; McMurtry, J.A.; Huffaker, C.B. Ecology of mites and their natural enemies. A review. III Biology, ecology, pest status, and host plant relations of tetranychids. Hilgardia 1972, 41, 345–432. 3. Bostanian, N.J. The relationship between winter egg counts of the European red mite Panonychus ulmi (Acari: Tetranychidae) and its summer abundance in a reduced spray orchard. Exp. Appl. Acarol. 2007, 42, 185–195. 4. McEnroe, W.D. The role of the digestive system in the water balance of the two-spotted spider mite. Adv. Acarol. 1963, 1, 225–231. 5. Leisering, R.; Beitrag, O. Beitrag zum phytopatologischen Wirkungsmeechanismus von Tetranychus urticae. Pflanzenschutz 1960, 67, 525–542. 6. Mothes, U.; Seitz, K.A. Functional microscopic anatomy of the digestive system of Tetranychus urticae (Acari: Tetranychidae). Acarologia 1981, 22, 257–270. 7. Noggle, G.R. The organization of plants. In Introductory Plant Physiology; Noggle, G.R., Fritz, G.J., Eds.; Prentice Hall: Englewood Cliffs, New Jersey, 1983; 9–38. 8. Mangum, W.A. Honey bee biology: The third annual report on the coexistence of my North Carolina bees with varroa mites. Am. Bee J. 2009, 149, 63–65. 9. Mullens, B.A. Temporal changes in distribution, prevalence and intensity of northern fowl mite (Ornithonyssus sylviarum) parasitism in commercial caged laying hens, with a comprehensive economic analysis of parasite impact. Vet. Parasitol. 2009, 160, 116–133.

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10. Diuk-Wasser, M.A. Field and climate-based model for predicting the density of host-seeking nymphal Ixodes scapularis, an important vector of tick-borne disease agents in the eastern. U. S. Global Ecol. Biogeogr. 2010, 19, 504–514. 11. Vercruysse, J. World Association for the Advancement of Veterinary Parasitology (W.A.A.V.P.) guidelines for evaluating the efficacy of acaricides against (mange and itch) mites on ruminants [electronic resource]. Vet. Parasitol. 2006, 136, 55–66. 12. Smissaeret, H.R.; Voerman, S.; Oostenbrugge, L.; Reenooy, N. Acetylcholinesterases of organophosphate-susceptible and resistant spider mites Tetranychus urticae. J. Agric. Food Chem. 1970, 18, 66–75. 13. Hoy, M.A.; Conley, J.; Robinson, W. Cyhexatin and fenbutatin-oxide resistance in Pacific spider mite (Acari: Tetranychidae) stability and mode of inheritance. J. Econ. Entomol. 1988, 81, 57–64. 14. Hollingsworth, R.M.; Ahammadsahib, K.I.; Gadelhak, G.; McLaughlin, J.L. New inhibitors of Complex I of the mitochodrial electron transport chain with activity as pesticides. Biochem. Soc. Trans. 1994, 22, 230–233. 15. Walsh, D.B.; Zalom, F.G.; Shaw, D.V.; Welch, N. C. Effect of strawberry plant physiological status on the translaminar activity of avermectin B1 and its efficacy on the two-spotted spider mite Tetranychus urticae Koch (Acari: Tetranychidae). J. Econ. Entomol. 1996, 89 (5), 1250–1253. 16. Leeuwen, T.V.; Dermauw, W.; Tirry, L.; Vontas, J.; Tsagkarakou, A. Acaricide resistance mechanisms in the two- spotted spider mite Tetranychus urticae and other important Acari. Insect Biochem. Mol. Biol. 2010, 40, 563–572. 17. Burg, R.W., and E.O. Stapley. Isolation and characterization of the producing organism. In W.C. Campbell (ed) Ivermectin and Abamectin. Springer-Verlag, New York, N.Y. 1989. pp. 24–32. 18. Leigh, T.F. Cotton. In W. Helle and M.W. Sabelis (eds.) World Crop Pests: Spider Mites. Elsevier Press, Amsterdam, the Netherlands. 1990. pp. 349–358.

2 Endocrine Disruptors Introduction ...................................................................................................... 9 Mechanisms of Action

Hormones, Reproductive Aspects, and EDCs ............................................. 11 Screening Assays and Biomarkers for EDCs .............................................. 13 Guidelines for Regulatory Purposes ............................................................ 14 Overview of EDC Exposure .......................................................................... 16 Human Exposure Effects • Complex EDC Mixtures • Some Examples of Animal Exposure Effects

Vera Lucia S.S. de Castro

Dealing with Environment EDC Emission ................................................. 19 Conclusion ......................................................................................................20 References ........................................................................................................ 21

introduction Endocrine-disrupting chemicals (EDCs) refer to anthropogenic compounds that are able to mimic, antagonize, alter, or modify normal hormonal activity. Dichlorodiphenyl-trichloroethane (DDT), an insecticide first produced on a wide scale in 1945, was used extensively during the 1960s and 1970s and was the first chemical found to be estrogenic. Subsequently, other organochlorine insecticides such as dieldrin, endosulfan, and methoxychlor were found to be estrogenic. Endocrine-disrupting chemicals include environmental estrogens such as o,p-DDT, endosulfan, non-planar polychlorinated biphenyl (PCB), octyl-and nonylphenols, the antiandrogens such as vinclozolin and DDE, and the thyroid hormone disrupters such as fenvalerate and benzene hexachloride.[1] Endocrine-disrupting chemicals are a significant public health concern since these compounds interfere with normal function of pathways responsible for both reproduction and development and can affect the endocrine system, interfering in the production or action of hormones or compromising sexual identity, fertility, or behavior.[2,3] Besides, many of them are persistent in the environment, can be found in waters and sediments, and are easily transported long distances in the atmosphere.[4] In recent years, numerous studies have suggested that many environmentally persistent chemicals have a potential to disrupt normal functions of the endocrine system. The field of endocrine disrupters, such as the special susceptibility of the developing organism and early induction of latent effects, has come a long way since its initial impetus in 1991.[5] Specially, the possible effects of EDCs on early events of proper gonadal development—which is dependent on intercellular signaling mechanisms— deserve attention since the early steps in mammalian sexual development are vulnerable to genetic and environmental perturbation.[6] Exposure to EDCs is associated with dysfunctions of metabolism, energy balance, thyroid function, and reproduction, and an increased risk of endocrine cancers. These multifactorial disorders can occur through molecular epigenetic changes induced by exposure to EDCs early in life, the expression of which may not manifest until adulthood.[7] Effects attributed to the EDCs include developmental 9

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demasculinization and feminization in reptiles, mammals, amphibians, fish, and birds; reduced fecundity in reptiles, birds, and fish; and possibly increased breast cancer rates and reduced sperm counts in humans.[1] Since hormones, in synergy with genes, are responsible for sex-related differences in anatomical, physiological, and behavioral traits, even if EDCs are present in minute amounts in environment, their effects in male and female physiology could be greater than before expected. They might also prejudice the sex steroid hormone–induced integrated physiological responses in women and men. In addition, differences in male and female susceptibility to EDCs could be present even if there is still scarce information available on this aspect.[8]

Mechanisms of Action Several EDCs may work by multiple mechanisms, including uncharacterized mechanisms of action. Because of cross talk between different components of the endocrine systems, effects may occur unpredictably in endocrine target tissues other than the system predicted to be affected. A few modes of action could contribute to the same outcome, including aromatase inhibition, antiestrogenicity, testosterone biosynthesis disruption, and antiandrogens that alter upregulation of aromatase in the target regions within the brain. More complex biological responses to EDCs will generally represent combinations of several physiological processes integrated through multiple biological pathways.[9,10] Endocrine disrupters may interfere with the functioning of hormonal systems in at least three possible ways: 1) by mimicking the action of a naturally produced hormone, producing similar but exaggerated chemical reactions in the body; 2) by blocking hormone receptors, preventing or diminishing the action of normal hormones; and 3) by affecting the synthesis, transport, metabolism, and/or excretion of hormones, thus altering the concentrations of natural hormones.[11] The first characterized mechanism of action of EDCs is to act directly as ligands to steroid hormone nuclear receptors (NRs), in particular, estrogen, androgen, and thyroid NRs. Nuclear receptors are a class of proteins found within cells. In response to the presence of hormones, these receptors work in concert with other proteins to regulate the expression of specific genes by a conformation change. Schematically, NRs may be classified into four classes according to their dimerization and DNA-binding proprieties.[10] Cross talk between NR-mediated and other signal-transduction pathways is an important aspect of NR-based regulation. This so-called genomic or genotropic signaling is normally slow and sustained, taking hours before biological outcomes become manifest. For example, in the classic view of estrogen action, the effects of 17β-estradiol (E2) were thought of as mediated by the NRs estrogen receptor α and β, acting as ligand-dependent transcription factors, thereby regulating gene expression by binding estrogen response elements in the DNA.[12] Another type of NR cross talk that has recently been recognized is the non-genomic action of several NRs. Some non-genomic actions of NR ligands are apparently mediated through membrane receptors that are not part of the NR superfamily.[10] For example, it has become clear that E2 can also rapidly and transiently trigger a variety of second messenger signaling events, including the induction of cyclic adenosine monophosphate (cAMP) and adenylate cyclase; the mobilization of intracellular calcium; and the stimulation of PI3K, PKB, and Src with consequent activation of the extracellular-regulated kinases Erk1 and Erk2 in the Src/Ras/Erk cascade. All these effects are believed to be mediated through a membrane-associated or cytosolic estrogen receptor (ER) and have, therefore, been termed nongenomic or extranuclear actions of E2.[12] The cellular activities of estrogens and xenoestrogens are the result of a combination of extranuclear (non-genomic) and nuclear (genomic) events and highlight the need to take non-genomic effects and signaling cross talk into consideration when screening for environmental estrogen.[12] Disruption of the endocrine system by xenobiotic compounds is consistently reported in humans and wildlife and is a matter of concern worldwide. A great variety of natural or synthetic chemicals such as

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EDCs are thought to exert an acute effect at different levels of the thyroid cascade. It is consensual that EDCs probably act by interfering with thyroid hormone (TH) synthesis, cellular uptake, and metabolism, at the level of TH receptors and also TH transport, by binding to thyroid hormone distributor proteins (THDPs). The TH transport system in particular may be quite susceptible to EDCs as many chemicals are structurally related to THs and may bind THDPs and disturb homeostasis of extracellular TH levels or even cellular uptake.[13] Steroid hormone synthesis is controlled by the activity of several highly substrate-selective cytochrome P450 enzymes and a number of steroid dehydrogenases and reductases. Cytochrome P450 monooxygenases (CYPs) form a large group of enzymes found in most organisms from bacteria to mammals and can be grouped into 281 families. According to their function CYPs can be classified into enzymes metabolizing xenobiotics and enzymes that are part of key biosynthetic pathways, with narrow substrate specificity. Particularly, aromatase (CYP19), the enzyme that converts androgens to estrogens, has been the subject of studies into the mechanisms by which chemicals interfere with sex steroid hormone homeostasis and function, often related to (de)feminization and (de)masculinization processes.[14,15] After all, several findings suggest that responses to EDCs cannot be assumed to be monotonic across a wide dose range and that change can occur in response to extremely low concentrations. In particular, low-dose effects may be mediated by endocrine-signaling pathways, evolved to act as powerful amplifiers, with the result that large changes can occur in response to extremely low concentrations. Dose–response relationship, however, is perhaps one of the most controversial issues in EDC studies. Reports on non-linearities in dose–response functions are highly controversial and the subject of intense research: non-monotonic, linear, and even threshold responses are all possible outcomes of lowdose exposure.[16] A non-monotonic response decreases testing efficiency and multiplies the time and other resources necessary to understand the potential hazard posed by a chemical. Because the issue of low-dose effects of EDCs was based on unknown and unexpected mechanisms, the actual features of these effects were not readily resolved.[17] Low-dose effects of EDCs are based on unknown and unexpected mechanisms. Recent developments in the biological sciences, including homeostatic regulatory disturbance and epigenetic response, have aided in clarifying the mechanisms underlying the low-dose issue. Elucidating the xenobiotic effects of EDCs requires development of systems toxicology, i.e., deciphering the toxicity mechanisms underlying homeostatic regulatory disturbances.[17,18]

Hormones, Reproductive Aspects, and eDcs In vertebrates, the ability to attain reproductive competence in adulthood involves the organization of a complex, steroid sensitive network in hypothalamic–preoptic–limbic brain regions during critical developmental windows. This process includes the establishment of the hypothalamic neural network of gonadotropin-releasing hormone (GnRH) cells, together with their regulatory inputs from other neuronal and glial cells in the brain, which enable feedback effects of steroid hormones on pulsatile GnRH release and the preovulatory GnRH/LH–luteinizing hormone surge in females. The a natomical development of this steroid-sensitive hypothalamic network occurs early in life, typically the late embryonic and early postnatal period in mammals, and its organization is important to the attainment and activation of appropriate reproductive functions in adulthood. Importantly, this same early developmental period is also a critical period for sexual differentiation of hypothalamic–limbic neural networks that must be organized perinatally to enable proper behavioral activation in adulthood. During mammalian development, the fetal organism is exposed to its own gonadal hormones, placental steroids, and maternal hormones that may cross the placental barrier. There are sex differences in exposures to androgens and estrogens that appear to underlie normal reproductive neuroendocrine development. Aberrations in these developmental patterns in females can cause masculinization (acquisition of a male-typical trait) or defeminization (loss of a female-typical trait) and in males may

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cause feminization or demasculinization (comparably defined). Perinatal hormones have permanent imprinting effects on the hypothalamus, manifested early on as morphological sex differences in the brain and manifested much later on as physiological and behavioral differences between the sexes.[19–21] Besides, androgens and estrogens can play a special role in the development of sexually dimorphic behaviors.[22] Some populations exposed to chemicals from industrial accidents or chemical misadventures are of particular interest. Data from these select populations with higher levels of exposure than the common population seem to suggest that some of these chemicals have a role in genitourinary development as endocrine disrupters. Furthermore, animal studies of EDC effects on genitourinary development have confirmed that changes occur with exogenous manipulation of steroid levels or hormone receptors. These findings in animals have led to observational and epidemiological studies in humans to document a link between environmental exposure and human disease.[23] Global declines in semen quality were suggested to be associated with enhanced exposure to environmental chemicals that act as endocrine disrupters as a result of increased use of pesticides, plastics, and other anthropogenic materials. A significant body of toxicology data based upon laboratory and wildlife animals studies suggests that exposure to certain endocrine disrupters is associated with reproductive toxicity, including the following: 1) abnormalities of the male reproductive tract (cryptorchidism, hypospadias); 2) reduced semen quality; and 3) impaired fertility in the adult.[24] Recently, there has been increasing concern about the potential for environmental EDCs as fungicides to alter sexual differentiation in mammals. In this direction, observations demonstrate that vinclozolin (a systemic dicarboximide fungicide) can affect embryonic testicular cord formation in vitro. This transient in utero exposure to the fungicide increases apoptotic germ cell numbers in the testis of pubertal and adult animals. This effect is correlated with reduced sperm motility in the adult and putative effects on spermatogenic capacity later in adult life. In conclusion, transient exposure to this fungicide during the time of testis differentiation alters testis development and function.[25] A higher prevalence of cryptorchidism and hypospadias was found in areas with extensive farming and pesticide use and in sons of women working as gardeners. Recently, a relation has been reported between cryptorchidism and persistent pesticide concentration in maternal breast milk.[26] Other commonly used fungicides, such as the azoles, may also act as endocrine disrupters in vivo. They showed endocrine-disrupting potential when tested for endocrine-disruptive effects using a panel of in vitro assays. Overall, the imidazoles (econazole, ketoconazole, miconazole, prochloraz) were more potent than the triazoles (epoxiconazole, propiconazole, tebuconazole). The critical mechanism in vitro seems to be disturbance of steroid biosynthesis.[27] Regarding in vivo effects, many of the commonly used azole fungicides act as endocrine disrupters, although the profile of action varies. Common features for azole fungicides are that they increase gestational length, virilize female pups, and affect steroid hormone levels in rat fetuses and/or dams.[28] For example, prochloraz causes reproductive malformations in androgen-dependent tissues of male offspring of exposed rats.[29] Also, tebuconazole has been found to demasculinize the male offspring and to possess some of the same endocrine effects as prochloraz. These effects strongly indicate that one major underlying mechanism for the endocrine-disrupting effects of azole fungicides is disturbance of key enzymes like CYP17 involved in the synthesis of steroid hormones.[28] Also, triazole-induced male reproductive toxicity includes disruption of testosterone homeostasis. Elevated serum testosterone, increased testis, weights and anogenital distance, and hepatomegaly indicative of altered liver metabolism of steroids are the key events consistent with this mode of action.[30] Developmental exposure to triazole fungicides such as propiconazole, myclobutanil, and triadimefon can adversely impact reproduction in the female rat.[31] In this way, epoxiconazole and ketoconazole may be fetotoxic, increasing postimplantation loss and late resorptions.[32] Aside from triazoles, the organic fungicide fenarimol possesses estrogenic properties[33] and acts both as an estrogen agonist and as an androgen antagonist.[34] In addition, fenarimol affects rat aromatase activity in vivo, inhibiting estrogen biosynthesis in rat microsomes[35] and in human tissues.[3]

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This compound also affects other enzymes of the cytochrome P450 gene family that are involved in the metabolism of steroids.[36] Induction of reactive oxygen species (ROS) by environmental contaminants and associated oxidative stress also have a role in defective sperm function and male infertility, although there are some controversial data. This is evidence for the existence of a link between endocrine-mediated and ROS-mediated adverse effects of environmental contaminants on male reproduction. Another link is the antioxidant enzyme superoxide dismutase, which has been shown to have a superoxide scavenging effect as well as act as an alternate regulatory switch in testicular steroidogenesis.[37] Endocrine-disrupting chemicals can also impact female fertility by altering ovarian development and function, purportedly through estrogenic, antiestrogenic, and/or antiandrogenic effects. These compounds may also cause transgenerational effects by targeting oocyte maturation and maternal sex chromosomes.[38] In girls, earlier age at menarche was reported after exposure to PCBs, polybrominated biphenyls, persistent pesticides (DDT), and phthalate esters. However, several other studies found no effect of these compounds on age at menarche. In boys, exposure to PCBs, PCDFs, or the pesticide endosulfan was associated with delayed puberty or decreased penile length. Much of the results found in population studies are in accordance with experimental studies in animals. However, the mixture of different components with antagonistic effects (estrogenic, antiestrogenic, antiandrogenic) and the limited knowledge about the most critical window for exposure (prenatal, perinatal, and pubertal) may hamper the interpretation of results.[39] In human and rodent models, EDCs also interfere with the development of cognition and behaviors. In this way, fenvalerate is a potential EDC and is a candidate environmental risk factor for cognitive and behavioral development, especially in the critical period of development.[40] Also, prenatal phthalate exposure was associated with childhood social impairment in a multiethnic urban population.[41] Recently, the interference of EDCs with receptors regulating metabolism has been proposed especially in relation with the etiology of metabolic diseases such as obesity and diabetes. In particular, the harmful action of EDCs on normal adipocyte development, homeostatic control of adipogenesis, early energy balance, and, in turn, body weight has been demonstrated. Much remains to be studied about the endocrine pathways responding to EDC exposure, especially those controlling feeding behavior, as their impairment represents a real risk factor for metabolic and feeding disorders.[42]

Screening Assays and Biomarkers for eDcs Environmental stresses as presence of EDCs due to human activities are increasingly likely to pose habitat disturbances that could have potential deleterious effects on physiological function in vertebrates. These effects could result in major impacts on the life cycles of organisms, affecting morphology, physiology, and behavior. However, because animals live in diverse habitats, there is variation in susceptibility to disruption of response systems to environmental cues. While some populations of vertebrates, from fish to mammals, temporarily resist environmental stresses and breed successfully, many others show varying degrees of failure, sometimes resulting in population decline.[43] The development of targeted bioassays in combination with adequate chemical analyses is important for EDC risk assessment. It is acknowledged that EDCs can affect humans and animals at low exposure levels and that responses to EDCs are in many cases complex, activating a range of different molecular events, e.g., receptor agonism/antagonism and enzyme induction, in multiple hormone systems. As a result, regulatory testing for these effects and evaluating the results is complicated.[44,45] In the typical case of assessing human risk, a scientifically justified validation could only mean an experimentally validated mechanistic link between EDC assay results and human susceptibilities at environmental exposures, sustained by reliable sensitivity and specificity benchmarks. Analytical methods have long been used to determine concentrations of chemical residues that persist in environment and accumulated in biota. Although they are useful, the development of EDC screening and monitoring procedures may help in the establishment

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of EDC exposure and biological responses. In this way, several in vitro and in vivo procedures have been proposed to screen and monitor individual EDCs or their mixtures.[45] As an in vitro model, the use of the bovine ovarian follicle has already been recommended as a valuable instrument to unravel reproductive events in women due to the similarities in ovarian follicular dynamics and endocrine control.[46] As an in vivo test, some external biomarkers of prenatal androgen disruption may be used, including the anogenital distance and the juvenile nipple/areola number. The anogenital distance is defined as the distance between the genital papilla and the anus; male rodents have an anogenital distance that is approximately twice the length as that of females. Areolae are dark areas surrounding the nipple bud and. their presence as measured at postnatal day 2–3 is indicative of adult nipples. Adult female rats typically have 12 nipples, whereas males have none. Both of these biomarkers vary with prenatal exposure to androgens or antiandrogens in females and males, respectively. Reduction of anogenital distance and/or retention of nipples in male rats is indicative of prenatal exposure to antiandrogens.[47] In this context, a wide spectrum of potential biomarkers also could be applied to the study of endocrine disruption in the aquatic environment. In fish, they include changes in hormone titers (steroid hormones, thyroid hormones), abnormal gonad development, low gamete viability, and alterations in some enzyme activities (i.e., aromatases) and protein levels (i.e., vitellogenin, zona radiata proteins, spiggin). Likewise, evidence is slowly growing that indicates that gamete development and vitellogenesis of marine bivalve mollusks are targets of EDCs.[48] On the other hand, although it is known that aquatic invertebrates contain different classes of steroids,[49] a clear cause–effect relationship between exposure and specific responses for most EDCs is far from being established.[48] In crustacean populations, the attribution of endocrine toxicity to observed disturbances requires the identification of definitive biomarkers of such toxicity. Mortality, reduced fecundity, lowered recruitment, and impaired growth all might serve as indicators of endocrine disruption in crustaceans; however, such end points are indicative of adversity involving a variety of mechanisms. An exception to this premise is excess males in parthenogenic branchiopod populations that normally exist predominantly as females.[50] Other organisms, such as amphibians, may be used to study the endocrine system and can serve as sentinels for detection of the modes of action of EDCs. Recently, amphibians are being reviewed as suitable models to assess (anti)estrogenic and (anti)androgenic modes of action influencing reproductive biology as well as (anti)thyroidal modes of action interfering with the thyroid system.[51] Biochemical end points can also be useful biomarkers since environmental toxicants can trigger biological effects at the organism level only after initiating biochemical and cellular events. The cellular response to stress is characterized by the activation of genes involved in cell survival to counteract the physiological disturbance induced by physical or chemical agents. As an example, Hsps are suitable as an early-warning bioindicator of environmental hazard by various pollutants such as EDCs, because of their sensitivity to even minor changes in cellular homeostasis and their conservation along the evolutionary scale.[52] In addition, a combined testing strategy, considering both markers of endocrine/ hormonal maturation and behavioral end points under hormonal control, may evidence even subtle perturbations of the neuroendocrine homeostasis, which often go undetected.[53]

Guidelines for Regulatory Purposes In recent years, under the current European Union chemical regulation REACH (Registration, Evaluation, Authorization and Restriction of Chemicals), which revised plant protection product and biocide directives,[54] evaluation of endocrine-disrupting properties of chemicals has become a regulatory effort. The initial framework for regulatory purposes has been revised by the Endocrine Disrupters Testing and Assessment (EDTA) Task Force at its meetings to reflect the Organization for Economic Cooperation

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and Development (OECD) member countries’ views. The conceptual framework agreed upon by the EDTA6 in 2002 is not a testing scheme but rather a toolbox in which the various tests that can contribute information for the detection of the hazards of endocrine disruption are placed. The toolbox is organized into five compartments or levels, each corresponding to a different level of biological complexity (for both toxicological and ecotoxicological areas). Even though the conceptual framework may be full of testing tools, this does not imply that they all will be needed for assessment purposes. Tools will be added as they are validated in future. The conceptual framework is subject to further elaboration and discussion as the work on endocrine disrupters proceeds.[55] The OECD adopted in 2007 the uterotrophic bioassay as a standardized screening test with international regulatory acceptance. This assay may be used to screen for estrogenic properties of chemicals. However, generally, EDCs are handled as such only if their endocrine-disruption potential has been previously identified via, for example, academic research or is indicated by effects observed in required toxicity tests.[44] The Endocrine Disruptor Screening Program (EDSP) of the United States Environmental Protection Agency (EPA) has been working to reach a consensus validation on a battery of screens and long-term tests for endocrine disrupters.[45] The Endocrine Disrupter Screening and Testing Advisory Committee (EDSTAC) was established by the EPA in 1996 as a federal advisory committee to provide advice in developing and implementing new screening and testing procedures for endocrine effects as mandated by the U.S. Congress (through the Food Quality Protection Act of 1996) in response to public concern.[56] The ED-STAC assesses the current state of the science and assists the agency in developing an endocrine screening program. The EDSTAC consists of scientists and others representing various interests, including advocates of the endocrine-disruption theory and the regulated community. The EDSTAC concluded that the assays necessary to determine the potential endocrine activity of chemical substances varied significantly in their degree of development and validation. Several screens had an extensive history, e.g., the uterotrophic and the Hershberger screens, but others were only partially developed or were only hypothetically useful as screens, e.g., the amphibian developmental screen and the fish gonadal recrudescence screen. The fundamental validation principles are to clearly state the purpose and biological basis for the assay and to verify the performance of the assay against validation criteria using a common set of test chemicals across multiple laboratories.[57] At the same time, EDSTAC recommended that EPA develop an extensive program that would subject all chemicals to screening and testing for estrogenic, antiestrogenic, androgenic, antiandrogenic, and thyroid effects in both humans and wildlife. Specifically, EDSTAC recommended, among other things, that the EPA do the following: 1) adopt a two-tiered, hierarchical testing and evaluation framework; and 2) initiate a research program, composed of both basic and applied research, to develop, standardize, and validate the necessary endocrine test methods. The EPA’s EDSP was implemented in 2009–2010 with the issuance of test orders requiring manufacturers and registrants of 58 pesticide active ingredients and 9 pesticide inert/high-production-volume chemicals to evaluate the potential of these chemicals to interact with the estrogen, androgen, and thyroid hormone systems. Despite this great effort, numerous questions and uncertainties remain as to the usefulness and limitations of the specific assays. Understanding the tests’ strengths and limitations is critical for interpretation of the screening results and for decision making based on those results.[57,58] During the time EDSTAC was meeting, OECD began collaborating with its member countries, including the U.S. EPA, to develop internationally harmonized test guidelines.[48] Although the EPA and OECD endocrine screening and testing methods have been substantially harmonized, the framework of OECD’s endocrine screening and testing program differs significantly from EPA’s two-tiered EDSP. The EPA screening will entail evaluation of responses in the Tier 1 Endocrine Screening Battery, consisting of 11 distinct in vitro and in vivo assays. The OECD framework provides the flexibility to enter and exit at any level depending on information needs and encourages the maximal use of all existing relevant information that may be equally predictive and reduce vertebrate testing. The screening results are collectively intended to identify chemicals for which subsequent Tier 2 testing is necessary. Tier 2

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testing uses test methods that encompass reproduction and developmental life stages in several species to provide data on adverse effects and dose response for risk assessment.[57] In the years that the EPA worked on developing, standardizing, and validating the EDSTACrecommended assays and implementing the EDSP, significant advances have been made in both computational and molecular technologies for discerning potential endocrine activity.[57] Accordingly, there are efforts to model EDC effects using computational approaches by the development and validation of mechanistically based computational models of hypothalamic–pituitary–thyroid (HPT); hypothalamic–pituitary–gonadal (HPG); hypothalamic–pituitary adrenal (HPA) axes in ecologically relevant species to better predict accommodation and recovery of endocrine systems.[59]

overview of eDc exposure Human exposure effects Models for estimating human exposure to endocrine disruptor (ED) pesticides are an important risk management tool. Many of them are harmful at very low doses, especially if exposure occurs during sensitive stages of development, producing effects that may not manifest for many years or that affect descendants via epigenetic changes. The main requirement for the use of such models is more quantitative data on the sources and pathways of human ED pesticide exposure. Quantifying the risks posed by the different routes of exposure will play an important part in designing and implementing effective risk mitigation for ED pesticides. In fact, it is difficult to assess the relative importance of some routes of exposure because no data sets that would allow these to be calculated are available. Pesticide exposure from the use of pesticides for medicinal purposes and exposure from cut flowers and ornamental plants both need to be quantified, and better data sets are required for pesticide exposure from spray drift, home use, municipal use, and travel.[60] Food and water are both chronic exposure routes affecting the entire population. Food residues are currently thought to be the most important exposure pathway, for although residue levels present in food tend to be below the maximum residue levels permitted by law, they do result in constant measurable low-level exposure.[51] Food as a major xenobiotic and heavy metal exposure route to humans is studied intensively. More than 100 chemicals have been identified as antiandrogens, including certain phthalates, widely used as plasticizers, pesticides, and various other chemicals found in food and consumer products.[61] Indeed, typical food contaminants, like pesticides, dioxins, PCBs, methylmercury, lead, etc., are well characterized in food. In contrast, the role of food and beverage packaging as an additional source of contaminants has received much less attention, even though food packaging contributes significantly to human xenobiotic exposure. Especially, EDCs in food packaging are of concern since even at low concentrations, chronic exposure is toxicologically relevant. Thus, non-intentionally added substances migrating from food contact materials need toxicological characterization.[62] Some chemicals used in food processing have an environmental endocrine-disrupting effect that affects reproduction in wildlife. For example, bisphenol-A is a monomer of polycarbonate plastics and a constituent of epoxy and polystyrene resins, which are used in the food cans and found as a contaminant not only in the liquid of the preserved foods but also in the water autoclaved in the cans. This chemical is also released from polycarbonate flasks during autoclaving. Moreover, it has been reported that significant amounts of bisphenol-A are detected in the saliva of dental patients treated with fissure sealants. The exposure to low doses of this chemical was reported to affect the rate of growth and sexual maturation, hormone levels in blood, reproductive organ function, fertility, immune function, enzyme activity, brain structure, brain chemistry, and behavior.[63] Another important route of EDC exposure is occupational. Relatively high levels of exposure to environmental endocrine disrupters in the form of pesticides occur among people working in agriculture. Some pesticides are able to influence the synthesis, storage, release, recognition, or binding of

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hormones, which may lead to alterations in reproductive hormone levels. The issue of male infertility caused by occupational exposure is pertinent worldwide. A significant increase in the incidence of male infertility has been described in the international literature. Part of this effect may result from synthetic toxic substances acting on the endocrine system, many of which are routinely used in work processes. However, progress is needed in the knowledge of possible effects of exposure on male fertility since monitoring these effects requires sufficient time for the manifestations to occur. Such progress will allow the development of preventive measure within the field of workers’ health.[64] Apart from EDC effects on males, several studies on occupational exposure to pesticides and adverse effects on human reproduction have been performed, including end points such as prolonged time to pregnancy, spontaneous abortion or stillbirth, low birth weight, and developmental disorders.[65]

complex eDc Mixtures Concerns increase when humans are exposed to mixtures of similar-acting EDCs and/or during sensitive windows of development. It is difficult to predict biological effects directly from the composition of pollutant mixtures. In addition to simple additive effects, interactions between different chemicals in a mixture may result in either a weaker (antagonistic) or a stronger (synergistic) combined effect than would be expected from knowledge about the toxicity and mode of action of each individual compound. Such interactions may take place in the toxicokinetic phase (i.e., processes of uptake, distribution, metabolism, and excretion) or in the toxicodynamic phase (i.e., effects of chemicals on the receptor, cellular target, or organ). A chemical mixture may contain a number of xenoestrogens enhancing the response of endogenous estrogens, or it may contain xenoantiestrogens that inhibit the normal action of endogenous estrogens.[66] Substances of concern include certain phthalates, pesticides and chemicals used in cosmetics and personal care products. A lack of knowledge about relevant exposure scenarios presents serious obstacles for better human risk assessment. A d isregard for combination-effect studies may lead to underestimations of risks. In this way, the study of EDC mixture effects by developing biomarkers that capture cumulative exposure to endocrine disrupters is needed.[67] Doses of endocrine-disrupting pesticides that appear to induce no effects on gestation length, parturition, and pup mortality when alone induced marked adverse effects on these end points together with other pesticides. They can also affect the sexual differentiation of offspring.[68] Chemicals that act on different fetal tissues via diverse cellular mechanism of action may produce additive effects. This fact indicates that the current framework for conducting cumulative risk assessments should not only consider including chemicals from different classes with the same mechanism of toxicity but also include chemicals that disrupt differentiation of the same fetal tissue at different sites in the androgen signaling pathway.[5] Compounds that act by disparate mechanisms of toxicity to disrupt the dynamic interactions among the interconnected signaling pathways in differentiating tissues produce cumulative dose-additive effects, regardless of the mechanism or mode of action of the individual mixture component.[69] Predictive approaches are generally based on the mathematical concepts of concentration addition and independent action, both predicting the toxicity of a mixture based on the individual toxicities of the mixture components.[26] In this sense, a combination of five pesticides with dissimilar mechanisms of action produced greater androgen sensitive end-point responses than would be expected using response-addition modeling.[70] Deltamethrin, methiocarb, prochloraz, simazine, and tribenuron-methyl are all commonly used for agricultural and horticultural purposes. In vivo, the levator ani/bulbocavernosus muscle and adrenal gland weight changes indicated that the pesticides had an accumulating effect that was not observed for the individual pesticides. Several pesticide-induced gene expression changes were observed, indicating that these may be very sensitive antiandrogenic end points. In another study,[71] dexamethasone appeared to exacerbate the reproductive anomalies induced by in utero exposure of male rats to dibutyl phthalate.

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In a recent study, male Sprague Dawley rats were sub-chronically exposed to single doses of dibutyl phthalate, single doses of benzo(a)pyrene, and combined doses of both EDCs. Significant adverse effects were observed on the reproductive system, including decreased sperm count, increased production of abnormal sperm, changes in serum testosterone levels, and irregular arrangements of the seminiferous epithelium. It is also observed that biochemical analyses showed that the activities of superoxide dismutase and glutathione peroxidase decreased after exposure to these EDCs. Therefore, the data suggest that exposure to them, in either separate or combined doses, can affect the reproductive system of male rats adversely via oxidative stress-related mechanisms.[72] Thus, assessment of risks posed by chemicals causing reproductive effects and protection of future generations are important public health tasks. To determine the levels of significant human exposure to a given chemical and associated health effects, the Agency for Toxic Substances and Disease Registry’s (ATSDR’s) toxicological profiles examine and interpret available toxicological and epidemiological data. The ATSDR categorizes the health effects according to their seriousness as serious (effects that prevent the organism from functioning normally or can cause death), less serious (changes that will prevent an organ or organ system from functioning in a normal manner but will not necessarily lead to the inability of the whole organism to function normally), or minimal (effects that reduce the capacity of an organ or organ system to absorb additional toxic stress but will not necessarily lead to the inability of the organ or organ system to function normally). The ATSDR uses the highest no-observed-adverse-effect level or the lowest-observed-adverse-effect level (LOAEL) in the available literature to derive a healthbased guidance value called a minimal risk level (MRL). An MRL is defined as “an estimate of the daily human exposure to a substance that is likely to be without an appreciable risk of adverse, non-cancer effects over a specified duration of exposure.” Minimal risk levels based on reproductive and endocrine effects were described in a review by Pohl et al.[73]

Some examples of Animal exposure effects There is widespread exposure to EDCs, which can disrupt the reproduction and development of various non-target organisms. Effects of EDCs have been shown by observed adverse reproductive and  developmental effects. Indeed, most studies of potential EDC effects are based on indirect evidence of endocrine disruption rather than defined endocrine pathways. Some domestic mammals may come into contact with EDCs by sewage exposure. As an example, sewage sludge is sometimes recycled to arable land or pasture and contains large amounts of a variety of pollutants, including EDCs and heavy metals, derived from industrial, agricultural, and domestic sources. A demasculinizing effect of exposure to higher pollutant concentrations with respect to exploratory sheep behavior was observed.[74] These observations demonstrate the need to take into account the effects of pollutant combinations, even at very low, environmental concentrations, and further highlight the usefulness of ethotoxicology for the study of biological effects of environmental pollutants. Endocrine-disrupting chemicals have been found in sewage effluent in low concentrations (ng/L). Some of these estrogens bind with estrogen receptors in exposed organisms and have the potential to exert effects at extremely low concentrations. Data from laboratory experiments support the hypothesis that EDCs in the aquatic environment can impact the reproductive health of various fish species, but evidence in the aquatic environment is still weak and needs a dependable method or indicator to assess reproduction of fish in situ. The link between endocrine disruption and reproductive impairment that cause an ecologically relevant impact on the sustainability of fish populations remains to be better understood.[75] Surface waters are the main sink of EDCs, which are mainly of anthropogenic origin. Thus, aquatic organisms, especially lower vertebrates such as fish and amphibians, are the main potential targets for EDCs, being at direct or indirect risk via ingestion and accumulation of EDCs via exposure or the food chain. These compounds may play an important role in the decline of the amphibian population.[51] Several incidents in the wildlife population strongly correlated decreased reproductive capacity

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with exposure to specific industrial chemicals, and the organisms may be viewed as sentinels of human health effects. Reported reproductive disorders in wildlife have included morphologic abnormalities, eggshell thinning, population declines, sex reversal, impaired viability of offspring, altered hormone concentration, and changes in sociosexual behavior.[76] The ED are prevalent over a wide range of chemicals in the aquatic ecosystems, most of them being resistant to environmental degradation and considered ubiquitous contaminants.[48,77] Some imidazole (prochloraz, imazalil) and triazole (epoxiconazole) agricultural fungicides induced oocyte maturation in rainbow trout. Prochloraz, epoxiconazole, and imazalil strongly potentiated the induction of oocyte maturation by gonadotropin in a dose-dependent manner.[78] Above all, prochloraz caused responses consistent with aromatase inhibition, although there were indications that the fungicide may also be disturbing the balance between estrogens and androgens via effects elsewhere in the steroidogenic pathway.[77] In U.K. rivers, a widespread feminization of wild fish was observed involving contributions from both steroidal estrogens and xenoestrogens and from other yet-unknown contaminants with antiandrogenic properties. The widespread occurrence of feminized male fish downstream of some wastewater treatment works has led to substantial interest from ecologists and public health professionals. This concern stems from the view that the effects observed have a parallel in humans and that both phenomena are caused by exposure to mixtures of contaminants that interfere with reproductive development.[79] Some authors reported the occurrence of fish feminization as well as reproduction and development interference with other aquatic organisms,[80] although there is no universally accepted bioassay or chemical technique to quantify EDCs in the aquatic environment.[81] Endocrine-disrupting chemicals can also promote disrupting effects in vitro on ovarian follicular cells exposed to environmentally relevant doses of mixtures of persistent organic pollutants extracted from marine and freshwater ecosystems.[66] Population studies have revealed alterations in crustacean growth, molting, sexual development, and recruitment that are indicative of environmental endocrine disruption. However, environmental factors other than pollution (i.e., temperature, parasitism) also can elicit these effects and definitive causal relationships between endocrine disruption in crustacean field populations, and chemical pollution is generally lacking.[50] Also, temperature and photoperiod are the two most important environmental cues in the regulation of the annual cycles of circulating sex steroid hormones and reproduction in fish. Thus, these variables may alter the endocrine-disruption effects induced by EDCs.[82] In contrast to mammals and birds, the mechanisms underlying sex determination and differentiation in fishes vary widely and are changeable or labile in response to environmental parameters. These environmental parameters include temperature, behavioral cues and demographic structure of the local population, and EDCs. Understanding the gender similarities and differences in how organisms respond following exposure to environmental chemicals is important to determine the relative risk of these agents to wildlife and human populations. Given the central role of sex steroid hormones in the sex determination and sexual differentiation of fishes, amphibians, and reptiles, future research that includes sex as a factor is recommended. Thus, the risk assessment can address the probable gender differences in effects from exposure to chemicals in the environment.[83]

Dealing with environment eDc emission Municipal wastewater contains a complex mixture of EDCs originating from different sources. A number of organic pollutants, such as polycyclic aromatic hydrocarbons, PCBs, and pesticides, are resistant to degradation and represent an ongoing toxicological threat to both wildlife and human beings. Furthermore, recently, wastewater sludge has been subjected to reuse for production of valueadded products. These facts have heightened the need for novel and advanced bioremediation techniques to effectively remove EDCs from a variety of contaminated environmental media including water, wastewater sludge, sediments, and soils. One possibility to solve this problem is the use of microbial potential to degrade or detoxify EDCs and other toxic intermediates.[80,84]

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Also, there are physical methods such as absorption by activated carbon and rejection by membranes to remove EDCs. However, pollutant removal from wastewater is a process with high energy consumption, where cost and efficiency are the key considerations for their application. Biodegradation processes have proven to be the most cost-effective.[85] Water companies became aware of the endocrine-disruption problem when a survey confirmed the observation by anglers of hermaphrodite fish in wastewater treatment plant lagoons after being exposed to significant levels of persistent man-made chemicals. The evolving regulatory context related to micropollutants in the environment may have a decisive impact on wastewater management and requires an increased knowledge of the fate of micropollutants during wastewater treatment. Advanced treatments such as oxidation (ozone) are known to be able to enhance the removal of micropollutants, but technical, economic, and environmental risk/benefit evaluations must be performed before implementing such additional processes. In any case, the reduction of the pollution at the source, i.e., upstream of the wastewater treatment plant, represents the most sensible option, which should be promoted.[86] Although numerous studies have investigated degradation of individual EDCs in laboratory or natural waters, chemical-based analytical methods cannot represent the combined or synergistic activities between water quality parameters and/or the EDC mixtures at environmentally relevant concentrations since natural variations in water matrices and mixtures of EDC in the environment may confound analysis of the treatment efficiency. In conjunction with standard analytical approaches, bioanalytical assessments of residual estrogenic activity in treated water will enable estimates of the interactions and/ or combined estrogenic activity among mixtures of EDCs and the water matrix in natural water.[87] By contrast, the agricultural sector, a significant user of veterinary pharmaceuticals, has no such treatment—compounds are deposited straight to the ground in dung and urine or washed from hides in the case of topical applications. There has been little research as to whether any of these compounds leach into and persist in local soil and aquatic ecosystems. The extent to which the active ingredients of any of these chemicals (and their metabolites) leach into pastures, soil, runoff, and groundwater is a matter for field research. Also, much spraying of pesticides as herbicides and insecticides is done by ground crews. In such circumstances, it is not known whether they react with each other as well as pesticides and herbicides, forming further compounds which, either acting individually or in combination, could adversely affect bacteria, fungi, and higher organisms.[88] Above all, a ranking system that could be customized for specific geographical locations will aid public policies in prioritizing EDCs that need monitoring and removal of aquatic sources as drinking water.[89] The establishment of simple but integrative screening assays for regulatory purposes is allowed by a strong correlation between xenoestrogen exposure and reproductive impairment. In  fact, molecular screening assays could contain a battery of molecular targets allowing a more comprehensive approach in the identification of endocrine-disrupting compounds in fish and vertebrates in general.[90] Different assays can be successfully employed as a battery of assays to screen environmental water samples for estrogenicity. The results obtained from this battery of assays should be interpreted as a firsttier screen for estrogenicity. Samples that test positive should be further investigated using second-and third-tier screens with routine sampling in order to monitor rivers for estrogenicity.[91] Complementarily, a fugacity-based model may be applied to simulate the distribution of EDCs in reservoirs of recycled waste-water,[92] or a fugacity-hydrodynamic model may be used for predicting the concentrations of the organic pollutants in surface water.[93]

conclusion Endocrine-disrupting chemicals can cause a wide range of reproductive damage and developmental, growth, immune, and behavior effects even in low doses and by different mechanisms of action. They encompass a variety of chemical classes, including hormones, plant constituents, pesticides, compounds

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used in the plastics industry and in consumer products, and other industrial by-products and pollutants. Some of them are widely dispersed in the environment. Exposure to EDCs can occur through direct contact with these chemicals or through ingestion of contaminated water, sediment, air, soil, and food and consumer products. In humans, it is difficult to definitively link a particular EDC with a specific effect because the studies have inconclusive results. However, fetuses and embryos, whose growth and development are highly controlled by the endocrine system, are more vulnerable to exposure and may suffer reproductive abnormalities. The timing of exposure is also presumed to be critical, since different hormone pathways are active during different stages of development. Perinatal exposure, in some cases, can lead to permanent alterations that may be overt in adulthood. Compared with humans, the evidence that wildlife has been affected adversely by exposures to EDCs is extensive. Available evidence seems to indicate that endocrine disruption caused by xenobiotics is primarily an ecotoxicologic problem. These chemicals may be extremely challenging for aquatic organisms and mammals that have a large habitat and that consume fish from many different areas throughout their lives. Low concentrations of endocrine disrupters can have synergistic effects in various organisms as amphibians. For removal of these compounds from aquatic sources, the most cost-effective process is biodegradation. In spite of the need to manage the environmental, human health, and economic impacts of EDCs, most attention is focused on pharmaceutically active chemicals instead of those for agricultural use. The impact of these latter compounds is understudied. The legal approach has been improved by new test protocols. Progress has been made in the identification and quantification of a wide array of chemicals with endocrine-active properties, especially those that persist and bioaccumulate in organisms and their environment. Studies with mammals have shown that exposure to endocrine-active compounds during early development may result in adverse health impacts that are not realized until adulthood. However, from a regulatory perspective, the ability of animals to recover from chemical insults is problematic because it complicates efforts to establish acceptable levels of exposure. Consequently, research to define the limits and biological cost recovery using standardized test designs is needed.[94] However, exposure complexities, including transient and low-concentration exposure to EDCs, maternal metabolism of bioaccumulated EDCs, varying vulnerability and response by developmental stage, poorly understood exposure sources, mixtures and synergies, and cultural, social, and economic patterns, make it difficult for science to make solid exposure determinations. While there has been a great deal of research and effort in context with the hazard assessment and regulation of EDCs, there are also remaining uncertainties and issues. These include animal rights concerns due to significant increases in the use of animals to fulfill testing requirements; associated needs for alternative testing concepts such as in vitro, in silico, and modeling approaches; and lack of understanding of the relevance of exposure of humans and wildlife to EDCs.[95] Given the dynamic nature of the endocrine system, future efforts in the study of EDCs need more focus on the timing, frequency, and duration of exposure to these chemicals.

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43. Wingfield, J.C.; Mukai, M. Endocrine disruption in the context of life cycles: Perception and transduction of environmental cues. Gen. Comp. Endocrinol. 2009, 163, 92–96. 44. Beronius, A.; Rudén, C.; Hanberg, A.; Håkansson, H. Health risk assessment procedures for endocrine disrupting compounds within different regulatory frameworks in the European Union. Regul. Toxicol. Pharmacol. 2009, 55, 111–122. 45. Gori, G. Regulating endocrine disruptors. Regul. Toxicol. Pharmacol. 2007, 48, 1–3. 46. Campbell, B.K.; Souza, C., Gong, J.; Webb, R., Kendall, N.; Marsters, P., Robinson, G.; Mitchell, A.; Telfer, E.E.; Baird, D.T. Domestic ruminants as models for the elucidation of the mechanisms controlling ovarian follicle development in humans. Reprod. Suppl. 2003, 61, 429–43. 47. Wilson, V.S.; Blystone, C.R.; Hotchkiss, A.K.; Rider, C.V.; Gray, L.E., Jr. Diverse mechanisms of antiandrogen action: impact on male rat reproductive tract development. Int. J. Androl. 2008, 31, 178–187. 48. Porte, C.; Janer, G.; Lorusso, L.C.; Ortiz-Zarragoitia, M.; Cajaraville, M.P.; Fossi, M.C.; Canesi, L. Endocrine disruptors in marine organisms: Approaches and perspectives. Comp. Biochem. Physiol., Part C 2006, 143, 303–315. 49. Lafont, R.; Mathieu, M. Steroids in aquatic invertebrates. Ecotoxicology 2007, 16, 109–130. 50. LeBlanc, G.A. Crustacean endocrine toxicology: A review. Ecotoxicology 2007, 16, 61–81. 51. Kloas, W.; Lutz, I. Amphibians as model to study endocrine disrupters. J. Chromatogr. A 2006, 1130, 16–27. 52. Morales, M.; Planelló, R.; Martfnez-Paz, P.; Herrero, O.; Cortés, E.; Martfnez-Guitarte, J.L.; Morcillo, G. Characterization of Hsp70 gene in Chironomus riparius: Expression in response to endocrine disrupting pollutants as a marker of ecotoxicological stress. Comp. Biochem. Physiol. 2011, 153, 150–158. 53. Calamandrei, G.; Maranghi, F.; Venerosi, A.; Alleva, E.; Mantovani, A Efficient testing strategies for evaluation of xenobiotics with neuroendocrine activity. Reprod. Toxicol. 2006, 22, 164–174. 54. Danga, Z.C.; Ru, S.; Wang, W.; Rorije, E.; Hakkert, B.; Vermeire, T. Comparison of chemicalinduced transcriptional activation of fish and human estrogen receptors: Regulatory implications. Toxicol. Lett. 2011, doi:10.1016/j.toxlet.2010.12.020. 55. OECO Conceptual Framework for the testing and Assessment of Endocrine Disrupting Chemicals, available at http://www.oecd.org/document/580.3343,en_2649_34377_2348794_1_1_1_1.00.html (accessed May 2011). 56. Juberg, D.R. An evaluation of endocrine modulators: Implications for human health. Ecotoxicol. Environ. Safety 2000, 45, 93–105. 57. Borgert, C.J.; Mihaich, E.M.; Quill, T.F.; Marty, M.S.; Levine, S.L.; Becker, R.A. Evaluation of EPA’s Tier 1 Endocrine Screening Battery and recommendations for improving the interpretation of screening results. Regul. Toxicol. Pharmacol. 2011, 59, 397–411. 58. Vogel, J.M. Perils of paradigm: Complexity, policy design, and the Endocrine Disruptor Screening Program. Environ. Health: Global Access Sci. Source 2005, 4, 2, doi:10.1186/1476-069X-4-2. 59. Nichols, J.W.; Breen, M.; Denver, R.J.; Distefano, J.J.; Edwards, J.S.; Hoke, R.A.; Volz, D.C.; Zhang, X. Predicting chemical impacts on vertebrate endocrine systems. Environ. Toxicol. Chem. 2011, 30 (1), 39–51. 60. McKinlay, R.; Plant, J.A.; Bell, J.N.B.; Voulvoulis, N. Calculating human exposure to endocrine disrupting pesticides via agricultural and non-agricultural exposure routes. Sci. Total Environ. 2008, 398, 1–12. 61. Kortenkamp A.; Faust M. Combined exposures to anti-androgenic chemicals: Steps towards cumulative risk assessment. Int. J. Androl. 2010, 33, 463–474. 62. Muncke, J. Exposure to endocrine disrupting compounds via the food chain: Is packaging a relevant source? Sci. Total Environ. 2009, 407, 4549–4559. 63. Narita, M.; Miyagawa, K.; Mizuo, K.; Yoshida, T.; Suzuki, T. Prenatal and neonatal exposure to low-dose of bisphenol-A enhance the morphine-induced hyperlocomotion and rewarding effect. Neurosci. Lett. 2006, 402, 249–252.

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64. Queiroz, E.K.R.; Waissmann, W. Occupational exposure and effects on the male reproductive system. Cadernos de Saúde Pública 2006, 22 (3), 485–493. 65. Bretveld, R.W.; Hooiveld, M.; Zielhuis, G.A.; Pellegrino, A.; van Rooij, I.A.L.M.; Roeleveld, N. Reproductive disorders among male and female greenhouse workers. Reprod. Toxicol. 2008, 25, 107–114. 66. Gregoraszczuk, E.L.; Milczarek, K.; Wojtowicz, A.K.; Berg, V.; Skaare, J.U.; Ropstad, E. Steroid secretion following exposure of ovarian follicular cells to three different natural mixtures of persistent organic pollutants (POPs). Reprod. Toxicol. 2008, 25, 58–66. 67. Kortenkamp A. Low dose mixture effects of endocrine disrupters: implications for risk assessment and epidemiology. Int. J. Androl. 2008, 31, 233–240. 68. Jacobsen, P.R.; Christiansen, S.; Boberg, J.; Nellemann C.; Hass, U. Combined exposure to endocrine disrupting pesticides impairs parturition, causes pup mortality and affects sexual differentiation in rats. Int. J. Androl. 2010, 33, 434–442. 69. Rider, C.V.; Furr, J.R.; Wilson, V.S.; Gray, L.E., Jr. Cumulative effects of in utero administration of mixtures of reproductive toxicants that disrupt common target tissues via diverse mechanisms of toxicity. Int. J. Androl. 2010, 33, 443–462. 70. Birkhoj, M.; Nellemann, C.; Jarfelt, K.; Jacobsen, H.; Andersen, H.R.; Dalgaard, M; Vinggaard, A.M. The combined antiandrogenic effects of five commonly used pesticides. Toxicol. Appl. Pharmacol. 2004, 201, 10–20. 71. Drake, A.J.; van den Driesche, S.; Scott, H.M., Hutchison, G.R.; Seckl, J.R.; Sharpe, R.M. Glucocorticoids amplify dibutyl phthalate-induced disruption of testosterone production and male reproductive development. Endocrinology 2009, 150, 5055–5064. 72. Chen, X.; An, H.; Ao, L.; Sun, L.; Liu, W.; Zhou, Z.; Wanga, Y.; Cao, J. The combined toxicity of dibutyl phthalate and benzo(a)pyrene on the reproductive system of male Sprague Dawley rats in vivo. J. Hazard. Mater. 2011, 186 (1), 835–841. 73. Pohl, H.R.; Luukinen, B.; Holler, J.S. Health effects classification and its role in the derivation of minimal risk levels: Reproductive and endocrine effects. Regul. Toxicol. Pharmacol. 2005, 42, 209–217. 74. Erhard, H.W.; Rhind, S.M. Prenatal and postnatal exposure to environmental pollutants in sewage sludge alters emotional reactivity and exploratory behaviour in sheep. Sci. Total Environ. 2004, 332, 101– 108. 75. Mills, L.J.; Chichester, C. Review of evidence: Are endocrine-disrupting chemicals in the aquatic environment impacting fish populations? Sci. Total Environ. 2005, 343, 1– 34. 76. Fox, G.A. Wildlife as sentinels of human health effects in the Great Lakes-St. Lawrence Basin. Environ. Health Perspect. 2001, 109 (Suppl. 6), 853–861. 77. Kinnberg, K.; Holbech, H.; Petersen, G.I.; Bjerregaard, P. Effects of the fungicide prochloraz on the sexual development of zebrafish (Danio rerio). Comp. Biochem. Physiol., Part C 2007, 145, 165–170. 78. Monod, G.; Rime, H.; Bobe, J.; Jalabert, B. Agonistic effect of imidazole and triazole fungicides on in vitro oocyte maturation in rainbow trout (Oncorhynchus mykiss). Mar. Environ. Res. 2004, 58, 143–146. 79. Jobling, S.; Burn, R.W.; Thorpe, K.; Williams, R.; Tyler, C. Statistical modeling suggests that antiandrogens in effluents from wastewater treatment works contribute to widespread sexual disruption in fish living in English rivers. Environ. Health Perspect. 2009, 117 (5), 797–802. 80. Barnabé, S.; Brar, S.K.; Tyagi, R.D.; Beauchesne, I.; Surampalli, R.Y. Pre-treatment and bioconversion of wastewater sludge to value-added products—Fate of endocrine disrupting compounds. Sci. Total Environ. 2009, 407, 1471–1488. 81. Nelson, J.; Bishay, F.; van Roodselaar, A.; Ikonomou, M.; Law, F.C.P. The use of in vitro bioassays to quantify endocrine disrupting chemicals in municipal wastewater treatment plant effluents. Sci. Total Environ. 2007, 374, 80–90.

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82. Jin, Y.; Shu, L.; Huang, F.; Cao, L.; Sun, L.; Fu, Z. Environmental cues influence EDC-mediated endocrine disruption effects in different developmental stages of Japanese medaka (Oryzias latipes). Aquat. Toxicol. 2011, 101, 254–260. 83. Orlando, E.F.; Guillette, L.J., Jr. Sexual dimorphic responses in wildlife exposed to endocrine disrupting chemicals. Environ. Res. 2007, 104, 163–173. 84. Robinson, B.J.; Hellou, J. Biodegradation of endocrine disrupting compounds in harbour seawater and sediments. Sci. Total Environ. 2009, 407, 5713–5718. 85. Liu, Z.; Kanjo, Y.; Mizutani, S. Removal mechanisms for endocrine disrupting compounds (EDCs) in wastewater treatment—Physical means, biodegradation, and chemical advanced oxidation: A review. Sci. Total Environ. 2009, 407, 731–748. 86. Janex-Habibi, M.; Huyard, A.; Esperanza, M.; Bruchet, A. Reduction of endocrine disruptor emissions in the environment: The benefit of wastewater treatment. Water Res. 2009, 43, 1565–1576. 87. Chen, P.; Rosenfeldt, E.J.; Kullman, S.W.; Hinton, D.E.; Linden, K.G. Biological assessments of a mixture of endocrine disruptors at environmentally relevant concentrations in water following UV/H2O2 oxidation. Sci. Total Environ. 2007, 376, 18–26. 88. Fisher, P.M.J.; Scott, R. Evaluating and controlling pharmaceutical emissions from dairy farms: A critical first step in developing a preventative management approach. J. Cleaner Prod. 2008, 16, 1437–1446. 89. Kumar, A.; Xagoraraki, I. Pharmaceuticals, personal care products and endocrine-disrupting chemicals in U.S. surface and finished drinking waters: A proposed ranking system. Sci. Total Environ. 2010, in press 90. Scholz, S.; Mayer, I. Molecular biomarkers of endocrine disruption in small model fish. Mol. Cell. Endocrinol. 2008, 293, 57–70. 91. Swart, J.C.; Pool, E.J.; van Wykb, J.H. The implementation of a battery of in vivo and in vitro bioassays to assess river water for estrogenic endocrine disrupting chemicals. Eco-toxicol. Environ. Saf. 2011, 74, 138–143. 92. Cao, Q.; Yu, Q.; Connell, D.W. Fate simulation and risk assessment of endocrine disrupting chemicals in a reservoir receiving recycled wastewater. Sci. Total Environ. 2010, 408, 6243–6250. 93. Zhang, Y.; Song, X.; Kondoh, A.; Xia, J.; Tang, C. Behavior, mass inventories and modeling evaluation of xenobiotic endocrine-disrupting chemicals along an urban receiving wastewater river in Henan Province, China. Water Res. 2011, 45, 292–302. 94. Nichols, J.W.; Breen, M.; Denver, R.J.; Distefano, J.J., III; Edwards, J.S.; Hoke, R.A.; Volz, D.C.; Zhang X. Predicting impacts on vertebrate endocrine systems. Environ. Toxicol. Chem. 2011, 30 (1), 39–51. 95. Hecker, M.; Hollert, H. Endocrine disruptor screening: Regulatory perspectives and needs. Environ. Sci. Eur. 2011, 23, 15, doi:10.1186/2190-4715-23-15.

3 Herbicides Herbicide Discovery.......................................................................................27 Herbicide Classification ................................................................................. 28 Method or Timing of Application • Chemical Structure or Mode of Action

Malcolm Devine

Components of Herbicide Action ................................................................28 Herbicide Selectivity ...................................................................................... 28 Herbicide Mode of Action............................................................................. 29 Herbicide Resistance in Weeds .....................................................................29 Herbicide-Resistant Crops ............................................................................30 Safety and Environmental Fate of Herbicides.............................................30 References .........................................................................................................31

Herbicide Discovery Traditionally, herbicides have been discovered through large chemical synthesis and screening programs. Given the high cost of discovery and development of new compounds, efforts have focused on herbicides that target major weeds in major world crops (e.g., corn, rice, wheat, soybean, cotton). However, these compounds often find uses in minor crops, also. The process normally involves a step wise progression from greenhouse screening on a few crop and weed species, to more extensive indoor testing, and eventually to field trials in many locations examining the interactions between different weed species, soil types, weather conditions, etc. Toxicological testing and formulation improvement proceed concurrently, to ensure that regulatory and efficacy requirements are met. A recent innovation is the use of combinatorial chemistry to identify lead compounds. Rather than being synthesized and evaluated in isolation, compounds are produced and screened as mixtures. When combined with in vitro screens (activity testing at the biochemical or cellular level) rather than wholeplant assays, this can allow the testing of 20- to 50-fold more compounds per year than a traditional herbicide discovery program. “Rational” discovery of herbicides involves identification of a candidate herbicide target site in the plant, followed by design of inhibitors that specifically block that target. While this has led to the discovery of some novel enzyme inhibitors, to date no commercial herbicides have been discovered through this approach. One difficulty is that compounds predicted to have high activity might not penetrate the tissue satisfactorily or may be rapidly degraded inside the tissue. Finally, herbicides can be developed from bacterial, fungal, or plant toxins. One commercial herbicide, glufosinate (= phosphinothricin), was developed from the bacterial toxin bialaphos from Streptomyces hygroscopicus.

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Herbicide classification Method or timing of Application Preplant incorporated herbicides are applied to the soil surface and mechanically incorporated into the upper 5–10 cm of the soil, in order to minimize photodecomposition and volatilization losses. Preemergence herbicides are applied to the soil surface and often rely on precipitation or soil moisture to transport them to the plant root or shoot for uptake. Postemergence herbicides are applied to exposed foliage after the plants have emerged.

chemical Structure or Mode of Action Herbicides within the same structural family usually have the same mode of action, with varying degrees of activity or selectivity depending on the structural variations between compounds. However, herbicides from different chemical families can have the same mode of action (see later).

components of Herbicide Action To be effective, herbicides must penetrate the tissue and reach the target site in sufficiently high concentrations to block its activity. The overall process of herbicide action can be separated into the following components:[1] Absorption: Herbicides can be absorbed by the roots or directly into the leaves. Root uptake occurs through mass flow of herbicide in soil moisture and is a function of root distribution in the soil, soil moisture status, the physical properties of the soil, and the behavior of the herbicide in the soil. Foliar absorption occurs following application to the leaves. To facilitate entry through the cuticle (the waxy layer on most leaf surfaces), herbicides are usually formulated with an array of inert ingredients including surfactants, emulsifiers, etc. Once inside the tissue, further penetration through cell walls and membranes usually occurs by simple diffusion. Translocation: Long-distance transport of herbicides in the plant can occur in the xylem and/or phloem. In some instances, distribution through the plant is a critical component of overall activity. The amount of translocation depends on the plant stage of development, the physicochemical properties of the herbicide, and the rate at which herbicide injury slows down the movement of endogenous compounds. Metabolic degradation: Plants have evolved various enzyme systems to detoxify potentially harmful compounds. Fortuitously, some of these enzymes (e.g., cytochrome P450 monooxygenases, glutathione S-transferases) can degrade herbicides to inactive compounds. Interaction at the target site: Finally, all herbicides must interfere with some critical process in the plant. In most cases this involves binding to a protein (usually a functional enzyme, or a transport or structural protein) so that it cannot carry out its normal function. Over time, through the combined effect of this direct action and other indirect actions, the plant dies.

Herbicide Selectivity Most herbicides are selective, that is, they kill some plant species but not others. A few herbicides, on the other hand, are nonselective, and kill essentially all species. Selectivity is normally based on one of the following:[1] • Failure to absorb the herbicide, due to either selective placement (e.g., directed spray on weeds growing between the crop rows) or failure of the herbicide to reach the roots of deep rooted crops, while killing shallow rooted weeds.

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• Enhanced rate of metabolic degradation of the herbicide in the crop. This is the most common basis of selectivity of agricultural herbicides. After entry into the plants, the crop metabolizes the herbicide to inactive compounds, whereas degradation does not occur or is slower in the susceptible weeds. • Differential sensitivity of the target site. This is particularly common in the case of herbicide resistant weeds (see later).

Herbicide Mode of Action Herbicides kill plants by interfering with an essential process in the plant. The major modes of action of herbicides, the biochemical target sites, and some examples of chemical groups that interfere with those targets, are shown in Table 1. In addition, many herbicides have been identified that interact with other, unique target sites in plants. However, most of the herbicides that have been developed over the past 50 years target about 15 distinct molecular targets.

Herbicide Resistance in Weeds The repeated use of herbicides can lead to the development of herbicide-resistant weed populations. The  use of herbicides per se does not create herbicide resistance. Rather, the continuous selection pressure through herbicide use creates a niche, allowing resistant individuals to survive and increase in TABLE 1

Modes of Action of Major Herbicide Groupsa

Mode of Action

Target Site

Representative Chemical Groups

Inhibition of amino acid biosynthesis Branched chain amino acids

Acetolactate synthase

Glutamine synthesis Aromatic amino acid biosynthesis

Glutamine synthetase Enolpyruvylshikimate phosphate synthase

Sulfonylureas, imidazolinones, triazolopyrimidines Glufosinate Glyphosate

Photosynthesis

Chlorophyll synthesis Carotenoid synthesis

Qb or D1 protein PS I electron acceptor Pigment biosynthesis Protoporphyrinogen oxidase Phytoene desaturase and others

Fatty acid synthesis

Acetyl-coA carboxylase

Fatty acid elongation

“Elongase” complex Cell division β-Tubulin

Photosynthetic electron transport (PS II) Photosynthetic electron transport (PS I)

S-Triazines, phenylureas, benzonitriles Bipyridiliums (Nitro) diphenylethers Aminotriazole, clomazone

Lipid biosynthesis

Spindle formation

Aryloxyphenoxypropionates, cyclohexanediones Thiocarbamates Dinitroanilines, carbamates

Other Auxin disruption Homogentisate biosynthesis a

From Devine.[1]

Auxin binding proteins (?) 4 hydroxyphenylpyruvate dioxygenase

Phenoxyacetic acids, benzoic acids Isoxazoles

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the population as susceptible weeds are killed. Thus, starting with an initial population of perhaps one resistant weed in a population of 106–109, resistance builds up over time until the resistant weeds become predominant in the field. Herbicide resistance in weeds was first observed in the late 1950s, but was a minor problem until the mid-1970s, when resistance to triazine herbicides became a widespread concern. Since then the occurrence of resistance has increased dramatically, with >200 cases now reported.[2] In most cases, resistance is due to a point mutation in the gene coding for the herbicide target site.[3] This alters the structure of the target protein in such a way that it can still perform its natural function, but herbicide binding is reduced. This type of resistance often confers cross-resistance to herbicides in the same chemical family or mode-of-action group, although exceptions exist and each case must be analyzed separately. However, target-site mutations do not alter the sensitivity of the weed to herbicides with other mechanisms of action. Resistance can also be conferred by elevated activity of the enzyme(s) responsible for herbicide degradation. These weeds can be cross-resistant to other herbicides with different mechanisms of action. Again, the possibilities for cross-resistance have to be analyzed on a case-by case basis. Although herbicide resistance has become widespread, in almost all cases alternative control methods are available, through the use of other herbicides, changes in cropping or tillage practices, or some combination of these. Avoidance and management of herbicide resistance has become an integral part of good farming practices in modern agriculture.

Herbicide-Resistant crops A recent development in selective weed control has been to create resistance to certain herbicides in crops where it did not exist previously. While this has been done primarily to extend the market share of certain products, it offers farmers the advantage of broad-spectrum weed control in crops with a single herbicide application.[4] In some cases this has substantially reduced the total amount of herbicide required in a single season. Herbicide-resistant crops can be produced by three methods:[4] 1. Making crosses between the crop (sensitive to the herbicide) and a related, resistant species. This method was used to develop triazine-tolerant canola (Brassica napus). 2. Selecting resistant cells in tissue culture, through random mutation or by selecting somaclonal variants, and regenerating resistant plants from these cells. Corn lines resistant to the herbicide sethoxydim were generated in this way. 3. Transfer of an herbicide-resistance gene through genetic engineering. The gene is identified in an unrelated species (often a bacterium), cloned, and transferred into the crop species of interest. These procedures were used to develop canola varieties resistant to the nonselective herbicides glufosinate and glyphosate. Herbicide-resistant crops have greatly facilitated weed control, but present some new research questions that have had to be addressed. These include the likelihood and long term ecological consequences of gene flow to related species, and the need to control volunteer plants in the following season(s). These issues do not present insurmountable obstacles but, again, need to be dealt with on a case-by-case basis.

Safety and environmental Fate of Herbicides Environmental safety is of prime concern in the development of new herbicides. This includes an understanding of herbicide toxicology, safe handling and application procedures, and environmental behavior and fate. In most countries, approval of herbicides by the relevant decision making bodies is dependent on the registrant satisfying the regulatory requirements imposed by those countries.

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Toxicological requirements vary from country to country, but usually include data on oral and dermal toxicity in a range of species, in tests of varying duration. Based on the data collected, maximum residue levels are established in food or food products. Data from these tests and field experiments are used to establish maximum application doses and safe intervals between product application and harvest or grazing. Data may also be required on effects on nontarget organisms and ecosystems that may be exposed to low herbicide doses. Appropriate handling procedures are an important aspect of herbicide safety. The use of appropriate safety clothing (gloves, coveralls, masks, etc.), more benign formulations (e.g., dispersible granules rather than wettable powders), etc., contribute to reduced applicator exposure. Recently, novel formulations have been developed that further reduce applicator exposure when adding products to the spray tank. Herbicide drift immediately after spraying can be a source of off-site contamination, resulting in injury to adjacent sensitive crops and other species. Various measures, including spraying only under calm conditions, use of wind deflectors, and avoiding very small droplets, can substantially reduce the risk of spray drift. Herbicides in soil are lost by a combination of microbial and chemical degradation, plant uptake, and, in some instances, leaching or surface run-off. Stringent environmental regulations have been introduced in many countries to minimize the possibility of groundwater contamination. Herbicide residues in soil can provide extended weed control, but also may limit crop rotation options in future seasons. Field research is conducted to establish the risk of such carryover and the effects on future crops.

References 1. Devine, M.D.; Duke, S.O.; Fedtke, C. Physiology of Herbicide Action; Prentice-Hall: Englewood Cliffs, NJ, 1993; 441. 2. Heap, I.M. International survey of herbicide-resistant weeds. http://www.weedscience.com (accessed April 2000). 3. Devine, M.D.; Eberlein, C.V. Physiological, Biochemical and Molecular Aspects of Herbicide Resistance Based on Altered Target Sites. In Herbicide Activity: Toxicology, Biochemistry and Molecular Biology; Roe, R.M., Burton, J.D., Kuhr, R.J., Eds.; IOS Press: Amsterdam, 1997; 159–185. 4. Herbicide-Resistant Crops. Agricultural, Environmental, Economic, Regulatory and Technical Aspects; Duke, S.O., Ed.; CRC Press: Boca Raton, FL, 1996; 420.

4 Herbicides: Non-Target Species Effects Introduction .................................................................................................... 33 History, Types of Herbicides, and Their Use............................................... 34 Benefits to Agriculture and Forest Management ....................................... 34 Herbicide Use and Exposure to Primary Producers .................................. 35 Herbicide Toxicity to Humans and Animals ..............................................37 Toxicity to Plants and Effects to Terrestrial Habitats ................................. 38 Effects at the Species Level • Population/Community Level • Ecosystem/Trophic Level

Céline Boutin

Factors Interacting with Herbicide Effects .................................................. 41 Measuring Phytotoxicity................................................................................42 Conclusion ......................................................................................................43 Acknowledgments ..........................................................................................44 References ........................................................................................................44

introduction Pesticides such as inorganic chemicals (e.g., sulfur, arsenic, or other metal compounds) have been used in agriculture for centuries to protect crops. In more recent times, especially after World War II, organic pesticides have been discovered and increasingly used to suppress unwanted plants, insects, and other organisms that interfere with crop production. The main pesticides used are categorized by their target organisms: insecticides used to kill or suppress insects, fungicides used for pathogens, and herbicides for weeds. In this entry, we are mostly concerned with herbicides because of their considerable use, especially in North America, and their phytotoxic effects. The definition accepted by the Weed Science Society of America is that an herbicide is “a chemical substance or cultured organism used to kill or suppress the growth of plants.”[1] Phytotoxicity refers to the capability of herbicides or other pesticides to exert toxic effect on plant growth, reproduction, and survival. By extension and in addition to terrestrial and aquatic vascular plants, it usually includes other primary producers such as algae and cyanobacteria, which will not be considered in this entry. This entry will first consider the history, types, and main uses of herbicides. There are advantages and disadvantages of using herbicides in agriculture and forest management. Benefits have long been established; however, the limitations and undesirable effects attributed to herbicide use are still debated. The environmental impact will be examined, including environmental exposure, phytotoxicity, and toxicity to different trophic levels, as well as the multitudinous factors to take into account in risk assessment. Lastly, the techniques and

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limitations of the phytotoxicity assessments currently used to determine environmental risk evaluations will be discussed. Mitigation measures and alternatives to herbicide use will be discussed in conclusion.

History, types of Herbicides, and their Use DNOC or 4,6-dinitro-o-cresol, developed in 1935, was the first organic herbicide used to control weeds. This herbicide was also used as an insecticide, fungicide, and a defoliant and was shown to be toxic to animals.[2] It is no longer used in many countries. In the 1940s, several phenoxy herbicides were discovered and many are still used today, including 2,4-D and MCPA [(4-chloro-2-methylphenoxy)acetic acid]. In the 1950s, other herbicides appeared on the market, including diuron, diquat, paraquat, and triallate. In the 1960s, the triazine herbicides were developed and are still considerably used, especially atrazine. Glyphosate, which is still the most widely used herbicide worldwide, was developed in the 1970s. In the 1980s, low-dose high-efficacy herbicides such as sulfonylureas and imidazolinones were found. Since the late 1990s, a new method for controlling weeds has emerged with the development of herbicide tolerant crops, i.e., genetically engineered crops resistant to glyphosate and, to a lesser extent, glufosinate ammonium. More genetically modified herbicide tolerant crops will no doubt be engineered or bred in the future. Herbicides exhibit different mechanisms of action.[1,3] They include disruption of photosynthesis (uracils, substituted ureas, and triazines), inhibition of lipid biosynthesis (carbamothioates such as EPTC [s-ethyl dipropylthiocarbamate], triallate, clethodim, fluazifop, and metolachlor), inhibition of cell division (dinitroanilines such as trifluralin and pendimethalin), plant hormone mimics (the phenoxy herbicide 2,4-D, MCPA, the benzoic acid dicamba, and the picolinic acid picloram), inhibition of amino acid biosynthesis (glyphosate, sulfonyl ureas including metsulfuron methyl and chlorsulfuron, imidazolinones including imazethapyr), blockage of carotenoid biosynthesis (clomazone), and disruption of cell membranes (acifluorfen and bipyridylium compounds such as diquat and paraquat). Other herbicides such as glufosinate ammonium act by inhibiting glutamine synthetase, thus leading to a complete breakdown of ammonia metabolism in affected plants. Uncouplers of oxidation phosphorylation, such as the widely used bromoxynil, interfere with plant respiration. Many herbicides act primarily on systems unique to plants, e.g., photosynthesis, but some herbicides act at more than one site of action. Undoubtedly, the secondary mode of action of some herbicides could explain their relatively high toxicity to animals (see below). In the case of some herbicides, the precise mode of action is unknown and exact molecular sites of action remain to be determined.[4] The number of herbicides listed in the Weed Science Society of America reached 374 in 2010.[5] In Canada, there are 500 pesticide active ingredients (the ingredient to which the pesticide is attributed) and 7000 pesticide formulated products (mixture containing one or several active ingredients and formulants) available since many formulated products contain a mixture of active ingredients.[6] The number of herbicide active ingredients registered in Canada amounts to approximately 125.[6] The majority of herbicide use occurs in agriculture where in modern practices they dominate weed control practice. Approximately 90% of areas planted with corn, cotton, potato, wheat, and soybean were sprayed with herbicides in the United States in 2004.[1] While the benefits of using herbicides from an agronomic perspective are well known, the undesirable effects of herbicides on the environment have not always been considered carefully. Regardless of the method of application, it is generally accepted that misplacement will take place through drift at the time of application or through runoff, leaching, and volatilization from soil or plants or from particles moving with contaminated soil after application has occurred.

Benefits to Agriculture and Forest Management Benefits of herbicide to agriculture, forest management, and other agronomic applications leave no doubt as to their utility, although some applications for cosmetic reasons (e.g., domestic use, horticulture, golf courses) are more questionable. The negative relationship between crop yield and weed density

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is well established.[1] Weeds are well-adapted species that compete with crops in disturbed environments and will reduce crop yield depending on their germination timing and densities, growth patterns, and growth rates. However, there may be cases where weeds can be beneficial to crops. As an example, field experiments were conducted to evaluate the effects of nicosulfuron and imazethapyr, a sulfonyl urea and an imidazolinone herbicide, respectively, for the control of johnsongrass (Sorghum halepense L.), a weed difficult to control in corn fields.[7] It was noted that corn vigor was greatly reduced in plots where these two herbicides were applied and where johnsongrass was reduced. In order to verify if there was an unwanted effect of the two herbicides on corn, small plots were sprayed with the two herbicides separately. Two other treatments were also included in the experimental design: control of johnsongrass via mechanical means and no treatment. It was noticed that in plots where johnsongrass was removed, corn was more prone to being attacked by the maize dwarf mosaic and maize chlorotic dwarf viruses, which are transmitted by aphids or leafhoppers. Results revealed that in treated plots, virus disease was increased because the preferred host johnsongrass was suppressed. Furthermore, there was no change in yield between treatments. Crop yield was reduced in treated plots (whether through mechanical or chemical treatments) due to increased virus severity, while in non- treated plots, reduced corn yield was due to more competition from johnsongrass.

Herbicide Use and exposure to Primary Producers The amount of herbicide active ingredients used worldwide exceeded 950 million kg in 2007, of which 39% was herbicides.[8] Herbicides are used most extensively in North America. For instance, in 2003, herbicides accounted for 77% of total pesticide sales in Canada, while fungicides represented 9%, insecticides 8%, and other products 6%.[6] In the United States, 47% of pesticides used were herbicides, while insecticides and fungicides amounted to 8% and 6%, respectively.[8] In Europe, fungicides are used extensively mainly due to the use of sulfur fungicides on vineyards by France, Italy, Spain, and Greece.[9] Herbicides constitute less than 35% of pesticides used in Europe, but nevertheless, this is approximately 75,000 tons of applied active ingredients.[10] Of note, it appears that these European data underestimated the actual amount of pesticides used.[9] France is the biggest herbicide user followed by Germany and the United Kingdom. In other parts of the world, insecticides constitute the bulk of the pesticides used.[11] Large amounts of herbicides are annually used in terms of hectares sprayed and in terms of quantity per hectares. In Canada, in excess of 25 million ha are sprayed with atleast one herbicide on an annual basis.[12] In the United States, 110 million ha were treated with herbicides in the late 1980s,[13] and in 1997, 209 million kg of herbicide active ingredients were used (Olszyk et al.[14] and references therein). Total land area devoted to agriculture worldwide is considerable,[15] especially in Europe where around 70% of the land is cropland or pastureland.[16] As a result, total areas treated with herbicides reach large figures. Herbicide application is highly dependent on the types of crops, the prevailing climate conditions, the land use, and other factors such as topography. In the United Kingdom where 77% of the land is used for agriculture, pesticide use was also intensive at 5.8 kg/ha in 1988[17] but seemed to have declined in 1999 at 2.7 kg/ha (1.4 kg-ai of herbicides), not taking into account the underestimation mentioned previously.[9] By contrast, in France, the intensity of pesticide used increased from 4.4 kg-ai/ha in 1988 to 6.0 kg-ai/ha in 1999, including 2.0 kg-ai/ha of herbicides. This is well above the average for European member states, which used approximately 4.5 kg-ai/ha of pesticides and 1.3 kg-ai/ha of herbicides. In the Netherlands, 11.8 kg-ai/ha of pesticides are used on an annual basis (mostly non-herbicides).[18] On the other hand, in Canada, pesticide use is low at 0.9 kg/ha (largely herbicides).[6] The United States uses 23% of all the pesticides used in the world, and 28% of all herbicides. These figures, however, dissimulate important facts. In Canada, where less than 10% of the land is devoted to agriculture, two ecoregions are particularly important in this regard, the Prairies of western Canada and the Mixed- wood Plains of the Great Lakes–Saint Lawrence corridor. The Prairies, which constitute the largest agricultural areas in Canada (5.1% of the land),[19] has been almost totally modified

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to satisfy increasing needs for more cropping areas. Consequently, there are almost no pristine grassland prairies left where native plant and animal communities can survive. The Mixedwood Plains are a smaller area in southern central Canada (1.5% of the land cover) where 50% of the Canadian population lives and where agriculture is also very intense. In this region, a large portion of the Carolinian forest and the mixed-wood forest has disappeared to satisfy human needs. In the United States, agriculture covers 48% of the land and is concentrated in the mid-west regions[20] where original ecosystems have largely vanished. The same pattern is repeated in Europe where mostly seminatural habitats remain in most countries. What remains of these ecosystems interspersed in a sea of intensively cultivated land can be greatly impacted by herbicide use. Herbicide exposure to non-target environments can occur when application is performed with aircraft, mist-blower, and ground applications through overspray or via spray drift, vapor drift, revolatilization from soil and plants, runoff, or dust particles moved by wind or water.[14] Depending on the equipment and prevailing weather conditions during application, the amount of sprayed herbicide that will deposit in hedgerows and other field edges from multiple consecutive spray tracks can reach 1% to 10% of the application rate within 10 m of a single swath with ground equipment (Boutin and Jobin[21] and references therein), and much more with mist-blower sprayers and aerial applications.[22] Herbicides can travel a considerably longer distance with aerial equipment applications, for example, 500 m downwind from the source.[23,24] With ground application, it was found that herbicides could cover long distances. A study was undertaken to assess the protection afforded by buffer zones from herbicide drift (Boutin and Baril, unpublished data). Surveys of nontarget plants situated in two small woodlots adjacent to crop fields were conducted prior to (May) and after (May, June, and July) herbicide application. Vegetation was surveyed for community composition and symptoms of herbicidal impact. The experimental work was conducted in southwestern Ontario, Canada, under normal field operation conditions, for soybean sprayed with imazethapyr in 1993, corn with dicamba in 1994, and wheat with MCPA in 1996, following the usual rotation in southwestern Ontario. The buffer zone was defined as a 12 m wide seeded strip of crop, upwind of a woodlot, where herbicides were not applied. Each treatment consisted of four transects, divided into sampling points, at 1 m, 2 m, 4 m, 8 m, 16 m, and 32 m distances into the woodlots. Herbicide application occurred in the early morning or evening, when no precipitation was forecasted, when wind speed was at or less than 8 km/hr, and when the direction of the wind was across the soybean field into the woodlot. Results revealed that herbicides could move up to 32 m into woodlots (Figure 1). Up to 43% of the vegetation of one quadrat (of the 23 species) in the woodlots showed visual effects characteristics of herbicidal injury: discoloration, bleaching, epinasty, yellow or brown spots, etc. Effects were less pronounced in transects abutted to buffer zones. The plants most affected [e.g., raspberry (Rubus idaeus L.), goldenrod (Solidago canadensis L.), and ash tree (Fraxinus spp.)] were species of open areas growing in the first few meters of the woodlots. In some cases, effects lasted for more than 2 mo. Vapor drift can also migrate a long way (Franzaring et al.[25] and references therein) causing recurrent sublethal effects on native plants not only in bordering seminatural habitats but also in more remote habitats. Presence of airborne herbicides in the atmosphere has been reported in Europe[26] and North America.[27] In the Netherlands, it was found that non-target vegetation was repeatedly exposed to small amounts of herbicides (Franzaring et al.[25] and references therein). Pesticides used in agriculture in southern Canada, including some herbicides, have been reported in the  arctic environments.[28] Of the 10 chemicals surveyed by Hoferkamp et al.,[29] 9 were detected in the arctic. Traveling distances for these chemicals ranged from 55 km to 12,100 km. Persistence in the environment of these herbicides can explain their presence in the arctic where they are transported via the air or by dust. Undoubtedly, the large number of herbicides available for use, the geographical extent of their use on different crops, and the quantity applied together with the method of application suggest a high probability of exposure to primary producers and other wildlife.

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FIGURE 1 Number of plants showing herbicide effects in two woodlots (L and H) surveyed during 3 years prior to and after herbicide spray under normal field operations. Fields were sprayed with the herbicides imazethapyr in 1993, dicamba in 1994, and MCPA in 1996. Plants were surveyed in quadrats situated between 1 and 32 m from crop fields. A 12 m buffer zone within fields was used in half of the fields while the other half was sprayed right to the field edge. Four transects were placed for each treatment. In total, 116 species were inventoried and 35 species were found to be sensitive to herbicides during the course of the 3 years the study lasted.

Herbicide toxicity to Humans and Animals Herbicides are generally not very toxic to humans or to animals, including mammals, birds, amphibians, and invertebrates, as well as to microbial organisms. There are some exceptions, however, notably paraquat, which is very toxic to humans. It is still used extensively in the United States[30] and many developing countries. A large epidemiological study suggested that the herbicides pendimethalin and EPTC may be linked with human pancreatic cancer.[31] Sulfonyl ureas, which comprise a large number of low-dose herbicides widely used in agriculture, are known to affect humans as they are used in medicine

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to treat Type 2 diabetes and other illnesses.[32] They act by increasing insulin release from the beta cells in the pancreas, and side effects such as incidence of hypoglycemia have been reported. As early as 1976, the herbicide diquat was shown to have an adverse effect on body size and pigmentation of Xenopus laevis, the South African clawed frog (family: Pipidae; subfamily: Xenopinae). Moreover, when treated with diquat and the fungicide nabam, the deleterious effect was more pronounced.[33] More recently, the herbicide atrazine has been implicated in a number of reports demonstrating effects on animals as an endocrine disruptor.[34] Similarly, the formulated product containing glyphosate is widely known to affect amphibians.[34] Glyphosate formulation has been implicated in a synergistic effect with a trematode species on fish.[35] Survival of juvenile fish was unaffected by exposure to glyphosate alone or by an infectious trematode parasite alone. However, simultaneous exposure to infection and glyphosate significantly reduced fish survival. Spinal malformations of juvenile fish were also enhanced when both stressors were present. A species of snail acts as the vector between the fish and the trematode. Glyphosate at high concentration killed all the snails, but at moderate concentrations, the snail produced more trematodes than in control and low-concentration groups.[35] This elegant experiment demonstrated the intricate interactions between different components of ecosystems. Herbicides are categorized for their effect on unwanted plants, yet a number of them demonstrate toxicity to microbial organisms. A recent study revealed the control activities of glyphosate against rust diseases (Puccinia striiformis and Puccinia triticina) on glyphosate-resistant wheat and soybeans.[36] Control was equivalent to that of registered fungicides. Similarly, diquat was investigated for its potential to control the bacterial infection Columnaris disease (Flavobacterium columnare) affecting several fish species and was found to reduce Columnaris infection.[37] These findings demonstrate that herbicides can be directly toxic not only to plants but also to different types of life form at various trophic levels, including humans.

toxicity to Plants and effects to terrestrial Habitats Herbicides are especially designed to control weeds and therefore are of particular concern for unwanted effects on desirable plants. However, several types of pesticides can injure plants since pesticides are mostly classified by target organisms. For example, the widely used insecticides chlorpyrifos, diazinon, carbaryl, malathion and others can cause injury to several crops and ornamental species.[38] Several fungicides belonging to the benzimidazole chemical class were shown to be toxic to several plants.[39] Nevertheless, this entry will concentrate on herbicides. Effects of herbicides on crops and weeds are well documented for obvious agronomic and economic reasons. Unwanted effects on native species is usually assessed through regulatory processes using crops as surrogate species (see below) and via studies performed sporadically by research scientists. By and large, herbicides affect native species at different levels. Herbicides can alter biochemical and developmental processes in plants as well as plant morphology. Habitats within agroecosystems can experience modifications in their species abundance, composition, and diversity when subjected to herbicides. Plants as terrestrial primary producers constitute the basis of terrestrial ecosystems. Herbicide effects on them can have cascading effects at other trophic levels, on overall biodiversity, and on ecosystem functions. These will be addressed in turn.

effects at the Species Level Herbicides can modify biochemical processes in plants and as a result can increase plant susceptibility to pests and diseases. Most of the work demonstrating these effects has been performed with crop plants, but there is no reason to believe that native plants would not be affected in the same manner. The incidence of mildew on spring wheat was enhanced by three different herbicides as a result of stress induced by metabolic interference.[40] Wheat treated with 2,4-D was higher in protein content resulting in a proliferation of aphids.[41] The concentration of nitrogenous compounds in crops was enhanced by

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sublethal doses of several phenoxy, triazine, and uracil herbicides,[42] sometimes leading to an increase in pathogens and pests on corn[43] and rice.[44] Symbiotic processes with plants such as nodulation and mycorrhizae, which are vital for the biochemical activities of most terrestrial plants, can be greatly modified by herbicides. Nodulation and nitrogen fixation was disrupted by herbicides resulting in deleterious effects on growth and reproduction in crops such as dry bean (Phaseolus vulgaris), soybean (Glycine max), broad bean (Vicia faba), and peanut (Arachis hypogea) (Schnelle and Hensley[45] and references therein). Mycorrhizal activities were affected by MCPA sprayed on (Pisum sativum) with ensuing decreases in growth observed.[46] Herbicides can exert anomalous effects on plant developmental processes. They have been shown to affect seed production and seed germination. It was demonstrated as early as 1948 that applications of 2,4-D caused a delay in seed germination and growth of wheat plants sufficient to favor an increase of wireworm damage.[47] It has been known for some time that when some of these herbicides are applied to cereal crops late in their growth stage, just before seed formation, the plants produced far fewer seeds. Greenhouse experiments recently revealed that at typically used rates, dicamba and picloram reduced all or nearly all seed production while 2,4-D was much less effective.[48] Further field test experiments supported these greenhouse results.[49] Research with glyphosate showed that depending upon the plant species, application rate, and the timing of application, effects on seed production, seed germination, and seedling development have been observed on a large number of plant species from various families.[50] Glyphosate produced an inhibitory effect on pollen germination and seed formation when applied at the flower bud stage of goldenrod (S. canadensis).[51] Germination, emergence, and plant establishment of native Australian plant species were impeded by the herbicide fluazifop-p-butyl.[52] Plants of Stellaria media treated with the herbicide glufosinate ammonium produced seeds with reduced germination and emergence.[53] Herbicides can instigate unexpected effects on plants that have not been studied as part of the registration package because they are not relevant from an agronomical viewpoint. Sulfonylurea herbicides are selective herbicides that act by inhibiting the enzyme acetolactate synthase, which catalyzes the synthesis of the three branched-chain amino acids valine, leucine, and isoleucine. They are very potent herbicides applied at exceedingly low doses, in the order of a few grams per hectare. When applied at the fraction of the recommended label rate at the onset of flower formation, however, they were found to reduce the reproductive outputs of several species with significant reproductive damage occurring with only scant visible symptoms on leaves. Cherry trees sprayed at doses as low as 0.2% of the field application rate of the sulfonylurea chlorsulfuron showed a significant reduction in the production of fruits, with almost no observable damage to vegetative parts.[54] Low doses of metsulfuron methyl produced important injury on the vegetative biomass when crop and native species were sprayed at the seedling stage, but plants sprayed at later stages showed considerable reduction in the reproduction.[55] Potato bulking was greatly affected by low rates of sulfometuron methyl whereas few visible symptoms were observed on the vegetative parts.[56] Under field situations, it was found that berry production in hawthorn (Crataegus monogyna Jacq.) was severely affected by average spray drift concentrations higher than 2.5% of the label rate of the sulfonyl urea metsulfuron methyl (0.1 g-ai/ha) and that the effect was still observed 1 year after the spray event.[57,58] Imidazolinones are another chemical family classified as low-dose herbicides. They also inhibit the enzyme acetolactase synthase and as such have been implicated in unforeseen effects. Potato (Solanum tuberosum L. from the Solanaceae family) tuber size and quality were most detrimentally affected by exposure to imidazolinone herbicides, as low as 0.1 times the recommended field rate, when exposure occurred during tuber bulking, as compared to exposure at seedling emergence or at tuber initiation.[59] While reductions in the weight and the overall yield of sensitive species were problematic, another major concern was the potential lowering of fruit quality, which could lead to significant economic losses in the case of crops.[56,59] Effects on native species have not been investigated, but it is easy to

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stipulate that drift of imidazolinones to non-target habitats could elicit a reduction in weight of sensitive species, including effects on underground and reproductive parts. This may lead to species of the Solanaceae family and other families to produce fruits and storage organs that are less attractive to wildlife. Damage to or modification of plant morphology, including epinasty, inhibition of leaf expansion, and stem and root distortion, has been reported in weeds and crops, particularly for phenoxy herbicides.[60] Herbicides can cause morphological deformation with unforeseen repercussions. For instance, it was possible to demonstrate that morphological deformations of the flowering parts in Arabidopsis thaliana (L.) Heynh. induced by some herbicides prevented pollination and seed production.[61] Bud and flower abortion has been reported for sulfonylureas (see above). Such effects, usually undetected, can have long term adverse impacts, particularly on monocarpic species.

Population/community Level Herbicide use in modern agriculture alters species abundance, composition, and diversity in non-target habitats. Effects of low doses of herbicides on plants (grown in pots and placed at different distances from the spray swath) have been reported.[62,63] Marshall and Bernie[64] showed that several of the broadleaved species found in field margins were susceptible to the six different herbicides tested separately in pots. In the field, Marrs et al.[65] assessed effects of spray drift in relation to plant damage and yield for a range of plant species of conservation interest in Britain after applying each of six herbicides with a standard agricultural hydraulic ground sprayer. They observed lethal effects at 2–6 m from field edges and damage at greater distances, although damaged plants were able to recover within the growing season. Effects on seedlings were observed up to 20 m.[66] The long-term effect of such damage to plants remains unknown. In North America, work performed by Jobin et al.[67] showed that recurrent applications of herbicides had a long-term effect on plant populations inhabiting hedgerows and woodlot edges adjacent to crop fields. In Britain, several native arable weeds are considered endangered due to destruction of their habitats and extensive use of agrochemicals.[68] In the Netherlands, it was calculated that 9.5% of all pesticides applied was dispersed outside croplands.[18] Drift scenarios together with herbicide toxic effects investigated using bioassays and taking into account distances from spray events were used to estimate impact on biodiversity. In 2005, 41% of the linear landscape features near cropland were affected. This was an improvement since, in 1998, 59% of the area was affected. Natural areas located within farming regions were also affected by herbicide displacement in 31% and 11% of the area in 1998 and 2005, respectively. Measures in place such as unsprayed buffer zones and better equipment as well as reduced reliance on herbicides were largely responsible for this decline in unwanted effects on plant diversity. Also in the Netherlands, small plot experiments used to investigate the effects of the herbicide fluroxypyr on plants monitored for 3 years showed a decline in diversity and biomass.[69] Change in plant populations was also noticed.

ecosystem/trophic Level Alterations by herbicides on primary producers through effects on morphology, physiology, phenology, species composition, diversity, and abundance can resonate considerably to other trophic levels. The best documented study implicating herbicide repercussions was conducted in Britain over several decades. The grey partridge (Perdix perdix) has been surveyed since 1933 in the margins of crop fields, and it was found that numbers declined by 80% between 1952 and the mid-1980s.[70] Studies conducted from the 1960s led to the conclusion that the use of herbicides and, to a lesser extent, insecticides precipitated the decline of grey partridge populations. Although partridges are largely herbivorous, newly hatched chicks feed largely on arthropods during the first 2–3 weeks of their lives. The falling number of grey partridges in agricultural land was attributed to declining chick survival early in the season due to weed removal by herbicides. A reduction in weed diversity and density on which insects feed and

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inhabit caused a food shortage at this very crucial period of the year.[71] Removal of field margins was also a contributing factor. Many subsequent studies have shown that as a result of herbicidal effects, cover and diversity of flowering plant species are reduced in crop fields and field margins, thus subsequently reducing the resources available to flower- visiting insects and other arthropods.[72–74] Likewise, abundant floral diversity was found to be the prevailing factor related to high Lepidopteran diversity in farmland habitats (Boutin et al.[75] and references therein).

Factors interacting with Herbicide effects Toxicity to plants and other organisms may be exacerbated by the chemical and physical properties of a compound, namely, volatility, mobility, and persistence in the environment. For example, atrazine is a widely used product in North America applied for the pre- or postemergence control of broadleaf weeds and grasses, predominately on corn crops. It is a selective compound with systemic activity resulting in chlorosis and eventual plant death. The primary mode of degradation of atrazine is through hydrolysis. It has been reported to persist in soil, water, and aquifers for years.[76] In a study conducted in Germany, it was found that the herbicide atrazine was still present 22 years after its application in soil.[77] It is regularly found in groundwater in the United States and Canada where it is widely used. Unquestionably, the persistence of atrazine explains its presence and accumulation in the environment and long-term effects.[78] Most herbicides are highly water soluble and therefore can move away from crop fields with rainwater. Volatile compounds can elicit damage to non-target plants; clomazone applied in crop fields produced striking bleaching effects that caused the U.S. regulatory agencies to consider wild plants in their risk assessments.[79] Persistent herbicides may cause long-lasting and unexpected contamination. Clopyralid is a growthregulator-type herbicide used for the control of broad-leaved species. It was used on sugar beet that was later fed to cattle with the resulting manure containing sufficient residues to contaminate crops on which it was subsequently spread.[80] The persistence of clopyralid in compost and mulches was also demonstrated in the United States.[81] The same problem occurred with the related compound aminopyralid, an auxinic herbicide registered for the control of broad-leaved weeds on grassland and rangeland. Grass treated with aminopyralid persisted in the silage for more than 1 year. Cattle or horses fed with the hay produced contaminated manure toxic to receiving crops.[82] Herbicide toxic effects may vary depending on the characteristics of the receiving environment. Residues of the sulfonylurea chlorsulfuron have been detected and caused crop damage as long as 7 years after application in Alberta, Canada.[83] Sulfonylureas are known for their increased activity at soil pH above neutrality, which is often observed in soils in western Canada.[84] Temperature may also exert effects on persistence and thus on long-term toxicity. In Canada and elsewhere where winters are long and cold, many pesticides, including chlorsulfuron, will take longer to degrade.[85] Direct toxicity of herbicides to primary producers and other organisms is typically assessed without taking into account confounding stressors present under normal agriculture practices or occurring in natural conditions. For instance, high fertilizer application in intensive agriculture is the norm. Few studies on the misplacement of fertilizer in non-target adjacent areas have been published, but Rew  et  al.[86] in Britain demonstrated that off-target deposition could vary from 2% to 133.3% of the application rate in field margins, depending on the type of machinery used. Some species will thrive under high nutrient conditions, including grasses.[69] Field margins where intensive agriculture prevails tend to become simplified and dominated by grasses because both herbicides and fertilizers tend to suppress broad-leaved species.[21] Grasses are wind pollinated and, as a consequence, pollinators are reduced or eliminated. A plethora of unforeseen events and interactions can occur when herbicides are released in the environment. Measuring phytotoxicity is a first step for assessing unwanted effects in the environment.

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Measuring Phytotoxicity Herbicides (and other pesticides) are special contaminants because they are intentionally released in the environment. Therefore, before they are registered for use, risk assessment of herbicide phytotoxicity is conducted using data submitted by registrants. In current guidelines, phytotoxicity tests designed to assess the impact to terrestrial plants are conducted under controlled greenhouse conditions, using crop plants growing individually in pots, and effects are assessed at the juvenile stage, usually before reproduction occurs.[87,88] These tests are used for risk assessment of native plants growing within communities exposed to variable outdoor conditions at various phenological stages (Figure 2). Therefore, there is a large gap in regulatory testing between actual phytotoxicity tests performed with selected nonnative species under artificial conditions and risk assessment conducted for protection of native plants in natural habitats. Ecological risk assessment for effects on native plants usually takes into account both herbicide use patterns and herbicide characteristics. Elements that are typically not considered are biotic and abiotic factors, which will affect plant sensitivity to herbicides. Current phytotoxicity testing assumes similar effects at different phenological stages and does not take into account recurrent exposure of sublethal doses. In the past, phytotoxicity testing was designed for the protection of agronomically relevant species from accidental herbicide drift; therefore, crop species were appropriately used in testing. However,

FIGURE 2 Schematic representation of measured and observed herbicide phytotoxicity testing. Phytotoxicity is measured following routine test guidelines. Subsequent ecological risk assessment for effects on native plants usually takes into account herbicide use pattern such as how (aerial or ground applications), when (time of the year and day), and where (type of terrains) herbicides are applied, which will determine to some extent exposure through drift, runoff, or overspray. Herbicide persistence, mobility, and volatility will increase exposure. Elements represented in dashed boxes are usually not considered and determined the observed phytotoxicity under natural conditions. Biotic factors include competition or interactions with other plant species, effects of diseases, and predation on plant health. Wind, low or high temperature, drought or flooding, and soil properties may affect plants and modify sensitivity to herbicides. Phytotoxicity testing assumes similar effects at different phenological stages and does not take into account recurrent exposure of sublethal doses.

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it has been increasingly recognized that the vegetation bordering or in the vicinity of crop fields was important for the native plant species and for the other trophic levels relying on primary producers for food, habitat, and shelter. In studies aimed at comparing crop and noncrop species, results indicated that plant sensitivity is both herbicide and species dependent and that no obvious pattern emerged with numerous native and crop plant species tested with a number of herbicides.[89,90] For some herbicides, results with crops would mean underprotection of native species, while for other herbicides, testing with crop species would entail overprotection. However, these studies were limited to a narrow taxonomic range in line with the crop species usually tested, e.g., short-lived species. It was also found that the numerous native species selected for the studies were, for the most part, easy to grow and maintain in the greenhouse. These results suggest that testing should cover a broader range of native species in toxicity testing for regulatory purposes. Further results indicated that a large variability in herbicide response existed among crop varieties and native plant ecotypes.[89,91,92] The number of species needed to be tested to encompass the range of sensitivity of a given herbicide has not been resolved. Boutin and Rogers[93] showed that the range of species sensitivity increases with an augmentation of numbers of species tested up to 40, suggesting that the number of species (n = 6 or 10) tested in current guidelines is insufficient. These studies and others indicated that the large emphasis placed on collecting very precise EC25 or EC50 values (effective concentration or dose causing a 25% or 50% effect on test plants) may be problematic and erroneous and that extrapolation factors should be determined in order to alleviate some of the uncertainty in testing that is currently not considered. Tests in current guidelines are required for plants at the two- to four-leaf stage, the assumption being that the juvenile stage is more sensitive than the adult period. However, natural vegetation bordering crop fields consists of plants at various phenological stages. It was found in several studies that indeed the young stage is very sensitive to herbicides.[66,94] Conversely, plants at later stages can show higher sensitivity to some herbicides.[54,94] For practical reasons, phytotoxicity tests are conducted under ideal greenhouse conditions using plants growing singly in pots and devoid of intra- or interspecific competition. This contrasts with plants growing within natural communities subjected to the occasional drought or flood, continuous wind, attacks from herbivores or pathogens, and other abiotic or biotic elements that interfere with herbicide effects. It is of course impossible and impractical to study the interactions of all these factors. In the past, studies comparing greenhouse and field tests elicited contradictory results with species being more or less sensitive when tested in greenhouses as compared to field tests.[53,69,95–97] A few studies have examined the direct effect of herbicide use of communities of native plants[21,64–67] (Figure 1), and although responses were extremely variable, most species showed some effects. There is considerable uncertainty in current phytotoxicity testing, and further studies are required to monitor herbicide effects on terrestrial native plants.[98]

conclusion Herbicides are used extensively, mostly in agriculture, with sometimes unforeseen environmental effects in non-target areas. Although measuring phytotoxicity can be straightforward when conducted on known crop plants under simple laboratory conditions (in greenhouses), determining phytotoxicity is a multifaceted affair as demonstrated by the numerous factors interacting under natural conditions (Figure 2). The linkage between measured phytotoxicity in greenhouses and observed phytotoxicity in nature implies the inclusion of environmental biotic and abiotic factors as well as considerations of test conditions, herbicide characteristics, level of exposure, and use pattern (Figure 2). It was also shown that phytotoxicity can have consequences at all trophic levels, and this is seldom taken into account. Ecological and agronomic needs are often conflicting. Weed plants want to be suppressed in favor of crops, yet they are necessary to sustain some wildlife in otherwise desolate agrarian systems. This wildlife can in turn provide vital ecological services (e.g., pollinators for crops or biological control).

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Mitigation measures to reduce phytotoxicity on desirable plants, including the elimination of aerial application, implementation of unsprayed buffer zones, more advanced equipment, and avoidance of adverse spray conditions (e.g., high wind, temperature inversion, or forecast rain), can substantially reduce herbicide impacts on the environment and are increasingly considered by the farming communities and regulators. Nevertheless, there appears to be an excessive dependence on herbicides for the control of weeds. New uses for herbicides are still being developed, including the desiccation of crops at the end of the growing season with foliar contact herbicides to facilitate harvest or the use of herbicide-resistant, genetically modified crops. Alternatives to herbicides exist and are widely used in organic farming. They comprise mechanical weed control using tillage or mowing, non-mechanical weed control through crop rotation or the planting of companion crops, and biological weed control using parasites, predators, or pathogens. Another complementary measure that can be promoted in conventional farming involves Integrated Pest Management (IPM), a management system that uses all available tactics to manage pest and weeds including crop rotation, host-resistant varieties, mechanical and physical controls, and chemical control. Weed control in agriculture should and can be made compatible with wildlife values.

Acknowledgments The author wishes to thank D. Carpenter for useful comments on the manuscript and B. Daigneault for assistance in formatting. Funding was provided by Environment Canada (including Pesticide Science Fund) for the experimental work presented by the author and collaborators.

References 1. Zimdahl, R.L. Fundamentals of Weed Science, 3rd Ed.; Academic Press: San Diego, CA, 2008. 2. Matsumoto, M.; Hirose, A.; Ema, M. Review of testicular toxicity of dinitrophenolic compounds, 2-sec-butyl-4, 6-dinitrophenol, 4,6-dinitro-o-cresol and 2,4-dinitrophenol. Reprod. Toxicol. 2008, 26 (3–4),185–190. 3. Duke, S.O. Overview of herbicide mechanisms of action. Environ. Health Perspect. 1990, 87, 263–271. 4. Walsh, M.J.; Maguire, N.; Powles, S.B. Combined effects of wheat competition and 2,4-D amine on phenoxy herbicide resistant Raphanus raphanistrum populations. Weed Res. 2009, 49, 316–325. 5. Anonymous. Common and chemical names of herbicides approved by the Weed Science Society of America. Weed Sci. 2010, 58 (4), 511–518. 6. Brimble, S.; Bacchus, P.; Caux, P.-Y. Pesticide utilization in Canada: A compilation of current sales and use data, GoC Cat. No. En4–56/2005E; Environment Canada: Gatineau, QC, 2005. 7. Eberwine, J.W.; Hagood, E.S. Effect of johnsongrass (Sorghum halepense) control on the severity of virus diseases of corn (Zea mays). Weed Technol. 1995, 9, 73–79. 8. United States Environmental Protection Agency, 2001, available at http://www.epa.gov/pesticides/ pestsales/ (accessed May 2011). 9. IRENA (Indicators Reporting on the Integration of Environmental Concerns into Agricultural Policy). Fact Sheet no. 9. Consumption of Pesticides, 1999. 10. Eurostat. Statistical books. The use of plant protection products in the European Union. Data 1992–2003. Statistical books. 2007 edition; 222 pp. 11. Wilson, C.; Tisdell, C. Why farmers continue to use pesticides despite environmental, health and sustainability costs. Economics, Ecology and the Environment, Working Paper no. 53; School of Economics, University of Queensland: Brisbane, Australia, 2000. 12. Statistics Canada. Census of Agriculture. Ministry of Supply and Services Canada, 1987. 13. National Research Council. Alternative Agriculture National Research Council: Washington, DC, 1989; 448 pp.

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35. Kelly, D.W.; Poulin, R.; Tompkins, D.M.; Townsend, C.R. Synergistic effects of glyphosate formulation and parasite infection on fish malformations and survival. J. Appl. Ecol. 2010, 47 (2), 498–504. 36. Feng, P.C.C.; Clark, C.; Andrade, G.C.; Balbi, M.C.; Caldwell, P. The control of Asian rust by glyphosate in glyphosate-resistant soybeans. Pest Manage. Sci. 2008, 64 (4), 353–359. 37. Darwish, A.M.; Mitchell, A.J. Evaluation of diquat against an acute experimental infection of Flavobacterium columnare in channel catfish, Ictaluruspunctatus (Rafinesque). J. Fish Dis. 2009, 32 (5), 401–408. 38. Thomson, W.T. Agricultural Chemicals. Book 1 Insecticides; Thomson Publications: Fresno, CA, 1985–1986. 39. van Lersel, M.W.; Bugbee, B. Phytotoxic effects of benzimidazole fungicides on bedding plants. J. Am. Soc. Hortic. Sci. 1996, 121 (6), 1095–1102. 40. Heitefuss, P. Assessment of herbicide effects on interactions of weeds, crop plants, pathogens and pests. In Field Methods for the Study of Environmental Effects of Pesticides, BCPC Monograph No. 40; Greaves, M.P., Greig-Smith, P.W., Smith, B.D., Eds.; British Crop Protection Council: Thornton Heath, U.K., 1988; 265–274. 41. Adams, J.B.; Drew, M.E. Grain aphids in New Brunswick. III. Aphid populations in herbicidetreated oat fields. Can. J. Zool. 1965, 43, 789–794. 42. Ries, S.K. Subtoxic effects on plants. In Herbicides; Audus, L.J., Ed.; Academic Press: London, 1976; Vol. 2, 313–344. 43. Oka, I.N.; Pimentel, D. Herbicide (2,4-D) increases insect and pathogen pests on corn. Science 1976, 193 (4249), 239–240. 44. Ishii, S.; Hirano, C. Growth responses of larvae of the rice stem borer to rice plants treated with 2,4-D. Entomol. Exp. Appl. 1963, 6, 257–262. 45. Schnelle, M.A.; Hensley, D.L. Effects of pesticides upon nitrogen fixation and nodulation by dry bean. Pestic. Sci. 1990, 28 (1), 83–88. 46. Garcia-Romera, I.; Ocampo, J.A. Effect of the herbicide MCPA on VA mycorrhizal infection and growth of Pisum sativum. J. Plant Nutr. Soil Sci. 1988, 151 (4), 225–278. 47. Fox, W.B. 2,4-D as a factor in increasing wireworm damage to wheat. Sci. Agric. 1948, 28 (9), 423–424. 48. Rinella, M.J.; Haferkamp, M.R.; Masters, R.A.; Muscha, J.M.; Bellows, S.E.; Vermeire, L.T. Growth regulator herbicides prevent invasive annual grass seed production. Invasive Plant Sci. Manage. 2010, 3 (1), 12–16. 49. Rinella, M.J.; Masters, R.A.; Bellows, S.E. Growth regulator herbicides prevent invasive annual grass seed production under field conditions. Rangeland Ecol. Manage. 2010, 63 (4), 487–490. 50. Blackburn, L.G.; Boutin, C. Subtle effects of herbicide use in the context of genetically modified crops: A case study with glyphosate (Roundup®). Ecotoxicology 2003, 12, 271–285. 51. Guo, S.L.; Jiang, H.W.; Fang, F.; Chen, G.Q. Influences of herbicides, uprooting and use as cut flowers on sexual reproduction of Solidago canadensis. Weed Res. 2009, 49 (3), 291–299. 52. Rokich, D.P.; Harma, J.; Turner, S.R.; Sadler, R.J.; Tan, B.H.; Fluazifop-p-butyl herbicide: Implications for germination, emergence and growth of Australian plant species. Biol. Conserv. 2009, 142 (4), 850–869. 53. Riemens, M.M.; Dueck, T.; Kempenaar, C. Predicting sublethal effects of herbicides on terrestrial non-crop plant species in the field from greenhouse data. Environ. Pollut. 2008, 155 (1), 141–149. 54. Fletcher, S.J.; Pfleeger, T.G.; Hillman, C.R. Potential environmental risks associated with the new sulfonylurea herbicides. Environ. Sci. Technol. 1993, 27 (10), 2250–2252. 55. Boutin, C.; Elmegaard, N.; Kjaer, C. Toxicity testing of fifteen non-crop plant species with six herbicides in a greenhouse experiment: Implications for risk assessment. Ecotoxicology 2004, 13 (4), 349–369.

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56. Pfleeger, T.; Olszyk, D.; Plocher, M.; Yilma, S. Effects of low concentrations of herbicides on full-season, field-grown potatoes. J. Environ. Qual. 2008, 37, 2070–2082. 57. Kjær, C.; Strandberg, M.; Erlandsen, M. Metsulfuron spray drift reduces fruit yield of hawthorn (Crataegus monogyna L.). Sci. Total Environ. 2006, 356, 228–234. 58. Kjær, C.; Strandberg, M.; Erlandsen, M. Effects on hawthorn the year after simulated spray drift. Chemosphere 2006, 63 (5), 853–859. 59. Eberlein, C.V.; Guttieri, M.J. Potato (Solanum tuberosum) response to simulated drift of imidazolinone herbicides. Weed Sci. 1994, 42 (1), 70–75. 60. Tottman, D.R.; Davies, E.L.P. The effect of herbicides on the root system of wheat plants. Ann. Appl. Biol. 1978, 90 (1), 93–99. 61. Ratsch, H.C.; Johndro, D.J.; McFarlane, J.C. Growth inhibition and morphological effects of several chemicals in Arabidopsis thaliana (L.) Heynh. Environ. Toxicol. Chem. 1986, 5 (1), 55–60. 62. Breeze, V.G.; van Rensburg, E. Vapour of the free acid of the herbicide 2,4-D is toxic to tomato and lettuce plants. Environ. Pollut. 1991, 72 (4), 259–267. 63. Marrs, R.H.; Frost, A.J. A microcosm approach to the detection of the effects of herbicide spray drift in plant communities. J. Environ. Manage. 1997, 50 (4), 369–388. 64. Marshall, E.J.P.; Bernie, J.E. Herbicide effects on field margin flora. Proceedings of the BCPC International Congress—Weeds; British Crop Protection Council: Farnham, Surrey, U.K., 1985; 1021–1028. 65. Marrs, R.H.; Williams, C.T.; Frost, A.J.; Plant, R.A. Assessment of the effects of herbicide spray drift on a range of plant species of conservation interest. Environ. Pollut. 1989, 59 (1), 71–86. 66. Marrs, R.H.; Frost, A.J.; Plant, R.A. Effects of herbicide spray drift on selected species of nature conservation interest: The effects of plant size and surrounding vegetation structure. Environ. Pollut. 1991, 69 (2–3), 223–235. 67. Jobin, B.; Boutin, C.; DesGranges, J.-L. Effects of agricultural practices on the flora of hedgerows and woodland edges in Southern Québec. Can. J. Plant Sci. 1997, 77, 293–299. 68. Marshall, E.J.P. The impact of landscape structure and sown grass margin strips on weed assemblages in arable crops and their boundaries. Weed Res. 2009, 49 (1), 107–115. 69. Kleijn, D.; Verbeek, M. Factors affecting the species composition of arable field boundary vegetation. J. Appl. Ecol. 1997, 37 (2), 256–266. 70. Sotherton, N.W.; Dower, J.W.; Rands, N.R.W. The effects of pesticide exclusion strips on faunal populations in Great Britain. Ecol. Bull. 1988, 39, 197–199. 71. Potts, G.R. The effects of modern agriculture, nest predation and game management on the population ecology of partridges (Perdix perdix and Alectoris rufa). Adv. Ecol. Res. 1980, 11, 2–79. 72. Lagerlöf, J.; Stark, J.; Svensson, B. Margins of agricultural fields as habitats for pollinating insects. Agric., Ecosyst. Environ. 1992, 40, 117–124. 73. Longley, M.; Sotherton, N.W. Factors determining the effects of pesticides upon butterflies inhabiting arable farmland. Agric., Ecosyst. Environ. 1997, 61 (1), 1–12. 74. Holzschuh, A.; Steffan-Dewenter, I.; Kleijn, D.; Tscharntke, T. Diversity of flower-visiting bees in cereal fields: effects of farming system, landscape composition and regional context. J. Appl. Ecol. 2007, 44 (1), 41–49. 75. Boutin, C.; Baril, A.; McCabe, S.K.; Martin, P.A.; Guy, M. The value of woody hedgerows for moth diversity on organic and conventional farms. Environ. Entomol. 2011, 40 (3): 560–569. 76. Lazorko-Connon, S.; Achari, G. Atrazine: Its occurrence and treatment in water. Environ. Rev. 2009, 17, 199–214. 77. Jablonowski, N.D.; Köppchen, S.; Hofmann, D.; Schäffer, A.; Burauel, P. Persistence of 14C-labeled atrazine and its residues in a field lysimeter soil after 22 years. Environ. Pollut. 2009, 157 (7), 2126–2131. 78. Graymore, M.; Stagnitti, F.; Allinson, G. Impacts of atrazine in aquatic ecosystems. Environ. Int. 2001, 26, 483–495.

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79. Poster, J. Command herbicide. The rookie battles controversy. Crops Soils Mag. 1986, 39 (1), 9–11. 80. Eagle, D.J. Agrochemical damage to U.K. crops. Pestic. Outlook 1990, 1 (2), 14–16. 81. Nilsson, H.; Aamisepp, A. Persistence in plants and transfer of clopyralid (3,6-dichloropicolinic acid) through plant remains. In Weeds and Weed Control, 25th Swedish Weed Conference, Vol. 1 Reports. Uppsala, Sweden, 1984. 82. http://www.pesticides.gov.uk/garden.asp?id=2480 and http://www.guardian.co.uk/environment/ 2008/jun/29/food.agriculture (accessed May 2011). 83. Moyer, J.R.; Esau, R.; Kozub, G.C. Chlorsulfuron persistence and response of nine rotational crops in alkaline soils of southern Alberta. Weed Technol. 1990, 4 (3), 543–548. 84. Blair, A.M.; Martin, T.D. A review of the activity, fate and mode of action of sulfonylurea herbicides. Pestic. Sci. 1988, 22 (3), 195–219. 85. Walker, A. Simulation of herbicide persistence in soil. II. Simazine and linuron in long-term experiments. Pestic. Sci. 1976, 7 (1), 50–58. 86. Rew, L.J.; Theaker, A.J.; Froud-Williams, R.J.; Boatman, N.D. Nitrogen fertilizer misplacement in field boundaries. Aspects Appl. Biol. 1992, 30, 203–206. 87. Organisation for Economic Co-operation and Development (OECD). Terrestrial Plant Test: Vegetative Vigour Test. New TG 227. Seedling Emergence and Seedling Growth Test. Updated TG 208, 2006. 88. United States Environmental Protection Agency (USEPA). Ecological Effects Test Guidelines: Terrestrial Plant Toxicity—Vegetative Vigor, OPPTS 850.4250, EPA 712-C-96- 163 and Early Seedling Toxicity Test, OPPTS 850.4130, EPA 712–C–96–347. Washington, DC, 1996. 89. White, A.L.; Boutin, C. Herbicidal effects on non-target vegetation: Investigating the limitations of current pesticide registration guidelines. Environ. Toxicol. Chem. 2007, 26 (12), 2634–2643. 90. Carpenter, D.; Boutin, C. Sublethal effects of the herbicide glufosinate ammonium on crops and wild plants: Short-term effects compared to vegetative recovery and plant reproduction. Ecotoxicology 2010, 19 (7), 1322–1336. 91. Boutin, C.; White, A.L.; Carpenter, D. Measuring variability in phytotoxicity testing using crop and wild plant species. Environ. Toxicol. Chem. 2010, 29 (2), 237–242. 92. Bhatti, M.A.; Felsot, A.S.; Al-Khatib, K.; Kadir, S.; Parker, R. Effects of simulated chlorsulfuron drift on fruit yield and quality of sweet cherries (Prunus avium L.). Environ. Toxicol. Chem. 1995, 14 (3), 537–544. 93. Boutin, C.; Rogers, C.A. Pattern of sensitivity of plant species to various herbicides—An analysis with two databases. Ecotoxicology 2000, 9 (4), 255–272. 94. Boutin, C.; Lee, H.B.; Peart, T.E.; Batchelor, S.P.; Maguire, R.J. Effects of the sulfonylurea herbicide metsulfuron methyl on growth and reproduction of five wetland and terrestrial plant species. Environ. Toxicol. Chem. 2000, 19 (10), 2532–2541. 95. Dalton, R.L.; Boutin, C. Comparison of the effects of glyphosate and atrazine herbicides on nontarget plants grown singly and in microcosms. Environ. Toxicol. Chem. 2010, 29 (10), 2304–2315. 96. Damgaard, C.; Mathiassen, S.K.; Kudsk, P. Modeling effects of herbicide drift on the competitive interactions between weeds. Environ. Toxicol. Chem. 2008, 27 (6), 1302–1308. 97. Clark, J.; Ortego, L.S.; Fairbrother, A. Sources of variability in plant toxicity testing. Chemosphere 2004, 57 (11), 1599–1612. 98. Boutin, C.; Aya, K.L.; Carpenter, D.; Thomas, P.J.; Rowland, O. Phytotoxicity testing for herbicide regulation: shortcomings in relation to biodiversity and ecosystem services in agrarian systems. Sci. Total Environ. 2011, 40 (3): 560–569.

5 Insecticides: Aerial UltraLow-Volume Application

He Zhong

Introduction ....................................................................................................49 Right Place .......................................................................................................49 Right Time ....................................................................................................... 50 Right Dose ....................................................................................................... 50 Novel Application Technology ..................................................................... 51 Insecticide Residue Monitoring.................................................................... 51 Conclusion ...................................................................................................... 51 Acknowledgments .......................................................................................... 52 References ........................................................................................................ 52

introduction Mosquito control is necessary in order to protect public health from mosquito-borne diseases such as West Nile virus, eastern equine encephalitis, St. Louis encephalitis, malaria, and dengue. Aerial ultralow-volume (ULV) application of mosquito insecticides is one of the most effective techniques for controlling adult mosquitoes and preventing mosquito-borne diseases.[1] During ULV application, large insecticide droplets may sometimes result in unwanted mortality to nontarget organisms.[2–8] For many years, nontarget mortality caused by mosquito adulticiding was usually accepted as a “casualty of war.” With the advancement of mosquitocide residue monitoring and new spray technologies, we now recognize that “casualties” can be reduced to a minimum. The important factors that contribute to mosquito control efficacy and associated nontarget mortality are insecticide deposition,[7–9] droplet size,[1,10] spray time,[10] application dose,[1,10] topography,[10] and weather conditions (such as wind velocity, direction, and atmospheric stability).[10] Due to the complexity of aerial ULV applications, it may be difficult to achieve optimal mosquito reduction without nontarget mortality during an aerial spray mission. However, control efficacy can be increased, and nontarget mortality can be minimized, if aerial application is conducted at the right place (by increasing retention time of mosquitocide droplets in the air in order to enhance their contact with flying mosquitoes),[11,12] at the right time (dusk, or night when adult mosquitoes are actively flying),[10] and at the right dose (proper insecticide concentration in the air to kill mosquitoes but not nontarget organisms).[1,9]

Right Place Current aerial spray technology is capable of targeting a specific zone where adult mosquitoes are actively flying. The droplet size of aerially applied insecticides governs downwind dispersal and subsequent impingement on targets.[1] Smaller droplets [5–25 pm volume median diameter (VMD)] remain aloft longer and offer a better probability for impingement on flying adult mosquitoes.[1] Larger insecticide 49

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droplets (more than 100 pm VMD) will deposit on the ground sooner after application, thereby reducing the likelihood of contact with flying mosquitoes. Insecticide that is deposited on the ground not only is wasted but also may adversely affect nontarget organisms.[7,8] Environmental contamination can be reduced by adopting application techniques that maintain droplets in the air and promote controlled downwind movement of the insecticide cloud while minimizing ground deposition (particularly in environmentally sensitive areas).[7,9] This concept is different from agricultural applications, where insecticide deposition is needed to coat the surface of crops. Moreover, agricultural applications try to reduce insecticide drift away from the target zone, such as a crop or field, rather than maximize droplet suspension in the air column.

Right time The best time of the day for mosquito adulticide applications (also called the “spray window”) is at dusk, dawn, or nighttime when mosquitoes are actively flying, and this is when the optimal control efficacy will be achieved.[1,10] Applying an insecticide in that period will reduce impact on daytime active nontarget organisms like honeybees, dragonflies, and butterflies. These spray windows are the time periods when daytime nontargets are resting and, therefore, protected from insecticide exposure. Mosquito control operation in such spray windows will protect many daytime active nontargets from being adversely impacted by insecticide dispersal clouds. Spraying at the right time also means spraying under optimal meteorological conditions, currently recognized as 3–10 miles/hr wind velocity without the presence of a temperature inversion. Understanding, as well as achieving, “optimal” meteorological conditions for mosquito spraying is often difficult.[1] Although aerial mosquito application technology can accurately calibrate the amount of insecticide delivered using the nozzle systems. Once released, the aerosol is in the hands of Mother Nature. The spray cloud, as it is carried by wind and influenced by gravity, starts its journey to the ground from an altitude of 100300 ft. Wind velocity, direction, and atmospheric stability greatly affect the distribution of the spray cloud.[1] Also, downwind movement and deposition of insecticide residue can vary greatly from one spray mission to another.[11,12] As wind speed increases, the impingement force of spray droplets will increase on targets and nontargets. This situation creates considerable variation in control efficacy and can often influence whether the effects on nontargets are minimal or substantial.[13]

Right Dose Increasing the application rate may not always improve the level of control but will generally increase the risk of nontarget mortality due to escalating exposure levels. In reality, it is sometimes very difficult to apply the proper application rate to achieve adequate mosquito control without causing nontarget mortality. Nontargets’ differential tolerance to insecticides may be the result of physiological as well as geographical differences within and among those organisms. The larger body size of some nontarget species may increase their tolerance levels to the insecticide. Natural topographic barriers, such as trees, bushes, and grasses, may provide refuge for them to escape lethal exposure from the insecticide aerosol.[8] Conversely, adult mosquitoes in vegetated areas may also be protected by the physical barriers,[1] and if the application rate is increased to compensate for this, adverse effects to nontargets may occur. To determine the proper application rate or “right dose,” studies are needed to determine the correlation of the insecticide concentration in the air column with adult mosquito mortality[11,12] and that of ground deposition concentrations with nontarget mortality.[7–9] The insecticide concentration at its final destination, whether at an airborne target or on the ground, is referred to as the terminal insecticide concentration (TIC), which is different from and can be influenced by the initial application dose. TIC is influenced by many environmental variables and therefore needs to be frequently monitored. If the TIC is adequate to kill the majority of adult mosquitoes and low enough that it spares nontargets of concern, the application dose will be considered appropriate. In this way, TIC critically affects control efficacy

Aerial Ultra-Low-Volume Application

51

and nontarget impact. Determination of TIC during routine application of mosquitocides provides a mechanism to assess or cross-compare control efficacy and impact on nontargets during aerial mosquito control missions. This process will ensure the proper application dose to achieve the delicate balance between effective mosquito control and minimal nontarget impact.

novel Application technology Mosquito control programs worldwide continue to develop novel application technologies to increase control efficacy and lessen damage to nontargets. In the late 1990s, James Robinson’s group at Florida’s Pasco County Mosquito Control District led an effort to develop a high-pressure nozzle system to deliver small insecticide droplets (80 μm VMD)[11] and dramatically reduced the mortality of caged fiddler crabs, Uca pugilator (Bosc), from 80% to 0 in Collier County, Florida.[7] Another field trial with honeybees (Apis  mellifera L.) conducted during routine nighttime aerial adult mosquito control missions in Manatee County, Florida, further demonstrated the advantage of the new high-pressure system. Honeybees that clustered outside of beehive entrances were exposed to naled sprays for mosquito control. The larger insecticide droplets produced by flat-fan nozzles killed more than 90% of the bees clustered outside of the hives and resulted in an average of 35% reduction in honey yield at the end of the season.[8] In contrast, bee mortality and average honey yield were not significantly different compared with hives similarly exposed to the smaller droplets of the high-pressure nozzle system.[9] Several spray nozzles such as air-assisted nozzles, high-pressure nozzles, and Micronair nozzles— capable of delivering smaller spray droplets (10,000 mg/kg (metal dependent)

Industrial activities, smelting

50 mg/kg

Manufacturing, processing

5). Both groups of compounds are very persistent to decomposition processes, which explains why they are strong bioaccumulators, although dioxins have a UV–VIS absorption spectrum that results in significant absorption from solar radiation. Some dioxins have a half-life time in the troposphere of a few days.

Pesticides Pesticides are used to remove pests and, due to their direct use in nature, have probably been the most criticized environmental contaminants. Usage of dichlorodiphenyltrichloroethane (DDT) and relatedinsecticides accelerated during the 1940s and the subsequent decades until environmental doubt occurred in the mid-1960s. Since 1970, DDT has been banned in most industrialized countries, but it is still used in developing countries, for instance, India, where it has resulted in very high body concentrations in the Indian population. All chlorinated hydrocarbon insecticides are banned in most industrialized countries due to their persistence and ability to bioaccumulate (Kow is high). Pesticides can be divided into the following classes depending on their use and their chemical structure: • Herbicides comprise carbamates, phenoxyacetic acids, triazines, and phenylureas. • Insecticides encompass organophosphates, carbamates, organochlorines, pyrethrins, and pyrethroids. • Fungicides are dithiocarbamates, copper, and mercury compounds. See also the entries covering these two metals.

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FIGURE 3

Managing Global Resources and Universal Processes

Molecular structure and names of five common pesticides.

The pesticides are an extremely chemically diverse group of substances, as they only have in common their toxicity to pests. A few of the most important molecules are shown in Figure 3. They are mostly produced synthetically although the natural pesticide pyrethrin has achieved commercial success. Chlorohydrocarbons are strongly bioconcentrated, as already emphasized. In addition, they are very toxic to a wide range of biota, particularly to aquatic biota. Organophosphates are almost equally toxic to biota, but due to these compounds’ lack of persistence, higher solubility in water, and bioaccumulation capacity, they are still in use. Carbamates are relatively water soluble and have limited persistence. They are, however, toxic to a wide range of biota. They act by inhibiting cholinesterase. The pyrethins have a complex chemical structure and a high molecular weight. Thus, they are poorly soluble in water and tend to be lipophilic. They are, however, readily degraded by hydrolysis. They are more attractive to use than most of the other pesticides due to their very low mammalian toxicity. Phenoxyacetic acid is a very effective herbicide but contains trace amounts of tetrachloro-dibenzo-dioxin. Pesticides are banned in organic agriculture where they are replaced by other methods, for instance, mechanical and biological methods (use of predator insects).

Polycyclic Aromatic Hydrocarbons PAHs are molecules containing two or more fused 6C aromatic rings. They are ubiquitous contaminants of the natural environment, but growing industrialization has increased environmental concern about these components. Two common members are naphthalene and benzo(a)pyrene (see Figure 4). PAHs are usually solids, with naphthalene (lowest molecular weight) having a melting point of 81°C.

Toxic Substances

FIGURE 4

143

Molecular structure and names of five common PAHs.

The natural sources of PAHs in the environment are forest fires and volcanic activity. The anthropogenic sources are coal-fired power plants, incinerators, open burning and motor vehicle exhaust. As a result of these sources, PAHs commonly occur in air, soil, and biota. They are lipophilic compounds able to bioaccumulate. The low-molecular- weight compounds are moderately persistent, while for example benzo(a)pyrene, with a higher molecular weight, persists in aquatic systems for up to about 300 weeks. They are relatively toxic to aquatic organisms and have LC50 values for fish in the range of 0.1–10 mg/L. The major environmental concern of PAHs is that many PAHs are carcinogenic. It is has been shown[5] that benzo(a)pyrene is an endocrine disrupter and that many more PAHs have the environmental adverse effect of disturbing the hormone balance of nature. Human exposure to PAHs occurs through tobacco smoking as well as through compounds in food and the atmosphere.

organometallic compounds Organometallic compounds are compounds having metal carbon bonds, where the carbon atoms are part of an organic group. The best known example is probably tetraethyl lead, which is used as an additive to gasoline. It has now been phased out of use in many countries—all industrialized countries—due to its

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environmental consequences. Organometallic compounds can be formed in nature from metal or metal ions, for instance, di- methylmercury, or are produced for various purposes, as catalysts (e.g., organoaluminum), as pesticides (e.g., organoarsenic and organotin compounds), as stabilizers in polymers (e.g., organotin compounds), and as gasoline additives (e.g., organolead compounds). Organometallic compounds exhibit properties that are different from those of the metal itself and inorganic derivatives of the metal, for instance, a relatively higher toxicity than the metals. Most organometallic compounds are relatively unstable and undergo hydrolysis and photolysis easily. Most organometallic compounds have weakly polar carbon–metal bonds and are often hydrophobic. They therefore only dissolve in water to a small extent and are readily sorbed onto particulates and sediments. The most harmful organometallic compounds from an environmental point of view are organomercury, organotin, organolead, and organoarsenic, which are all very toxic to mammals.

Detergents (and Soaps) Detergents and soaps contain surface active agents (surfactants) that are classified according to the charged nature of the hydrophilic part of the molecule: • • • •

Anionic: negatively charged Cationic: positively charged Nonionic: neutral, but polar Amphoteric: a zwitterion containing positive and negative charges

They are produced and consumed in large quantities and are mostly discharged into the sewage system and end up in wastewater plants. The early surfactants contained highly branched alkyl hydrophobes that were resistant to biodegradation. These surfactants are largely obsolete today, having been replaced by linear alkyl benzene sulfonates (LAS) and other biodegradable surfactants. The toxicity to mammals is generally low for all surfactants, while the toxicity to aquatic organisms is relatively high (LC50 from about 0.1 to about 77 mg/L). The toxicity will generally increase with the carbon chain length (see Figure 5).

FIGURE 5 Log LC50 is plotted versus the number of carbon atoms in the chain for LASs. As seen, increased chain length implies increased toxicity.

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145

Many surfactants bind strongly to soils and sediments, which implies that, to the extent that they are not biodegraded in a biological treatment plant, they will mainly be found in the sludge phase.

Synthetic Polymers and Xenobiotics Applied in the Plastics industry Synthetic polymers and xenobiotics applied in the plastics industry form a very diverse group of compounds from a chemical viewpoint. Synthetic polymers are useful (plumbing, textiles, paint, floor, covering, and as the basic material for a wide spectrum of products) because they are resistant to biotic and abiotic processes of transformation and degradation. These properties, however, also cause environmental management problems associated with the use of these components. In addition, several xenobiotic compounds are used as additives, softeners, stabilizers, and so on in synthetic polymer to improve their properties. Some of these additives are very toxic and may cause other and additional environmental problems; for instance, phthalates are widely used in the plastics industry and it has been demonstrated that phthalates have effects as endocrine disrupters. After use, synthetic polymers are usually incinerated together with industrial and household garbage (solid waste). The presence of poly-vinyl-plastic (PVC) will imply that hydrochloric acid and, to a certain extent, dioxins are formed, but this is strongly dependent on the incineration conditions. As it is difficult to separate different types of plastics, it has been discussed to phase out the use of PVC, but due to the unique properties of PVC, this has not yet been decided.

References 1. Loganathan, B.G.; Lam, P.K.S. Global Contamination Trends of Persistent Organic Chemicals; CRC Press: Boca Raton, London, and New York, 2012; 639 pp. 2. Newman, M.C.; Unger, M.A. Fundamentals of Ecotoxicology, 2nd Ed.; CRC Press: Boca Raton, London, and New York, 2003. 3. Hoffman, D.J.; Rattner, B.A.; Burton, G.A., Jr.; Cairns, J., Jr. Handbook of Ecotoxicology; Lewis Publ./CRC Press: Boca Raton, London, and New York, 1994. 4. Schuurmann, G.; Markert, B. Ecotoxicology; John Wiley: New York, 1998; 902 pp. 5. Andersen, H.R. Examination of Endocrine Disruptors. Thesis, DFU, Copenhagen University, 1998.

Natural Elements and Chemicals

II

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14 Allelochemics

John Borden

Terminology ....................................................................................................149 Natural Occurrence ........................................................................................149 Practical Applications ....................................................................................152 Conclusion ...................................................................................................... 153 References........................................................................................................ 153

terminology To clarify in part the emerging maze of newly discovered message-bearing chemicals, or semiochemicals (Gk. semeion, sign or signal), the term allelochemic (Gk. allelon, of each other) was coined in 1970 to embrace any semiochemical with interspecific activity.[1] Thus allelochemics are distinguished from pheromones (Gk. pherum, to carry; horman, to excite) that convey a message between organisms of the same species. Three categories of allelochemics are commonly recognized. Kairomones (Gk. kairos, opportunistic) are allelochemics that provide an adaptive advantage to the perceiver. In most cases there is no benefit, or even harm, to the emitter, for example, the attraction of predators to the odor of their prey. The evolution of such chemicals as true biological signals would be disadaptive, and therefore unlikely. Therefore some semantic purists remove all evolutionary implications pertaining to kairomones by referring to them as infochemicals. Allomones (Gk. allos, other) are allelochemics that convey an adaptive advantage to the emitter. The repellent odor of an alarmed skunk is often used as an example. However, in an evolutionary sense it may also be adaptive for the receiver to be able to detect and avoid the skunk’s odor, for example, for a predator not to be “tagged” with an aroma that warns potential prey of its presence. Therefore, a skunk’s odor is more aptly termed a synomone (Gk. syn, with), an allelochemic that conveys a mutual advantage to both the emitter and the receiver. Most allelochemics have a releaser effect, in which behavioral responses are evoked. However, they may also have a primer effect, in which a physiological or biochemical function is stimulated or inhibited.

natural occurrence Table 1 provides a small window on the thousands of allelochemic interactions that occur in nature. The compounds that mediate these interactions are equally diverse (Figure 1). Very few interactions are mediated by a single compound; most involve relatively simple blends; and some, for example, floral fragrances, comprise dozens of compounds in a single blend. Although Table 1 provides examples of allelochemic interactions among terrestrial plants, arthropods, and vertebrates, many are also found among aquatic organisms, and examples occur in all Kingdoms and Phyla. Kairomonal interactions include the attraction of many species of phytophagous insects to their host plants, entomophagous insects to their prey or to insects that they parasitize, and blood-feeding diptera to their vertebrate hosts.[1–3] They also include the avoidance by prey species of odors associated 149

150 TABLE 1

Managing Global Resources and Universal Processes Examples of Natural Occurrence and Function of Allelochemicsa

Type of Allelochemic and Source

Example

Kairomone: Plants

Insects

Vertebrates

Allomone: Plants

Spiders Insects

Synomone: Plants

Insects

a

Attraction of Colorado potato beetles, Leptinotarsa decemlineata, to 6-carbon leaf volatiles, e.g., (E)-2-hexen1-ol-(2), from solanaceous plants Attraction of ambrosia beetles, Trypodendron lineatum, Gnathotrichus sulcatus and G. retusus, to ethanol (1) from moribund coniferous trees, logs, and stumps Attraction and stimulation of oviposition by onion maggots, Delia antigua, in response to mono- and disulfides, e.g., dipropyl disulfide (3) from onions, Allium cepa Employment of host tree kairomones, e.g., a-pinene (8) from conifers and a-cubene (10) from elms as synergists of aggregation pheromones that mediate mass attack of trees by bark beetles e.g., the Douglas-fir beetle, Dendroctonus pseudotsugae, and the smaller European elm bark beetle, Scolytus multistriatus, respectively Stimulation in a parasitic chalcidoid wasp, Trichogramma evanescens, of searching for and oviposition in corn earworm eggs, Helicoverpa zea, by tricosane (15) in moth scales adhering to newly laid eggs Attraction of predaceous clerid beetles, Thanasimus and Enoclerus spp. to aggregation pheromones, e.g., ipsenol (19), ipsdienol (20), and frontalin (22) of their bark beetle hosts Attraction of blood-feeding mosquitoes to CO2 (23) exhaled by mammals Attraction of tsetse flies, Glossina spp., to volatiles, e.g., acetone (24) and 1-octen-3-ol (25) from bovine animals on which they feed Avoidance by voles, Microtus spp., of volatile chemicals in the urine of mustellid predators, e.g., 2-propylthiotane (26) and 3-propyl-1,2-dithiolane (27) from short-tailed weasels, Mustella erminea Inhibition of germination of growth of one species of plant by allelopathic chemicals produced in the leaves, roots, or other tissues of another plant, e.g., by juglone (7) leaching from the leaves of black walnut trees, Juglans nigra Disruption of growth, metamorphosis, or reproduction of insect herbivores by producing insect hormones or hormone analogues, e.g., juvabione (11), a juvenile hormone mimic in true firs, Abies spp. Emission of moth sex pheromones, e.g., (Z)-9-tetradecenyl acetate (18) by bolas spiders, Mastophora spp., to attract male moths as prey Mimicking the cuticular recognition compounds, e.g., 11-methyl pentacosane (16) of larval ants by caterpillars of lycenid butterflies, syrphid fly larvae, and scarab beetle larvae, thereby gaining entry into and acceptance within ant nests, where they prey on ant brood Avoidance of nonhost plants by insects in response to volatiles emitted by the plants, e.g., repellency of coniferophagous bark beetles, e.g., the mountain pine beetle, Dendroctonus ponderosae, to conophthorin (13) (also a repellent pheromone of cone and twig beetles) in the bark of birches, Betula spp. Repellency of black bean aphids, Aphis fabae, to methyl salicylate (4) in the volatiles of nonhost plants Tritophic interaction in which corn plants, Zea mays, respond to volicitin (14) (a kairomone) in the saliva of beet armyworm caterpillars, Spodoptera exigua, feeding on them by producing specific blends of volatiles (synomones), e.g., (E)-4, 8-dimethyl-1,3,7-nonatriene (9) that attract females of a parasitic wasp, Cotesia marginiventris, which in turn oviposit on the feeding caterpillars Attraction of honey bees, Apis mellifera, to multicomponent blends of floral volatiles, e.g., geraniol (12), of many species of angiosperm plants, with mutual benefit of pollination to the plant and a pollen and nectar source to the bees Antagonists in the blends of moth sex pheromones that repel males of related species, ensuring reproductive isolation even though the major pheromone components are attractive to males of both species, e.g., (Z)-9-tetradecanal (17), a pheromone component of threelined leafrollers, Pandemis limitata, that inhibits response of oblique banded leafroller males, Choristoneura rosaceana, to threelined leafroller females Mutual repellency between aggregation pheromones of two species of bark beetles, e.g., (R)-(–)-ipsdienol (20) produced by pine engravers, Ips pini, and (S)-( – )-ipsenol (19) produced by California fivespined ips, I. paraconfusus, reserving the host phloem resource for the first-arriving species, and avoiding interspecific exploitative competition

Numbers in parentheses correspond with numbered compounds in Figure 1.

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151

FIGURE 1 Structural formulae of compounds given in Table 1 exemplifying some of the chemical diversity among allelochemics as follows: primary alcohol (1, 2), secondary alcohol (25), disulfide (3), aromatic ester (6), unsaturated ester (18) monoterpene (8), sesquiterpene (9, 10), sesquiterpenoid (11), terpene alcohol (12,  19,  20), terpene ketone (21), spiroacetal (13), fatty acid derivative conjugated to an amino acid (14), straight chain hydrocarbon (15),  branched hydrocarbon (16), unsaturated aldehyde (17), bicyclic ketal (22), atmospheric gas (23),  ketone (24), thiotane (26), and thiolane (27).

with predators, a phenomenon found in five animal phyla, but curiously not yet among terrestrial insects.[4] Sometimes more than one type of semiochemical may be involved; for example, the aggregation of bark beetles necessary to mass attack and kill a tree is mediated by a blend of aggregation pheromones synergized by host tree kairomones. Allomonal interactions may employ “trickery”,[5] for example, bolas spiders that emit moth sex pheromones that lure mate-seeking male moths to their death, and many species of myrmecophiles (ant lovers) that gain access to ant nests by chemically mimicking the cuticular recognition compounds of ants on which they prey. The most well-known allomones are released by or are contained in plants, and have a primer effect. Allelopathic allomones are often leached from the leaves (or other parts) of plants of one species, and inhibit the germination of seeds or growth of plants in other species that could be potential competitors.[6] Some plants may also produce defensive allomones against insect herbivores, for example, hormones or analogues of hormones that disrupt the growth and metamorphosis of their insect enemies. Among the many examples of synomones are repellents that ensure reproductive isolation between closely related species of insects, or mitigate against the occurrence of interspecific exploitative competition for a limited host resource. Often synomones are the same compounds as one or more of the

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components that convey a pheromonal message, for example, to attract mates or to aggregate on or near a food source. There is increasing evidence that host-seeking phytophagous insects not only use kairomones to find their host plants, but also use synomones to avoid nonhost plants on which their fitness would be greatly reduced. Synomonal floral scents provide a mutual benefit to flowering plants that gain from pollination by insects that in turn are attracted to a nutritious nectar or pollen source. Another type of allelochemic interaction involves three trophic levels and the action of both kairomonal and synomonal stimuli.[3] In one remarkable example of this type of tritrophic interaction, corn plants being fed on by beet armyworm caterpillars are exposed to minute amounts of a kairomone called volicitin in the insect’s saliva. Volicitin has a primer effect, eliciting the plant to synthesize a specific blend of volatile synomonal compounds that attract females of a parasitic wasp. The wasp oviposits in the beet armyworm larvae, benefiting by finding its host, and in turn providing an advantage to the plant by parasitizing and killing the caterpillar.

Practical Applications Knowledge about the natural occurrence and role of allelochemics opens up a huge, but relatively untapped, potential for exploiting them as pest management tools.[2] In some cases the knowledge itself is important. Plant breeders may seek varieties of plants that contain or release chemicals that deter feeding or development by phytophagous insects. Plants with allelopathic characteristics, for example, Eucalyptus spp., may be useful in landscaping to reduce weed problems. Species or varieties rich in attractive kairomones may be used as trap crops for various insect pests. In the production of transgenic agricultural crops it is critical not to lose the capacity for tritrophic interaction that will ensure parasitism of herbivorous insects, lest the genetically modified plants be more vulnerable to insect pests than unmodified plants. In other cases, the capacity to use allelochemics as pest management tools may be demonstrated, but technological, economic, or social limitations may prevent their use. Allelopathic allomones from plants are under consideration for development as a new class of biodegradable herbicides.[7] But to date none can compete with conventional chemical herbicides with regard to ease of synthesis, capacity for formulation, efficacy, and/or safety. If used widely, some allelo-pathogens may pose an unacceptable threat to environmental or human health. Recent investigations show considerable promise for using nonhost volatiles to “disguise” herbivorous host plants or trees as nonhosts, but commercial formulations have not yet appeared on the market, in part because of the challenge and expense of registering those allelochemic products as pesticides. Similar problems beset the development and use of predator volatiles to protect plants from damage by herbivorous vertebrates such as deer and voles. Despite the above limitations, a few kairomones have found widespread commercial use.[2,8] Among them is methyl engenol, which is used worldwide for capturing tephritid fruit flies, both for detection of unwanted introductions and for direct suppression of populations in a lure and kill tactic employing an insecticide-laced substrate baited with methyl engenol. A lure and kill tactic is also used effectively for control of tsetse flies that are drawn by acetone and 1-octen-3-ol baits to insecticide-treated “target” traps that simulate the silhouette of a large vertebrate. Other applications combine kairomones with pheromones. A combination of phenethyl proprionate and methyl engenol with the sex pheromone of the Japanese beetle is used in many thousands of traps in the United States each year. Similarly the kairomones ethanol and α-pinene have been used since 1981 in combination with aggregation pheromones in commercial mass trapping programs for three species of ambrosia beetles in British Columbia. Allelochemics may also find use in the future in the application of “push-pull” tactics, in which one repellent volatile treatment is used to protect a plant or group of plants from attack by insects (push), and another attractive treatment is used in baited traps or trap plants to pull the insects away. One outstanding example of a successful push-pull application saved a rare stand of endangered Torrey pines in California from being killed by the California five spined ips. Two repellent synomones, verbenone, produced by western pine beetles, and (–)-ipsdienol, produced by pine engravers, were deployed inside the

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153

uninfested portion of the stand, and traps baited with attractive aggregation pheromone were arrayed in an adjacent area of beetle-killed pines. Over 86 weeks beginning in May 1999, 330,717 beetles were captured.

conclusion Despite many studies, most natural allelochemic interactions are yet to be discovered. The adoption of allelochemics as pest management tools has been limited. However, there is great potential for judicious selection and commercial development of allelochemics, particularly in integrated pest management programs that will combine a number of alternative ecologically based tactics with the reduced use of conventional chemical pesticides.

References 1. Whittaker, R.; Feeney, P. Allelochemics: chemical interactions between species. Science 1971, 171, 757–770. 2. Metcalf, R.; Metcalf, E. Plant Kairomones in Insect Ecology and Control; Chapman and Hall: New York, 1992; 168. 3. Vet, L.; Dicke, M. Ecology of infochemical use by natural enemies in a tritrophic context. Annu. Rev. Entomol. 1992, 37, 141–172. 4. Kats, L.; Dill, L. The scent of death: chemosensory assessment of predation risk by prey animals. Ecoscience 1998, 5, 361–394. 5. Stowe, M.; Turlings, T.; Loughrin, J.; Lewis, W.; Tumlinson, J. The Chemistry of Evesdropping, Alarm, and Deceit. In Chemical Ecology: The Chemistry of Biotic Interaction; Eisner, T., Meinwald, J., Eds.; National Academy Press: Washington, DC, 1995; 51–65. 6. Zindahl, R. Fundamentals of Weed Science; Academic Press: New York, 1993; 135–146. 7. Cutler, G. Allelopathy for Weed Suppression. In Pest Management: Biologically Based Technologies; Lumsden, R., Vaughn, J., Eds.; American Chemical Society: Washington, DC, 1993; 290–302. 8. Borden, J. Disruption of Semiochemical-Mediated Aggregation in Bark Beetles. In Insect Pheromone Research: New Directions; Cardé, R., Minks, A., Eds.; Chapman and Hall: New York, 1997; 421–438.

15 Aluminum Introduction .................................................................................................. 155 Chemistry of Al ............................................................................................ 156 Al Uses and Production ............................................................................... 156 Sources of Al in Soil ..................................................................................... 157 Al Minerals • Forms of Soluble Al in Soil • Mobilization of Al in Soil

Alleviation of Al in Soils .............................................................................. 158 Sources of Al in Water ................................................................................. 159 Acid Deposition • Acid Mine Drainage and Acid Sulfate Soils • Aluminum from Water Purification

Johannes Bernhard Wehr, Frederick Paxton Cardell Blamey, Peter Martin Kopittke, and Neal William Menzies

Alleviation of Al in Water............................................................................ 160 Sources of Al in the Atmosphere ................................................................ 160 Sources of Al in Foodstuff, Cosmetics, Pharmaceuticals, and Workplaces................................................................................................161 Environmental Toxicology of Al ................................................................ 162 Al Toxicity in Plants • Al Toxicity in Aquatic Organisms • Al Toxicity in Humans

Conclusion .................................................................................................... 164 Acknowledgments ........................................................................................ 164 References ...................................................................................................... 164

introduction Aluminum (Al) is a metallic element characterized by its low density (2.7 g/cm3) and resistance to corrosion due to the formation of a protective oxide layer on the surface. It is the most common (8%) metallic element of the earth’s crust and the third most common element (after oxygen and silicon) on earth (Wehr, Blamey, and Menzies 2007). The element reacts strongly with oxygen-containing ligands and never occurs naturally as pure metal. It forms both octahedral and tetrahedral coordination compounds and occurs only in the trivalent oxidation state. In soil, Al-containing minerals dissolve at low pH, especially below pH 4.5, and release Al into the soil solution and aquatic environment. Al can also occur in the atmosphere as aeolian dust or ash. Soluble forms of Al can have toxic effects on plants, humans, as well as soil and aquatic organisms, but alleviation of Al toxicity can be achieved by increasing the pH to above 4.5. Aluminum present in water and food is not readily taken up by humans. Nano-sized Al particles in the environment may result in some effects on plants and animals but more research is needed. The chemistry and uses of Al are discussed, as well as the sources and alleviation of Al toxicity in soils, water, and atmosphere. Finally, the environmental toxicology of Al to plants, microorganisms, and animals is briefly discussed.

155

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chemistry of Al Dissolution of Al minerals at low pH releases octahedral Al3+, which is hexa-coordinated with water molecules (Martin 1996). Soluble Al3+ is classified as a hard acid due to its small ionic radius of 0.053 nm and reacts strongly with “hard” ligands such as oxygen and fluoride (Martin 1996, Martell et al. 1996). It also forms stable complexes with didentate and multidentate ligands (chelate effect) (Martell et al. 1996, Salifoglou 2002). As the pH of an acidic Al solution is increased, especially >4, hydrolysis of Al occurs, giving rise to a series of Al-hydroxy species [AlOH2+, AlOH2+, and Al(OH)3] (Martin 1996), which may undergo aggregation (polymerization) via OH bridges if the Al concentration is sufficiently high (>0.3 mg/L) (>10 μM) (Furrer, Trusch, and Muller 1992). As the pH is raised above pH 6, Al changes its coordination number to 4 and yields tetrahedral aluminate (AlOH4–) (Martin 1996). Depending on the pH and Al concentration in solution, hydrolyzed Al species may aggregate and form polymeric (polycationic) Al species such as Al2(OH)24+, Al3(OH)45+, and Al8(OH)20(H2O)x4+, the “gibbsite fragment” model forms, Al6(OH)12(H2O)126+ through Al54(OH)144(H2O)3618+, Al13O4(OH)24(H2O)127+ (Al13), and Al2O8Al28(OH)56(H2O)2618+ species (Bertsch and Parker 1996, Brown et al. 1985, Orvig 1993, Sarpola et  al. 2006, Wang and Muhammed 1999). The Al13 polycation in solution is highly toxic to root elongation (Bertsch and Parker 1996). While many more Al polymers have been proposed, the experimental evidence in support of these species is limited.

Al Uses and Production Due to its corrosion resistance, light weight, and excellent thermal and electrical conductivity, the metal is extensively used in the building and construction industries (window frames, doors, external cladding, A/C ducts, thermal insulation), automotive (engine blocks, car bodies), shipping (hulls) and aerospace industries (aircraft bodies), power lines, and food packaging (cans and other containers, foil). Aluminum salts and compounds are used for water purification, in which alum (KAl(SO4)2.12H 2O) is important, as catalysts in the chemical industry, and as ingredients in cosmetics (antiperspirants), pharmaceuticals (antacids, vaccine adjuvant), and foods (baking powder, spreading agent) (Table 1). The main producer of bauxite ore is Australia (30% of world production, 88 Mt in 2017), with China, Brazil, Guinea, Jamaica and India producing lesser quantities (Resources-and-EnergyQuarterly 2018) (Table 2). The main producers of alumina in 2017 were Australia and China (WorldAluminum.org 2019), whereas smelting (refining of bauxite) is occurring in regions with cheap electrical energy (Table 2).

TABLE 1

Global Uses of Al (2018)

Use Transport and manufacturing Packaging Construction Electrical Consumer durables Machinery Other

(%) 41 20 14 8 7 7 3

Source: https://archive.industry.gov.au/Office-of-the-Chief-Economist/ Publications/ResourcesandEnergyQuarterlyMarch2018/documents/ Resources-and-Energy-Quarterly-March-2018-Aluminium-aluminaand-bauxite.pdf (accessed 29 March 2019).

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Aluminum TABLE 2 2017

Worldwide Production of Alumina and Primary Al Metal in

Geographic Region Africa and Asia (excl China) North America South and Central America East and Central Europe Western Europe China Oceania Rest of World Total

Alumina (kt)

Primary Al Metal (kt)

8,382 3,033 12,713 4,499 5,890 70,699 20,783 6,391 132,400

10,776 3,950 1,378 3,999 3,776 35,905 1,817 1,800 63,404

Source: www.world-aluminium.org/statistics/primary-aluminium-production (accessed 29 March 2019).

Sources of Al in Soil Al Minerals The main Al-containing primary minerals are feldspars and micas. Several gemstones (ruby, sapphire, tourmaline) also contain Al. The primary Al-containing minerals weather initially to 2:1 layer aluminosilicate clay minerals (e.g., vermiculite, montmorillonite, smectite), which upon further weathering form 1:1 layer aluminosilicates such as kaolinite. Further weathering of clay minerals leads to leaching of silica and base cations (calcium and magnesium), leaving behind hydrous aluminum oxide (e.g., gibbsite, boehmite, and diaspore) in the form of bauxite, which is an important aluminum ore (Wehr, Blamey, and Menzies 2007).

Forms of Soluble Al in Soil In acidic soils (pH < 4.5), aluminosilicate clay minerals are unstable and dissolve, releasing the trivalent Al3+ cation. The solubility of Al increases three orders of magnitude for every unit decrease in solution pH: at pH 4.5 the soluble Al concentration is around 0.09 mg/L (3.5 µM), decreasing to 95 ng/L (3.5 nM) at pH 5.5 (Menzies, Bell, and Edwards 1994). The trivalent Al3+ ion is toxic to plant roots at concentrations of 0.1–1 mg/L (5–50 µM) (Brown et al. 2008, Horst, Wang, and Eticha 2010, Poschenrieder et al. 2008). The Al3+ ion is readily complexed by soil organic matter and the soil organic matter controls the Al availability in most soils containing sufficient organic matter (Adams et al. 2000, Brown et al. 2008, Lofts et al. 2001, Simonsson 2000, Skyllberg 1999, Guo et al. 2007). In mineral soils low in organic matter, the availability of Al is determined by the cation exchange capacity, ionic strength, and pH (Guo et al. 2007). The polymeric Al species (e.g., Al13) are metastable with a half-life for Al13 of several hundred hours (Etou et al. 2009, Furrer, Gfeller, and Wehrli 1999) which implies that the species undergo depolymerization or crystallization with time. The Al13 species is also rhizo-toxic to plant roots, with concentrations between 0.1 and 2 mg/L (0.1–2 μM) inhibiting root growth (Kopittke, Menzies, and Blamey 2004, Bertsch and Parker 1996). It is generally accepted that sulfate, phosphate, fluoride, silicate, and organic acids prevent or reverse Al13 formation (Bertsch and Parker 1996, Casey 2005, Masion et al. 1994, Yamaguchi et al. 2003, Kerven, Larsen, and Blamey 1995) and crystalline gibbsite may induce crystallization and depolymerization of Al13 (Sanjuan and Michard 1987). Since sulfate and silicate ions are prevalent even in acid soils, the natural occurrence of Al13 in soil solution is considered unlikely (Bertsch and Parker 1996, Gerard, Boudot, and Ranger 2001, Hiradate, Taniguchi, and Sakurai 1998). However, there is evidence that Al13 may form in the cell wall of plant roots (Kopittke, Menzies, and Blamey 2004, Masion and Bertsch 1997, Xia and Rayson 1998) even if the conditions in bulk solution are not favorable for Al13 formation. A report by Hunter and Ross (Hunter and Ross 1991) claiming that up

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to 30% of acid forest soils contained Al13 could not be replicated by other research groups. In alkaline soils, the aluminate anion [Al(OH)4–] can form but this species is not very toxic to plants (Kinraide 1991, Kopittke, Menzies, and Blamey 2004) but may be toxic to aquatic organisms (Griffitt et al. 2008, Sjostedt et al. 2009).

Mobilization of Al in Soil Aluminosilicate and aluminum sesquioxide clay minerals slowly dissolve at pH < 4.5 and release Al into the soil solution. Currently, more than 30% of potentially arable soils are acidic and affected by Al toxicity (von Uexküll and Mutert 1995). Soil acidification and consequential Al toxicity can be exacerbated by anthropogenic acid deposition (acid rain), use of ammonium-containing fertilizers, use of legumes in crop rotations, and the removal of crops from agricultural land (von Uexküll and Mutert 1995, Lesturgez et al. 2006). In unfarmed soils, atmospheric acid inputs are the main causes of acidification (Lapenis et al. 2004, Bergholm, Berggren, and Alavi 2003), whereas acidification in fertilized soils is mainly caused by nitrification of ammonium N-fertilizer, especially when the N is not taken up by the plant and is leached from the soil (Bergholm, Berggren, and Alavi 2003, Guo et al. 2010). Aluminum in soil solution decreases turn-over of soil organic matter by limiting microbial degradation of the stable Al-organic matter complexes formed in soil, leading to an increase in soil organic matter (Scheel, Dorfler, and Kalbitz 2007, Scheel et al. 2008, Rasmussen, Southard, and Horwath 2006, Takahashi and Dahlgren 2016, Miyazawa et al. 2013). Aluminum can also be released from soil minerals in strongly acidic conditions found in acid sulfate soils (Dent and Pons 1995, Faltmarsch, Astrom, and Vuori 2008), monosulfidic black ooze (Bush, Fyfe, and Sullivan 2004) and acid mine drainage (Liang and Thomson 2009, Pu et al. 2010). Acid sulfate soils are limited in occurrence to permanently or temporarily waterlogged low-lying areas (up to 5 m above sea level) (Dent and Pons 1995), whereas organic matter in irrigation channels can lead to the formation of monosulfidic black ooze (Bush, Fyfe, and Sullivan 2004). Exposure of these sulfide-containing materials to air leads to the formation of sulfuric acid and release of Al (Dsa et al. 2008, Liang and Thomson 2009, Pu et al. 2010, Soucek, Cherry, and Zipper 2003). Fluoride ions (F–), present as an impurity in phosphate fertilizers, can form strong complexes with Al and the resultant lesser charged complexes, AlF2+, AlF2+, and AlF3, are mobile in soil solution (Martin 1996, Martinent-Catalot et al. 2002) and can be either removed from soil by drainage or seepage water, or taken up by plants (Manoharan et al. 2007) and alleviate Al toxicity (Yang et al. 2016). Organic acids form complexes with Al and can mobilize Al in the soil (Takahashi et al. 2008, Lange, Solberg, and Clarke 2006a). In the field, movement of organic acid–Al complexes gives rise to podzolization of soils and downward movement of Al until conditions are favorable for dissociation of Al complexes deeper in the profile or export to waterways (Bardy et al. 2007).

Alleviation of Al in Soils Since the solubility of Al in soil solution is lowest around pH 6–7 (McBride 1994), controlling the pH is the most effective way in reducing potential Al toxicity in ecosystems. The pH value to which acid soils need to be limed depends upon the plant species, but is often near 5.5. Compounds to increase the pH of acid soils include calcitic (CaCO3) or dolomitic (CaMgCO3) limestone (Machacha 2004), slaked lime [Ca(OH)2] (Sun et al. 2000), wood ash (Materechera and Mkhabela 2002), alkaline poultry manure (Mokolobate and Haynes 2002, Tang et al. 2007), charcoal (Steiner et al. 2007), biochar (Dai et al. 2017), fly ash (Morikawa and Saigusa 2002, Tarkalson et al. 2005), ground basalt (Panhwar et al. 2016), alkaline slag (Shi et al. 2017, Li et al. 2015) or waste cement (Morikawa and Saigusa 2002). Any waste product with high ash alkalinity can be used to ameliorate acid soil by increasing the pH (Naramabuye and Haynes 2006). Surface application of liming agents with subsequent incorporation into the soil (ploughing, harrowing, etc.,) is commonly used. The mobility of limestone in the soil

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159

profile is very low (Conyers et al. 2003, Godsey et al. 2007, Scott et al. 2007). Therefore, rapid alleviation of subsoil acidity can only be achieved by deep placement of limestone (deep ripping). This, however, is currently not economically possible for most low-value broadacre crops. Depending on the severity of soil acidity, and the buffer capacity of the soil as determined by the cation exchange capacity of the soil, between 0.1 and 1 ton lime-stone/ha/yr are needed to maintain soil pH in agricultural production systems (Scott et al. 2007, Machacha 2004). Numerous methods have been developed to determine the required liming rate based on laboratory soil analyses. The liming of soil needs to be repeated every few years depending on the land management system. To overcome the low mobility of limestone in soils, higher rates of gypsum can be used instead. The Ca from gypsum moves more readily down the soil profile (Liu and Hue 2001) and the increase in Ca and ionic strength can lower the Al toxicity, despite increasing the overall concentration of Al (as AlSO4+, which is less toxic than Al 3+). Liming has greater beneficial effect on soil bacterial communities than on soil fungal communities (Mota et al. 2008, Nelson and Mele 2006). Application of organic matter (green manure, farm yard manure) can lower Al toxicity by complexing Al (Qin and Chen 2005, Vieira et al. 2008). However, application of organic matter requires higher application rates (>10 tons/ha) than limestone (1 ton/ha) and is, therefore, not as effective as limestone (Raboin et al. 2016). Humic acid has been shown to either precipitate or bind Al13 (Yamaguchi et al. 2004) and Al3+, (Matthias, Maurer, and Parlar 2003, Shoba and Chudnenko 2014), thereby reducing root growth inhibition. Biochar can also adsorb Al-OH complexes and reduce rhizotoxicity in crops (Qian, Chen, and Hu 2013). Sewage sludge can also be alkaline and the presence of organic acids in sludge can bind and immobilize Al (Lopez-Diaz, Mosquera-Losada, and Rigueiro-Rodriguez 2007). Alkaline bauxite refinery waste (red mud) poses a challenge for plant growth due to the high Fe oxide concentration leading to phosphate immobilization, absence of diverse microbial communities, and lack of organic matter (Wehr, Menzies, and Fulton 2006). However, the presence of aluminate in these waste materials is not limiting to plant root growth (Kopittke, Menzies, and Blamey 2004).

Sources of Al in Water Acid Deposition Atmospheric gases such as NO2, SO2, and CO2 dissolve in rainwater and form acid rain (Lapenis et al. 2004, Lawrence 2002). The atmospheric acid inputs in Europe are considered to be around 0.2–4 kmol/ha/yr (de Vries, Reinds, and Vel 2003). The acids can either dissolve Al minerals in soil (see above) or dissolve sediments in water bodies receiving acidic water. Furthermore, acid deposition can release Al complexed to organic matter. Introduction of gaseous emission standards in the 1980s in Europe, and later in the United States, has lowered acid rain, with a consequential decrease in Al in surface water (Lange, Solberg, and Clarke 2006b, Skjelkvale et al. 2001, Vuorenmaa et al. 2018). Podzolization of lateritic soils was found to export vast quantities of Al to waterways (Bardy et al. 2007). The concentration of Al in drainage water from forest watersheds is highly variable, however (Lange, Solberg, and Clarke 2006b). In coastal areas subjected to atmospheric salt deposition, the salt can displace Al from cation exchange sites, leading to an increase in Al in drainage water (Lange, Solberg, and Clarke 2006b). Concentrations of Al in groundwater are governed by pH since Al solubility is pH dependent: less than 0.01 mg/L Al is observed at pH 7, which increases to 51 mg/L at pH < 4 (Fest et al. 2007).

Acid Mine Drainage and Acid Sulfate Soils Sulfidic mine wastes and acid sulfate soils release acid when exposed to air, resulting in leaching of acid drainage water and soluble Al into water bodies. While there are indications that the high Fe concentration in acid mine drainage can protect aluminosilicate clay mineral from dissolution and minimize Al release (Dubikova et al. 2002), no such mechanism has been reported for acid sulfate soils. At Trinity

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inlet in North Queensland, Australia, Al concentrations in drainage water from acid sulfate soils were measured in the range 4.2–10 mg/L (Hicks, Bowman, and Fitzpatrick 1999).

Aluminum from Water Purification Water clarification involving the settling of fines (suspended colloids and microorganisms) may be accomplished with alum (potassium aluminum sulfate). Since drinking water will be near neutral (pH  6.5–8.5 (NHMRC 2004)), solubility of Al is low. The European, Australian, New Zealand, and WHO regulations for drinking water propose a threshold concentration of Al of 0.1-0.2 mg/L (Schaefer and Seifert 2006). Traces of Al are found in raw and alum-treated potable water (Cech and Montera 2000, Srinivasan and Viraraghavan 2002). The amount of Al taken up from drinking water has been estimated at 1%–2% of total Al uptake per day (Yokel, Hicks, and Florence 2008, Stauber et al. 1999, Willhite, Ball, and McLellan 2012), and less than 2% of Al in drinking water is bioavailable (Willhite, Ball, and McLellan 2012, Yokel and McNamara 2001). Water treatment residues (alum sludges) have been used to immobilize phosphate and heavy metals in soils and wastes (Mahdy et al. 2009, Oladeji, Sartain, and O’Connor 2009). Land application of these sludges has not resulted in elevated Al concentrations in plants grown in these soils (Oladeji, Sartain, and O’Connor 2009), especially when the sludges were aged to reduce Al availability (Agyin-Birikorang and O’Connor 2009).

Alleviation of Al in Water Raising the pH of water to pH 6–7 will precipitate Al and establish an equilibrium concentration of Al determined by the solubility product of Al-hydroxide in water. At pH < 4.5, solubility of Al minerals increases rapidly, releasing Al3+ into the waterbody and at pH > 8, aluminate anions are formed (Wehr, Blamey, and Menzies 2007). Therefore, pH adjustment of water to near neutrality is the most effective method to decrease Al concentrations in water (Fest et al. 2007). The Al remaining in potable water after purification can only be removed by ion exchange resins (Othman, Abdullah, and Abd Aziz 2010).

Sources of Al in the Atmosphere Atmospheric dust derived from raised soil particles is the main source of Al in the atmosphere, but volcanic activity can be a source also. In the atmosphere, Al occurs generally as silicate, sulfate, or oxide compounds. Flue ash and flue gases can also contribute Al to the atmosphere (Kabata-Pendias and TABLE 3

Concentrations of Al in the Atmosphere

Region South Pole Greenland Shetland Islands Norway Germany Japan China North America Central America South America

Al Concentration (µg/m3) 0.0003–0.0008 0.24–0.38 0.06 0.03 0.16–2.90 0.04–10.60 2.00–6.00 0.60–2.33 0.76–0.88 0.46–15.00

Source: Adapted from Kabata-Pendias and Kabata 2001 and Wang et al., 2007.

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Kabata 2001). Dust contains around 2.6% Al (Schussler, Balzer, and Deeken 2005). Concentrations of Al in the atmosphere vary between geographic locations (Table 3). Inhalation of Al from the atmosphere has been estimated as 4–20 μg Al/day (Yokel, Hicks, and Florence 2008).

Sources of Al in Foodstuff, cosmetics, Pharmaceuticals, and Workplaces The concentration of Al in foodstuff is generally very low since little Al is taken up by plants and translocated to the aboveground parts (Table 4) (Chen et al. 2008, Fung et al. 2009, Mueller, Anke, and IllingGuenther 1998, Ertl and Goessler 2018, Schaefer and Seifert 2006). However, the concentration of Al in black tea is high (Table 4), but most of the Al in infused tea is complexed with phenolics and organic acids, though addition of lemon or lime juice may increase Al availability (Fung et al. 2009, Krewski et al. 2007). The bioavailability of Al from black tea is around 0.4% (Yokel and Florence 2008). Preparation of acidic foodstuff in Al containers or covering with Al foil may increase Al in foodstuff (Ertl and Goessler 2018, Turhan 2006, Ranau, Oehlenschlager, and Steinhart 2001). The addition of sodium aluminum phosphate as a spreading agent to processed foodstuff (e.g., cheese spread) can also contribute to the dietary Al intake (Yokel, Hicks, and Florence 2008, Schaefer and Seifert 2006). Salts of Al are also used as a food additive for clarification of beverages, as anticaking agent, as baking powder, and to enhance color stability (Stahl et al. 2018, Aguilar et al. 2008, Schaefer and Seifert 2006). The Al in food contributes around 95% of daily oral Al uptake and the daily dietary Al intake has been estimated as 4–16 mg, of which 0.1%–0.3% is bioavailable (Krewski et al. 2007, Yokel, Hicks, TABLE 4

Concentration of Al in Foodstuff

Foodstuff

Al Concentration (mg/kg DW)

Grains Pulses Fruit Vegetables Herbs Spices Meat Fish Dairy Tea

1–135 3–16 0.5–20 1–270 8–280 6–695 0.5–16 0.5–10 1–16 900–1000

Source: Adapted from Kabata-Pendias and Kabata 2001, Yi and Cao 2008, Fung et al. 2009, Ertl and Goessler 2018, Mueller, Anke, and Illing-Guenther 1998, Schaefer and Seifert 2006, Chen et al. 2008.

TABLE 5

Estimated Sources and Daily Exposure to Al, its Bioavailability, and Estimated Daily Intake

Environmental Exposure Air Industrial air Water Food Cosmetics Vaccines Antacids

Exposure (μg/day)

Availability (%)

Intake (μg/day)

2–200 25,000

2 2

0.04–4

200–1,000 8,000–16,000

11; may oxidize to Cr(VI) by O2 Soluble complexes and chelates in which water molecules of hydration surrounding Cr(H2O)63+ are displaced by carboxylic acid and N-containing ligands; formation is pH and concentration dependent; blue–green– purple colors, depending on ligand binding Cr(III) Fully protonated form of Cr(VI) formed at pH < 1; see Figure 2 for key Eh values for redox Form of Cr(VI) that predominates at 1 < pH < 6.4; yellow; see Figure 2 for key Eh values for redox Form of Cr(VI) that predominates at pH > 6.4; yellow; see Figure 2 for key Eh values for redox Form of Cr(VI) that predominates at pH < 3 and in concentrated solutions (>1.0 mM); rapidly reverts to HCrO4– or CrO42– upon dilution or pH change

3+ 6

198 TABLE 2

Managing Global Resources and Universal Processes Solubility in Water at pH 7 of Selected Chromium Compounds

Oxidation State Chromium (III) (trivalent chromium)

Chromium (VI) (hexavalent chromium)

Compound Name

Formula

Approximate Solubility (mol Cr/L) 10–12 10–17 10–20

Chromium(III) hydroxide Chromium(III) oxide Chromite

Cr(OH)3 (am) Cr2O3 (Cr) FeO·Cr2O3 (Cr)

Chromium chloride

CrCl3

Highly soluble

Chromium sulfate

Cr2(SO4)3

Highly soluble

Chromium phosphate

CrPO4

Chromium fluoride

CrF3

10–10 1.2 × 10–3

Chromium arsenate

CrAsO4

10–10

Potassium chromate Sodium chromate Calcium chromate

K2CrO4 Na2CrO4 CaCrO4

3.2 5.4 0.14

Barium chromate

BaCrO4

1.7 × 10–3

“Zinc yellow” pigment

3ZnCrO4·K2CrO4·Zn(OH)2H2O

8.2 × 10–3

Strontium chromate

SrCrO4

5.9 × 10–3

Lead chromate

PbCrO4

1.8 × 10–6

Chromium(VI) jarosite

KFe3(CrO4)2(OH)6 (Cr)

10–30

solubility product (Ksp).[17,18] In the pH range of 5.5–8.0, Cr(III) reaches minimum solubility in water due to this hydrolysis and precipitation reaction, an important process that controls the movement of Cr(III) in soils enriched with industrial wastewaters and solid materials. Under strongly acidic conditions (pH  11), particularly in response to adding base to solutions of soluble salts of Cr(III), e.g., CrCl3, Cr(NO3)3, or Cr2(SO4)3. Other anions besides OH– coordinate with Cr(H2O)63+ and displace water molecules of hydration to form sparingly soluble compounds and soluble chelates (Table 2). In water treatment facilities and in natural waters, phosphate (H2PO4–, HPO42–, PO43–), arsenate (H2AsO4–, HAsO42–) and fluoride (F–) may form low solubility compounds with Cr(III). Organic complexes of Cr(III) with carboxylic acids (e.g., citric, oxalic, tartaric, fulvic) remain soluble at pH values above which Cr(OH)3 forms. By increasing the solubility of Cr(III) in neutral and alkaline waters, such organic complexes enhance the potential for absorption of Cr(III) by cells. Stable, insoluble complexes of Cr(III) also form with humic acids and other high molecular aggregate weight organic moieties in soils, sediments, wastes, and natural waters.[19] With the exception of chromium(VI) jarosite (Table 2), Cr(VI) compounds are more soluble over the pH range of natural waters than are those of Cr(III), thereby leading to the greater concern about the potential mobility and bioavailability of Cr(VI) than Cr(III) in natural waters. The alkali salts of Cr(VI) are highly soluble, CaCrO4 is moderately soluble, and PbCrO4 and BaCrO4 are only sparingly soluble. In colloidal environments containing aluminosilicate clays and (hydr)oxides of Al(III), Fe(II,III), and Mn(III,IV) (e.g., in soils and sediments), Cr(VI) anions may be adsorbed similarly to SO42–. Low pH and high ionic strength promote retention of HCrO4– and CrO42– on positively charged sites, especially those associated with colloidal surfaces dominated by pH-dependent charge. Such electrostatic adsorption may be reversible, or the sorbed Cr(VI) species may gradually become incorporated into the structure of the mineral surface (chemisorption). Recently precipitated Cr(OH)3 can adsorb Cr(VI) or incorporate Cr(VI) within its structure as it forms, thereby forming a Cr(III)–Cr(VI) compound.[20]

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199

oxidation-Reduction chemistry of chromium in natural Waters The paradox of the contrasting solubilities and toxicities of Cr(III) and Cr(VI) in natural waters and living systems is complicated by two reduction–oxidation (electron transfer) reactions: Cr(III) can oxidize to Cr(VI) in soils and natural waters, and Cr(VI) can reduce to Cr(III) in the same systems, and at the same time. Understanding the key electron transfer processes (redox) and predicting environmental conditions governing them are central to treatment of drinking water, wastewaters, and contaminated soils, and to predicting the hazard of Cr in natural systems.[21] The metaphor of a seesaw (Figure 1) is useful in picturing the undulating nature of the changes in Cr speciation in water due to oxidation of Cr(III) and reduction of Cr(VI). A balance for the two redox reactions is achieved in accordance with the quantities and reactivities of reductants and oxidants in the system [e.g., organic matter and Mn(III,IV) (hydr)oxides, as modulated by pH and pe as a master variables].[15,22] The thermodynamics (energetics predicting the relative stability of reactants and products of a chemical reaction) of interconversions of Cr(III) and Cr(VI) compared to other redox couples can be used to predict the predominance of Cr(III) or Cr(VI) in water supplies (Figure 2). Certain electronpoor species may act as oxidants (electron acceptors) for Cr(III), especially soluble forms of Cr(III), in the treatment of water supplies or in soils enriched with Cr(III) (Nieboer and Jusys, Figure 3).[19] Examples are those above the bold line for Cr(VI)–Cr(III) on the Eh–pH diagram: Cl2, OCl–, H2O2, O3, and MnOOH. In contrast, electron-rich species may donate electrons to electron-poor Cr(VI) and reduce it to Cr(III): Fe2+ [or Fe(0)], H2S, H2, ascorbic acid (and organic compounds, generally), and SO2. Sunlight may affect the kinetics of both oxidation and reduction reactions for Cr, a relevant fact for natural processes in lakes and streams and for treatment technologies for drinking water purification. Depending on pH, temperature, and the concentrations of oxidants and reductants, Cr(VI)-to-Cr(III) ratio in natural waters may be predicted.

FIGURE 1 Seesaw model depicting a balance of the oxidation of Cr(III) by Mn(III,IV) (hydr)oxides and the reduction of Cr(VI) by organic compounds, with the pH acting as a sliding control (master variable) on the seesaw to set the redox balance for given quantities and reactivities of oxidants and reductants. The equilibrium quantity of Cr(VI) in the water is indicated by the pointing arrow from the fulcrum.

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FIGURE 2 Eh–pH diagram illustrating the stability field defined by Eh (redox potential relative to the standard hydrogen electrode, SHE) and pH for Cr(VI) and Cr(III) at 10–4 M total Cr. The vertical dashed lines indicate semiquantitatively the pH range in which Cr(OH)3 is expected to control Cr(III) cation activities in the absence of other ligands besides OH–.

FIGURE 3 Eh–pH diagram showing potential oxidants for Cr(III) in natural waters as dashed lines above the bold Cr(VI)–Cr(III) line and potential reductants for Cr(VI) below the line. Each line for an oxidant (first species of the pair) and reductant (second species) combination represents the reduction potential (in mV) at a given pH established by that oxidant–reductant pair (e.g., O3–O2). The oxidant member of a pair for a higher line is expected to oxidize the reductant member of the lower line, thereby establishing the area and species between the lines as thermodynamically favored to exist at chemical equilibrium.[23]

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201

conclusion Hexavalent chromium remaining in industrial by-products may contaminate soils, surface waters, and groundwater that are supplies for domestic uses, irrigation, and industrial processes. Prediction of the likelihood of Cr(III) oxidation and Cr(VI) reduction occurring is important for water treatment and for establishing health-based regulations and allowable limits for Cr(VI) and Cr(III) in water supplies. In agricultural soil–plant–water systems, Cr(VI) added in irrigation water or formed via oxidation of Cr(III) will reduce to Cr(VI) if electron donors (e.g., Fe2+, H2S, and organic matter) and Eh–pH conditions are sufficiently reducing (Bartlett and James[8] and James and Bartlett[20]; Figure 2). If not reduced, Cr(VI) may leach from surface soils to subsoils and groundwater. Therefore, prediction of Cr bioavailability and mobility in natural waters must consider redox reactions of this heavy metal.

References 1. US Environmental Protection Agency (EPA). Basic Information about Chromium in Drinking Water, 2010. Available at http://www.epa.gov/safewater/contaminants/basicinformation/chromium. html#four. (Accessed on May 13, 2012). 2. California Department of Public Health (DPH). Chromium-6: Timeline for Drinking Water Regulations, 2009. Available at: http://www.cdph.ca.gov/certlic/drinkingwater/Pages/Chromium6 timeline.aspx. (Accessed May 13, 2012). 3. International Agency for Research on Cancer (IARC). Chromium and chromium compounds. IARC Monogr. Eval. Carcinog. Risks Hum. 1990, 49, 49–256. 4. Stout, M.D.; Herbert, R.A.; Kissling, G.E.; Collins, B.J.; Travlos, G.S.; Witt, K.L.; Melnick, R.L.; Abdo, K.M.; Malarkey, D.E.; Hooth, M.J. Hexavalent chromium is carcinogenic to F344/N rats and B6C3F1 mice after chronic oral exposure. Environ. Health Perspect. 2009, 117 (5), 716–722. 5. National Toxicology Program (NTP). Toxicology and carcinogenesis studies of chromium picolinate monohydrate (CAS No. 27882-76-4) in F344/N rats and B6C3F1 mice (feed studies). TR556. 2010. Research Triangle Park, N.C. 6. Anderson, R.A. Essentiality of chromium in humans. Sci. Total Environ. 1989, 86, 75–81. 7. Kimbrough, D.E.; Cohen, Y.; Winer, A.M.; Creelman, L.; Mabuni, C.A. Critical Assessment of Chromium in the Environment. Crit. Rev. Environ. Sci. Technol. 1999, 29 (1), 1–46. 8. Bartlett, R.; James, B. Behavior of chromium in soils: III. Oxidation. J. Environ. Qual. 1979, 8, 31–35. 9. James, B.R.; Petura, J.C.; Vitale, R.J.; Mussoline, G.R. Oxidation–reduction chemistry of chromium: Relevance to the regulation and remediation of chromate-contaminated soils. J. Soil Contam. 1997, 6 (6), 569–580. 10. Nriagu, J.O. Production and uses of chromium. In Chromium in the Natural and Human Environments; Nriagu, J.O., Nieboer, E., Eds.; Wiley-Interscience: New York, 1988; 81–103. 11. Barnhart, J. Chromium chemistry and implications for environmental fate and toxicity. J. Soil Contam. 1997, 6 (6), 561–568. 12. Ball, J.W.; Nordstrom, D.K. Critical evaluation and selection of standard state thermodynamic properties for chromium metal and its aqueous ions, hydrolysis species, oxides, and hydroxides. J. Chem. Eng. Data 1998, 43, 895–918. 13. Bartlett, R.J. Chromium redox mechanisms in soils: Should we worry about Cr(VI)? In Chromium Environmental Issues; FrancoAngeli: Milan, 1997; 1–20. 14. James, B.R. Remediation-by-reduction strategies for chromate-contaminated soils. Environ. Geochem. Health 2001, 23 (3), 175–179. 15. James, B.R. The challenge of remediating chromium-contaminated soils. Environ. Sci. Technol. 1996, 30, 248A–251A.

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16. Fendorf, S.; Wielenga, B.W.; Hansel, C.M. Chromium transformations in natural environments: The role of biological and abiological processes in chromium(VI) reduction. Int. Geol. Rev. 2000, 42, 691–701. 17. Rai, D.; Saas, B.M.; Moore, D.A. Chromium(III) hydrolysis constants and solubility of chromium(III) hydroxide. Inorg. Chem. 1987, 26, 345–369. 18. Sass, B.M.; Rai, D. Solubility of amorphous chromium(III)–Iron(III) hydroxide solid solutions. Inorg. Chem. 1987, 26, 2228–2232. 19. Nieboer, E.; Jusys, A.A. Biologic chemistry of chromium. In Chromium in the Natural and Human Environments; Nriagu, J.O., Nieboer, E., Eds.; Wiley-Interscience: New York, 1988; 21–80. 20. James, B.R.; Bartlett, R.J. Behavior of chromium in soils: VII. Adsorption and reduction of hexavalent forms. J. Environ. Qual. 1983, 12, 177–181. 21. James, B.R. Redox phenomena. In Encyclopedia of Soil Science, 1st Ed.; Lal, R., Ed.; Marcel Dekker, Inc.: New York, 2002; 1098–1100. 22. James, B.R.; Brose, D.A. Oxidation - Reduction Phenomena. In Handbook of Soil Sciences: Properties and Processes. 2nd edn.; Huang, P.M., Li, Y., Summer, M.E., Ed.; CRC Press: Boca Raton, 2012; 14-1–14-24. 23. James, B.R. Chromium. In Encyclopedia of Water Science; Stewart, B.A.; Howell, T., Ed., Marcel Dekker: New York, 2003; 75–79.

20 Cobalt and Iodine Soil Cobalt ..................................................................................................... 203 Forms of Cobalt in Soils • Plant Availability of Soil Cobalt

Cobalt Deficiency ........................................................................................ 204 Soil Iodine...................................................................................................... 205 Forms of Iodine in Soils • Plant Availability of Soil Iodine

Ronald G. McLaren

Iodine Deficiency ......................................................................................... 205 Conclusions ...................................................................................................206 References ......................................................................................................206

Soil cobalt The Co concentration in soils depends primarily on the parent materials (rocks) from which they were formed and on the degree of weathering undergone during soil development.[2] Cobalt tends to be most abundant as a substituent ion in ferromagnesian minerals, and therefore has relatively high concentrations in mafic and ultramafic rocks (rocks containing high or extremely high proportions of ferromagnesian minerals). Conversely, Co concentrations are relatively low in felsic rocks (rocks containing large amounts of silica-rich minerals) such as granite, and in coarse-textured quartz-rich sedimentary rocks (sandstones). Higher concentrations of Co may be associated with finer textured sediments (shales) in which Co has become surface adsorbed by, or incorporated into, secondary layer silicates by isomorphous substitution.[3] Typical concentrations of Co reported in different rock types are shown in Table 1. As a result of the large range in Co concentrations of soil parent materials, and variation in the degree of weathering, total soil Co concentrations also vary widely. However, the mean values reported for agricultural soils from many countries appear to have a somewhat restricted range of between approximately 2 and 20 mg/kg (Table 1).

Forms of cobalt in Soils Cobalt in soils, whether released from parent materials during soil development, or derived from anthropogenic contaminant sources, occurs in several different forms or associations. Cobalt may be present as 1) the simple Co2+ ion, or as complexes with various organic or inorganic ligands in the soil solution; 2) exchangeable Co2+ ions; 3) specifically adsorbed Co, bound to the surfaces of inorganic soil colloids (clays and oxides/hydrous oxides of Al, Fe, and Mn); 4) Co complexed by soil organic colloids; 5) Co occluded by soil oxide materials; and 6) Co present within the crystal structures of primary and secondary silicate minerals.[6] In some soils, there appears to be a particularly strong association between Co and manganese (Mn) oxides, especially in soils where Mn oxides occur as distinct nodules or coatings.[7,8]

203

204 TABLE 1

Managing Global Resources and Universal Processes Cobalt and Iodine Concentrations in Rocks and Soils Co Concentration (mg/kg)

I Concentration (mg/kg)

Rock type Ultramafic (e.g., serpentinite) Mafic (e.g., basalt, gabbro) Intermediate (e.g., diorites) Felsic (e.g., granites, gneiss) Sandstones Shales/argillites Limestones Soils

100−300 30−100 1−30   Fe/Mn oxide bonded REEs.[9] The formation of bridged hydroxo complexes is probably the dominant sorption mechanism to clay minerals.[14] Clay type, pH, CEC, organic matter, and amorphous iron content regulate the adsorption kinetics of REEs.[1,2,9] Langmuir and Freundlich equations were found to describe precisely the absorption of REEs in soils.[15]

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translocation of Rare earth elements in Soils The question is still open whether the use of REEs in industry and agriculture may result in a pollution of soils, plants, and groundwater. In leaching experiments under controlled conditions with 141Ce and 147 Nd, these elements were abundant only in the top soil layer because of their strong adsorption.[9] For field conditions a translocation depth of 15 is considered as “sodic,” and in Australia, the criterion is >6. These differences are owing to the influence of several soil factors including salt levels, pH, organic matter, and clay mineralogy on the adverse effects of ESP on soil properties. Sodic soils are widespread in arid and semiarid regions of the world extending up to 30% of the total land area (Table 1). Use of saline water, including waste and effluent waters containing sodium salts, for irrigation induces sodicity in soils. Sodicity is a latent problem in many saltaffected soils where deleterious effects on soil properties are evident only when salts are leached below a threshold level.[2] While soil salinity reduces plant growth, directly affecting physiological functions through osmotic and toxicity effects on plants, sodicity causes deterioration of soil physical properties indirectly affecting plant growth and survival. Sodic soils are subjected to severe structural degradation and exhibit poor soil-water and soil-air relationships; these properties (Table 2) adversely affect root growth, thereby restricting plant production and making it difficult to work in soils when they are wet or dry.

Yield Decline in Sodic Soils Sodic soils make the paddocks prone to waterlogging, poor crop emergence and establishment, gully erosion, and in some instances tunnel erosion. Because of the heterogeneity in the accumulation of sodium by soil particles, these symptoms may be observed only in certain parts of the paddock. Generally, patchy growth and barren patches are visible in a number of spots in a paddock, while the rest of the field may look normal. However, the effects of sodicity are fully realized in the harvested yield. The actual yield obtained in sodic soils is often less than half of the potential yield expected on the basis of climate,

211

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Managing Soils and Terrestrial Systems TABLE 1

World Distribution of Sodic Soils

Continent

Country

North America

Canada U.S.A.

6,974 2,590

South America

Argentina Bolivia

53,139 716

Africa

Brazil

362

Chile

3,642

Algeria Angola

129 86

Botswana

670

Cameroon

671

Chad Ethiopia

5,950 425

Ghana

118

Kenya

448

Liberia Madagascar

44 1,287

Namibia

1,751

Niger

1,389

Nigeria

5,837

Somalia

4,033

Sudan

2,736

Tanzania South Asia

Area of Sodic Soils (000 ha)

583

Zimbabwe

26

Bangladesh India

538 574

Iran

686

North and Central Asia

China U.S.S.R.

437 119,628

Australasia

Australia

339,971

Source: Bui et al.[1]

particularly rainfall and evapotranspiration.[3,4] Relative yield of cereals grown in dryland sodic soils in Australia in relation to average root zone ESP is given in Figure 1. Swelling and dispersion of sodic aggregates destroy soil structure, reduce the porosity and permeability of soils, and increase the soil strength even at low suction (i.e., high water content). These adverse conditions restrict water storage and transport. Soils are, therefore, either too wet immediately after rain or too dry within a few days for optimal plant growth. Thus, the range of soil water content that does not limit plant growth and function (“nonlimiting water range”) is very small.[5] Dense, slowly permeable sodic subsoils reduce the supplies of water, oxygen, and nutrients needed for obtaining maximum potential yield. During the rainy season, even with prolonged ponding of water on the surface, only a small increase in water content occurs in subsoil. The low porosity leads to slow internal drainage and water redistribution within the profile.[6] This reduction in water storage causes crop water stress during prolonged dry periods. Subsoil as a source of water and nutrients becomes more important in dryland cropping regions than in irrigated soils.

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Sodic Soils: Properties TABLE 2

Physical and Chemical Properties of a Typical Sodic Soil Profile in South Australia

Properties

0–20 cm

20–40 cm

40–100 cm

8.9 3.8 0.6 14.6 2.8 22.0 1.2

9.2 4.9 0.3 24.5 4.5 38.5 2.6

Chemical properties pH1:5 (water) ECe (dS/m) Organic carbon (%) Exchangeable sodium (%) CaC03 (%) Boron (mg/kg) Water soluble Al(OH)4− (mg/kg)

7.9 0.4 1.2 6.2 0.1 1.2 0.0 Physical properties Spontaneously dispersed clay (%) 1.2 Swelling (mm/mm) 0.04 Hydraulic conductivity at saturation (mm/day) 22.8 Penetrometer resistance at lOOkPa suction (MPa) 1.8 Aeration porosity (%) 9.7 2.0 Bulk density (Mg/m3) Final infiltration rate in the field (mm/hr) 0.2

FIGURE 1

8.6 0.18 4.5 4.2 4.8 2.2

9.4 0.20 2.3 4.8 3.9 2.3

Relative yield of cereals grown in Australian sodic soils in relation to average root zone ESP.

Salt Accumulation in Root Zones of Sodic Soils Soils with sodic subsoils are characterized by moderate to high exchangeable sodium and, in many cases, with high pH (>8.5) where carbonate and bicarbonate minerals are present. Subsoil sodicity restricts drainage beyond the root zone and as a result salts accumulate in this zone. The concentration of accumulated salts fluctuates with rainfall pattern, input of salt from agronomic practices, and soil weathering, as schematically explained in Figure 2. Dryland salinity or “seepage salinity” in many countries is associated with rising saline groundwater tables. However, the extent of subsoil salinity, also called “transient salinity,” not associated with saline groundwater is large in many landscapes dominated by subsoil sodicity. A relationship between rainfall, subsoil ESP, and ECe for northeastern Australian soils has been reported.[7] In dryland regions with annual rainfall between 250 and 600 mm, sodic subsoils have an ECe between 2 and 20 that can dramatically affect crop production through osmotic effects during dry periods. Laboratory measured ECe will increase several folds under field conditions as the soil layers dry in between rainy days. The combination of poor water storage and osmotic stress enhance water stress to crops under dryland cropping.

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Root Zone constraints in Dryland Sodic Soils Multiple problems occur in soils with subsoil sodicity. Soil compaction, crusting, and induration of subsoil layers re quire “physical” reclamation. Sodicity, salt accumulation, and alkaline pH require “chemical” reclamation. All of these conditions cause, in addition to water stress, macro- and micronutrient deficiency, and toxicity owing to Na+, Cl−, HCO3−, CO32−, B, Al(OH)4−, and others. Low organic matter and biological activity compound these problems encountered in sodic subsoils.

Management of Dryland Sodic Soils Major criteria in increasing productivity in dryland sodic soil are improved water storage and transport in the root zone and crop water use efficiency. More information is available on agricultural management in sodic soils that is more relevant to irrigated lands.[6] Reclamation procedures involving high costs are prohibitive in dryland regions because of low benefit/cost ratio. Diagnosis of multiple problems with large variations, vertically and horizontally across the paddock, is primarily important. Gypsum is the most commonly used compound to reclaim sodic soils. Subsoil reclamation may involve higher rates of gypsum application or deep placement of gypsum by deep ripping or deep plowing. Salt-tolerant plant species may alleviate subsoil salinity. Plants that can tolerate ion toxicity such as boron, carbonate, sodium, and chloride have also been identified. Strategies to improve subsoil fertility may include 1) mechanical means of placing nutrients deeper in the profile; 2) using nutrient sources of lower or higher mobility; 3) using deep-rooted legumes to fix nitrogen at depths; and 4) selection of plant species and genotypes better suited to acquiring nutrients from subsoils. Future research is needed on developing plants that modify the rhizosphere and adapt to edaphic conditions. Farming systems should be developed to prevent accumulation of salts and toxic elements in the root zone of sodic soils.

FIGURE 2

Salt accumulation in sodic subsoils.

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215

conclusions Soils with a high proportion of exchangeable sodium are considered as sodic soils. On wetting these soils, clay particles swell and disperse degrading soil structure, and on further drying soils become dense. Poor water storage and restricted movement of air and water in soil profile lead to yield decline of many crops. Sodic subsoils cause accumulation of salts in the rootzone. In dryland sodic soils multiple problems such as salinity and toxic elements in addition to sodicity occur making management decisions difficult. Soil amelioration with gypsum and choice of tolerant crops are useful in farming these soils.

References 1. Bui, E.N.; Krogh, L.; Lavado, R.S.; Nachtergaele, F.O.; Toth, T.; Fitzpatrick, R.W. Distribution of sodic soils: the world scene. In Sodic Soils: Distribution, Properties, Management and Environmental Consequences; Sumner, M.E., Naidu, R., Eds.; Oxford University Press: New York, 1998; 19–33. 2. Rengasamy, P.; Olsson, K.A. Sodicity and soil structure. Aust. J. Soil Res. 1991, 31, 821–837. 3. French, R.J.; Schultz, J.E. Water use efficiency of wheat in a mediterranean-type environment. 1. The relation between yield, water use and climate. Aust. J. Agric. Res. 1984, 35, 765–775. 4. Rengasamy, P. Sodic soils. In Methods for Assessment of Soil Degradation; Lal, R., Blum, W.H., Valentine, C., Stewart, B.A., Eds.; CRC Press: New York, 1997; 265–277. 5. Letey, J. The study of soil structure: science or art. Aust. J. Soil Res. 1991, 29, 699–707. 6. Oster, J.D.; Jayawardane, N.S. Agricultural management of sodic soils. In Sodic Soils: Distribution, Properties, Management and Environmental Consequences; Sumner, M.E., Naidu, R., Eds.; Oxford University Press: New York, 1998; 125–147. 7. Shaw, R.J.; Coughlan, K.J.; Bell, L.C. Root zone sodicity. In Sodic Soils: Distribution, Properties, Management and Environmental Consequences; Sumner, M.E., Naidu, R., Eds.; Oxford University Press: New York, 1998; 95–106.

26 Soil Degradation: Global Assessment

Ahmet Çilek, Suha Berberoğlu, Erhan Akça, Cenk Dönmez, Mehmet Akif Erdoğan, Burçak Kapur, and Selim Kapur

Introduction .................................................................................................. 217 Background ................................................................................................... 217 Europe ............................................................................................................ 219 Asia ................................................................................................................. 221 Africa.............................................................................................................. 223 South America .............................................................................................. 224 North America, Central America, and the Caribbean Islands ............... 226 Australia.........................................................................................................228 Conclusions ................................................................................................... 230 References ...................................................................................................... 231

introduction One of the major challenges facing humanity in the coming decades is the need to increase food production in limited soil sources to cope with the ever-increasing population, which is indeed the greatest threat for food/soil security.[1] Although estimates on the prevalence on land degradation vary from 1 to 6 billion ha,[2] Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES)[3] reported that the worldwide degraded land is reached to 75% which now threatens 3.2 billion of people well-being in spite of the strict laws enacted for environmental protection. This is an unpalatable but important indicator of high external input dependent on global food, fiber and biofuel production, and a corresponding increase in soil/land damage, reflecting the expensive agricultural production.[4] Global efforts are in an increasing trend for achieving sustainable soil use in the world and the Global Soil Partnership established in 2012 as a mechanism to develop a strong interactive partnership and enhanced collaboration and synergy of efforts between all stakeholders for the protection of the soils against main soil degradation threats.[5] It is based upon the concept of soil multifunctionality and recognizes the various soil functions relevant for global economy and the environment. These functions are under threats of losses in soil organic carbon and soil biodiversity, soil sealing, nutrient imbalance, salinity-alkalinization, and pollution that are severely limiting soil functionality and therefore socioeconomic growth. Soils in the world are recognized as being the result of millennia of human interaction with the landscape and have never been as fragile as today.[6]

Background Soil can even degrade without actually eroding. It can lose its nutrients and soil biota, and can become damaged by waterlogging and compaction. Erosion is only the most visible part of degradation, where the forces of gravity, water flow, or wind actively remove soil particles. Rather than taking the classical 217

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view that soil degradation was, is, and will remain an ongoing process mainly found in countries of the developing world, this phenomenon should be seen as a worldwide process that occurs at different scales and different time frames in different regions.[7] The causes of biophysical and chemical soil degradations are enhanced by socioeconomic interventions, which are the main anthropogenic components of this problem, together with agricultural mismanagement, overgrazing, deforestation, overexploitation, loss of soil organic carbon soil sealing, and pollution as the main reasons for erosion and chemical soil degradation.[5] Soil degradation, the threat to “soil security,” is ubiquitous across the globe in its various forms and at varying magnitudes, depending on the specific demands of people and the inexorably increasing pressures on land. Europe provides many telling examples of the fragile nature of soil security and the destructive consequences of a wide range of soil degradation processes. Asia, Africa, and South and North America are not only partly affected by the non-resilient impacts of soil degradation but also experiencing subtler destruction of soils via political developments, which seek to provide temporary relief and welfare in response to the demands of local populations. Several diverse data sets were used in this research (Table 1). The modeling of land degradation adopted here is based on soil types, water and wind erosion, land cover, terrestrial net primary productivity (NPP), percent tree cover, aridity, and chemical deterioration. The global soil map derived by the United States Department of Agriculture/Natural Resources Conservation Service (USDA-NRCS, Soil Survey Division) was used as one of the inputs in this study. This map was produced in September 2005 and was based on a reclassification of the Food and Agriculture Organization of the United Nations Educational, Scientific and Cultural Organization (FAO-UNESCO) Soil Map of the World combined with a soil climate map. The soil map shows the global distribution of major soil types. The scale of the global soil map is 1:5,000,000. The global water erosion vulnerability map was based on a reclassification of the global soil climate map and global soil map of the USDA-NRCS. The global soil climate map comprises the major biome types at the global scale, which was derived using global land cover maps of the European Space Agency. The global water erosion vulnerability map consisted of four vulnerability classes of the water. These classes were also based on soil climate and soil classification of the USDA-NRCS, Soil Survey Division. This map was produced in 1998 with a scale of 1:5,000,000. The global wind erosion vulnerability map was derived using soil and soil climate maps of the USDA-NRCS, Soil Survey Division. The map was 1:5,000,000 scale and comprised four vulnerability classes. The global land cover map was derived by ESA Climate Change Initiative—Land Cover led by UCLouvain. The land cover map was derived using AVHRR HRPT (1992–1999), SPOT-Vegetation (1999–2012), and PROBA-V (2013–2015) at 1 km spatial resolution, and comprises 22 land cover classes TABLE 1

The Data Sets Used in the Modeling of Land Degradation

Data Sets Global soil map Global water erosion vulnerability Global wind erosion vulnerability Global saline/alkaline domains Global land cover map Net primary production Percent tree cover Aridity map Chemical deterioration map

Source of Data United States Department of Agriculture (USDA) USDA USDA USDA ESA Climate Change Initiative—Land Cover led by UCLouvain (2017). The Land Cover CCI Climate Research Data Package (CRDP) NASA Earth Observations Geospatial Information Authority of Japan, Chiba University, and collaborating organizations CGIAR-CSI (International Research Center and CGIAR – Consortium for Spatial Information) LADA

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defined according to the UN Land Cover Classification System. This set of cover types included 11 categories of natural vegetation covers.[8] The global NPP map indicated the net amount of carbon from the atmosphere into green plants. NPP is an important ecological variable that measures the energy input to the biosphere and terrestrial carbon dioxide assimilation and shows ecosystem performance. The Global NPP map was gathered from The Oak Ridge National Laboratory Distributed Active Archive Center (ORNL DAAC) for biogeochemical dynamics, which is one of the National Aeronautics and Space Administration-Earth Observing System Data and Information System (NASA-EOSDIS) data centers managed by the Earth Science Data and Information System (ESDIS) Project. This map was derived by using the NASA-CASA (Carnegie Ames Stanford Approach) NPP Model Approach. The percent tree cover map was gathered from the Global Land Cover Facility/University of Maryland. It illustrates the global cover of woody vegetation, on a continuous scale from 0% to 100%. Advanced very high-resolution radiometer (AVHRR) images were used to derive the percent tree cover map at 1 km spatial resolution. The percent tree cover was estimated using the linear regression method. This map comprises three classes, namely, the (1) 10%–80% tree cover, (2) non-vegetated land, and (3) tree cover less than 10%.[9] The global aridity index map was gathered from the -Consultative Group for International Agricultural Research – Consortium for Spatial Information (CGIAR-CSI). Aridity is expressed as a function of precipitation, potential evapo-transpiration (PET), and temperature, and is classified according to the climatic zones proposed by the UNEP.[10] The Aridity Index is used to quantify the precipitation deficit over atmospheric water demand. The Aridity Index map was derived using MODIS images with 1 km spatial resolution. The climate data sets were obtained from the WorldClim data set,[11] and the map was estimated with the following equation: Aridity Index ( AI) = MAP / MAE, where MAP is the mean annual precipitation and MAE is the mean annual evapotranspiration. The chemical deterioration map was obtained from the Land Degradation Assessment in Drylands (LADA) organization. The chemical deterioration map used in this study comprised four main classes: (1) fertility decline and reduced organic matter content (not by erosion)—leaching, nutrient mining, oxidation, and volatization (N); (2) acidification—decreased soil pH; (3) soil pollution—soil contamination with heavy metals and toxic components; and (4) salinization/alkalinization—a net increase in salt and sodium contents of the (top) soil leading to a productivity decline. Standardization and weighting of these inputs were the two critical steps while modeling the land degradation within a Geographical Information System (GIS) environment. The inputs were scaled between 0 and 1 in standardization using sigmoidal functions. The fuzzy approach was used in standardization to evaluate the possibility that each pixel belongs to a fuzzy set by evaluating any of a series of fuzzy set membership functions. The sigmoidal, J-shaped, and linear functions were controlled by four points ordered from low to high on the measurement scale. Factor-by-factor weighting was done (i.e., weak NPP will have the higher score and risk than dense NPP areas) following standardization. All factors were evaluated through overlapping each other.

europe The major problems concerning the soils of Europe are the quality and quantity loss of soils due to erosion, sealing/urbanization, flooding, loss of fertility due to deep ploughing, removing crop residues, mono-culture, decline in organic carbon and biodiversity along with local and diffuse soil contamination especially in industrial and urban areas, and soil acidification (Figure 1).[12,13] Salinity, although not widely distributed in Europe, is seen in some parts of Western and Eastern Europe,[14] whereas urbanization and construction of infrastructure at the expense of fertile land are widespread in Europe,

220

FIGURE 1

Managing Soils and Terrestrial Systems

Soil degradation map of Europe.

particularly in the Benelux countries, France, Germany, and Switzerland, and such effects are most conspicuously destructive along the misused coasts of Spain, France, Italy, Greece, Turkey, Croatia, and Albania. The EU average soil sealing from 2000 to 2006 was 3%, but this figure raises to approximately 14% in Ireland, Cyprus, and Spain.[15] The drastic increase in the rate of urbanization since the 1980s is now expected to follow the Plan Bleu,[16] which seeks to create beneficial relationships between populations, natural resources, the main elements of the environment, and the major sectors of development in the Mediterranean Basin and to work for sustainable development in the Mediterranean region. The very appropriate term “industrial desertification” remains valid for the once-degraded soils of Eastern Europe, under the pressure of mining and heavy industry, as in the Ukraine where such land occupies 3% of the total land area of the country.[17] There are three broad zones of “natural” erosion across Europe, including Iceland: (1) the southern zone (the Mediterranean countries); (2) a northern loess zone comprising the Baltic States and part of Russia; and (3) the eastern zone of Slovenia, Croatia, Bosnia-Herzegovina, Romania, Bulgaria, Poland, Hungary, Slovakia, the Czech Republic, and Ukraine (Figure 1). Seasonal rainfalls are responsible for severe erosion due to overgrazing and the shift from traditional crops. Erosion in Southern Europe is an ancient problem and still continues in many places, with marked on-site impacts and with significant decreases in soil productivity as a result of soil thinning. The northern zone of high-quality loess soils displays moderate effects of erosion with less intense precipitation on saturated soils. Local wind erosion on light textured soils is also responsible for the transportation of agricultural chemicals used in the intensive farming systems of the northern zone to adjacent water bodies, along with eroded sediments. The high erodibility of the soils of the eastern zone is exacerbated by the presence of large statecontrolled farms that have introduced intensive agriculture at the expense of a decrease in the natural vegetation. Contaminated sediments are also present in this zone, particularly in the vicinity of former industrial operations/deserts, with high rates of erosion in Ukraine (41% of the total land area) and Russia (57% of the total land area) whose agricultural land has been subjected to strong water and wind erosion ever since the beginning of industrialization.[17] Localized zones of likely soil contamination through the activities of heavy industry are common in northwestern and central Europe as well as Northern Italy, together with more scattered areas of known and likely soil contamination due to the intensive use of agricultural chemicals.[21] Sources of

221

Soil Degradation: Global Assessment 45,00,000 40,00,000

Area (km2)

35,00,000 30,00,000 25,00,000 20,00,000 15,00,000 10,00,000 5,00,000 0

FIGURE 2

Very low

Low

Moderate Degradation

High

Very high

Distribution of soil degradation rates in Europe.

contamination are especially abundant in the “hot spots” associated with urban areas and industrial enclaves in the northwestern, southern, and central parts of the continent (Figure 1). Acidification through deposition of windborne industrial effluents and aerosols has been a long-standing problem for the whole of Europe; however, this is not expected to increase much further, especially in Western Europe, as a result of the successful implementation of the emission-control policies over 40 years.[18] The desertification of parts of Europe has been evident for some decades, and the parameters of the problem are now becoming clear, with current emphasis on monitoring of the environmentally sensitive areas[19] on selected sites, seeking quality indicators for (1) soil, (2) vegetation, (3) climate, and (4) human management throughout the Mediterranean Basin. Apart from the human factor, these indicators are inherent. Nevertheless, the majority of the lands/soils in Europe are under the threat of low and moderate degradation levels as compared with other continents that have suffered extensive and very high degradation, especially through erosion, salinity, and loss of nutrients (Figure 2).

Asia The most severe aspect of soil degradation on Asian lands has been desertification due to the historical, climatic, and topographic character of this region as well as the political and population pressures created by the conflicts of the past 500 years or so. At least US $10 billion is lost annually as a result of losses resulting from land degradation in India, Pakistan, Bangladesh, Iran, Afghanistan, Nepal, Sri Lanka, and Bhutan. FAO, UNDP, and UNEP estimated that this figure is equal to 2% of the region’s Gross Domestic Product, which is about 7% of the value of its agricultural output.[20] Salinization caused by the rapid drop in the level of the Aral Sea and the waterlogging of rangelands in Central Asia due to the destruction of the vegetation cover by overgrazing and cultivation provide the most striking examples of an extreme version of degradation—desertification caused by the misuse of the land. Soil salinity has had severe effects at the northern and western parts of the Caspian Sea, north of India and Pakistan, Laos, Myanmar, and Thailand[21] mainly due to the shift to irrigated agriculture and destruction of the natural vegetation (Figure 3). A major threat to the Asian environment has occurred through the accumulation of soluble salts, mainly deposited from saline irrigation water or through mismanagement of available water resources, as in the drying Aral Sea and the Turan lowlands as well as the deterioration of the oases in Turkmenistan, with excessive abstraction of water in Central Asia (Figure 3).[21] Soil salinity costs Uzbekistan about US $1 billion annually.[21] Moreover, as soil salinity decreases plant productivity, it has negative effect on soil organic carbon inputs which lead to low soil organic carbon contents.

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FIGURE 3

Soil degradation map of Asia.

The drylands of the Middle East have been degrading since the Sumerian epoch, with excessive irrigation causing severe salinity and erosion/siltation problems,[22] especially in Iraq, Syria, and Saudi Arabia. Iraq has been unique in the magnitude of the historically recorded buildup of salinity levels, with 8.5 million ha saline land, which is 64% of the total arable land surface (i.e., 90% of the land in the southern part of that country). The historical lands of Iran, Pakistan, Afghanistan, India, and China are also subject to ancient and ongoing soil/land degradation processes, which are subtle in some areas but evident and drastic in others (Figure 3). All ongoing contemporary loss of soil and land in the Asian continent is at an increasing trend even to the presence of the extensive untouched natural forest areas of the north of the continent (Figure 4). 2,50,00,000

Area (km2)

2,00,00,000

1,50,00,000

1,00,00,000

50,00,000

0

Very low

Low

Moderate Degradation

FIGURE 4

Distribution of soil degradation rates in Asia.

High

Very high

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223

Africa Africa’s primary past and present concern has been the loss of soil productivity by nutrient depletion, i.e., the decreasing NPK levels (in kilograms per hectare) in cultivated soils following the exponential growth in population and the resulting starvation and migrations at high levels in the northern, central, and southern parts of the continent (Figure 5). Intensification of land use to meet the increased food demands combined with the mismanagement of the land leads to the degradation of the continental soils. Central African forests are converted to low productive agricultural lands and are used for charcoal and fuelwood production. Moreover, urbanization and uncontrolled mining are other major threats to African forests. This poses the ultimate question of how the appropriate sustainable technologies that will permit the increased productivity of the soils can be identified. This problem is illustrated by the example of the Sudan, where nutrient depletion has steadily increased through

FIGURE 5

Soil degradation map of Africa.

224

Managing Soils and Terrestrial Systems 1,20,00,000

1,00,00,000

Area (km2)

80,00,000

60,00,000

40,00,000

20,00,000

0

FIGURE 6

Very low

Low

Moderate Degradation

High

Very high

Distribution of soil degradation rates in Africa.

more mechanized land preparation, planting and threshing without the use of inorganic fertilizers, and legume rotations. Thus, aggregate yields have been falling as it became more difficult to expand the cultivated area without substantial public investments in infrastructure. In Burkina Faso, and in the other lands of Africa located around the southern and northern fringes of the Saharan desert and the desert margins, the decreased infiltration and increased runoff causing erosion is a further consequence of repeated cultivation. Thus, the technological measures to be identified for Burkina Faso and similarly to other African countries must be smart agriculture. This includes development of water retention technologies, while polyculture/rotations with proper manuring and fertilization for cost-efficient provisions of N and P by preferably green manuring are major actions to be taken. These actions in the Sudan and likewise other parts of Africa will provide balanced management of soil moisture, nutrients, and organic matter by enhancing C-sequestration in soil, which is the main goal for sustainability based on the sustainable management of the land, to ensure the security of both the soil and the global climate.[23,24] Salinity is progressively menacing the lands of the African continent by the increased salt accumulation in the soils of the Sahel (especially in Libya) and parts of the Sahara, which is irrigated and/or excess irrigated by fossil water. The land area of the African continent that is highly and very highly affected by land/soil degradation is estimated to be above 17 million km 2, 1.7 times larger than 10 million km 2 Europe continent, which is mainly caused by depletion of the nutrients and increase of soil salinity (Figure 6). The slight erosion determined on the Saharan lands reflects a mere reality, i.e., the ultimate fate of the land depends on strict conservation measures, as well as detailed research on any fossil water considered for use. Only by such means can stable ecosystems be conserved in such water–soil poor regions.

South America Water and wind erosion on the eastern and western coastal/ inland areas of the South American continent are the dominant soil degradation processes in these regions and have caused the loss of the topsoil at alarming rates due to the prevailing climatic and topographic conditions. Almost as important is the

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loss of nutrients from the Amazon Basin[13] (Figure 7). Graesse et al.[25] calculated 17% of new cropland and 57% of new pastureland replaced forests in Latin America from 2001 to 2013. These deforestations and overgrazing effects caused the degradation of 576 million ha of potential agricultural land, an area almost equal to the very high, high, and moderately high degraded land/soil areas (Figure 8). Another important factor has been the ever-increasing introduction of inappropriate agricultural practices derived from the so-called imported technology, which has not been properly adapted to indigenous land-use procedures. The traditional methods of permitting the land to recover naturally have been almost totally abandoned and has been replaced by unsuitable technological measures designed to maintain production levels (temporarily) and to overcome the loss of soil resilience, thus increasing chemical inputs. The rapid industrialization/urbanization of the limited land resources in the Caribbean has been expelling agricultural communities to remote and marginal regions that are at present rich in

FIGURE 7

Soil degradation map of South America.

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Managing Soils and Terrestrial Systems 1,00,00,000 90,00,000 80,00,000

Area (km2)

70,00,000 60,00,000 50,00,000 40,00,000 30,00,000 20,00,000 10,00,000 0

Very low

Low

Moderate

High

Very high

Degradation

FIGURE 8

Distribution of soil degradation rates in South America.

biodiversity and biomass—a major global C sink. Moreover, large-scale livestock herding in Central and South America is also a major threat to soil security and has been responsible for degrading 35% of the pasture lands of Argentina.

north America, central America, and the caribbean islands The most prominent outcome of soil/land degradation (or more correctly desertification) in the United States is exemplified by the accelerated dust storm episodes of the 1930s—the Dust Bowl years, marked by the “Black Blizzards,” which were caused by persistent strong winds, droughts, and overuse of the soils. These resulted in the destruction of large tracts of farmland in the south and central United States. Recently, salinization has become an equally severe problem in the western part of the country (Figure 9) through the artificial elevation of water tables by extensive irrigation, with associated acute drainage problems. An area of about 10 million ha in the west of the United States has been suffering from salinity-related reductions in yields, coupled with very high costs in both the Colorado River basin and the San Joaquin Valley.[26] Unfortunately, new irrigation technologies such as the Center-Pivot irrigation system (developed as an alternative to the conventional irrigation systems that caused the salinity problems) has caused a decline of the water table levels in areas north of Lubbock, Texas, by around 30–50 m, leading to a dramatic decrease (by 50%) in the thickness of the well-known Ogallala aquifer. In some areas, this has been followed by ground subsidence, which is an extreme form of soil structure degradation, i.e., loss of the physical integrity of the soil. Loss of topsoil, as the result of more than 200 years of intensive farming in the United States, is estimated to vary from 25% to 75% and exceeds the upper limit in some parts of the country.[27,28] The United States provides good examples of the difficulties involved in erosion control, with its large-scale intensive agriculture damaging soil structure and also increasing the erosion of susceptible soils. This problem could be greatly mitigated by strict enforcement of the no-till system. Such no-till areas have increased tremendously in the United States and are forecast to increase in linear fashion.[29] Conservation farming is practiced in only about half of all US agricultural land and on less than half of the country’s most erodible cropland. Conservation farmers are encouraged to use only the basic types of organic fertilizers, such as animal and green manure together with compost, mulch farming, improved pasture management, and crop rotation to conserve soil nutrients. In Alaska, scientists have identified a

Soil Degradation: Global Assessment

FIGURE 9

227

Soil degradation map of North and Central America and the Caribbean islands.

number of events that may have been the cause of a severe and rapid soil erosion, where the northern ice sheet is diminishing and consequently causing the increase of water temperatures, which means the increased development of storm events, more precipitations, and higher waves with a potential of eroding some 14 m of coastline each year, which is a much higher rate than the 6.1 m erosion recorded in the mid-1950s.[30] Jones et al. state that this might be a new era in ocean–land interactions and the reshaping of the Arctic coastline by erosion. Furthermore, they claim that the warming sea-surface temperatures and the rising sea level will most probably act to reduce the permafrost-dominated coastline and thaw the ice-rich bluffs and consequently increase erosion and degrade Arctic landscapes together with the long-standing cultural sites. Canada is a large country where half of the 68 million ha of available land is cultivated, with an average farm size of 450 ha. It is reported that Canada has experienced annual soil losses on the prairies (through wind and water erosion) that are similar to the Asian steppes, amounting, respectively, to 60 and 117 million tons. These annual rates are much higher than the rate of soil formation, resulting in an annual potential grain production loss of 4.6 million tons of wheat. With regard to primary soil salinity, during historic times, the prairies have experienced steady increases, related partially to increasing groundwater levels. Major problems of secondary salinity are estimated to affect 2.2 million ha of land in Alberta, Saskatchewan, and parts of Manitoba, with an immense economic impact each year. Central America is drastically suffering from soil erosion, where almost 26% of the territory is strongly affected via this irreparable menace caused mainly by overgrazing, which almost adds up to a total of 58% of the overall degradation in the southern tip of North America (Figure 10). Concrete precautions are being taken in the whole region by implementing the “Quesungual” agro-forestry method, concerning the planting of trees between the dispersed natural forest tree communities, converting pastures to forests (World Bank Project), and implementing ambitious projects such as the Lake Managua Watershed drainage to avoid landslides—a major component of degradation in Nicaragua.[31]

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Area (km2)

1,00,00,000 80,00,000 60,00,000 40,00,000 20,00,000 0

Very low

Low

Moderate

High

Very high

Degradation FIGURE 10

Distribution of soil degradation rates in North and Central America.

The Caribbean islands and the lands of Central America have long faced a common fate of soil degradation via intense conversion of forests to cultivated soil, i.e., deforestation, causing flooding, landslides, and development of land surfaces prone to erosion and degrading the cycles of carbon and water—the wonders that sustain our living globe. Costa Rica loses about 860 million tons of valuable topsoil every year.[32] Earthquakes of the Caribbean and Central America were/are followed primarily by volcanic eruptions with immediate and dramatic changes in the environment, where lava buries soils that would be covered for at least a generation-long period before developing into fertile lands rich in nutrients. Drastic examples reveal the inappropriate soil use systems imposed by the colonial powers of Europe back in the 16th century based on the creation of extensive deforestation areas replaced by sugar cane fields, which has been the beginning of the soil degradation era, particularly in Haiti in the Caribbean islands. Precautions to combat the vast ongoing soil erosion in the Caribbean islands, especially in Haiti, seek the development of tree plantations and in particular bamboo crops to overcome the soildegradation- and political-unrest-oriented contemporary poverty in the country. The high rates of erosion, as sediment depositions in the sea, are also threatening one-third of the coral reefs around the Caribbean islands. High turbidity of inshore water and elevated algal cover on reefs are linked to the impacts of coastal development inducing erosion, with soil/sediment deposition being a major influence on the conditions of reefs.[33,34] The Caribbean Environment Program is active in the region, seeking to mitigate soil and land degradation by developing projects primarily on the “Best Management Practices in Agriculture” and “Integrating Watersheds and Coastal Areas Management—IWCAM” via the Global Environment Facility (GEF) included in National Programs of Action.[35]

Australia The Australian agricultural/soil resource base has been long endangered by the “business as usual” concept on the continent aiming to achieve temporary economic betterment,[36] which, in contrast, induced soil degradation. Identification of the different types of soil degradation in Australia reveals that erosion has been the main component, primarily via dust storms, which still are a serious problem, especially where cropping practices do not include retention of cover and minimum tillage methods. Water erosion effects are also particularly severe in areas of summer rainfall and topographic extremities (Figure 11). Remedial actions for this include the well-known measures of maintaining adequate

Soil Degradation: Global Assessment

FIGURE 11

229

Soil degradation map of Australia and New Zealand.

cover and changing prevailing altitudes towards stock management, storage feed, redesign of watering sites, and management of riparian areas. Part of the excess salinity in Australia is of primary origin and was retained in the subsoil by trees, which have now been cleared to create soil surfaces for cropping and pastures, allowing penetration of water to the saline subsoil, then followed by abstraction from the water table, thus leading to serious problems in the southwest, east, and central-eastern parts of the country (Figure 11). About 30% of Australia’s agricultural land is sodic, creating poor physical soil conditions and impeded productivity. This problem can only be alleviated by massive revegetation programs and by taking extra care of the water table and plant cover. Despite the introduction of costly conventional measures for reclamation, salinity levels continue to increase across Australia in the dry and irrigated soils. The dryland salinity in the continent affects about 2.5 million ha of farmland and is expanding at a rate of 3%–5% yr−1 which may sum up to 8.8 million ha (33%) by 2050.[37] The retardation of the organic matter levels also requires remediation measures, with economically justified fertilizer use strategies to be utilized throughout the continent. Moreover, overgrazing has resulted in the impoverishment of plant communities and loss of habitats as well as the decline in the chemical fertility of the soil by progressive depletion of organic matter in the topsoil, followed by deterioration in soil structure. Acidification caused by legume-based (clover, etc.) mixed farming plus use of ammonia-based fertilizers threatens 55 million ha of Australian land, which is about one-half of the degraded land areas (high, moderate, and very high, respectively) of the continent (Figure 12). Liming seems to be the most effective present remedy, but this is costly, does not lead to rapid recovery, and is impractical for subsoil acidity. Thus, the precise remedies have yet to be developed for the conditions on this continent, utilizing careful, long-term monitoring and the experience of farmers to devise specific treatment and conservation procedures. Despite the common view that reflects the absence of soil loss in

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Managing Soils and Terrestrial Systems 60,00,000 50,00,000

Area (km2)

40,00,000 30,00,000 20,00,000 10,00,000 0

FIGURE 12

Very low

Low

Moderate Degradation

High

Very high

Distribution of soil degradation rates in Australia and New Zealand.

New Zealand due to the presence of the widespread forests, soil erosion has been a major threat to the natural resources as being accelerated by early colonial and contemporary deforestation followed by the unwise land-use practices such as overgrazing. Accelerated erosion is the most serious and the least reversible of soil degradation problems encountered in the country. The main forms of erosion in New Zealand are, namely, the mass, fluvial, surface, and stream-bank transportation of the soil in the hilly country, in the Marlborough and Manawatu regions, in the lands outside the hill country, and at the unstable areas cleared of tree cover. The long-standing experience and research provided the relevant bodies dealing with erosion in New Zealand, with the appropriate techniques to reduce the impacts of erosion in the pastoral lands. These techniques include the maintaining of the adequate vegetative cover (e.g., avoiding overgrazing and maintaining a dense pasture sward through regular applications of fertilizer and grass seed), the spaced or close tree planting, retiring the land from pasture, fencing off and planting riverbanks, and building debris dams to slow water flows in gullies. A similar system to the Australian “National Landcare Organisation” is being implemented in New Zealand in 1996 as the “New Zealand Landcare Trust” which is a nongovernmental organization facilitating sustainable land management (SLM) and biodiversity initiatives at a community-driven basis, in order to effective catchment management to improve the land and water quality of the country. The Landcare Trust seeks the assurance of implementing land management more sustainably.[38]

conclusions The state of soil/land degradation and its remediation through a multifunction–multi-impact approach has been defined through a driving force–pressure–state–impact–response (DPSIR) matrix developed by the European Environmental Agency[39] (Figure 13) leading to SLM[40–42] measures to be taken for the future. SLM is concerned with more soil/land-friendly farming practices that minimize the erosion potential of soils, together with the adoption by landholders of property management planning procedures that involve community actions, such as the Landcare Program of Australia.[38] Moreover, as Smyth and Dumanski[40] have stated, this approach combines socioeconomic principles with environmental concerns to enable production to be enhanced alongside a reduction in the level of risk to the sustenance of natural resources. Thus, degradation of soil and water quality, which have negative

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231

FIGURE 13 DPSIR framework applied to soil. Source: UNEP/EEA.[39]

effects on climate[43] and biodiversity, can be prevented or minimized in a manner that is acceptable to the farming community. The methods to be adapted for SLM via community actions include contour farming, terracing, vegetative barriers, and other land-use practices amalgamated with indigenous (traditional) technical knowledge (ITC)[44,45] as applied to farming and landscape preservation. The impetus for the use of ITK by scientists and local communities in creating new strategies for sustainable resource management was provided in the United Nations Conference on Environment and Development (UNCED) held in Rio de Janeiro (Brazil) in 2012. Thus, at the 13th meeting of the Parties to the United Nations Convention to Combat Desertification, 113 countries had agreed to specify solid targets with clear indicators, to rehabilitate degraded lands and reverse degradation, which currently affects more than third of the world’s prime land resources.[46] This approach was widely accepted and endorsed as the Land Degradation Neutrality approach for a primary solution to the degradation of global soil/land resources in the new UNCCD 2018–2030 Strategic Framework[47] in order to meet UN Sustainable Development Goals aimed to be achieved by 2030 for the global welfare and peace.

References 1. Alexandratos, N. World food and agriculture: the outlook for the medium and longer term. NAS Colloquium. Plants and Population: Is There Time? Beckman Center of the National Academy of Sciences: UC Irvine, 5–6 December 1998. 2. Gibbs, H.K.; Salmon, J. M. Mapping the world’s degraded lands. Applied Geography, 2015, 57, 12–21. 3. IPBES. Thematic Assessment of Land Degradation and Restoration. Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services: Medellin, 2018.

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4. Dowdall, C.M.; Klotz, R.J. Pesticides and Global Health: Understanding Agrochemical Dependence and Investing in Sustainable Solutions. Routledge: New York, 2016. 5. FAO and ITPS. Status of the World’s Soil Resources (SWSR)—Main Report. Food and Agriculture Organization of the United Nations and Intergovernmental Technical Panel on Soils: Rome, 2015. 6. Cherlet, M.; Hutchinson, C.; Reynolds, J.; Hill, J.; Sommer, S.; Von Maltitz, G. (Eds.). World Atlas of Desertification. Publications Office of the European Union: Luxembourg, 2018. 7. FAO and ITPS. Global Soil Organic Carbon Map (GSOCmap)—Technical Report. Food and Agriculture Organization of the United Nations and Intergovernmental Technical Panel on Soils: Rome, 2018. 8. Strahler, A.; Muchoney, D.; Borak, J.; Friedl, M.; Gopal, S.; Lambin, E.; Moody, A. MODIS land cover product algorithm theoretical basis document (ATBD). Boston University Official Web Page, 2019, available at http://modis.gsfc.nasa.gov/data/atbd/atbd_mod12.pdf (accessed August 2019). 9. DGS. Land cover and land use change. University of Maryland, Department of Global Geographical Sciences, 2019, available at https://geog.umd.edu/research/landingtopic/landcover-land-use-change (accessed August 2019). 10. Cherlet, M., Hutchinson, C., Reynolds, J., Hill, J., Sommer, S., von Maltitz, G. (Eds.), World Atlas of Desertification, Publication Office of the European Union, Luxembourg, 2018. 11. Hijmans, R.J.; Cameron, S.E.; Parra, J.L.; Jones, P.G.; Jarvis, A. Very high resolution interpolated climate surfaces for global land areas. International Journal of Climatology, 2005, 25, 1965–1978. 12. Panagos, P.; Imeson, A.; Meusburger, K.; Borrelli, P.; Poesen, J.; Alewell, C. Soil conservation in Europe: wish or reality? Land Degradation & Development, 2016, 27(6), 1547–1551. 13. Lal, R. Soil erosion by wind and water: problems and prospects. In: Soil Erosion Research Methods. Lal, R., Ed. Routledge: New York, 2017; 1–10. 14. Daliakopoulos, I.N.; Tsanis, I.K.; Koutroulis, A.; Kourgialas, N.N.; Varouchakis, A.E.; Karatzas, G.P.;, Ritsema, C.J. The threat of soil salinity: a European scale review. Science of the Total Environment, 2016, 573, 727–739. 15. Gardi, C.; Panagos, P.; Van Liedekerke, M.; Bosco, C.; De Brogniez, D. Land take and food security: assessment of land take on the agricultural production in Europe. Journal of Environmental Planning and Management, 2015, 58(5), 898–912. 16. UNEP/MAP. Mediterranean Strategy for Sustainable Development 2016–2025. Plan Bleu, Regional Activity Centre: Valbonne, 2016. 17. Stankevich, S.A.; Kharytonov, N.N.; Dudar, T.V.; Kozlova, A.A. Risk assessment of land degradation using satellite imagery and geospatial modelling in Ukraine. In: Land Degradation and Desertification—A Global Crisis; Kaswamila, A., Ed. IntechOpen, 2016, 53–77. 18. Crippa, M.; Janssens-Maenhout, G.; Dentener, F.; Guizzardi, D.; Sindelarova, K.; Muntean, M.; Van Dingenen, R.; Granier, C. Forty years of improvements in European air quality: regional policy-industry interactions with global impacts. Atmospheric Chemistry and Physics, 2016, 16, 3825–3841. 19. Kosmas, C.; Ferrara, A.; Briassouli, H.; Imeson, A. Methodology for mapping environmentally sensitive areas (ESAs) to desertification. In: The MEDALUS Project. Mediterranean Desertification and Land Use Report; Kosmas, C., Kirkby, M., Gecson, N., Eds. European Commission: Belgium, 1999; 31–47. 20. Khor, M. 2019. Land degradation causes $10 billion loss to South Asia annually. https://www. globalpolicy.org/global-taxes/49705-land-degradation-causes-10-billion-loss-to-southasia-%20annually-.html (accessed August 2019). 21. Ivushkin, K.; Bartholomeus, H.; Bregt, A.K.; Pulatov, A.; Kempen, B.; De Sousa, L. Global mapping of soil salinity change. Remote Sensing of Environment, 2019, 231, 111260. 22. Lowdermilk, W.C. Conquest of the Land through Seven Thousands Years. USDA, Soil Conservation Service, U.S. Government Printing Office: Washington, DC, 1997; 99, 1–30.

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23. Sanders, J.H.; Southgate, D.D.; Lee, J.G. The Economics of Soil Degradation: Technological Change and Policy Alternatives. SMSS Technical Monograph; Department of Agricultural Economics, Purdue University, 1995; 22, 8–11. 24. Masso, C.; Baijukya, F.; Ebanyat, P.; Bouaziz, S.; Wendt, J.; Bekunda, M.; Vanlauwe, B. Dilemma of nitrogen management for future food security in sub-Saharan Africa—A review. Soil Research, 2017, 55(6), 425–434. 25. Graesser, J.; Aide, T.M.; Grau, H.R.; Ramankutty, N. Cropland/pastureland dynamics and  the  slowdown of deforestation in Latin America. Environmental Research Letters, 2015, 10(3), 034017. 26. Barrow, C.J. Land Degradation. Cambridge University Press: Cambridge, 1994. 27. USDA-NRCS. Risk of Human Induced Desertification. NRSC, Soil Survey Division, World Soil Resources: Washington, DC, 1999. 28. USDA-NRCS. Global Desertification Vulnerability. NRSC, Soil Survey Division, World Soil Resources: Washington, DC, 1999. 29. Kern, J.S.; Johnson, M.G. Impact of Conservation Tillage Use on Soil and Atmospheric Carbon in the Contiguous United States. EPA/600/3-91/056. Environmental Research Laboratory: Corvallis, OR and Orlando, FL, 1991. 30. Jones, B.M.; Arp, C.D.; Jorgenson, M.T.; Hinkel, K.M.; Schmutz, J.A.; Flin, P.L. Increase in the rate and uniformity of coastline erosion in Arctic Alaska. Geophysical Research Letters, 2009, 36(3), L03503. Doi: 10.1029/2008GL036205. 31. Faber, D. A revolution in environmental justice and sustainable development: the political ecology of Nicaragua. In: Environmental Justice; Thompson, P., Ed. Routledge, 2017, 39-70. 32. Butler, R. Soil erosion and its effects. https://rainforests.mongabay.com/0903.htm, 2012 (accessed August 2019). 33. UNEP/CEP. Guidelines for Sediment Control Practices in the Insular Caribbean, CEP Technical Report 32; UNEP Caribbean Environment Program: Kingston, 1994. 34. DPNR/DEP and USDA/NRCS. Unified Watershed Assessment Report—United States Virgin Islands, Virgin Islands Department of Planning and Natural Resources in cooperation with USDA Natural Resources Conservation Service, Caribbean Area, St. Croix, USVI, 1998. 35. UNEP/GEF/Kalmar Högskola/Cimab. Global International Water Assessment (GIWA), Caribbean Islands Bahamas, Cuba, Dominican Republic, Haiti, Jamaica, Puerto Rico Regional Assessment 4, Kalmar Sweden. 2004. 36. Russell-Smith, J.; Sangha, K.K.; Costanza, R.; Kubiszewski, I.; Edwards, A. Towards a sustainable, diversified land sector economy for North Australia. In: Sustainable Land Sector Development in Northern Australia; Russell-Smith, J. James, G., Pedersen, H., Sangha, K.K. Eds. Taylor and Francis: Boca Raton, FL. 2018, 85-132. 37. Lambers, H. Dryland salinity: a key environmental issue in southern Australia. Plant and Soil, 2003, 257, v–vii. 38. Landcare Australia. Annual Report and Yearbook. Landcare Australia: Melbourne, 1998; 70. 39. UNEP/EEA. Soil degradation and sustainable development in Europe: a challenge for the 21st century. 4th Conference to the Parties, Bonn, December 2000. 40. Smyth, A.J.; Dumanski, J. FESLM: An International Framework for Evaluating Sustainable Land Management, World Soil Resources Report. No. 73. FAO: Rome, 1993. 41. Kapur, S.; Akça, E.; Zucca, C.; Berberoğlu, S.; Miavaghi, S.R. Anthroscapes: a robust basis for mapping land quality and sustainable land use patterns. In: Eastern Mediterranean Port Cities; Yenişehirlioğlu, F., Özveren, E., Selvi Ünlü, T., Eds. Springer: Cham, 2019, 63–77. 42. Kapur, S.; Eswaran, H.; Blum, W.E. (Eds.). Sustainable Land Management: Learning from the Past for the Future. Springer Science & Business Media: Heidelberg. 2010. 43. Watanabe, T.; Kapur, S.; Aydın, M.; Kanber, R.; Akça, E. (Eds.). Climate Change Impacts on Basin Agro-Ecosystems. Springer International Publishing: Cham, 2019.

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44. Eswaran, H.; Beinroth, F.; Reich, P. Biophysical considerations in developing resource management domains. Proceedings of the IBSRAM International Workshop on Resource Management Domains, Kuala Lumpur, August 26–29, 1996; Leslie, R.N., Ed. The International Board for Soil Research and Management (IBSRAM): Bangkok, 1996, 61–78. 45. Kapur, S.; Kapur, B.; Akça, E.; Eswaran, H.; Aydin, M. A. Research strategy to secure energy, water, and food via developing sustainable land and water management in turkey. Hexagon Series on Human and Environmental Security and Peace, 2009, 4 (Part V), 509–518. 46. UNCCD. Countries agree on a landmark 2030 strategy to save fertile lands https://www.unccd. int/news-events/countries-agree-landmark-2030-strategy-save-fertile-lands. 2017 (accessed August 2019). 47. UNCCD. 2030 agenda for sustainable development: implications for the United Nations convention to combat desertification: the future strategic framework of the convention. Conference of the Parties. Thirteenth Session. Ordos, China, 6–16 September 2017.

27 Soil Erosion and Carbon Dioxide Introduction .................................................................................................. 235 Eroded Soil Organic Carbon: State of the Knowledge ............................ 236 Assessment of Erosion-Induced Carbon Dioxide Fluxes from Existing Data ...........................................................................................236 Conceptual Framework and Description of the Mass Balance Approach

Pierre A. Jacinthe and Rattan Lal

Estimates of Erosion-Caused CO2 Emissions ........................................... 238 Conclusions ................................................................................................... 238 Acknowledgments ........................................................................................ 239 References ...................................................................................................... 239

introduction The projected warming of global climate and the degradation of land resources by erosion are among the most pressing environmental challenges of the modern era. The recent warming trend in the world’s climate is linked to the accumulation of radiatively active gases (CO2, N2O, CH4) in the atmosphere.[1,2,3] Atmospheric CO2 concentration has been increasing at a rate of 2.2 ppm C y–1,[4] which, if maintained, could profoundly alter the chemical composition and the energy balance of the Earth’s atmosphere. Although fossil fuel burning is the primary contributor to the current atmospheric CO2 buildup, it is estimated that agriculture and land-use change contributed 24% of total anthropogenic emissions of CO2.[5] In soils, evolution of CO2 occurs through soil respiration which includes root respiration and the microbial decomposition of crop residues and soil humus. The soil organic carbon (SOC) content is positively related to structural stability and negatively related to the susceptibility of soils to erosion.[6,7] Cultivation reduces the stability of aggregates,[6] thereby increasing the erodibility of soils.[8] The effect of SOC on susceptibility to erosion is best exemplified by the use of SOC content in determining the soil erodibility factor (K) in the universal soil loss equation (USLE).[7] Worldwide, an estimated 1.11 × 109 ha of land are affected by water erosion,[9] resulting in the annual transport of 20–25 × 109 Mg of sediment to rivers and oceans.[10,11] While much is known about the protective role of SOC against soil erosion, information is limited regarding the fate of SOC mobilized by erosion. It is estimated that as much as 1.1 PgC [1 Pg = 1015 g = billion metric ton] may be emitted into atmosphere by erosion by water.[12] In this entry, we deliberate that water erosion, as a mechanism for SOC removal, has a profound effect on the global CO2 budget, and we present an approach to assess erosion-induced CO2 emission into the atmosphere.

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eroded Soil organic carbon: State of the Knowledge Water erosion involves the detachment of soil aggregates, the release of aggregate-protected SOC, and the translocation and deposition of detached soil particles. Globally, estimates of SOC displaced in terrestrial ecosystems by water erosion are in the order 57 Pg of Cy–1.[13,14] There is considerable uncertainty regarding the fate of the eroded SOC, however. The Intergovernmental Panel on Climate Change (IPCC) echoed this uncertainty in its handbook for conducting national greenhouse gas inventory by stating that the net effect of erosion on CO2 evolution is unclear at the present time.[15] The fate of SOC mobilized by erosion can include redistribution, long-term storage in aquatic and terrestrial systems, and mineralization during transport and deposition leading to emission of CO2 (production of CH4 possible depending on oxidation status).[14] During transport, detached soil particles are hydrodynamically sorted and redistributed over the eroding landscape as determined by flow condition, surface roughness, particle size, and geomorphology. Between 80% and 90% of the total soil mass mobilized by water erosion remains near the site of erosion.[15] A fraction of the eroded SOC is exported to streams, rivers, and lakes. Several publications have documented the magnitude of these exports which are in the order of 0.6–0.8 Pg C y–1 [16,17] or more.[12,14] Eroded SOC in runoff is also retained in the low-lying portions of the eroding landscape, but the fate of SOC entrapped in these terrestrial deposits is also not fully understood. On the premise that the decomposition process may be O2-constrained, it has been suggested that entrapment of eroded SOC in terrestrial deposits can be a mechanism for C sequestration.[18,19] It is estimated that a sequestration of 0.6–1.5 Pg C y–1 may be possible via this mechanism.[19] Eroded soil materials typically contain twice as much (or more) C than the average topsoil.[18] In addition, these materials are enriched in labile C.[20] Given the propensity of these SOC fractions to be associated with the lighter and most erosion-prone soil particles, mineralization of eroded C during transport and deposition is a reasonable expectation. A study comparing CPMAS13C-nuclear magnetic resonance spectrometry data of eroded and deposited soil samples indicated that 70% of SOC in eroded soil could be decomposed during transport and deposition.[21] A recent review suggests a lower (20%) mineralization rate.[13] The information presented above clearly underscores the need for better quantitative assessments of the various fates of eroded SOC. Of particular interest is an assessment of water erosion’s contribution to atmospheric CO2. At the present time, there are few if any experimental data to support such an evaluation. However, an indirect determination of CO2 production during transport and deposition of eroded SOC is possible using literature data pertaining to the other pathways identified above.

Assessment of erosion-induced carbon Dioxide Fluxes from existing Data conceptual Framework and Description of the Mass Balance Approach In order to estimate CO2 production during transport and deposition of eroded SOC, a mass balance approach, described in Figure 1, is proposed. This approach assumes a steady-state equilibrium with respect to SOC content. Augmentation of the initial pool of SOC via humification of crop residues compensates for SOC losses, which include soil respiration, leaching, and SOC translocation with runoff waters. It follows that C0 + h ⋅C A = Ct + SR + E

(1)

SR = k ⋅Ct

(2)

E = LC + Sw + EW + SL + EL

(3)

EL = (C0 − Ct ) t + h ⋅C A − k ⋅Ct − ( LC + Sw + Ew + SL )

(4)

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FIGURE 1

A conceptual mass balance model of SOC evolution in soils.

where C0 = initial SOC stocks, h = humification rate, CA = annual residue addition, Ct = the SOC stocks after t years, SR = annual SOC mineralization, k = coefficient of SOC mineralization, LC = leaching loss of SOC, Sw = terrestrial C delivered to lakes and oceans, Ew = CO2 evasion from aquatic systems (lakes, rivers, streams), SL = C storage in terrestrial deposits, and EL = mineralization and CO2 evolution during transport and deposition of eroded SOC. The coefficients h and k are in units of y–1, C0 and Ct are in g C m–2, and the other parameters are in units of g C m–2 y–1. Using Eq. (4), CO2 evolution during transport and deposition of eroded SOC (EL) can be computed. Values for the parameters in Eq. (3), assembled from various regional and global scale studies, are summarized in Table 1. We used the mass balance model and data from several long-term studies of SOC evolution to estimate EL.

TABLE 1 Estimates of the Major Pathways of Soil Organic Carbon Exported from Agricultural and Forested Watersheds Description of SOC Loss Pathway Leaching of organic carbon in forest and cropland Terrestrial carbon delivered to lakes, rivers, and oceans Evasion CO2 from aquatic systems Organic carbon storage in terrestrial deposits

Symbol

Range (g C m–2 y–1)

LC SW EW Sl

0.5–5a 10–25b 0.2–3.6c 6–15d

Data from Guggenberger and Kaiser.[29] Terrestrial C delivery to the ocean [computed using global C input to the ocean (750 × 1012 g C/yr from Ludwig et al.[17]) and the total area of river draining to the ocean (89 × 1012 m2 from Stallard[19])] was added to accumulation rate of terrestrial C in freshwater systems and man-made reservoirs (2–10 g C m–2 y–1 from Mulholland and Elwood[30]). c Adapted from Spitzy and Leenheer.[31] d Estimate computed using global C storage in terrestrial sediments (0.6–15 × 1015 g C y–1 from Stallard[19]) and total area of river basins (1014 m2 from Meybeck[32]). a

b

238

FIGURE 2

Managing Soils and Terrestrial Systems

A hypothetical distribution of erosion-mobilized SOC among various possible fates.

estimates of erosion-caused co2 emissions Analyses of the data available in the literature show that estimates of CO2 production (EL) during transport and deposition of eroded SOC range between 6 and 52 g C m–2 y–1 (Table 2). These estimates are necessarily crude given that the values assigned to the model parameters were obtained from different studies. It is important to note, however, that our parameter estimates (e.g., SW, EW) were similar to direct measurements made in studies conducted in various geographical regions including the United States,[22] Europe, and New Zealand.[16] Assuming a sediment delivery ratio of 15%[11] and a C-enrichment ratio of 2 for eroded soil material,[18] the total C exported from a typical eroded field averages 30% of the total C mobilized during erosional events. Using the C export data in Table 1, estimates of total C mobilized by water erosion range between 56 and 168 g C m–2 y–1. Similar estimates (40 and 190 g C m–2 y–1) of SOC mobilization are also obtained by assuming that computed EL values represent 20% of the total SOC displaced by water erosion.[12] Using these ranges of erosion-induced SOC mobilization and the data in Table 1, a hypothetical distribution of eroded SOC among several pathways is presented in Figure 2. For example, for a soil containing 4–5 × 103 g C m–2, the loss and redistribution of 1%–5% of the total SOC might go unnoticed for some time. Removal of SOC via sheet erosion could be particularly difficult to detect in the short term, hence a possible dismissal of water erosion as an important mechanism for SOC loss.[23] However, field study data support the view that erosion is a significant contributor to SOC decline and sometimes contributes more to SOC depletion than soil respiration.[24,25] If the data in Table 2 are extrapolated to the global cropland (1.5× 109 ha),[26] erosion-induced global CO2 fluxes to the atmosphere ranging between 0.12 and 0.55 Pg C y–1 can be computed. These fluxes could represent up to one-third of the annual CO2 emission (1.6 Pg C y–1) due to deforestation and landuse change in the tropics.[13]

conclusions The information presented in this entry supports our contention that sizable amounts of CO2 can be produced as a result of SOC mobilization by water erosion. Moreover, water erosion could elicit profound disturbances in the plant–soil–atmosphere segment of the global C cycle through a cascading process involving nutrient depletion, decreased biomass production, and alteration of the radiative and thermal properties of soils. The strong control of temperature on CO2 emission[21] and the higher temperature

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TABLE 2 Estimates of CO2 Production during Transport and Deposition of Eroded Soil Organic Carbon Using a Mass Balance Approach References Buyanovsky et al.[33] Paul et al.[34]

Rasmussen et al.[35] Dumanski et al.[36]a

Lal[37]

Cropping System and Location Continuous wheat, Sanborn Field plots, Missouri, USA Corn–soybean rotation, long-term ecological research (LTER) plots, Michigan, USA Wheat–fallow system, Pacific Northwest, USA Grain- and wheat-fallow systems in the Canadian prairies (Saskatchewan, Alberta) and Ontario, Canada Continuous corn, with and without N fertilization, International Institute of Tropical Agriculture (IITA), Nigeria

Rate of Residue C Input

Rate of SOC Declinec (g C m–2 y–1)

Erosion-Induced CO2 Flux, EdL[33]

698

42

17

425

42

21

520–700b

0.6–2.8

6–52

245–326

9.6–38.3

6–14

225d

25

29–32

Compilation of several studies. This includes annual carbon input from wheat residue (252 g C m–2) and manure (rates: 448–2680 g C m–2). c (C − C )/t. 0 t d To compute EL, values for h and k provided in the reference were used. If not available, h (0.16–0.2) and k (0.01–0.02) were used. a

b

usually recorded in eroded soils[22] also indicate a shortened residence time of photosynthetically fixed C in the soil system. Although many of the effects remain to be evaluated experimentally, available information suggests a positive feedback of water erosion on atmospheric CO2 and climate warming.

Acknowledgments The authors gratefully acknowledge the technical support provided by Dr. J.M. Kimble of the National Soil Survey Center (NSSC) of Natural Resources Conservation Service (NRCS) at Lincoln, Nebraska.

References 1. IPCC. The Physical Science Basis: Summary for Policy Makers. Cambridge University Press: Cambridge, 2013. 2. IPCC. Global Warming Special Report. Cambridge University Press. Cambridge, 2018. 3. IPCC. Land Use, Land Use Change and Forestry; Special Report; Intergovermental Panel on Climate Change. Cambridge University Press: Cambridge, 2000. 4. WMO. Greenhouse Gas Bulletin. World Meterological Organization: Geneva, 2018. 5. Tubillo, F.N.; Salvatore, M.A.; Ferraro, A.F.; House, J. et al. The contribution of agriculture, forestry and other land use activities to global warming. Global Change Biology 2015, 21, 2655–2660. 6. Cook, G.D.; So, H.B.; Dalal, R.C. Structural degradation of 2 vertisols under continuous cultivation. Soil and Tillage Research 1992, 24 (1), 47–64. 7. Wischmeier, W.H.; Smith, D.D. Predicting Rainfall Erosion Losses: A Guide to Conservation Planning; Agriculture Handbook No. 537. United States Department of Agriculture: Washington, DC, 1978; 58. 8. Lal, R. Soil erosion on Alfisols in western Nigeria. I. Effects of slope, crop rotation and residue management. Geoderma 1976, 16, 363–375.

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9. Oldeman, L.R. The global extent of soil degradation. In Soil Resilience and Sustainable Land Use; Greenland, D.J., Szabolcs, I., Eds. CAB International: Wallingford, 1994; 99–118. 10. Milliman, J.D.; Syvitski, J.P.M. Geomorphic/tectonic control of sediment discharge to the ocean: the importance of small mountainous rivers. Journal of Geology 1992, 100 (5), 325–344. 11. Walling, D.E. The changing sediment loads of the world rivers. IAHS Publication 2008, 325, 323–338. 12. Lal, R. Soil erosion and the global carbon budget. Environment Intrenational 2003, 29, 437–450. 13. Lal, R. Global soil erosion by water and carbon dynamics. In Soil and Global Change; Lal, R., Kimble, J.M., Levine, E., Stewart, B.A., Eds. CRC/Lewis: Boca Raton, FL, 1995; 131–142. 14. Lal, R.; Griffin, M.; Apt, J.; Lave, L.; Morgan, M.J. Managing soil carbon: a response. Science 2004, 305, 1567–1567. 15. IPCC. Revised IPCC Guidelines for National Greenhouse Gas Inventories, Reference Manual, vol. 3. Intergovernmental Panel on Climate Change: Geneva, 1996. 16. Hope, D.; Billet, M.F.; Cresser, M.S. A review of the export of carbon in river water: fluxes and processes. Environmental Pollution 1994, 84 (3), 301–324. 17. Ludwig, W.; Amiotte-Suchet, P.; Probst, J.L. River discharges of carbon to the world’s oceans: determining local inputs of alkalinity and of dissolved and particulate organic carbon. Comptes Rendus de l’Acade’mie des Sciences, Se’rie II., Sciences de la Terre et des Plane’tes 1996, 323 (12), 1007–1014. 18. van Noordwijk, M.; Cerri, C.; Woomer, P.L.; Nugroho, K.; Bernoux, M. Soil carbon dynamics in the humid tropical forest zone. Geoderma 1997, 79 (1–4), 187–225. 19. Stallard, R.F. Terrestrial sedimentation and the carbon cycle: coupling weathering and erosion to carbon burial. Global Biogeochemical Cycles 1998, 12 (2), 231–257. 20. Tiessen, H.; Stewart, J.W.B. Particle-size fractions and their uses in studies of soil organic matter: II. Cultivation effects on organic matter composition in size fractions. Soil Science Society America Journal 1983, 47, 509–514. 21. Beyer, L.; Frund, R.; Schleuss, U.; Wachendorf, C. Colluvisols under cultivation in SchleswigHolstein. 2. Carbon distribution and soil organic matter composition. Journal of Plant Nutrition and Soil Science 1993, 156 (3), 213–217. 22. Moeller, J.R.; Minshall, G.W.; Cummins, K.W.; Petersen, R.C.; Cushing, C.E.; Sedell, J.R.; Larson, R.A.; Vannote, R. Transport of dissolved organic carbon in streams of differing physiographic characteristics. Organic Geochemistry 1979, 1, 139–150. 23. Schlesinger, W.H. Changes in soil carbon storage and associated properties with disturbance and recovery. In The Changing Carbon Cycle: A Global Analysis; Trabalka, J.R., Reichle, D.E., Eds. Springer-Verlag: New York, 1986; 194–220. 24. Slater, C.S.; Carleton, E.A. The effect of erosion on losses of soil organic matter. Soil Science Society America Proceedings 1938, 3, 123–128. 25. Gregorich, E.G.; Anderson, D.W. Effects of cultivation and erosion on soils on four toposequences in the Canadian prairies. Geoderma 1985, 36 (3–4), 343–354. 26. FAO. Production Yearbook, Vol. 52. Food and Agricultural Organization of the United Nations: Rome, 1998. 27. Fortin, M.C.; Rochette, P.; Pattey, E. Soil carbon dioxide fluxes from conventional and notillage smallgrain cropping systems. Soil Science Society of America Journal 1996, 60 (5), 1541–1547. 28. Sutherland, R.A.; Menard, T.; Perry, J.L.; Penn, D.C. The influence of rolled erosion control systems on soil temperature and surface Albedo: Part I. A greenhouse experiment. Land Degradation and Development 1998, 9 (2), 159–178. 29. Guggenberger, G.; Kaiser, K. Significance of DOM in the translocation of cations and acidity in acid forest soils. Journal of Plant Nutrition and Soil Science 1998, 161 (2), 95–99. 30. Mulholland, P.J.; Elwood, J.W. The role of lake and reservoir sediments as sinks in the perturbed global carbon cycle. Tellus 1982, 34 (5), 490–499.

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31. Spitzy, A.; Leenheer, J. Dissolved organic carbon in rivers. In Biogeochemistry of Major World Rivers; Degens, E.T., Kempe, S., Richey, J.E., Eds. SCOPE 42. John Wiley and Sons: Chichester, NY, 1991; 214–232. 32. Meybeck, M. Concentrations des Eaux Fluviales en Eléménts Majeurs et Apports en Solution aux Océans. Revue de Géologie Dynamique et de Géographie Physique 1979, 21, 215–246. 33. Buyanovsky, G.A.; Kucera, C.L.; Wagner, G.H. Comparative analyses of carbon dynamics in native and cultivated ecosystems. Ecology 1987, 68 (6), 2023–2031. 34. Paul, E.A.; Harris, D.; Collins, H.P.; Schulthess, U.; Robertson, G.P. Evolution of CO2 and soil carbon dynamics in biologically managed, row crop agroecosystems. Applied Soil Ecology 1999, 11 (1), 53–65. 35. Rasmussen, P.E.; Albrecht, S.L.; Smiley, R.W. Soil C and N changes under tillage and cropping systems in semiarid pacific northwest agriculture. Soil and Tillage Research 1998, 47 (3–4), 197–205. 36. Dumanski, J.; Desjardins, R.L.; Tarnocai, C.; Monreal, C.; Gregorich, E.G.; Kirkwood, V.; Campbell, C.A. Possibilities for future carbon sequestration in Canadian agriculture in relation to land use changes. Climatic Change 1998, 40 (1), 81–103. 37. Lal, R. Soil quality changes under continuous cropping for seventeen seasons of an alfisol in Western Nigeria. Land Degradation and Development 1998, 9 (3), 259–274.

28 Soil Quality: Carbon and Nitrogen Gases Carbon Gases ................................................................................................243 Carbon Dioxide • Methane

Philippe Rochette, Sean McGinn, and Reynald Lemke

Nitrogen Gases............................................................................................. 244 Nitric and Nitrous Oxides • Ammonia

Conclusions ................................................................................................... 247 References...................................................................................................... 247

carbon Gases Carbon is a major constituent of biomass and soil organic matter. However, more than 99% of global carbon is locked into sediments and fossil forms and is not available for biological processes. The small remaining active fraction of global carbon transits between atmospheric CO2, biomass and soil organic matter, and detritus in the so-called carbon cycle (Figure 1). The carbon cycle is driven by photosynthetic fixation of atmospheric CO2 by plants. In global terrestrial ecosystems, it is estimated that plant photosynthesis fixes more than 200 Gt of CO2 every year.[1] Eventually, similar amounts are returned to the atmosphere by the respiration of animals and by the aerobic heterotrophic decomposition of soil organic matter and plant litter. In the absence of oxygen and other electron acceptors, CH4 is the final product of soil organic matter decomposition. Human activities, including changes in land use and soil management, are contributing to unprecedented rapid increases in atmospheric CO2 and

FIGURE 1 The soil carbon cycle. In well-aerated soils, decomposition of soil organic matter produces carbon dioxide. Significant amounts of methane are produced only under highly anaerobic (waterlogged) soilas.

243

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CH4 concentrations, which may result in important modifications of the Earth’s climate. In dry soils, auto-oxidation of organic compounds can produce carbon monoxide (CO). Carbon monoxide can also be biologically oxidized to CO2 in moist but well-aerated soils.

carbon Dioxide Soil-surface emitted CO2, or soil respiration, is the sum of the CO2 produced by root respiration and heterotrophic decomposition of root exudates, soil organic matter and plant litter. Decomposition processes, while dominated by soil microbes, are the result of complex interactions between soil fauna, fungi, actinomycetes, and bacteria. During decomposition, complex molecules like cellulose, hemicellulose, proteins, and lignin are broken down into low molecular weight substances and oxidized to CO2 to produce energy and C for the growth of organisms.[2,3] The rate of decomposition is regulated by the quality and quantity of organic substrates and by physical/environmental properties of soil, such as temperature, moisture, and aeration.[4] When ecosystems are in equilibrium, soil CO2 emissions are the result of the natural recycling of nutrients and equal the amount of atmospheric CO2 fixed by plant photosynthesis. However, in the past 150 years, human activity has broken this equilibrium by burning fossil carbon reserves and by decreasing soil organic matter through land- use changes. During this period, the additional CO2 that entered the carbon cycle could not be completely absorbed by increases in autotrophic activity and atmospheric CO2 concentrations increased from 280 to 365 ppm. Several measures have been identified to reduce net CO2 emissions from soils. Among them, it is believed that modifications in agricultural and forest management practices could result in increased C storage in soils as organic matter. Converting cultivated land into natural ecosystems usually increases soil carbon content because of increased return of plant litter and decreased decomposition. Also, tillage breaks down aggregates and increases soil carbon losses by exposing organic matter to microbes that decompose it. Therefore, reducing tillage intensity, fallow frequency, and the amount of cultivated land (by increasing productivity), and increasing perennial crops in rotation and permanent grasslands are seen as potential management practices for mitigating global warming.[5]

Methane Methane is produced in soils by the decomposition of organic matter and by the reduction of CO2 under highly anaerobic environments. Such conditions are found in wetlands and in rice paddies that, together with landfills, contribute about half of the total emissions of anthropogenic CH4. Methane is stable in waterlogged soils and can be emitted to the atmosphere via diffusion, ebullition, and transport through plants. In the presence of oxygen, certain bacteria can oxidize CH4 to CO2 (Figure 1).[6] Human intervention can greatly influence CH4 production or consumption in soils. Flooding of soils in natural or agricultural ecosystems usually results in increased emissions[7] while drainage of wetlands can turn a source of methane into a sink.[8] Flooded rice fields are a major source of anthropogenic CH4. Management practices have been proposed to reduce CH4 emissions from rice paddies, including draining the fields during the growing season, replacing urea by other types of nitrogen fertilizers, and reducing the input of crop residues by using new cultivars and alternative cultural practices.[5] Reduced rates of CH4 oxidation in well-aerated soils have been observed following cultivation and addition of nitrogen fertilizers.[9]

nitrogen Gases The passage of nitrogen (N) through ecosystems can be represented as a “loop-within-a-loop” (Figure 2). Nitrogen enters ecosystems primarily through biotic or abiotic processes that “fix” (convert molecular nitrogen [N2] to biologically available forms) N2 from the atmosphere. Within the soil-plant system, N

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245

FIGURE 2 Schematic representation of the nitrogen cycle depicted as a “loop-within-a-loop.” The outer loop traces the passage of N from the atmosphere through ecosystems and back to the atmosphere, while the inner loop traces the interconversion of N between organic and inorganic forms within the soil–plant system.

undergoes a complex series of transformations, resulting in a continual transfer of N between inorganic and organic forms—the inner loop (Figure 2). Nitrous oxide and NOx are both produced and consumed during these transformations. With time, most of the N entering ecosystems returns to the atmosphere as N2, but an important fraction is emitted as gaseous NH3, NOx, and N2O. More than 70% of the estimated 18 Tg N2O–N entering the atmosphere each year is emitted by soils.[1,10] Above plant-canopy emissions of NOx from soils are probably in the range of 3.3–21 Tg N yr−1.[11,12] This exchange of N between the atmosphere and the soil-plant system— the outer loop—strongly mediates atmospheric concentrations of NH3, NOx, and N2O.

nitric and nitrous oxides The majority of soil-emitted NOx is nitric oxide (NO);[13] therefore, the rest of this discussion will focus on nitrogen gases other than nitrogen dioxide (NO2). Nitric oxide is derived primarily via autotrophic nitrification,[14,15] whereas N2O emissions can be a product of autotrophic nitrification,[15] denitrification,[16] or a combination of both.[14] Nitrous oxide, and particularly NO, may also be emitted from various chemical reactions collectively known as chemodenitrification.[17] The magnitude of N2O and NO emissions is strongly influenced by site-specific variables—most particularly soil attributes—and human intervention. Given that both gases are produced primarily during microbial transformations of inorganic N, the potential of a soil to produce and emit N2O or NO increases with increasing N availability.[18] Anthropogenic activities that increase the flow of N into the system also increase emissions of N2O and NO. Such activities include, for example, the cultivation of legume grain or forage crops,[19] and intensive use of nitrogen fertilizers.[20] Total N2O emissions tend to increase as soil organic carbon (SOC) content increases. This is understandable, since N turnover is closely tied to SOC turnover, and the amount of carbon (C) available to drive microbial processes is directly related to the quantity and quality of SOC. The relative availability of C and N is of particular importance when livestock manure or sewage sludge is added to soils. The addition of both C and N tend to increase N2O and NO emissions more than N alone, particularly in C limited systems.[21,22] Agricultural management practices that minimize soil disturbance tend to increase soil water status and soil bulk density, favoring the development of anaerobic conditions and higher N2O[4] but possibly lower NO emissions.[23] Conversely, equal or lower emissions of N2O have also been reported in no-tillage

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systems compared to conventional tillage systems.[24] In this instance, reduced soil disturbance appeared to alter N cycling, resulting in lower NO3− availability—hence lower N2O emissions—during the spring thaw period. In general, soil conditions or management that lead to accumulations of NH4+, NO3−, and particularly NO2−, favor gaseous N losses. This may happen if N is released during soil C mineralization at a time when plant uptake is absent or minimal, an example being bare soil fallow periods commonly employed in the semiarid regions of North America, or if fertilizer or organic N is added at a time when crop growth is not vigorous. Low soil pH levels (2mm diameter)

>6.5 and 20 total solids), semisolid (10%–20% total solids), slurry (4%–10% solids), or liquid (6% of the world land area (Figure 1). Salinization is a serious hazard in San Joaquin Valley, United States; Murray-Darling Basin, Australia; Euphrates Basin, Pakistan; Indo-Gangetic Basin, India; and Aral Sea Basin, Central Asia.[5] Global land area degraded by salt-affected soils may be 1 billion hectare (Bha),[6,7] and the world may be losing as much as 2000 ha/day.[5] Worldwide 20% of the cultivated land area and 33% of the irrigated lands are degraded because of salt-induced problems.[8,9] It is also estimated that salinized area may be increasing at the rate of 10% annually. Fifty percent of arable land may be salinized by 2050.[8] Crop growth and agronomic productivity in salt-affected soils is adversely affected[10] because of osmotic and drought stress,[11] degradation of soil structure especially in sodic soils, and alteration in soil enzymatic activity.[12] Salinity problems also affect greenhouse crops, mine spoils, and disposal areas. Part of the yield loss occurs on irrigated lands, which are otherwise highly productive. In the irrigation district of Ciudad Juarez Valley in Mexico, the productivity of 70% of the lands is hindered by salinity. On average, 25% of irrigated Mexican lands are salt-affected. The impact of yield losses is magnified at a local scale because salt-affected soils are not uniformly distributed, threatening the economic welfare of some regions and countries. In Bangladesh, 24% of the total land area is salt-affected with a rapid expansion

FIGURE 1 Distribution of salt-affected soils. The percentage of salt-affected soils on the total land surface is reported for each region of the world. Source: FAO-AGL.[4]

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during the last quarter of the 20th century from 3 million ha. In Australia, almost 20% of the land is salt-affected.[4] Salt-affected soils not currently used for crop production (e.g., salt flats) are a potential source of salts for salinization of surrounding fields. According to Oldeman, Hakkeling, and Sombroek,[13] the annual rate of loss of agricultural land by salinization, alkalinization, and waterlogging is about 1.5 million ha. Previous estimates reported 10 million ha of agricultural land loss per year.[14] The rate of increase is likely to vary widely across regions and from year to year, but all data suggest an increasing extent of salt-affected lands, especially where the natural resources are scarce and the growing human population increases pressure on soils. Changes in the global environment are likely to increase yield losses consequent to soil salinity. The predicted increase of the sea level from thermal expansion of seawater ranges from 15 to >50 cm by the year 2100.[15] The rise of seawater is likely to worsen salinity problems from tidal inundation of coastal lands. On the other hand, there is no evidence to suggest that elevated atmospheric CO2 would increase the level of salinity suitable for plant growth.[16] Increased CO2 is expected to increase plant growth only on nonsaline soils, magnifying yield losses due to soil salinity.

crop tolerance Soil salinity and sodicity limit the potential area of growth of sensitive crops. All plants are sensitive to salts at some concentration, but the limiting concentration varies with plant species and variety. Crop tolerance is defined in relation to the level of salinity causing yield losses. High salt concentration may lead to plant death and no yield. Decreased growth occurs above a soil critical salt or sodium concentration (tolerance threshold). Crop yield reductions in salt-affected soils result from alteration of various metabolic processes in plants under salt stress. Negative effects of excess of salts in the soil solution include increased osmotic pressure limiting water uptake (physiological drought), abnormal pH and ionic competition limiting nutrient uptake, and ionic toxicity. Total salinity effects always depend on the specific soil ionic composition. The negative response of plants to low water potential may prevail in saline soils, while single ion toxicity or nutritional imbalance may be particularly severe in sodic soils. Soil structural impedance of plant growth is common in the absence of sufficient Ca 2+. Structural problems derive from crusting, formation of compacted layers, poor infiltration, and poor permeability to water and air. In waterlogged soils, anaerobic conditions may enhance salinity stresses. In all plants growing on saline media, a part of the metabolic energy is diverted to ion transport, synthesis of organic osmolytes (contributing to osmotic adjustment to low water potential), and ion compartmentation preventing ion toxicity.[2] Osmotic adjustments in salt-tolerant plants need to follow seasonal changes in soil salinity and water availability. In general, plants that are more salt-tolerant tend to grow more slowly at low salinity levels than less-tolerant plants. Broadly adaptable plants can produce good yields where strong temporal changes of soil salinity occur in the soil. But such plants, tolerant to a wide range of salinities, perform less well at any salinity than less adaptable plants at their optimal salinity.[8] Reduced yield in tolerant plants can be related to (1) greater allocation of organic C in the roots of tolerant plants at the expense of leaf area, (2) decreased use of the solar radiation, and (3) low transpiration rate. Soil salinity interacts with climate and other edaphic factors in determining yield. Reduced availability of water in the soil is concomitant to higher evaporative demand in the leaves. Adaptation to soil salinity is a reduction in the capacity for photosynthetic carbon assimilation consequent to the necessity of minimizing evapotranspiration. In general, crops are less sensitive to salinity in glasshouse conditions than outdoors. At low soil fertility levels, plants may appear more tolerant to salinity than at high fertility levels if salinity is not the limiting factor of yield. On the other hand, nutrient deficiencies induced by salinity and sodicity can further reduce low yields due to low fertility. In natural conditions, soil salinity varies with depth and time. Plants extract water from the least saline parts of the soil surrounding the roots. Plants may be able to tolerate high salinity levels in the root zone if occurring only for short periods of time during the growing season. Soil salinity may peak

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TABLE 2

Classification of Crop Salt Tolerance Based on the Values of EC above Which Yield Loss Begins

Relative Tolerance Sensitive Moderately sensitive Moderately tolerant Tolerant Unsuitable

Critical EC (dS/m)

Examples of Crops

10.0

Beans Rice, corn Wheat Barley, cotton Most crops

Source: Ayers and Wescot.[17]

TABLE 3 Classification of Crop Exchangeable Sodium Tolerance. Growth Is Usually Already Affected at Lower Values of ESP and Crop Production Is Usually Excluded at ESP>30 Relative Tolerance Sensitive Semitolerant Tolerant

ESP (%)

Examples of Crops

40

Beans, corn Rice, wheat Barley, cotton

Source: Ayers and Wescot.[17]

at different stages of the crop, when plants may be more or less sensitive. Consequently, the critical point for yield reductions varies with duration of the salt stress and stage of development. Growth can be inhibited at any stage of the biological cycle. Severe reductions of yields can be due to low germination and limited early plant establishment. Yield reductions can be caused by reduced vegetative growth and/or by perturbation of the reproductive phases. Developmental shifts of relative salt tolerance are common during plant life and vary with cultivar and environment. Often, qualitative yield reductions enhance total yield losses. Irrigated drylands are at a high risk of salinization,[18] and there are a wide range of reclamation measures.[19,20] A common classification of crop salinity tolerance distinguishes five categories (Table 2). Sodium is not considered a limiting factor in the above classification. A classification of sodium tolerance of crops consists of three groups (Table 3). Sodium-tolerance tables are usually based on nutritional responses in absence of soil structural degradation, which generally excludes crop production at ESP >30.[9] These classifications are based on the agronomic criterion of comparing the relative yield of each crop on saline media to its yield on normal nonsaline media. Published tolerance tables provide a reference of relative salt tolerance of crops under irrigation without local climatic specifications. Such data refer to duration of salinity from late seedling stage to maturation with irrigation, fertilization, and pest control maintained optimal for each crop species and variety at the time of the experiment. Tolerance tables may apply to individual cultivars in the absence of interactions between soil salinity and other environmental factors.

conclusions Salt-affected soils decrease crop productivity at local and global scales. The extent of soil salt problems is likely to increase if no measures are taken to limit soil salinization and sodication. Management and reclamation of salt-affected soils require a combination of agronomic practices, including hydraulic, mechanical, amending, and cropping practices.[4] The selection of proper management practices is relative to the soil, water, crop, and human available resources. Improved crop salt tolerance can contribute to the remediation of soil salt problems because: • Extends the choice of crops that can grow at each soil salinity level • Allows irrigation using more saline water • Increases soil organic matter and improves soil structure.

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Consequently, tolerant crops may assist in establishing a proper water balance and reduce the need and cost of reclamation measures such as leaching. Information on crop response to salinity and sodicity needs to be updated in relation to changes of plant materials, agronomic practices, and climatic conditions. A range of values in the continuum of salt and sodium stresses would be better suited than a single critical value to represent the tolerance or sensitivity of a plant since the intensity of salt stress depends on many other factors determining yields.

List of Abbreviations EC electrical conductivity of the soil saturation extract in dS/m ESP exchangeable sodium percentage in % SAR sodium adsorption ratio = Na+/([Ca 2+ + Mg2+]/2)1/2, where concentrations are in mmol/L

References 1. SSSA. Glossary of Soil Science Terms. Soil Science Society of America: Madison, WI, 1997; 134 pp. 2. Läuchli, A., Epstein, E. Plant responses to saline and sodic conditions. In Agricultural Salinity Assessment and Management, Tanji, K.K., Ed.; ASCE Manuals and Reports No. 71; ASCE: New York, 1990; 113–137. 3. Szabolcs, I. Salt buildup as a factor of soil degradation. In Methods for Assessment of Soil Degradation, Lal, R., Blum, W.H., Valentine, C., Stewart, B.A., Eds.; Boca Raton, FL: CRC Press, 1998; 253–264. 4. FAO-AGL. Global Network on Integrated Soil Management for Sustainable Use of Salt-Affected Soils. FAO-AGL, Land and Plant Nutrition Management Service, 2000; http://www.fao.org/ag/ AGL/ag11/spush. 5. Collins, T. World Losing 2000 Hectare of Fair Soil Daily to Salt-Induced Degradation. INWEH-UNU, Hamilton, Canada, 2014. 6. Yensen, N.P., Halophyte uses for the twenty-first century. In Ecophysiology of High Salinity Tolerant Plants, Khan, M.A., Weber, D.J., Eds.; Dordrecht: Springer, 2008; 367–396. 7. Metternicht, G.I., Zinck, J.A. Remote sensing of soil salinity: potentials and constraints. Remote Sens. Environ. 2003; 85: 1–20. 8. Patel, B.B., Patel Bharat, B., Dave, R.S. Studies on infiltration of saline-alkali soils of several parts of Mehsana and Patan districts of north Gujarat. J. Appl. Technol. Environ. Sanitation. 2011; 1(1): 87–92. 9. Jamil, A., Riaz, S., Ashraf, M., Foolad, M.R. Gene expression profiling of plants under salt stress. Crit. Rev. Plant Sci. 2011; 30(5): 435–458. 10. Paul, D. Osmotic stress adaptions in rhizobacteria. J. Basic Microbiol. 2012; 52: 1–10. 11. Munns, R. Genes and salt tolerance: bringing them together. New Phytol. 2005; 167: 645–663. 12. Seckin, B., Sekmen, A.H., Turkan, I. An enhancing effect of exogenous mannitol on the antioxidant enzyme activities in roots of what under salt stress. J. Plant Growth Regul. 2009; 28: 12–20. 13. Oldeman, L.R., Hakkeling, R.T.A., Sombroek, W.G. Second Revised Edition. World Map of the Status of Human-Induced Soil Degradation. An Explanatory Note. Wageningen, NL: International Soil Reference and Information Center, 1991; 35 pp. 14. WCED. Our Common Future. Oxford: Oxford University Press, 1987; 383 pp. 15. Warrick, R.A., Le Prevost, C., Meier, M.F., Oerlemans, J., Woodworth, P.L. Changes in sea level. In Climate Change 1995, Houghton, J.T., Meira Filho, L.G., Callander, B.A., Harris, N., Kattenberg, A., Maskell, K., Eds.; Cambridge: Cambridge University Press, 1996; 363–405.

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16. Ball, M.C., Sobrado, M.A. Ecophysiology of mangroves: challenges in linking physiological processes with patterns in forest structure. Physiological Plant Ecology: The 39th Symposium of the British Ecological Society, Held at the University of York, 7–9 September 1998. Scholes, J.D., Barker, M.G., Press, M.C., Eds.; Oxford: Blackwell Science, 1999; 331–346. 17. Ayers, R.S., Wescot, D.W. Water Quality for Agriculture; FAO Irrigation and Drainage Paper 29, Rev. 1; Rome: FAO, 1989; 174 pp. 18. Squires, V.R., Glenn, E.P. Salinization, desertification, and soil erosion. In The Role of Food, Agriculture, Forestry and Fisheries in Human Nutrition 2011, Squires, V.R., Ed., Vol. III, Oxford: UNESCO EOLSS Publishers, 2004; 102–123. 19. Oster, J.D., Shainberg, I., Abrol, I.P. Reclamation of salt-affected soils. In Agriculture Drainage, Agronomy Monograph 38; Madison, WI: ASA, 1999; 659–692. 20. Gupta, R.K., Abrol, I.P. Salt-affected soils: their reclamation and management for crop production. In Advances in Soil Science, Lal, R., Stewart, B.A. Eds.; New York, NY: Springer. 1990; 223–288.

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67 Agricultural Runoff

Matt C. Smith, David K. Gattie, and Daniel L. Thomas

Introduction .................................................................................................. 581 Agricultural Runoff Quantity ..................................................................... 582 Soil Erosion and Associated Pollutants ..................................................... 582 Dissolved Pollutants .....................................................................................583 Control of Agricultural Runoff ...................................................................584 Conclusion ....................................................................................................584 References ...................................................................................................... 585

introduction Agricultural runoff is surface water leaving farm fields as a result of receiving water in excess of the infiltration rate of the soil. Excess water is primarily due to precipitation, but it can also be due to irrigation and snowmelt on frozen soils. In the early 20th century, there was considerable concern about erosion of farm fields due to rainfall. The concern was primarily related to the loss of valuable topsoil from the fields and the resulting loss in productivity. With the passage of the Federal Water Pollution Control Act Amendments of 1972, the potential for pollution of surface water features such as rivers and lakes due to agricultural runoff was officially recognized and an assessment of the nature and extent of such pollution was mandated.[1,2] Agricultural runoff is grouped into the category of nonpoint source pollution because the potential pollutants originate over large, diffuse areas and the exact point of entry into water bodies cannot be precisely identified (see Pollution: Point Sources, p. 2190). Non-point sources of pollution are particularly problematic in that it is difficult to capture and treat the polluted water before it enters a stream. Point sources of pollution such as municipal sewer systems usually enter the water body via pipes and it is comparatively easy to collect that water and run it through a treatment system prior to releasing it into the environment. Because of the non-point source nature of agricultural runoff, efforts to minimize or eliminate pollutants are, by necessity, focused on practices to be applied on or near farm fields themselves. In other words, we usually seek to prevent the pollution rather than treating the polluted water. Due to the great successes made in treating polluted water from point sources such as municipal and industrial wastewater treatment plants, the relative significance of pollution from agricultural runoff has increased. Agricultural runoff is now considered to be the primary source of pollutants to the streams and lakes in the United States. It is also the third leading source of pollution in U.S. estuaries.[3] The water pollutants that occur in agricultural runoff include eroded soil particles (sediments), nutrients, pesticides, salts, viruses, bacteria, and organic matter.

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Agricultural Runoff Quantity Agricultural runoff occurs when the precipitation rate exceeds the infiltration rate of the soil. Small soil particles that have been dislodged by the impact of raindrops can fill and block soil pores with a resulting decrease in infiltration rate throughout the duration of the storm. As the excess precipitation builds up on the soil surface it flows in thin layers from higher areas of the field towards lower areas. This diffuse surface runoff quickly starts to concentrate in small channels called rills. The concentrated flow will generally have a higher velocity than the flow in thin films over the surface. The concentrated flow velocity may become rapid enough to cause scouring of the soil that makes up the channel sides and bottom. The dislodged soil particles can then be carried by the flowing water to distant locations in the same field or be carried all the way to a receiving water body. If the quantity of flow and the velocity of flow are large enough, the rills can grow so large that they cannot be easily repaired by typical earth moving machinery. When this happens, the rill has become a gulley. The quantity of runoff from agricultural fields is not usually listed explicitly as a concern separate from the quality of the runoff. However, it should be considered because it transports the pollutants and can cause erosion of receiving streams due to excessive flows. If less runoff is allowed to leave a field, there is less flow available to transport pollutants to the stream. Also, if more water is retained on the field, there is likely to be a corresponding reduction in the amount of supplemental water that will need to be added through irrigation. Runoff quantity varies significantly due to factors such as soil type, presence of vegetation and plant residue, physical soil structures such as contoured rows and terraces, field topography, and the timing and intensity of the rainfall event. Some agricultural practices increase the infiltration capacity of the soil while other practices can result in decreases. The presence of vegetation and plant residues on a field reduce runoff due to improving and maintaining soil infiltration capacity. Actively growing plants also reduce the amount of water in the soil due to evapotranspiration, thus making more room for infiltrating water to be stored in the soil profile. Bare soils increase runoff because there is nothing except the soil surface to absorb the energy of the falling raindrops. The rain, therefore, dislodges soil particles that will tend to seal the surface and reduce infiltration.

Soil erosion and Associated Pollutants One of the primary pollutants in agricultural runoff is eroded soil. In 1975, 223 million acres of cropland produced 3700 million tons of eroded sediments or an average of 17 tons of soil lost per acre of cropland per year (see various Erosion entries, pp. 967–1103). It is estimated that cropland, pasture, and rangeland contributed over 50% of the sediments discharged to surface waters in 1977.[4] As noted above, the energy of raindrops can dislodge and transport soil particles. In the aquatic environment the eroded soil is called sediment. There are several concerns related to excessive sediments in aquatic systems. These include loss of field productivity, habitat destruction, reduced capacity in reservoirs, and increased dredging requirements in shipping channels. Eroded sediments represent a loss of fertile topsoil from the field, which can reduce the productivity of the field itself. Soil formation is an extremely slow process occurring over periods ranging from decades to centuries.[5] Possible results to a grower from excessive erosion of their fields include increasing fertilizer and water requirements, planting more tolerant crops, and possibly abandoning the field for agricultural production. A second concern is that many of these sediments are heavy and will settle out in slow moving portions of streams or in reservoirs. The settled sediments can dramatically alter the ecology of the streambed. Aquatic plants, insects, and fish all have specific requirements related to composition of the streambed for them to live and reproduce.[6] Sediments in reservoirs reduce the volume of the reservoir available to store water. This may result in reduced production of hydroelectric power, reduced water

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availability for municipal supply, interference with navigation and recreation, and increased dredging requirements to maintain harbor navigability. Another concern with eroded sediments is that they can transport other pollutants into receiving waters. The plant nutrient phosphorus, for example, is most often transported from the fields where it was applied as fertilizer by chemically bonding to clay minerals. Many agricultural pesticides also bond to eroded clays and organic matter. Once these chemicals have entered the aquatic ecosystem, many processes occur that can result in the release of the pollutants from their sediment carriers. Phosphorus, when released, can contribute to the eutrophication of lakes and reservoirs (see the entry Eutrophication, p. 1115). Pesticides and their degradation products can be toxic to aquatic life and must be removed from municipal water supplies. Erosion from animal agriculture such as feedlots and pastures can also result in the transport of sediments composed of animal manures (see the various Manure Management entries, pp. 1680–1695). These sediments can transport significant quantities of potential pathogens (viruses and bacteria). The animal manures are primarily organic in nature and can serve as a food source for natural bacteria in the receiving water. When these naturally occurring bacteria begin to utilize the organic matter in this way they may lower or deplete the water of dissolved oxygen as they respire and multiply. This use of oxygen by aquatic bacteria is known as biochemical oxygen demand (BOD). High levels of BOD can reduce stream oxygen level to the point that fish and other organisms that require dissolved oxygen suffer, die, or relocate, when possible, to more suitable habitats.[6]

Dissolved Pollutants Agricultural runoff can carry with it many pollutants that are dissolved in the runoff water itself. These may include plant nutrients, pesticides, and salts. Since these pollutants are dissolved in the runoff, control measures are most often aimed at reducing the volume of runoff leaving an agricultural field, or making the pollutants less available to be dissolved into the runoff water. One of the major pollutants of concern in agricultural runoff is the plant nutrient nitrogen. Nitrogen is a relatively cheap component of most fertilizers and is necessary for plant growth. Unfortunately, nitrogen in the form of nitrate is highly soluble in water. Thus nitrate can be easily dissolved in runoff water. Just as it does in an agricultural field, nitrogen can promote growth of aquatic vegetation. Excess nitrogen and phosphorus in runoff can lead to the eutrophication of lakes, reservoirs, and estuaries (see the entry Eutrophication, p. 1115). Nitrogen in the form of ammonia can be dissolved into runoff from pastures and feedlots. Ammonia is toxic to many aquatic organisms, thus it is important to minimize ammonia in runoff.[7] Many agriculturally applied pesticides are also soluble in water. They can be dissolved in runoff and transported into aquatic ecosystems where there is a potential for toxic effects. These pesticides must also be removed from drinking water supplies and, if concentrations are high or persistent, such treatment can be difficult and expensive. Stable, persistent pesticides can bioaccumulate in the food chain with the result that consumers of fish from contaminated waters might be exposed to higher concentrations than exist in the water itself.[8] Runoff from agricultural fields can contain significant concentrations of dissolved salts. These salts originate in precipitation, irrigation water, fertilizers and other agricultural chemicals, and from the soil minerals. Plants generally exclude ions of chemicals that they do not need. In this way, dissolved salts in irrigation water, for example, can be concentrated in the root zone of the growing crop. Runoff can redissolve these salts and transport them into aquatic ecosystems where some, naturally occurring selenium for example, can be toxic to fish and other wildlife.[9] Transport of fertilizers and pesticides from their point of application can result in significant environmental costs. This transport, or loss from the field, can also have significant negative economic impacts on the grower. Fertilizers lost from the field are not available to promote crop growth. Agricultural

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chemicals lost from the field, likewise, are not available to protect the plants from pests and diseases. In both cases the grower is paying for expensive inputs and paying to apply them. It is always in the growers’ and the environment’s best interests, therefore, to keep agricultural chemicals in the field where they are needed and where they were applied.

control of Agricultural Runoff One of the most direct methods of controlling pollution by agricultural runoff is to minimize the potential for runoff to occur. Other methods can be employed to reduce the amounts of sediments and dissolved chemicals in runoff. As a whole, management practices designed to minimize the potential for environmental damage from agricultural runoff are called best management practices (BMPs), (see the entry Nutrients: Best Management Practices, p. 1829). Many times, practices aimed at controlling one aspect of agricultural runoff are also effective at reducing other components. This is due to the interrelationships between runoff volume, erosion, transport, dissolution, and delivery. Maintaining good soil tilth and healthy vegetation can minimize runoff. This will promote increased infiltration and a resultant decrease in runoff. Other management practices such as terracing, contour plowing, and using vegetated waterways to convey runoff can result in decreased quantities of runoff by slowing the water leaving the field and allowing more time for infiltration to occur. Construction of farm ponds to receive runoff can result in less total runoff from the farm, lowered peak rates of runoff, and storage of runoff for use in irrigation or livestock watering.[2] Control of water pollution by the mineral and organic sediments and associated chemicals in agricultural runoff is most effectively achieved by reducing erosion from the field. The primary method of reducing erosion is by maintaining a vegetative or plant residue cover on the field at all times or minimizing areas of the field that are bare. Techniques utilized to accomplish these tasks include conservation tillage, strip tillage, and the use of cover crops (see the entry Erosion Control: Tillage and Residue Methods, p. 1081). Additional measures that can be employed at the edge of the field, or off-site include vegetative filter strips and farm ponds. Methods to control the loss of nitrogen and other plant nutrients from cropland include applying nitrogen in the quantity required by the crop and at the time the crop needs it (see the entry Nutrients: Best Management Practices, p. 1829). This requires multiple applications and can be difficult for tall crops. For this reason, most, or all, of the nitrogen required by the crop is often applied at planting. Nitrogen fertilizers have often been applied based on general recommendations for the type of crop to be grown. Since nitrogen fertilizers are relatively inexpensive, growers have tended to over apply rather than under apply. Soil tests can tell a grower how much nitrogen is already in the soil and how much needs to be applied for a specific crop. Efforts have been made to make the nitrogen less soluble by changing the form of nitrogen applied to the field so that it becomes available to the plants (and, thus available for loss in runoff) more slowly.[10] One method of controlling the loss of agricultural chemicals is to minimize their solubility in water. Another is to minimize their use through programs such as integrated pest management (IPM) where some crop damage is allowed until it reaches a point that it becomes economically justified to apply pesticides.[11] And a third approach is to make the chemicals more easily degraded so that they do their job and then degrade into other, less harmful, chemicals so that they do not stay around long enough to be influenced by runoff-producing rainfall events.

conclusion Agricultural runoff is one of the leading causes of water quality impairment in streams, lakes, and estuaries in the United States. It can transport large quantities of sediments, plant nutrients, agricultural chemicals, and natural occurring minerals from farm fields into receiving water bodies. In many cases the loss of these substances from the field represent an economic loss to the grower as well as a potential

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environmental contaminants. There are many methods by which the quantity of agricultural runoff can be reduced. Many of these methods are referred to generically as BMPs. Adoption of BMPs can also improve the quality (reduce contaminant concentrations) of the runoff that does leave the farm. By reducing the quantity and improving the quality of agricultural runoff, it will be possible to improve the water quality in our streams, river, lakes, and estuaries.

References 1. U.S. Environmental Protection Agency. EPA Releases Guidelines for New Water Quality Standards; 2002; http://www.epa.gov/history/topics/fwpca/02.htm (accessed July 2002). 2. Stewart, B.A.; Woolhiser, D.A.; Wischmeier, W.H.; Caro, J.H.; Frere, M.H. Control of Water Pollution from Cropland, Volume II—An Overview, EPA-600/ 2–75-026b; U.S. Environmental Protection Agency: Washington, DC, 1976. 3. U.S. Environmental Protection Agency. Nonpoint Source Pollution: The Nation’s Largest Water Quality Problem; 2002; http://www.epa.gov/OWOW/NPS/facts/point1.htm (accessed July 2002). 4. Leeden, Van der. The Water Encyclopedia; Lewis Publishers: Chelsea, MI, 1990. 5. Foth, H.D. Fundamentals of Soil Science, 8th Ed.; John Wiley and Sons, Inc.: New York, 1990. 6. Gordon, N.D.; McMahon, T.A.; Finlayson, B.L. Stream Ecology: An Introduction for Ecologists; John Wiley and Sons Inc.: New York, 1992. 7. Abel, P.D. Water Pollution Biology, 2nd Ed.; Taylor and Francis, Inc.: Bristol, 1996. 8. U.S. Environmental Protection Agency. The Persistent Bioaccumulators Project; 2002; http://www. epa.gov/chemrtk/persbioa.htm (accessed July 15, 2002). 9. U.S. Geological Survey. Public Health and Safety: Element Maps of Soils; http://minerals.cr.usgs. gov/gips/na/0elemap.htm#elemap (accessed July 15, 2002). 10. Owens, L.B. Impacts of soil N management on the quality of surface and subsurface water. In Soil Process and Water Quality; Lal, R., Stewart, B.A., Eds.; Lewis Publishers, Inc.: Boca Raton, FL, 1994. 11. U.S. Department of Agriculture. National Integrated Pest Management Network; 2002; http:// www.reeusda.gov/agsys/nipmn/ (accessed July 15, 2002).

68 Desertification

David Tongway and John Ludwig

Introduction .................................................................................................. 587 Desertification Redefined ............................................................................ 587 Assessing the Degree of Desertification ....................................................589 Procedures to Reverse Desertification ....................................................... 589 Monitoring Rehabilitation........................................................................... 590 Conclusion .................................................................................................... 590 References ...................................................................................................... 590

introduction The term desertification was coined in the 19708 to graphically represent the state of the Sahelian lands, on the southern margin of the Sahara Desert. This was a period when major drought accompanied by big increases in the human population served to cause the desert margins to apparently move into formerly more productive land.[1,2] The image of an encroaching desert is powerful and evocative and resulted in major international efforts to understand and deal with the problem. Since that time, the notion of desertification has been reworked in the light of additional information and improved conceptual frameworks to the extent that the desert is no longer seen as inexorably increasing in size, nor restricted to the Sahelian fringe of the Sahara.[3–5] Most rangeland areas in the world have suffered some sort of degradation due to the impact of disturbance regimes, and recent reviews[6] have shown the process to be not at all restricted to hot deserts or areas of high population density. This is not to deny, however, the major effects on the human populations using these lands, and no doubt, much human hardship has been endured. This entry describes a process whereby the degree of desertification can be assessed and then used to design restoration activities appropriate to local biophysical and socioeconomic constraints.

Desertification Redefined This entry focuses on the biophysical aspects of desertification. Traditionally, easy to measure structural and functional attributes of vegetation, such as species composition and productivity, were the means by which desertification was initially assessed. These methods served to show the effect of desertification but did not provide a predictive understanding of how to combat it. However, recent advances in landscape ecology and restoration technology[7,8,10] have led to generic and practical approaches to study the basic nature and reversibility of desertification. Principally, this involves treating the affected landscapes as biophysical systems, comprised of sequences of processes and feedback loops and summarized in a conceptual framework (Figure 1).[8,9] In this framework, vital resources such as water, nutrients, and topsoil are transported, utilized and cycled in space and time, and the processes are driven by processes such as runoff/runon, erosion/deposition, and plant litter de- composition.[8,11,12] Landscapes are said to be “functional,” or 587

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FIGURE 1 A conceptual framework summarizing landscape functioning. Numbers refer to the recommended sequence of assessing practical function. The disturbance regime shown is generic, and it may impinge on a number of landscape processes at the same time. Source: A modification of Figure 2 in Tongway and Ludwig[9] (reproduced with permission).

non-desertified, if resources are substantially retained within the system and utilization and cycling processes are efficient. “Dysfunctional” or desertified landscapes are characterized by the depletion of the stock of some vital resources and the continued flow of these resources out of the system. This mind-set emphasizes the system attributes of processes acting in space, over time, in relation to applied stress and disturbance, rather than just changes in lists of species, or yields of marketable commodities. The role of vegetation as a significant regulator of energy and resources is integral to this approach.[13] Desertification should be viewed as a continuum ranging from slight to severe, rather than as a simple step function (Figure 2).

FIGURE 2 Desertification as a continuum. The four new biophysical parameters (bold) are added to the two existing desertification descriptors (dotted box) to locate any given landscape on the continuum.

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Assessing the Degree of Desertification If field sites are characterized according to “resource regulation” capacity, not only can the degree of desertification be assessed, but also the critical pathway of resource loss may be identified. Tongway and Hindley[14] and Tongway and Ludwig[15] have designed and implemented monitoring programs to quickly provide information about biophysical processes and edaphic properties related to plant habitat favorability at both landscape and plot or patch scales. Typically, the initial analysis examines the fate of rainfall into infiltrated water and runoff water. The data gathered need an interpretational framework to facilitate generic application. Graetz and Ludwig[16] proposed that system response to desertification be represented by a four-parameter sigmoidal or logistic curve. The curve form acknowledges upper and lower asymptotic plateaux, at the non- and highly desertified ends of the spectrum, respectively, and a gradual transition between those plateaux. This approach permits questions about whether a system was “fragile,” i.e., easily made dysfunctional, with low restoration potential or “robust,” or rather able to withstand stress and disturbance with only low attenuation of biophysical processes (Figure 3). Importantly, this curve type enables thresholds and milestones to be predicted and quantified using field indicators.[17]

Procedures to Reverse Desertification Rehabilitation and restoration of desertified landscapes, under the functional biophysical system mind-set, require that processes that accumulate resources be reintroduced or augmented, thus providing a rational procedure in the repair and functional recuperation of desertified landscapes. The approach explicitly seeks to retain vital resources by ecological processes.[18–20] Once the analysis of resource regulation system has been completed, and the most affected process identified, remedial efforts can begin. For example, Rhodes and Ringrose-Voase[20] deduced that ponding water for extended periods on clay soils with modest swell/shrink properties would eventually result in soils with high infiltration and water store through sequences of swelling and shrinking processes permitting infiltration into greater soil depths. Recolonization and establishment of plants then began spontaneously, eventually cycling organic matter, so that open friable soils developed, colonized by soil macrofauna that further improved soil properties. Tongway and Ludwig[18,19] used piles of branches arranged on the contour in gently sloping country to trap water, sediment, and organic matter to effect substantial improvements to a range of soil properties, permitting perennial grasses to self-establish. In each of these cases, an analysis of the different underlying causes facilitated the selection of the most appropriate techniques to reverse

FIGURE 3 Examples of response curves for fragile and robust landscapes. The initial response of landscape function to stress and/ or disturbance is markedly different. A fragile landscape deteriorates with low applied stress and has a much lower base value (y0) than the more robust landscape. Four-parameter sigmoid curves of the form y = (y0 + a)/1 + e(−x−x0)/b provide four practical values reflecting the nature of the landscape. Critical thresholds (arrows) for each index of desertification can be determined by curve analysis.

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the observed desertification. Attempting to revegetate desertified areas by simply reseeding without understanding both the current and required edaphic properties needed for the desired vegetation mix frequently results in unexplained failure. In some instances, where the system function is close to the lower asymptote (Figure 3), simple treatments such as exclusion from grazing will be too slow or ineffective and active intervention may be needed to improve one or more functional processes.

Monitoring Rehabilitation It is important to monitor the progress of processes set in train by the rehabilitation activities. Essentially, the degradation curves such as those in Figure 3 need to be driven “in reverse.” The landscape assessment procedure proposed by Tongway and Hindley and Tongway and Ludwig[15,21] can also be used to follow the trajectory of improvement in ecosystem functioning. The procedures use simple, rapidly acquired indicators of processes of resource regulation. Data recording biota establishment and development are included in this procedure and interpreted in terms of the rising plane of delivery of goods and services to the whole system over time. It is important that monitoring should provide accurate information quickly and at low cost. Remotely sensed data, related to landscape function, is a cost-effective procedure,[22] and new products such as Google Earth may, in the future, provide synoptic assessment of restoration trends at coarse scales. The effect of rare stochastic events such as fires or storms may need to be assessed to see if the resultant stress and/or disturbance has set the system back beyond a critical threshold or not.

conclusion We have described an ecosystem-function-based set of data gathering processes by which the fundamental nature of desertification can be assessed and combated, using a framework that characterizes the biophysical status of the affected system. This can be simply expressed as “assessing the regulation of vital resources in space and time.” In deploying the procedure, community groups can be easily instructed to “read the landscape.” The procedure enables the user to design or adapt restoration or rehabilitation technologies appropriate to the problem at hand because of the predictive understanding acquired, rather than using a “recipe” from another type of landscape elsewhere. The informationgathering procedure can be adapted to a wide range of bioclimatic situations because it deals with the basic processes controlling the availability of vital resources to biota.

References 1. UNEP. United Nations Conference on Desertification: An Overview; United Nations Environment Program: Nairobi, Kenya, 1977. 2. UNEP. General Assessment of Progress in the Implementation of the Plan of Action to Combat Desertification 1978–84; United Nations Environment Program: Nairobi, Kenya, 1984. 3. Arnalds, O. Desertification: An appeal for a broader perspective. In Rangeland Desertification; Arnalds, O., Archer, S., Eds.; Kluwer Academic: Dordrecht, 2000; 5–15. 4. Chen, Z.; Xiangzhien, L. In People and Rangelands: Building the Future, Proceedings of the VI International Rangeland Congress, Townsville, Australia, July 19–23, 1999; Eldridge, D., Freudenberger, D., Eds.; VI International Rangeland Congress: Townsville, 1999; 105–107. 5. Tongway, D.; Whitford, W. In People and Rangelands: Building the Future, Proceedings of the VI International Rangeland Congress, Townsville, Australia, July 19–23, 1999; Eldridge, D., Freudenberger, D., Eds.; VI International Rangeland Congress: Townsville, 1999; 89–142. 6. Archer, S.; Stokes, C. Stress, disturbance and change in rangeland ecosystems. In Rangeland Desertification; Arnalds, O., Archer, S., Eds.; Kluwer Academic: Dordrecht, 2000; 17–38.

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7. Breedlow, O.A.; Voris, P.V.; Rogers, L.E. Theoretical perspective on ecosystem disturbance and recovery. In Shrub-Steppe: Balance and Change in a Semi-Arid Terrestrial Ecosystem; Rickard, W.H., Rogers, L.E., Vaughn, B.E, Liebetrau, S.F., Eds.; Elsevier: New York, 1988; 257–269. 8. Ludwig, J.A; Tongway, D.J. A landscape approach to rangeland ecology. In: Landscape Ecology Function and Management: Principles from Australia’s Rangelands; Ludwig, J., Tongway, D., Freudenberger, D., Noble, J., Hodgkinson, K., Eds.; CSIRO: Melbourne, Australia, 1997; 1–12. 9. Tongway, D.; Ludwig, J. Restoring Disturbed Landscapes: Putting Principles into Practice; Island Press: Washington DC, 2010. 10. Whisenant, S.G. Repairing Damaged Wildlands: A Process- Oriented, Landscape-Scale Approach; Cambridge Univ. Press: Cambridge, England, 1999; 312 pp. 11. Jorgenson, S.E.; Mitsch, W.J. Ecological engineering principles. In Ecological Engineering: An Introduction to Ecotechnology; Mitsch, W.J., Jorgensen, S.E., Eds.; John Wiley and Sons: New York, 1989; 21–37. 12. Schlesinger, W.H.; Reynolds, J.F.; Cunningham, G.L.; Hueneke, L.F.; Jarrel, W.M.; Virginia, R.A.; Whitford, W.G. Biological feedbacks in global desertification. Science 1990, 247, 1043–1048. 13. Farrell, J. The influence of trees in selected agroecosystems in Mexico. In Agroecology: Researching the Ecological Basis for Sustainable Agriculture; Gliessman, S.R., Ed.; Springer-Verlag: New York, 1990; 167–183. 14. Tongway, D.J.; Hindley, N.L. Landscape Function Analysis: Methods for Monitoring and Assessing Landscapes, with Special Reference to Minesites and Rangelands; CSIRO Sustainable Ecosystems:Canberra, 2004, available at http://www.csiro.au/services/EcosystemFunctionAnalysis. htm. 15. Tongway, D.; Ludwig, J. Restoring Disturbed Landscapes: Putting Principles into Practice; Island Press: Washington DC, 2010. 16. Graetz, R.D.; Ludwig, J.A. A method for the analysis of piosphere data applicable to range assessment. Aust. Rangeland J. 1978, 1, 126–136. 17. Tongway, D.; Hindley, N. Ecosystem Function Analysis of Rangeland Monitoring Data; 2000, available at http://www.nlwra.gov.au/atlas. 18. Tongway, D.J.; Ludwig, J.A. Rehabilitation of semi-arid landscapes in Australia. I. Restoring productive soil patches. Restor. Ecol. 1996, 4, 388–397. 19. Ludwig, J.A.; Tongway, D.J. Rehabilitation of semi-arid landscapes in Australia. II. Restoring vegetation patches. Restor. Ecol. 1996, 4, 398–406. 20. Rhodes, D.W.; Ringrose-Voase, A.J. Changes in soil properties during scald reclamation. J. Soil Conserv. Serv. NSW 1987, 43, 84–90. 21. Tongway, D.; Hindley, N. Assessing and monitoring desertification with soil indicators. In Rangeland Desertification; Arnalds, O., Archer, S., Eds.; Kluwer Academic: Dordrecht, 2000, 89–98. 22. Kinloch, J.E.; Bastin, G.N; Tongway, D.J. (2000). Measuring landscape function in chenopod shrublands using aerial videography. In the Proceedings of the 10th Australasian Remote Sensing and Photogrammetry Conference, Adelaide, Australia 21–25 August 2000, 480–491; Causal Productions: Adelaide, South Australia.

69 Desertification: Prevention and Restoration

Claudio Zucca, Susana Bautista, Barron Orr, and Franco Previtali

Introduction .................................................................................................. 593 Desertification: The Underlying Concepts................................................ 593 Prevention and Reversal .............................................................................. 594 Implementing Solutions in the Field: Lessons Learned .......................... 595 Integrated Evaluation of Prevention and Restoration Actions ............... 597 Participatory Approaches ............................................................................ 598 Conclusion .................................................................................................... 599 References ..................................................................................................... 600

introduction This entry provides a brief overview of the most significant approaches to prevent and reverse land degradation in drylands, with a special emphasis on conceptual frameworks and on methods to monitor and evaluate their impacts. The entry first discusses the main underlying concepts, in light of recent developments on the conceptual framework of desertification. The methodological developments related to integrated and participatory evaluation are then presented. Finally, the implementation of mitigation and restoration programs is addressed, with particular reference to the constraints and risk factors and to the practical lessons learned in the field.

Desertification: the Underlying concepts Desertification is defined here as “land degradation in drylands resulting from various factors, including climatic variations and human activities,” in conformity with the UNCCD (United Nations Convention to Combat Desertification).[1] Before the entry into force of the UNCCD in 1996, the term “ desertification” had been given a number of different definitions.[2–6] The UNCCD definition has been and still is the subject of scientific debate; the related evolving concepts have been reviewed by a number of papers.[7–12] The Millennium Ecosystem Assessment (MA) defines desertification as a persistent reduction in the provision of ecosystem services over an extended period.[13,14] The scientific discussion promoted in 2009 by the UNCCD in view of the First Scientific Conference of its CST (Committee on Science and Technology), building on the MA definition, resulted in a proposal to redefine desertification as “an end state of the process of land degradation; this process is expressed by a persistent reduction or loss of biologic and economic productivity of lands that are under use by people.”[15] The ongoing discussion also focuses on the integrated sets of indicators needed for monitoring and assessing desertification,

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and on the related conceptual frameworks that would help scientists, practitioners, and policy makers organize, use, and communicate the results of that monitoring.[16–18] While the definition of desertification continues to evolve, ecosystem services are increasingly seen as a unifying supporting concept. The MA states that “desertification results from a long-term unbalance between demand for and supply of ecosystem services” and that measurement of persistent reduction in the capacity of ecosystems to supply services provides a robust and operational way to quantify land degradation and desertification. The ecosystem services framework is increasingly thought to provide a basis to assess and value the impacts of land change and degradation, as well as the effects of the actions aimed at reversing it.[19–21] Desertification manifests itself through different forms and processes in different ecosystems and socioeconomic contexts. Its direct causes are many and can be generally ascribed to different forms of land mismanagement, such as overgrazing, deforestation, overuse of irrigation, and non-resourceconservative agriculture and forest practices.[1,7,17] These are at the origin of the major land degradation processes that are globally affecting the provision of ecosystem services, including water and wind erosion, soil salinization, loss of vegetation cover and diversity, and degradation of the hydrological cycles.[13] No satisfying estimates of the global extent and severity of desertification are available thus far; however, the new World Atlas of Desertification, which is based on multiscale integrated sets of indicators, provides global datasets that can be used to identify local or regional areas of concern.[22] Assessment and monitoring of desertification, as well as of the impact of the prevention and restoration interventions, still constitute a major research challenge[15] and for this reason are a primary focus of this entry.

Prevention and Reversal Policy and management responses to desertification can be grouped under two major classes: prevention and reversal.[13] The boundaries between these ones are vague, as in practice they form a continuum of potential prevention, mitigation, and restoration actions, to be adapted to particular sites and dynamics through adaptive management approaches (Figure 1). Through integrated land use planning, the optimal spatial mix of prevention and reversal responses can be identified, so that in net terms, desertification can be controlled. This approach, which was endorsed by the country Parties of the UNCCD in 2017, is known as land degradation neutrality (LDN).[23–25] Prevention actions can be considered as avoidance approaches either through proactive management efforts or through changes in land use and management currently leading to desertification. Prevention, however, is not enough to address the challenges posed by desertification. Vast areas of drylands are already severely degraded, with reduced plant cover and species diversity, falling

FIGURE 1

Continuum of actions to combat desertification and land degradation.

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productivity, depleted or eroded soils, and very low potential for spontaneous recovery of ecosystem functions, even when degradation forces are no longer stressing the system. Many of these systems have changed at a level at which restoration is the only viable option to recover ecosystem services that have been lost.[26–28] Examples of prevention and mitigation actions include measures to improve water management and agricultural practices. These are often referred to as soil and water conservation (SWC) or sustainable land management (SLM) practices. The FAO–LADA (Food and Agriculture Organization of the United Nations/Land Degradation Assessment in Drylands)[29] project and WOCAT (World Overview of Conservation Approaches and Technologies)[30] provide a framework for classifying and evaluating SLM actions. In view of the needs of the UNCCD, SLM has been recently defined as “land managed in such a way as to maintain or improve ecosystem services for human wellbeing, as negotiated by all stakeholders.”[21] Examples include long-term crop rotations with cereals/legumes; more efficient use of fertilizers; improvements in water-use efficiency; conservation-minded tillage methods; water-harvesting and water storage techniques; measures that protect soils from erosion, salinization, and other forms of soil degradation; improved crop–livestock integration, combining livestock rearing and cropping to allow a more efficient recycling of nutrients within the agricultural system; and in situ conservation of genetic resources and better resource use with efficient germplasm.[31] Creating viable livelihood alternatives, including the creation of economic opportunities in urban centers, can also help reduce current pressures on drylands. For extremely degraded lands, rehabilitation and restoration approaches often involve the improvement in the quantity and/or quality of vegetation cover through, for example, the establishment of seed banks, reintroduction of selected species, control of invasive species, and reforestation programs. Desertification is driven by a combination of proximate causes and underlying forces, including their interactions and feedback; these vary from region to region and change over time.[32] Approaches and strategies to prevent and reverse desertification need therefore to address the dynamic causal patterns and multiplicity of actors, factors, and scales involved. In general, developing the appropriate engagement between scientific and local environmental knowledge is critically important for efforts to prevent and reverse land degradation.[33] Desertification is framed within multiscale, coupled human–environmental dynamics, and so must be the approaches for desertification prevention and reversal.[34,35] The relationships between land degradation and its causal agents are non-linear and complex. Degradative and aggradative trajectories commonly exhibit thresholds and rapid shifts, as well as hysteresis dynamics, where the trajectories of degradation and recovery differ.[36–38] Understanding and monitoring these relationships are critical in the design of strategies to combat desertification. In addition, socioeconomic conditions impose limitations on the technology and inputs available. Therefore, the approaches to combat desertification should incorporate both current conditions and scenario projections of socioeconomic and environmental constraints and opportunities.[13] Finally, there is growing evidence that land degradation in desertification-prone areas is likely to increase with climate change.[22] Desertification is linked to biodiversity loss and global climate change through the regulation of water and carbon fluxes. Therefore, interlinkages in policy formulations aimed at combating desertification, mitigating the effects of climate change, and conserving biodiversity must be beneficially exploited by developing multifunctional strategies that address the three global environmental goals.

implementing Solutions in the Field: Lessons Learned Despite the availability of technological, institutional, and even financial resources, efforts to combat desertification often fail because of poor implementation. A list of “lessons learnt” is presented below to summarize a range of major constraints and risk factors that can hinder the successful implementation of the interventions, while highlighting the lessons learned and necessary improvements. Some of

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those issues stem from the concepts of adaptive management and multiscale human–environmental dynamics as introduced in the previous section. Others are more related to the issues of local participation and integrated assessment that will be discussed in the following section. Points 1 and 2 deal with the quality of the technical design and its degree of adaptation to local conditions and knowledge. Points 3–7 are related to the ability of the projects to cope with the long-term human–environmental dynamics. Points 8–11 are connected to the quality of the participatory processes implemented by the projects, while the last two points are linked to major, common organizational constraints. 1. Lack of awareness of actual land conditions. Sometimes, people (including decision makers) living in degraded areas do not realize that their land still maintains productive potential or perhaps are unaware of how that potential can be exploited. Preliminary land surveys should be done to support project design. These baseline assessments are necessary to guide subsequent actions. In some cases, “no-action” options could be considered. In the case of degraded rangelands, simple and cost-effective “self-learning” tests based on grazing exclusion and rotation can be proposed to local communities as a means to demonstrate the effects of pressure mitigation and sustainable management. Such tests allow hands-on learning and practical experience, and are much less risky than the direct introduction of often expensive “all or nothing” rehabilitation programs. 2. Schematic approach. Sometimes, large programs are extensively implemented by adopting schematic approaches, which are not able to adapt to specific local land characteristics and the needs of local people. In other cases, they target geographic areas where they are not necessary or inadvertently lead to negative side effects. Implementation protocols offering multiple technical solutions and integrating local stakeholder input on the perceived benefits (and unintentional consequences) are needed. 3. Lack of long-time planning in restoration. Often the long-term dynamics of the “restored human–environmental systems” are not fully considered. This is particularly relevant when the interventions are based on introducing fast-growing, income-generating alien species that may require a “renaturalization” strategy in the long term to balance ecological and social sustainability. 4. Inability to cope with natural crises. Especially in projects aimed at increasing plant cover, a poor or delayed wet season or recurrent droughts may cause the loss of part or all of the investment in a project. For example, this may happen in the case of fodder shrubs plantations, when farmers cannot avoid early grazing due to drought and unavailability of alternative feeding resources. Emergency/contingency funds, quick diagnostic and intervention mechanisms, and flexibility in project duration are necessary to mitigate against this risk. 5. Socioeconomic and demographic dynamics. Some projects have experienced labor shortfalls due to the out-migration of young people. In contrast, the return of people onto land previously “closed for restoration” may cause unsustainable pressure. Addressing economic constraints and associated demographic pressures such as migration prior to project implementation can reduce such kind of risks. 6. Market drivers. The dynamics of international and local market prices may completely and quite rapidly negate achievements produced by years of conservation programs. 7. Contrasting policies. A restoration initiative may be useful in practical terms and yet be overwhelmed by unrelated policies. A common example is farming incentives that come in conflict with the goals of mitigation programs, if not accompanied by adequate guidelines. 8. Passive community participation. Community participation should be strongly based on responsible awareness and sharing of project objectives, and participation in its planning, implementation, and evaluation. Cases where, for example, sectoral administration goals are implemented through prior agreement with land users (e.g., “I let you do on my land”) may require or even force farmer action but, in the end, may not represent or address the key concerns or needs of local people.

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9. Uncertain community commitment. Stakeholder engagement that fails to bring all parties to the table can have dire consequences. Comprehensive, balanced, and approved representation is not easy to obtain, but it is essential for success. The key question often raised is, “Who is committing on behalf of whom?” The commitment should include a community contribution or investment to cover the implementation costs, be linked to final results, and be based on taking ownership and responsibility after a project ends, rather than simply for the completion of individual tasks. Community commitment should be clearly defined in a way that can cope with changing community priorities. 10. Institutional commitment. Complex projects, in which the implementation is based on the support of local administration staff, need a formal and clear institutional commitment. This may take time and very rarely can be established before project approval: Projects should have an inception phase to set agreements. 11. Lack of transparency in the engagement process of the ultimate project beneficiaries. This is especially true (but not only) when the involvement of individual farmers or other land users in projects is not mediated by the community. Inappropriate or unrepresentative participation can lead to failure and loss of credibility. Transparent and objective selection mechanisms must be understood by all and be followed rigorously by project implementers. 12. Unrealistic project duration. Very often, the most common investment period imposed by donors (3–5 years) is too short to allow for adequate biophysical response and/or socioeconomic adaptation that would assure success. Of particular concern is enough time for effective “ inception or learning phases” for community members to deal with the necessity to adapt to changing conditions and needs, to cope with unforeseen events, and most of all to monitor and assess impacts and sustainability. 13. Spending and reporting difficulties. Spending rules, procurement procedures, “exotic” rules imposed by the donors, or strict local rules, not well known by the local project management, often lead to delays and underperformance. More in general, the lack of flexibility can exacerbate the effects of most of the above-mentioned difficulties. Finally, our capacity to design effective projects is often undermined by a lack of integrated assessment of the progress and success of the previous projects. Projects that are adequately implemented, monitored, and evaluated generate impact, and useful lessons that can feed into successful scaling strategies.[39]

integrated evaluation of Prevention and Restoration Actions The development of integrated biophysical and socioeconomic analytical methods for evaluating progress and success, along with a framework for knowledge sharing and transfer, is crucial to combating desertification.[34] Furthermore, monitoring and evaluation are needed to demonstrate the benefits of sustainable dryland management, establish cost-effective thresholds for the various management alternatives, and identify priority areas where actions could be most effective. In recent years, there have been a number of initiatives to develop common and comprehensive methodologies for assessing and evaluating the effectiveness of management and restoration programs.[40–44] These approaches focus on indicators that relate to ecosystem integrity and services, and human well-being (socioeconomic and cultural variables). Irrespective of the biophysical or socioeconomic attributes assessed, the selected indicators should be relevant, be sensitive to variations of environmental stress, have the capacity to respond to stress in a predictable manner, but also be simple and measurable with a reasonable level of effort and cost.[45,46] Soil conditions and vegetation cover and composition are the most common metrics used for evaluating mitigation and restoration actions.[47,48] However, used in isolation, these indicators cannot always reflect how well an ecosystem is functioning. During the last decade, a variety of functional assessment approaches that relate to the spatial pattern of vegetation have been proposed.[49–51]

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FIGURE 2 actions.

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Example of a multiscale integrated framework and indicators for evaluating mitigation and restoration

Some of these functional assessment methods also incorporate properties relative to the soil surface condition.[52,53] The theoretical framework for these approaches considers that landscapes occur along a continuum of functionality from highly patchy systems that conserve all resources to those that have no vegetation patches and leak all resources.[54] For semiarid ecosystems, it has been hypothesized that vegetation patchiness could be used as a signature of imminent transitions[55] and that changes in patch-size distributions may be a warning signal for the onset of desertification.[56] Evaluation frameworks must account for the cross-scale and social–ecological interactions affecting the response of degraded lands to mitigation/restoration actions.[13] A multiscale approach is always advisable. Farm- and project-scale assessments focus on local resources and productivity, and a private economic valuation perspective (market-priced goods and services), while landscape- and program-level indicators address environmental benefits and public/social welfare considerations (Figure 2). Because of the large spatial and temporal variability of ecosystems, particularly in drylands, it is critical to focus on “slow variables,”[57] so the assessment of long-term changes, and of the sustainability of land management, is not confused by short-term variations in land and socioeconomic conditions.[34] Assessment methods range from simple, qualitative assessments based on field observations to relatively complex protocols based on quantitative measurements of critical ecosystem and landscape attributes and socioeconomic surveys.[42] The development and accessibility of remote-sensing (RS) products have led some international bodies to recommend the integrated use of RS-based geospatial information with ground-based observations to assess land degradation.[13,15,58,59]

Participatory Approaches While substantive progress has been made in developing more standardized and more relevant environmental assessment and monitoring approaches that reflect human–ecological interactions, the adoption of evaluation results at the local level remains challenging.[35,60] Though some suggest that local interests or even national policies outside the control of evaluators may partially explain this, environmental assessments tend to be independent, unilateral, and top-down, with results being delivered post-assessment. Stakeholder engagement is rare, or all-too-often limited to a period of “public comment” immediately before and/or after the assessment. In the same way that adaptive governance and ecosystem-based management approaches require local knowledge and continual stakeholder engagement, so too must the corresponding monitoring and evaluation if the goal is for results to be truly embraced and used.[61–63] Adaptive, ecosystems-based management is data intensive and requires

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a commitment to a variety of data sources, including local knowledge, which in turn necessitates more attention to knowledge integration and methods of analysis at different scales. Adoption of best practices that in theory come from a well-executed evaluation of a given desertification mitigation or restoration action can fail due to poor or limited communication and knowledge exchange among the involved actors. Essentially ignoring local knowledge increases the risk of missing essential key local factors, opportunities, or constraints. Engaging stakeholders from the outset of an environmental assessment and maintaining the interaction throughout the process can result in the integration of local people and their perceptions into management, planning, and evaluation, helping develop feelings of ownership and representation while giving voice to locals in the process.[33,64,65] The potential benefits go beyond the assessment itself. Analysis of environmental conflict resolution processes suggests that ensuring all parties are at the table and are effectively engaged is directly correlated with often-sought outcomes like reaching an agreement, the quality of agreement, and improved working relationships among parties.[66,67] An additional benefit to a participatory approach to the evaluation and the exchange of ideas among stakeholders, including researchers, is the learning that takes place. The collective self-reflection through interaction and dialogue among the diverse set of stakeholders involved with or affected by environmental challenge and the assessment of associated responses can result in the coproduction of knowledge.[68–70] Social groups that develop a shared understanding of a challenge can build up the experience necessary to improve linkages between knowledge and the environment, cope with change, and enhance adaptation because social learning helps solidify knowledge systems made up of the relevant sets of actors, networks, or organizations.[71] While the majority of discussion about the benefits of encouraging social learning has been focused on improving adaptive management, it is clear that the wealth of information explored during an evaluation suggests that an assessment period is an ideal time to encourage stakeholder interaction and knowledge exchange. In this sense, the evaluation itself becomes a tool for outreach and inreach, where land users, natural resource managers, and scientists all stand to learn from and potentially benefit from each other’s insights. The process has the potential to empower individuals, build relationships, expand networks, and thereby enhance the relevance and impact of decision making.

conclusion Desertification is one of the major global environment and sustainable development challenges. It affects the livelihoods of millions of people, threatening human well-being in drylands.[13] That risk, when considered relative to the foreseeable impacts of climate change, stands to grow significantly in the future, with estimates suggesting that as many as 50 million people will be in peril of physical displacement in the next ten years.[72] This entry underscores a range of conceptual and practical issues influencing the design and implementation of interventions, with a special emphasis on conceptual frameworks and associated methods to monitor and evaluate impacts. Lessons learned in the field highlight several issues threatening the success of prevention/reversal actions. The capacity to design and implement effective restoration actions and other countermeasures to desertification is often undermined by a lack of assessment of the outcomes of previous projects. The assessment of actions is complex and requires conceptual tools such as human–environmental frameworks for integrated assessment and participatory approaches to foster social learning. These should become the basis for the development of multifunctional mitigation strategies and the formulation of interlinked policies to address desertification, climate change, and biodiversity. The development of participatory approaches in the assessment of interventions is critical to capacity building and knowledge exchange. Encouraging social learning and a true sense of ownership is essential for successful adaptive management.

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Other aspects, lying beyond the scope of the present entry, are crucial to combating desertification. Primarily, the assessment of intervention programs should involve the policy makers to promote the adoption of sustainable rural development policies and to counteract the socioeconomic and policy-driven dynamics of desertification.[73] As a second priority, future interventions should be increasingly oriented to income-generating actions to strengthen social and economic sustainability in concert with environmental sustainability.[74] Finally, it is important to keep in mind that practices and interventions can be considered as “good” or “best” only with reference to their suitability in relation to specific human–environmental contexts. In this regard, to develop strategies tuned to the changing land features, land evaluation techniques could be updated based on the integrated assessment principles, as suggested by the new “anthroscape” concept.[75]

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54. Ludwig, J.A.; Tongway, D.J. Viewing rangelands as landscape systems. In Rangeland Desertification; Arnalds, O., Archer, S., Eds.; Kluwer Academic Publishers: Dordrecht, 2000; 39–52. 55. Rietkerk, M.; Dekker, S.C.; de Ruiter, P.C.; van de Koppel, J. Self-organized patchiness and catastrophic shifts in ecosystems. Science 2004, 305 (5692), 1926–1929. 56. Kéfi, S.; Rietkerk, M.; Alados, C.L.; Pueyo, Y.; Papanas-tasis, V.P.; El Aich, A.; de Ruiter, P.C. Spatial vegetation patterns and imminent desertification in Mediterranean arid ecosystems. Nature 2007, 449 (7159), 213–217. 57. Carpenter, S.R.; Turner, M.G. Hares and tortoises: Interactions of fast and slow variables in ecosystems. Ecosystems 2000, 3 (6), 495–497. 58. ICCD/COP8/CST. Report of the Fifth Meeting of the Group of Experts of the Committee on Science and Technology. Addendum: Methodologies for the Assessment of Desertification at Global, Regional and Local Levels. ICCD/COP(8)/CST/2/Add.6, 2008. 59. Zucca, C.; Wu, W.; Dessena, L.; Mulas, M. Assessing the effectiveness of land restoration interventions in drylands by multitemporal remote sensing – A case study in Ouled Dlim (Marrakech, Morocco). Land Degrad. Develop. 2015, 26, 80–91. 60. Turner, B.L., II; Kasperson, R.D.; Matson, P.A.; McCarthy, J.J.; Corell, R.W.; Christensen, L.; Eckley, N.; Kasperson, J.X.; Luers, A.; Martello, M.L.; Polsky, C.; Pulsipher, A.; Schiller, A. A framework for vulnerability analysis in sustainability science. Proc. Natl. Acad. Sci. U.S.A. 2003, 100 (14), 8074–8079. 61. Slocombe, D.S. Defining goals and criteria for ecosystem-based management. Environ. Manage. 1998, 22 (4), 483–493. 62. Folke, C.; Hahn, T.; Olsson, P.; Norberg, J. Adaptive governance of social–ecological systems. Annu. Rev. Environ. Resour. 2005, 30, 441–473. 63. Curtin, R.; Prellezo, R. Understanding marine ecosystem based management: A literature review. Mar. Policy 2010, 34 (5), 821–830. 64. Johannes, R.E. Integrating traditional ecological knowledge and management with environmental assessment. In Traditional Ecological Knowledge: Concepts and Cases. International Program on Traditional Ecological Knowledge; Inglis, J., Ed.; Canadian Museum of Nature: Ottawa, Canada, 1993. 65. Brush, S.B.; Stabinsky, D. Valuing Local Knowledge: Indigenous People and Intellectual Property Rights; Island Press: Washington, DC, 1996. 66. Emerson, K.; Orr, P.J.; Keys, D.L.; McKnight, K.M. Environmental conflict resolution: Evaluating performance outcomes and contributing factors. Conflict Resolut. Q. 2009, 27 (1), 27–64. 67. Orr, P.J.; Emerson, K.; Keys, D.L. Environmental conflict resolution practice and performance: An evaluation framework. Conflict Resolut. Q. 2008, 25 (3), 283–301. 68. Bautista, S.; Llovet, J.; Ocampo-Melgar, A.; Vilagrosa, A.; Mayor, A.G.; Murias, C.; Vallejo, V.R.; Orr, B.J. Integrating knowledge exchange and the assessment of dryland management alternatives – A learning-centered participatory approach. J. Environ. Manag. 2017, 195 (1), 35–45. 69. Pahl-Wostl, C.; Hare, M. Processes of social learning in integrated resources management. J. Commun. Appl. Soc. 2004, 14 (3), 193–206. 70. Schusler, T.M.; Decker, D.J.; Pfeffer, M.J. Social learning for collaborative natural resource management. Soc. Nat. Resour. 2003, 16 (4), 309–326. 71. Steins, N.A.; Edwards, V.M. Platforms for collective action in multiple-use common-pool resources. Agric. Hum. Values 1999, 16 (3), 241–255. 72. Rechkemmer, A. Societal impacts of desertification: Migration and environmental refugees. In Facing Global Environmental Change: Environmental, Human, Energy, Food, Health and Water Security Concepts; Brauch, H., BerghofStiftung, K., Eds.; Springer: Berlin, 2009. 73. Zdruli, P.; Trisorio Liuzzi, G., Eds. Managing Natural Resources through Implementation of Sustainable Policies. Proceedings of the MEDCOASTLAND Euro-Mediterranean Conference, Beirut, Lebanon, Jun. 25–30, 2006; IAM: Bari, 2007.

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74. Zdruli, P.; Trisorio Liuzzi, G., Eds. Determining an Income-Product Generating Approach for Soil Conservation Management. Proceedings of the MEDCOASTLAND International Seminar, Marrakech, Morocco, Feb. 12–16, 2004; IAM: Bari, 2005. 75. Eswaran, H.; Berberoglu, S.; Cangir, C.; Boyraz, D.; Zucca, C.; Özevren, E.; Yazici, E.; Zdruli, P.; Dingil, M.; Dönmez, C.; Akça, E.; Çelik, I.; Watanabe, T.; Koca, Y.K.; Montanarella, L.; Cherlet, M.; Kapur, S. The anthroscape approach in sustainable land use. In Sustainable Land Management— Learning from the Past for the Future; Kapur, S., Eswaran, H., Blum, W.E.H., Eds.; SpringerVerlag: Berlin Heidelberg, 2011.

70 Erosion

Dennis C. Flanagan

Introduction ..................................................................................................605 Erosion by Wind .......................................................................................... 606 Erosion by Water ......................................................................................... 606 Gravity-Induced Erosion .............................................................................607 Erosion Assessment..................................................................................... 608 Erosion Impacts ........................................................................................... 609 Erosion Control and Soil Conservation ................................................... 609 Conclusions ................................................................................................... 610 References ...................................................................................................... 610

introduction Soil erosion is the detachment or breaking away of soil particles from a land surface by some erosive agent, most commonly water or wind, and subsequent transportation of the detached particles to another location. Usually, erosion occurs when a fluid (air or water) moves into and/or across a soil surface. Fluid and sediment particle impact forces, shear forces, and turbulence act to detach and lift soil into the fluid flow that then transports the particles away (Figure 1). The force of gravity moves detached soil particles downward, while cohesive forces between soil particles resist detachment and transport. Physical and chemical dispersion can disrupt cohesion and break soil aggregates into smaller and more easily transported particles. At some time and location away from the initial point of detachment, the sediment particles will eventually move back down to a state of rest on a soil surface, in a process known as deposition or sedimentation.

FIGURE 1

Soil erosion by wind, showing the three modes of movement (creep, saltation, and suspension).

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erosion by Wind Erosion by wind occurs when wind speed exceeds a certain critical or threshold value. Soil particles can be detached and moved through suspension, saltation, or creep (Figure 1). Suspension usually lifts the smallest soil particles (clays, silts, organic matter) so high into the air mass that they are easily kept in motion and can travel for long distances. Soil particles that move by creep are larger sand grains and aggregates that stay in contact with the soil surface at almost all times—their motion is often through rolling and bouncing. Saltating soil particles are usually moderate in size and, once detached, move in trajectories up into the air and then back down to the soil surface. Saltating particles often cause further detachment through abrasion by striking the soil surface with sufficient momentum to dislodge additional soil particles from the in situ soil mass.

erosion by Water The most common types of soil erosion by water are sheet and rill erosion on upland areas, channel and gully erosion in small watersheds, and stream channel and bank erosion in larger catchments. Sheet erosion is caused by the action of raindrops (Figure 2) and shallow overland flows that remove a relatively uniform depth (or sheet) of soil. Because of the uniform nature of the soil loss, it is often difficult to detect and gauge the extent of damage caused by sheet erosion. On the other hand, rill erosion occurs in well-defined and visible flow concentrations or rills (Figure 3). Soil detachment in rills is large because of flow shear stress forces acting on the wetted perimeter of the rill channel (Figure 4). Once detached, larger sediment particles move as bedload, rolling and bouncing down slope with the flow, and are almost always in contact with the soil (or bed) surface. Smaller sediment particles (silts and clays) are much easier to transport and travel in the rill channels as suspended load. Rills are also the major pathways for transporting away sediment that is detached by sheet erosion (also known as interrill detachment). By definition, rill channels are small enough to be obliterated by tillage and will not reform in exactly the same location. As one moves from smaller hillslopes to larger fields and watersheds, additional erosion processes come into play, because of the increasing amounts of runoff water. Gullies are incised erosion channels that are larger than rills and form in regions of large runoff flow concentration. Ephemeral gullies are a common type of erosion feature in many fields (Figure 5). They are small enough to be tilled over but will re-form in the same location owing to the convergent topography in small catchments. Runoff flows from large events can erode down through tilled soil layers, until a nonerodible layer is reached, and then the ephemeral gully channel will widen and soil detachment will decrease. Classical

FIGURE 2

Soil detachment by raindrop impact and shallow flow transport.

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FIGURE 3 Rill erosion, which is caused by concentration of flowing water, forms easily recognizable regions of detachment on a soil surface. (Courtesy M. Huhnke, Oklahoma State University.)

FIGURE 4

Soil detachment and transport in rills are largely because of flow shear forces.

gullies are larger erosion features that cannot normally be tilled across (Figure 6). The physical processes in classical gullies include other factors such as headcutting, seepage, sidewall sloughing, and clean-out of fallen sidewall materials. As the size of watersheds increases further, and streams increase in size and become perennial (because of subsurface water flows from springs and aquifers), the erosion processes in play change as well. Stream and channel erosion at these larger scales can include scouring of the channel beds as well as the contributions from the channel banks. Areas in streams may be in states of degradation, in which active detachment is lowering the level of the channel bed, or they may be in states of aggradation, in which sediment deposition is raising the bed level.

Gravity-induced erosion There are also less frequent but more extreme forms of gravity-induced erosion on steep slopes from saturated soils that can be exacerbated by events such as earthquakes. Large masses of land can slowly or rapidly slide down hills when the cohesive forces holding them in place fail (landcreep, landslide, debris flow, etc.).

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FIGURE 5

Typical ephemeral gully located in a soybean field in Indiana.

FIGURE 6 hillslopes.

Classical gullies in the Loess Plateau of China. Terrace farming is being used to stabilize some of the

These types of erosion events typically occur when large rainfall or snowmelt water depths saturate soil profiles and weaken their resistance to slip.

erosion Assessment Erosion is a serious problem within the United States and throughout the world. In 2015, the United States Department of Agriculture (USDA)-Natural Resources Conservation Service estimated that about 1.7 billion tons of soils are lost each year from nonfederal rural croplands because of sheet and rill erosion by water and erosion by wind.[1] Also, over 20% of cropland in the United States is eroding at excessive rates. These estimates are on the low side because erosion of other types (e.g., gully) and at other locations (urban lands, federal lands) were not included in this inventory. Throughout the world, FAO estimates that 16% of the total land area (21,960,000 km2 of 134,907,000 km2) is subject to significant risk of soil erosion.[2] In Asia, South America, and Africa, soil erosion rates are highest at an estimated average of 30–40 t/ha/yr, while in Europe and North America average rates are somewhat lower at about 17 tons/ha/yr.[3,4] A sustainable rate of soil loss (rate of soil loss is equal to rate of soil formation) is thought to be about 1 t/ha/yr.[3]

Erosion

609

Erosion assessment can be a difficult task to perform in the field, and monitoring of soil lost and transported by wind or water can be expensive and prone to measurement errors. Gullies are easy to recognize, while soil lost to sheet and rill erosion is hard to gauge. Sheet and rill erosion may be occurring on hill slopes that are adjacent to a gully and may actually contribute more sediment to runoff water than the gully itself. Visual assessment of rates of wind erosion losses can be even more difficult to perform. Mathematical equations or sets of equations have been developed and used since the mid-1900s to estimate the rates of soil loss caused by wind (“Wind Erosion Equation”[5]) or water (“Universal Soil Loss Equation”[6,7]). More recently, computer models are being applied to simulate soil erosion processes and to estimate detachment, transport, and deposition of sediment.[8,9,10,11]

erosion impacts Erosion has a range of impacts, both on-site and off-site. Soil loss removes fertile topsoil, organic matter, and nutrients, thus decreasing the tilth, water-holding capacity, and general productivity of a soil for on-site agricultural production. Regions of detachment can expand to dislodge and remove small crop seedlings, while regions of deposition can bury and kill small plants. In the case of wind erosion, the erosion process can damage fragile young seedlings through abrasion of plant tissue. When excessive detachment occurs, such as is the case with gully erosion, whole sections of fields may be destroyed or may become inaccessible to farmers and their equipment. Eroded sediment can cause a number of off-site problems, including deposition along windbreaks, ditches, and waterways. The deposited sediment may require costly dredging and removal operations. Nutrients and pesticides associated with sediments can also contaminate air and water bodies. Erosion by wind can cause massive dust storms that blind drivers and cause accidents, and sand particles can abrade and damage painted surfaces on buildings and vehicles. Some recent estimates are that the cost of combined on-site and off-site effects from soil erosion in the United States is as high as $44 billion per year.[3]

erosion control and Soil conservation Many nations have created government agencies or organizations to specifically deal with soil erosion problems and to interact with landowners to get conservation practices implemented on the landscape. In the United States, the Natural Resources Conservation Service assists in the implementation of soil conservation practices on agricultural lands, the Forest Service manages sediment delivery from forests and timber harvest roads, and the Bureau of Land Management manages soil loss on range and grazing lands. The Department of Defense is responsible for managing erosion and off-site sediment delivery from lands that it uses for military training activities. However, in some countries, efforts to address and minimize erosion problems are nonexistent or severely limited because of the poor economic conditions, failure to recognize the erosion threat, and/or the extreme magnitude of the soil erosion. A variety of soil conservation methods are available that can be applied on a landscape to minimize erosion problems caused by wind or water. Wind erosion can be controlled through the use of windbreaks, crop residues, and tillage to induce significant surface roughness. Control procedures for erosion by water need to be determined, based upon the types of active erosion processes. For example, if sheet and rill erosion is a major problem, then some type of conservation tillage practice that leaves large amounts of crop residues intact on the soil surface may be appropriate. However, if the water erosion problem is because of the large amounts of surface runoff concentrating in a field and forming an ephemeral gully, then crop residues may not be adequate; instead, permanent vegetative cover in a grass waterway may be necessary, along with appropriate engineering structures (e.g., drop-box). Erosion prediction models can be used to assist in selecting and designing appropriate conservation practices.

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Through the use of proper conservation planning and application of appropriate soil conservation methods, most erosion problems can be minimized or eliminated. This is critically important if the soil resource is to be preserved for continuous use in food and fiber production for current and future generations.

conclusions Erosion is a natural process of soil detachment and removal that can be greatly influenced through human activities (agriculture, construction, timber-harvesting, etc.). Use of proper erosion prediction technology and appropriate erosion control methodologies is critical if we are to sustain the soil resource for use by future generations.

References 1. U.S. Department of Agriculture. Summary Report: 2015 National Resources Inventory; Natural Resources Conservation Service: Washington, DC, and Center for Survey Statistics and Methodology, Iowa State University: Ames, Iowa, 2018; Available at: https://www.nrcs.usda.gov/ Internet/FSE_DOCUMENTS/nrcseprd1422028.pdf (accessed April 2019). 2. Food and Agricultural Organization of the United Nations. World Soil Resources Report 90; FAO Land and Water Development Division: Rome, 2000; Available at: http://www.fao.org/3/a-x7126e. pdf (accessed April 2019). 3. Pimental, D.; Harvey, C.; Resosudarmo, P.; Sinclair, K.; Kurz, D.; McNair, M.; Crist, S.; Shpritz, L.; Fitton, L.; Saffouri, R.; Blair, R. Environmental and economic costs of soil erosion and conservation benefits. Science 1995, 267, 1117–1123. 4. Barrow, C.J. Land Degradation; Cambridge University Press: Cambridge, 1991. 5. Woodruff, N.P.; Siddoway, F.H. A wind erosion equation. Proc. Soil Sci. Soc. Am. 1965, 29, 602–608. 6. Wischmeier, W.H.; Smith, D.D. Predicting Rainfall Erosion Losses—A Guide to Conservation Planning; Agriculture Handbook No. 537; USDA: Washington, DC, 1978. 7. Renard, K.G.; Foster, G.R.; Weesies, G.A.; McCool, D.K.; Yoder, D.C. Predicting Soil Erosion by Water: A Guide to Conservation Planning with the Revised Universal Soil Loss Equation (RUSLE); Agriculture Handbook No. 703; USDA: Washington, DC, 1997; 384 pp. 8. Flanagan, D.C.; Nearing, M.A., Eds. USDA-Water Erosion Prediction Project (WEPP) Hillslope Profile and Watershed Model Documentation; NSERL Report No. 10; National Soil Erosion Research Laboratory, USDA-Agricultural Research Service: West Lafayette, IN, 1995; 298 pp.; Available at: https://www.ars.usda.gov/midwest-area/west-lafayette-in/national-soil-erosionresearch/docs/wepp/wepp-model-documentation/ (accessed April 2019). 9. Flanagan, D.C.; Gilley, J.E.; Franti, T.G. Water Erosion Prediction Project (WEPP): Development history, model capabilities, and future enhancements. Trans. ASABE 2007, 50 (5), 1603–1612. 10. Hagen, L.J. A wind erosion prediction system to meet user needs. J. Soil Water Conserv. 1991, 46 (2), 106–111. 11. Wagner, L.E. A history of wind erosion prediction models in the United States Department of Agriculture: The Wind Erosion Prediction System (WEPS). Aeolian Res. 2013, 10, 9–24.

71 Erosion by Water: Erosivity and Erodibility

Peter I.A. Kinnell

Introduction ................................................................................................... 611 Variants of the USLE .................................................................................... 612 Rainfall Kinetic Energy................................................................................ 616 More Process-Based Models ....................................................................... 617 Effect of Particle Travel Rates on Sediment Composition and Erodibility ................................................................................................620 Conclusion .................................................................................................... 623 References ...................................................................................................... 623 Bibliography .................................................................................................. 625

introduction Conceptually, rainfall erosivity is the capacity of rain to produce erosion, whereas soil erodibility is the susceptibility of the soil to be eroded. Historically, the terms erosivity and erodibility were originally associated with the R and K factors in the Universal Soil Loss Equation (USLE), A = R   K   L   S  C   P

(1)

where A is the long-term (e.g., 20 years) annual average soil loss per unit area from sheet and rill erosion, R is the rainfall (erosivity) factor defined as the average annual value of the product of the total storm kinetic energy (E) and the maximum 30-min intensity (I30, twice the maximum amount of rain that falls in any 30-min period during a storm), K is the soil (erodibility) factor, L is the slope length factor, S is the slope gradient factor, C is the crop (vegetation) and crop management factor, and P is the conservation support practice factor.[1] Numerically, soil erodibility is the mass of soil eroded per unit of the erosive index. This means that numerical values of K can only be used when R is as it was originally defined as the average annual value of the product of E and I30. The USLE was designed to predict sheet and rill erosion from field-sized areas and only R and K have units. The L, S, C, and P factors each have values of 1.0 for the so-called “unit” plot, a bare fallow area 22.1 m long on a 9% slope with cultivation up and down the slope. Consequently, the soil loss for the “unit” plot (A1) is given by A1 = R   K

(2)

A = A1  L   S  C   P

(3)

and, for any other situation,

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Although, traditionally, K is calculated by dividing A1 by R, K can be perceived as acting similarly to the regression coefficient in the direct relationship between event soil loss on the unit plot (A1e) and EI30. However, regression analysis of the relationships between storm soil losses on the unit plot and EI30 tends to produce regression coefficients that may differ appreciably from those calculated using the traditional method. Consequently, the traditional approach should always be used in determining K values from runoff and soil loss plot data. To reduce the need to run long-term experiments to determine K values for soils where K is unknown, Wischmeier[2] developed a nomograph for determining K from soil properties for soils with less than 70% silt in the United States. Alternatively, K values in customary U.S. units for soils where the nomograph can be used may be obtained using

(

)

K = 2.1  X11.14 10−4 (12 − X 2 ) + 3.25( X 3 − 2 ) + 2.5( X 4 − 3) /100

(4)

where X1 is % silt multiplied by 100 – % clay, X2 is % organic matter, X3 is the soil structure code used in the U.S. soil classification, and X4 is the profile permeability code. A number of other equations have been developed for soils at various geographic locations (e.g., El-Swaify,[3] Young and Mutchler,[4] Loch,[5] and Zhang et al.[6]), but Eq. (4) is frequently used outside the United States without being validated for the soils involved. Division of the right-hand side of Eq. (4) by 7.59 will yield K values in SI units of t h /MJ/mm. The two-staged mathematical approach shown by Eqs. (2) and (3) results from the fact that the USLE is an empirically based model. It was developed in the 1960s and 1970s from more than 10,000 plot years of data. Mathematically, the Revised Universal Soil Loss Equation (RUSLE[7]) uses Eqs. (2) and (3) in the same way as the USLE, but changes were made to how some of the factors in the model are calculated. Originally, in the USLE, the events used to calculate R were restricted to those that produced more than 12.5 mm of rain or at least 6.25 mm of rain in 15 min. That rule was abandoned in the RUSLE when R values were calculated for the western part of the United States because it was argued that the rule had no appreciable effect on R values. Yu[8] noted that changing the threshold to zero increased the R factor by 5% in the tropical region of Australia. However, the abandonment of the rule in the RUSLE failed to recognize that that rule had been put in place as a means of discounting storms that tended to be nonerosive because they produced no runoff.

Variants of the USLe Eq. (2) operates on the assumption that a direct linear relationship exists between event erosion (Ae) and EI30. Although this assumption is appropriate at some geographic locations, it is not appropriate at others (Figure 1). Soil measured as a loss from the plots used to develop the USLE was discharged in runoff. When runoff occurs, the amount of soil discharged (Qs) in the runoff can be considered to be the product of the amount of water discharged (Q w) and the sediment concentration (cs), the amount of soil per unit quantity of water, Qs = Q w c

(5)

It follows from Eq. (5) that, in the USLE, the sediment concentration is directly proportional to EI30 divided by runoff amount. However, it has been observed[10] that the sediment concentration for an event at some geographic locations is correlated to the product of the kinetic energy per unit quantity of rain and I30 (Figure 2). This results in an event erosivity index (Re) that is given by Re = QRe EI 30

(6)

where QRe is the runoff ratio (runoff amount divided by rainfall amount) for the event. This index is more effective in accounting for event soil losses at some locations where the EI30 index does not work

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FIGURE 1 The relationships between event erosion (Ae) and EI30 obtained on bare fallow plots in the United States: (a) Plot 5 Experiment 3 at Holly Springs, Mississippi, and (b) Plot 5 Experiment 1 at Morris, Minnesota. Eff(ln) is the Nash–Sutcliffe[9] efficiency factor for logarithmic transforms of the data.

well (Figure 3). Because Re is not equal to EI30, USLE K factor values cannot be used in conjunction with Eq. (6). Soil erodibilities associated with Eq. (6) are calculated dividing A1 by the average annual sum of the product of QRe and EI30. A number of other event erosivity indices have been proposed. Williams[11] proposed a modification of the USLE to predict event sediment yield (SYe): SYe = 11.8( qe qp )

0.56

K   L   S  Ce   Pe

(7)

where qe is the volume of runoff (m3) for the event; qp is the peak flow rate (m3/sec); K, L, and S are standard USLE factors; and Ce and Pe are event C and P factors, respectively. This model is commonly known as the Modified Universal Soil Loss Equation (MUSLE). It should be noted the value of 11.8 in

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FIGURE 2 The relationship between event sediment concentration (ce) and the product of I30 and the kinetic energy per unit quantity of rain (E/re) for Plot 5 Experiment 1 at Morris, Minnesota.

FIGURE 3 The relationship between event erosion and the product of event runoff ratio (Q Re) and EI30 for Plot 5 Experiment 1 at Morris, Minnesota.

Eq.  (7) was generated for the specific situation where Williams[11] obtained measured data. Consequently, the general applicability of this value is questionable. In Agricultural Policy/Experimental eXtender Model (APEX),[12] SYe = X e K   L   S  Ce   Pe [ RKOF ]

(8)

X e = EI 30

(9a)

Xe is selected from

X e = 1.586 ( qe qpe )

0.56

DA0.12

(9b)

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X e = 0.65  EI 30 + 0.45( qe qpe )

(9c)

X e = 2.5( qe qpe )

0.5

(9d)

DA0.009

(9e)

0.33

X e = 0.79( qe qpe )

0.65

X e = b5 qeb 6  qeb 7  DAb 8

(9f)

where DA is drainage area expressed in hectares, b6‒b8 are user-selected coefficients,[12] and RKOF is the coarse fragment factor as defined by Simanton et al.[13] However, Eqs. (7) and (8) with Eqs. (9b‒9f) all use USLE K factor values and not ones that are associated with the different erosivity indices used and so do not conform with the mathematical modeling rules upon which the USLE model is based. Consequently, the MUSLE and APEX models are not valid variations of the USLE. In addition, any model that uses event erosivity index values that are calculated using runoff from a vegetated area or any area that is not cultivated up and down the slope will violate the mathematical rules if it uses USLE C and P factor values. Mathematical models like the USLE and its derivatives are largely designed to aid management decisions and operate at a level where spatial and temporal variations in the various forms of erosion (splash erosion, sheet erosion, rill erosion, interrill erosion) are not considered in any appreciable detail. However, rill erosion is driven by flow energy, while sheet and interrill erosion are associated more closely with rainfall kinetic energy. In order to better account for this, Onstad and Foster[14] used the equation Re = 0.5EI 30 + 0.5α ( qe qp )

0.333

(10)

Eq. (10) indicates that the sediment concentration for an event (ce) in the data set considered by Onstad and Foster[14] is given by ce = 0.5 [[ EI 30 /qe ]] + 0.5 [α ( qp ) [

0.333

/qe 0.666 ] ]

(11)

The value of α in Eq. (10) was set so that the long-term average value of Re produced by Eq. (10) was the same as the long-term average value produced when Re = EI30. Given that K is, by definition, the amount of soil loss per unit of R when Re = EI30, this enabled Eq. (10) to be used with USLE K values to predict soil loss from bare fallow areas. Although Williams et al.[15,16] indicated that Onstad and Foster[14] was the source of Eq. (9c), there is no provision in APEX to do the same and ensure that USLE K values can be used. Bagarello et al.[17] observed that Re = (QRe EI 30 )β

(12)

with β > 1.0 applied to soil losses from a number of simultaneously operating plots of different length (λ) established at the experimental station of Sparacia, Sicily. Subsequent analysis [Bagarello et al.[18]] established that β > 1.0 appeared to be the result of an interaction between event runoff (Qe) and slope length (λ) on sediment concentration so that Re = QRe EI 30 Qe 0.0207λ

(13)

when λ is in meters. Arguably, from a physical viewpoint, since slope length and gradient influence the erosive stress, they can be considered to be factors that influence erosivity, but the USLE model is not designed to

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model the physical processes themselves. It is designed to predict rainfall erosion based on climate, soil, topography, crops, and management factors. Normally, in the USLE model, the effect of slope length on soil loss is expressed through the L factor, but in the case of Eq. (13), L would remain at 1.0 irrespective of variations in slope length. The positive effect of runoff on sediment concentration may have been associated with the fact that the data were collected on a 15% slope. Rainfall erosion on much lower slope gradients is not associated with as high an erosion stress from runoff as likely to have been the case at the Sparacia site. In fact, the RUSLE makes provision for the cushioning effect of water depth on the energy of raindrop impact on low slopes. Because, as noted above, Re is not equal to EI30, USLE K factor values cannot be used in conjunction with Eqs. (12) and (13). Soil erodibilities associated with Eq. (12) are calculated dividing A1 by the average annual sum of (QRe EI30)β. β varies between geographic locations in Italy, and consequently, the units for the associated soil erodibilities vary geographically. That is inconsistent with the concept of erodibility in the USLE modeling system. However, Kinnell (2018) showed that Eq. (12) can be rewritten as Re   = a1  (QRe EI 30 )

β

(X1)

where a1 is the ratio of the average annual value of (QRe EI30) to the average annual value of (QRe EI30)β. a1 and β are related mathematically to each other, a1 being equal to 1.0 when β equals 1.0, less than 1.0 when β is greater than 1.0, and greater than 1.0 when β is less than 1.0. When Eq. (X1) is used, soil erodibilities determined by dividing A1 by average annual value of (QRe EI30) apply at all geographic locations irrespective of the value of β.

Rainfall Kinetic energy Data on storm rainfall kinetic energy are not measured at many geographic locations, and often, kinetic energy values are obtained indirectly. If data on storm rainfall intensities are available, then it is common for storm kinetic energies to be estimated from rainfall kinetic energy–intensity relationships. These relationships are based on raindrop-size data collected during rainfall events at a geographic location assumed to have rainfall characteristics that are consistent with those at the geographic location of interest. Various techniques have been employed to do this, and these have generated rainfall intensity– kinetic energy relationships which are often used at locations that lie outside the climatic zones where the measurements were originally made. In the USLE, the relationship recommended between the kinetic energy per unit quantity of rain (e) and rainfall intensity (I) for the United States is expressed by e = 916 + 331 log10 I , for I < 3  in. / h e = 1074, for I ≥ 3  in. / h

(14a) (14b)

where e is in units of foot-tons per acre per inch and I is in inches per hour, following analysis of dropsize data collected by Law and Parsons.[19] In the RUSLE, Eq. (14) is replaced by e = 1099(1  −0.72 exp[ −1.27 I ])

(15)

em = 0.29(1  −0.72 exp[ 0.05 I m ])

(16)

whose metric equivalent is

where em has units of megajoules per hectare per millimeter and Im has units of millimeters per hour. Eqs. (15) and (16) use the mathematical form proposed by Kinnell.[20,21] Yu[8] observed that replacing

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Eq. (14) by Eq. (15) reduced the R factor by about 10% in the tropical region of Australia. A number of other mathematical equations have been reviewed by van Dijk et al.[22] Some of these produce negative values of rainfall kinetic energy at low rainfall intensities. In RUSLE2 (USDA, 2013), em = 0.29(1  −0.72 exp[ 0.08 I m ])

(X2)

Both Eq. (16) and Eq. (X2) produce response curves that increase non-linearly from low values at low intensities with em values remaining close to 0.29 MJ/ha/mm for intensities beyond 75‒100 mm/h, but Eq. (X2) gives higher values of em below 75 mm/h. Eq. (X2) was adopted because it produced results that were more consistent with the USLE equation (Nearing et al., 2017). However, short-term values measured during rainstorms vary greatly about the values produced by Eq. (X2) or any other rainfall intensity‒em relationship. Also, although Eqs. (15), (16), and (X2) were developed for use in the United States, they have been used in other parts of the world without validation. In reality, storm kinetic energies calculated from rainfall intensity–kinetic energy relationships are often not close to actual storm kinetic energies but are, in effect, numerical values that are biased toward high-intensity rainfall at the expense of low-intensity rainfall. In many geographic locations, there is a lack of rainfall intensity data so that it is not possible to determine storm kinetic energies using rainfall kinetic energy–intensity relationships. Event or daily rainfall amounts are more commonly recorded. One approach that has been used in a number of geographic areas such as Canada,[23] Finland,[24,25] Italy,[26] and Australia[27] considers that EI30 is related to a power of event rainfall amount (Xp), EI 30 = a1 X Pb1

(17)

where a1 and b1 are empirical constants. The value of a1 may show seasonal variation.[25,28] Given that, in the context of the criteria associated with the USLE/RUSLE model, daily rainfall provides a reasonable proxy for storm rainfall amount[29]; Eq. (17) provides a practical approach to extending observed EI30 values to areas where appropriate rainfall intensity data are lacking. Although determining R is a primary requirement for modeling soil loss using the USLE approach, temporal variations in erosivity during the year are required in order to account for the effects of crops and crop management on erosion. C = 1 when soil is bare, and C = 0 when vegetation completely protects the soil against erosion. Consequently, how erosivity and the protective effect of crops vary during the year are factors that are taken into account in determining C for various agricultural systems. In RUSLE2, a factor known as “erosivity density” is used (USDA, 2013). Erosivity density is EI30 divided by storm rainfall amount. While rainfall storm rainfall amounts are highly variable in time and space, erosivity densities are more readily mapped in time and space than EI30 itself. Because storm energy is highly influenced by the amount of rain energy that occurs when I30 is measured, storm erosivity densities tend to be highly influenced by variations in I30.

More Process-Based Models Although the USLE/RUSLE is the most widely used method of predicting soil losses from the land worldwide, rainfall erosion results from various forms of erosion (splash erosion, sheet erosion, rill erosion, interrill erosion) that are driven by different forces so that there is no absolute measure of either rainfall erosivity or soil erodibility. Consequently, more process-based models have been developed in order to predict the contributions of the various forms of erosion more directly. Often, the forms vary in a topographic sequence with splash erosion dominating erosion at the upper end of a slope and sheet erosion further down before areas of rill and interrill erosion. Particles detached at the top of the slope may be transported by a number of different transport mechanisms before being finally discharged from the eroding area. The Water Erosion Prediction Program in the United States generated the WEPP

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model, a more process-based model than the USLE/RUSLE designed to model the spatial contributions of rill and interrill erosion in agricultural landscapes.[30] Flow shear stress (τ) is used as the erosivity factor in rill erosion model, Dr = kr (τ − τ c ) (1 − qsr /Tr )

(18)

where Dr is rill detachment, kr is the rill erodibility factor, τc is the critical shear stress that has to be exceeded before detachment occurs, qsr is the sediment load in the flow, and Tcr is the sediment load at the transport limit. For erosion to occur, particles must be plucked from within the soil surface where they are held by cohesion and interparticle friction. Detachment is the term used to refer to this process. For detachment in rills to occur, the flow must have shear stress that exceeds τc. Also, for erosion to occur, detached particles must be transported away from the site of detachment. Flows are known to have a limited capacity to transport soil material and that limit is represented by Tr in Eq. (18). Consequently, the term 1‒qsr/Tr causes Dr = 0 when the sediment load in the rill equals the transport limit. As a result, Dr may vary along the length of a rill. Interrill erosion contributes to qsr so that rill erosion may be completely suppressed if the discharge of sediment from the interrill areas is high enough. Originally, the erosivity factor in the WEPP interrill model was assumed to be the square of rainfall intensity (I) so that Di = ki I 2

(19)

where Di is interrill detachment and ki is the interrill soil erodibility factor. A series of rainfall simulation experiments[31] were undertaken to determine ki values for soils in the United States. Subsequent analysis[32] of the data generated by these experiments established that it was more appropriate to use Di = kiq qi I

(20)

where qi is the runoff rate from the interrill area and kiq is the interrill erodibility factor. The ki values determined using Eq. (19) cannot be used in Eq. (20).[32] Di is determined using data on the amounts of soil material discharged from interrill areas. The use of the term “detachment” in relation to Di is not absolutely correct because the transport processes involved in moving soil material over interrill areas are not 100% efficient. Also, detachment varies with raindrop size and velocity but, in using Eq. (20), that effect is ignored. However, provision is made in WEPP to take account of differences in drop-size distribution associated with different rainfall simulators in experiments designed to determine values of kiq. Eqs. (18) and (20) provide the basic equations that enable WEPP to model the contributions of rill and interrill erosion to soil loss more directly than when the USLE/RUSLE model is used but, in WEPP, the spatial distribution of rill and interrill areas needs to be well defined. WEPP was originally designed to operate in situations where ridge tillage is used. With ridge tillage, interrill erosion is well defined as being on the ridge slopes and rill erosion as being in the furrows. In other situations, the distributions of rill and interrill areas are more difficult to specify. In rangelands, broad areas of surface water flow occur more commonly than rills so that splash and sheet erosion dominate. An alternative to the WEPP cropland interrill erosion model, DSS = kSS I 1.05 qSS 0.59

(21)

where Dss is the “detachment” associated with splash and sheet erosion, kss is the erodibility factor, and qss is the runoff rate from the associated eroding surface, has been developed for use in rangelands.[33] EUROSEM is another erosion model that is more process based than the USLE/RUSLE. In EUROSEM, detachment by raindrop impact (Dr) is modeled using

Erosion by Water: Erosivity and Erodibility

Dr =

kD  KE exp( − zh ) ρs

619

(22)

where kD is an index of detachability that is determined experimentally, ρs is particle density, KE is the rainfall kinetic energy of the raindrops impacting the ground, z is a factor that varies with soil texture, and h is the mean depth of the water layer on the soil surface. Detachment by flow is modeled using Df = β w   v s ( C T − C )

(23)

where β is an index of detachability, w is flow width, vs is particle settling velocity, C is the sediment concentration, and Ct is the sediment concentration at the transport capacity of the flow. The coefficients used in Eq. (22) result from measurement of soil material transported by splash under ponded conditions in the experiments of Torri et al.[34] where increases in water depth were observed to reduce splash erosion. The effect of water depth on splash erosion results from (1) dissipation of raindrop energy in the water layer and (2) the effect of water depth on splash trajectories. Considering that splash does not transport 100% of the material detached by raindrops impacting water, Eq. (22) does not actually model detachment. Also, erosion by rain-impacted flows where sediment is transported by rolling, saltation, and suspension in the flow is the real focus of Eq. (22), and the effect of flow depth on erosion by rain-impacted flows is quite different from its effect on splash erosion.[35] As a general rule, the values of soil erodibility factors have to be determined experimentally, although there are cases where they are predicted from soil properties. In areas where detachment results from raindrop impact, erodibility is affected by modification of the soil surface generated by the impacts. Raindrop impacts on surfaces not covered by water may break soil aggregates and generate surface crusts that affect soil erodibility. In areas where soil crusts occur, particles are held more tightly within the soil surface than in areas that are not crusted and this reduces detachment. With splash erosion, loose particles sit and wait on the soil surface between drop impacts. The transport efficiency of splash increases with slope gradient but, over time, a layer of loose particles builds up on the soil surface, and energy has to be expended in moving them before detachment can occur. Consequently, this layer provides a degree of protection against detachment. When raindrops impact a soil surface covered by water, the protection provided by loose particles sitting on the surface is in addition to that provided as the result of the dissipation of raindrop energy in the water layer. Although splash erosion may dominate erosion for considerable periods during a rainfall event, rainimpacted flows are usually more important in moving soil material across the soil surface because the transport mechanisms in rain-impacted flows are much more efficient than splash transport. In rainimpacted flows, particles move across the soil surface by rolling, saltation, and complete suspension. Shallow low-velocity flows often do not have the capacity to cause particles to move by rolling and saltation by themselves, but rolling and saltation can be stimulated to occur when raindrops impact the soil surface through the flow. Under these circumstances, each rolling or saltation event is of limited duration and is associated with individual raindrop impacts. Figure 4 illustrates how raindrop and flow factors influence the detachment and transport processes associated with the erosion of fine particles, silt, and sand by rainfall. Particles moving by raindrop-induced rolling and saltation move across the surface at rates that depend on raindrop size, impact frequency, particle size, and density and the velocity of the flow. Consequently, depending on the rain and flow conditions, particles in rain-impacted flows travel across the surface at virtual velocities that vary from near zero to the velocity of the flow. In effect, particles are winnowed from the soil surface at various rates. Large gravel particles may not move at all and so, over time, soils that have high gravel contents may become highly resistant to erosion through the formation of erosion “pavements.” Particles moving by raindrop-induced rolling and saltation sit on the soil surface between drop impacts and so provide a degree of protection against detachment as described in the case of splash erosion. As a result, soil resistance to erosion may vary considerably during a rainfall event. If H is the degree of protection provided by the loose material, then all the material

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FIGURE 4 Schematic of how variations in raindrop kinetic energy and flow shear stress affect the detachment and transport of silt, sand, and fine particles during rainfall erosion. Ec is the critical raindrop kinetic energy required to cause raindrop detachment (RD). Its variation during rain with no runoff signifies changes in resistance to detachment caused by, for example, the development of a soil crust. Its variation with flow shear stress signifies the cushioning effect of increasing water depths. τc (bound) is the critical shear stress required to cause flow detachment (FD). τc (loose) is the critical shear stress required to cause flow-driven saltation (FDS). FS is continuous suspension in the flow. RIS is raindrop-induced saltation. ST is splash transport.

discharged from the soil surface comes from the layer of loose material when H = 1.0. Consequently, it follows from Eq. (20) that, for any given rainfall event, Di = ( kiq.U (1 − H ) + Hkiq.P )qi I

(24)

where kiq.U is the interrill erodibility factor when no loose material exists on the surface and kiq.P is the soil erodibility factor when loose material completely protects the soil surface from detachment. As noted earlier, detachment of soil particles from a cohesive surface varies with cohesion and interparticle friction so that factors such as the development of surface crusts can cause kiq.U to vary with time. The soil erodibility term in Eq. (24) is kiq.U(1‒H) + H kiq.P, and values of kiq obtained in rainfall simulator experiments like those undertaken by Elliot et al.[31] lie between kiq.U and kiq.P. Where exactly they do lie is unknown because H is unknown.

effect of Particle travel Rates on Sediment composition and erodibility As noted above, erodibilities have units of soil loss per unit of the erosivity index used in the model that is being considered in the analysis of the data. Factors such as cohesion, particle size, and aggregate stability have been observed to influence these erodibilities. Data on the physical and chemical nature of the soil involved may, in some cases, be used to predict erodibility, but often little attention is given to the composition of the sediment discharged in experiments undertaken to determine erodibility. The composition of the soil transported from an eroding area by rain-impacted flow during an experiment tends to be finer than the original soil (e.g., Meyer et al.,[36] Miller and Baharuddin,[37] Palis,[38] and Parsons[39]), and particle travel rate is one of the factors influencing sediment composition. Fast-moving

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621

FIGURE 5 Enrichment factor curves for sediment discharged at 2 and 60 min when beds of sand were eroded by rain-impacted flows in experiments undertaken by Walker et al.[40] using 2.7 mm raindrops. Source: Extracted from Figure 3 of Walker et al.[40]

particles detached at the top of an eroding area arrive at the downstream boundary well before slowmoving particles detached at the same place at the same time. Consequently, once sediment transport in rain-impacted flow starts, the fine material dominates the early discharge of sediment, but the composition of the sediment becomes coarser with time as more of the slower-moving particles reach the downstream boundary. If the particles are stable, then at the steady state, the composition of the sediment discharge must be the same both physically and chemically as the original soil. This is demonstrated by the results of laboratory experiments on erosion of sand by rain-impacted flows undertaken by Walker et al.[39] In these experiments, 3-m-long sloping beds of sand were eroded for 1 hour by rain-impacted flows generated by artificial rain at three rainfall intensities (45, 100, and 150 mm/h). The data shown in Figure 5 were generated by rain made up of a single drop size (2.7 mm) falling on rain generated flows over beds of sand with two different slope gradients (5% and 0.5%). The least erosive situation is the case where 45 mm/h rain fell when the slope gradient was 0.5%. The most erosive situation is the case where 150 mm/h rain fell when the gradient was 5%. The enrichment factor is the ratio of the proportion of the material in the sediment discharge to the proportion in the original material. In all cases, the discharge of sediment was dominated by fine material at 2 min and, in some cases, coarse material at 1 h (Figure 5). In the most erosive situation (150 mm/h rain falling on 5% slope), the sediment composition at 1 hour was close to the composition for the steady state, the composition that occurs when the

FIGURE 6 Schematic representation of how detachment and transport processes associated with rainfall erosion may vary spatially on planar inclined bare soil surfaces.

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enrichment factor for all particle sizes equals 1.0. Sediment composition generated by the rain-impacted flows varies with the intensity of the rain, the slope gradient, slope length, and time because particles of different size and density travel at different rates. The differential rate of transport of particles in rain-impacted flows has consequences with respect to erodibility because the actual area contributing to the soil loss does not stabilize until the slowest-moving particles detached at the farthest point from the downslope boundary at the start of the rainfall event are discharged. This fact is seldom considered when experiments are undertaken to determine erodibility. Eq. (24) is essentially targeted at situations where raindrop-induced saltation (RIS) controls sediment discharge in situations like that illustrated in Figure 6a. Particles of silt and sand may travel over the soil surface in more than one mode before being discharged. For example, as may be perceived from Figure 6b, they may leave the point of detachment traveling in splash (ST), then move further downslope by RIS, and finally by flow-driven saltation (FDS) as flow energy increases down the slope. Transitions between RIS and FDS may vary in time and space during a rainfall event and will depend on the intensity of the rain, the infiltration characteristics of the soil, the length of the slope, and the

FIGURE 7 Brink (downstream boundary) flow velocities (b) for bare soil of various lengths inclined at 9% produced by the rainfall-runoff model described by Moore and Kinnell[41] and the rainfall intensities recorded during a rainfall event at the Ginninderra Experiment Station, Canberra, Australia (a).

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623

FIGURE 8 Amounts of coal, sand, and fine particles discharged for the rainfall and runoff conditions shown in Figure 7 when a mechanistic model of erosion by rain-impacted flow was used by Kinnell.[42]

slope gradient. The effect of the transition can have an appreciable effect on both soil loss and the composition of the sediment discharged. Figure 7b shows modeled flow velocities at the brink (discharge boundary) of planar bare soil surfaces of various length on a 9% slope resulting from the rainfall event shown in Figure 7a. Figure 8 shows the loss of materials of various size and density from those bare soil surfaces when a mechanistic model of erosion by rain-impacted flow was used by Kinnell.[42] For the surfaces up to 15 m in length, particles larger than 0.1 mm in size traveled over the whole length by RIS. Under these conditions, more of the 0.11 mm sand was lost during the event than the 0.46 mm coal. On the 20-m-long area, the 0.46 mm coal, which had been traveling slower than the 0.11 sand on shorter areas, traveled for a short period of time by FDS during the high-intensity bursts of rain, and, as a result, more of the 0.46 mm coal was lost than the 0.11 mm sand. On areas 25 m long and more, FDS also contributed to the discharge of the 0.11 mm sand, but the amount lost was always less than the amount of the 0.46 mm coal. Little regard is given to the effect of temporal and spatial changes in transport mechanism when experiments are undertaken to determine values for soil erodibility factors in sheet and interrill erosion areas even though these changes may have a major impact on the amount of soil lost from an area.

conclusion Although conceptually, rainfall erosivity is the capacity of rain to produce erosion, whereas soil erodibility is the susceptibility of the soil to be eroded, the factors controlling the erosive stress applied to the soil surface and the factors influencing the resistance of the soil to them vary in time and space in complex ways. In all existing predictive models of rainfall erosion, numerous simplifications and assumptions have to be made for practical reasons, and, as a result, these models do not have the capacity to deal with such complexity. Consequently, erosivity and erodibility values are specific to the model in which they are parameterized and to the scale that the model operates within.

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3. El-Swaify, S.A.; Dangler, E.W. Erodibilities of selected tropical soils. In Soil Erosion: Prediction and Control; Foster, G.R., Ed.; Soil Conservation Society of America: Ankeny, IA, 1976; 105–114. 4. Young, R.A.; Mutchler, C.K. Erodibility of some Minnesota soils. J. Soil Water Conserv. 1997, 32, 180–182. 5. Loch, R.J.; Slater, B.K.; Devoil, C. Soil erodibility (km) values for some Australian soils. Aust. J. Soil Res. 1998, 36, 1045–1056. 6. Zhang, K.L.; Shu, A.P.; Xu, X.L.; Yang, Q.K.; Yu, B. Soil erodibility and its estimation for agricultural soils in China. J. Arid Environ. 2008, 72, 1002–1011. 7. Renard, K.G.; Foster, G.R.; Weesies, G.A.; McCool, D.K.; Yoder, D.C. Predicting soil erosion by water: A guide to conservation planning with the Revised Universal Soil Loss Equation (RUSLE); U.S. Department of Agriculture Agricultural Handbook. No. 703; US Department of Agriculture: Washington, DC, 1997. 8. Yu, B. A comparison of the R-factor in the Universal Soil Loss Equation and the revised Universal Soil Loss Equation. Trans. ASAE 1999, 42, 1615–1620. 9. Nash, J.E.; Sutcliffe, J.E. River flow forecasting through conceptual models. Part 1—A discussion of principles. J. Hydrol. 1970, 10, 282–290. 10. Kinnell, P.I.A. Event soil loss, runoff and the Universal Soil Loss family of models: A review. J. Hydrol. 2010, 385, 384–397. 11. Williams, J.R. Sediment-yield prediction with universal equation using runoff energy factor. In Present and Prospective Technology for Predicting Sediment Yield and Sources, Publ. ARS-S-40. US Dept. Agric.: Washington, DC, 1975; 244–252. 12. Williams, J.W.; Izaurralde, R.C.; Steglich, E.M. Agricultural Policy/Environmental Extender Model Theoretical Documentation, BRC Report # 2008–17; Blackland Research and Extension Center: Temple, TX, 2008. 13. Simanton, J.R.; Rawizt, E.; Shirley, E.D. Effects of rock fragments on erosion of semi-arid rangeland soils. In Erosion and Productivity of Soils Containing Rock Fragments; Krai, D.M., Hawkins, S.L., Eds.; Soil Science Society of America: Madison, WI, 1984; 65–67. 14. Onstad, C.A.; Foster, G.R. Erosion modelling on a watershed. Trans. ASAE 1975, 18, 288–92. 15. Williams, J.R.; Jones, C.A.; Dyke, P.T. A modelling approach to determining the relationship between erosion and productivity. Transactions of the American Society of Agricultural Engineers 1984, 27, 129–144. 16. Williams, J.R.; Jones, C.A.; Dyke, P.T. The EPIC model and its application. In Proceedings of International Symposium on Minimum Data Sets for Agrotechnology Transfer, March 21–26, 1983. ICRISAT Center: India, 1984; 111–121. 17. Bagarello, V.; Ferro, V.; Giordano, G. Testing alternative erosivity indices to predict event soil loss from bare plots in Southern Italy. Hydrol. Process. 2010, 24, 789–797. doi:10.1002/hyp.7538. 18. Bargarello, V.; Di Stefano, C.; Ferro, V.; Kinnell, P.I.A.; Pampalone, V.; Porto, P.; Todisco, F. Predicting soil loss on moderate slopes using an empirical model for sediment concentration. J. Hydrol. 2011, 400, 267–273. 19. Law, J.O.; Parsons, D.A. The relation of raindrop size to intensity. Trans., Am. Geophys. Union 1943, 24, 452–460. 20. Kinnell, P.I.A. Rainfall intensity-kinetic energy relationships for soil loss prediction. Soil Sci. Soc. Am. J. 1981, 45, 153–155. 21. Kinnell, P.I.A. Rainfall energy in Eastern Australia: Intensity-kinetic energy relationships for Canberra. A.C.T. Aust. J. Soil Res. 1987, 25, 547–553. 22. van Dijk, A.I.J.M.; Bruijnzeel, L.A.; Rosewell, C.J. Rainfall intensity-kinetic energy relationships: A critical literature appraisal. J. Hydrol. 2002, 261, 1–23. 23. Bullock, P.R.; de Jong, E.; Kiss, J.J. An assessment of rainfall erosion potential in southern Saskatchewan, from daily rainfall records. Can. Agric. Eng. 1987, 29, 109–115.

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24. Rekolainen, S.; Posch, M. Adapting the CREAMS model for Finnish conditions. Nord. Hydrol. 1993, 24, 309–322. 25. Posch, M.; Rekolainen, S. Erosivity factor in the Universal Soil Loss Equation estimated from Finnish rainfall data. Agric. Sci. Finl. 1993, 2, 271–279. 26. Bagarello, V.; D’Asaro, F.V. Estimating single storm erosion index. Trans. ASAE 1994, 37, 785–791. 27. Yu, B.; Rosewell, C.J. An assessment of a daily rainfall. Erosivity model for New South Wales. Aust. J. Soil Res. 1996, 34, 139–152. 28. Richardson, C.W.; Foster, G.R.; Wright, D.A. Estimation of erosion index from daily rainfall amount. Trans. ASAE 1983, 26, 153–160. 29. Hoyes, N.; Waylen, P.R.; Jaramillo, A. Seasonal and spatial patterns of erosivity in a tropical watershed of the Colombian Andes. J. Hydrol. 2005, 314, 177–191. 30. Flanagan, D.C.; Nearing, M.A., Eds. USDA-Water Erosion Prediction Project: Technical documentation, NSERL Rep. No.10; National Soil Erosion Research Laboratory: West Lafayette, IN, 1995. 31. Elliot, W.J.; Liebenow, A.M.; Laflen, J.M.; Kohl, K.D. A compendium of soil erodibility experiments. Publication No. 3. USDA-ARS National Soil Erosion Research Laboratory: West Lafayette, IN, 1989. 32. Kinnell, P.I.A. Interrill erodibilities based on the rainfall intensity-flow discharge erosivity factor. Aust. J. Soil Res. 1993, 31, 319–332. 33. Wei, H.; Nearing, M.A.; Stone, J.J.; Guertin, D.P.; Spaeth, K.E.; Pierson, F.B.; Nichols, M.H.; Moffett, C.A. A new splash and sheet erosion equation for rangelands. Soil Sci. Soc. Am. J. 2009, 73, 1386–1392. 34. Torri, D.; Sfalanga, M.; Del Sette, F. Splash detachment: Runoff depth and soil cohesion. Catena 1987, 14, 149–155. 35. Kinnell, P.I.A. Raindrop impact induced erosion processes and prediction: A review. Hydrol. Process. 2005, 19, 2815–2844. 36. Meyer, L.D.; Harmon, W.C.; McDowell, L.L. Sediment sizes eroded from crop row sideslopes. Trans. ASAE 1980, 23, 891–898. 37. Miller, W.P.; Baharuddin, M.K. Particle size of interrilleroded sediments from highly weathered soils. Soil Sci. Soc. Am. J. 1987, 51, 1610–1615. 38. Palis, R.G.; Okwach, G.; Rose, C.W.; Saffigna, P.G. Soil erosion processes and nutrient loss. I. The interpretation of enrichment ratio and nitrogen loss in runoff sediment. Aust. J. Soil Res. 1990, 28, 623–639. 39. Parsons, A.J.; Abrahams, A.D.; Luk, S.-H. Size characteristics of sediment in interrill overland flow on a semiarid hill-slope, Southern Arizona. Earth Surf. Process. Landforms 2006, 16, 143–152. 40. Walker, P.H.; Kinnell, P.I.A.; Green, P. Transport of a noncohesive sand mixture in rainfall and runoff experiments. Soil Sci. Soc. Am. J. 1978, 42, 793–801. 41. Moore, I.D.; Kinnell, P.I.A. Kinematic overland flow—Generalization of Rose’s approximate solution, Part II. Hydrology 1987, 82, 351–362. 42. Kinnell, P.I.A. The influence of raindrop induced saltation on particle size distributions in sediment discharged by rain-impacted flow on planar surfaces. Catena 2009, 78, 2–11.

Bibliography Kinnell, P.I.A. Determining soil erodibility for the USLE-MM erosion model. Catena 2018, 163, 424–426. Nearing, M.A.; Yin, S.; Borelli, P.; Polyakov, V.O. Rainfall erosivity: An historical review. Catena 2017, 157, 357–362. USDA. Draft Science Documentation: Revised Universal Soil Loss Equation Version 2 (RUSLE2). 2013. https://www.ars.usda.gov/ARSUserFiles/60600505/RUSLE/RUSLE2_Science_Doc.pdf

Index Page numbers followed by f and t indicate figures and tables, respectively

A ABAG (German USLE), 311 Abiotic factors, 352 Accelerated erosion, 30, 133–134 Accelerated soil erosion, 134 Accelerated solvent extraction (ASE), 70 Acid mine drainage (AMD), 270, 363, 365 Acid rock drainage (ARD), 363, 365, 368 Acid sulfate soils (ASSs) assessment, 367 avoidance strategy, 367 biochemical processes, 363 distribution, knowledge of, 366–367 global issues, 364–365 management approaches, 365–366 neutralization of acidity, 368 occurrence, 363 oxidation prevention, 367 Active fractions, 106 Advanced very high-resolution radiometer (AVHRR), 219 Africa, soil degradation in, 223–224 AGNPS, see Agricultural NonPoint Source model (AGNPS) Agricultural activities with C sequestration, 169, 169f global climate change, 167–168 Agricultural energy electricity generation technologies, 14, 15, 15t energy audit, 9–10 farming systems, 10 conservation agriculture, 11–12 irrigation methods, 12–13 machinery operation, 10–11 overview of, 9 Agricultural land loss, soil erosion, 121 Agricultural NonPoint Source model (AGNPS), 294 Agricultural on-site impacts, 290 Agricultural production in Central Andes, 259–260, 260f

salt-affected soils, 90 accumulated salts removal, 90 saline environment adaptation, 90–91 salt accumulation prevention, 90 Agricultural runoff category of pollutants, 581 controlling pollution, 584 definition, 581 dissolved pollutants, 583–584 quantity, 582 soil erosion and eroded sediments, 582–583 Agricultural soils carbon and nitrogen cycles, 100–101, 101f nitrogen losses from, 106–108 nitrous oxide emission from agriculture, 3–4 biomass burning, 5 calculation, 6 denitrification, 4 fertilizer consumption, 6 flooded soils, 5 nitrification, 4 overview of, 3 practices to decrease, 7 organic substrate quality, 105–107, 107f soil organic matter, 106 overview of, 99 phosphorus, 111, 116 chemical speciation of, 111–112 mobility of, 112 spatial speciation of, 112–115, 114f, 115f soil microbial activity, 101–102, 102f organic substrates distribution, 103–106 soil temperature, 102 soil texture, 103 soil water content, 102, 103 Agricultural water management fuzzy mathematical programming, 375–376 general model, 378–380 IDFCCP, 385–389, 387t

627

628 Agricultural water management (cont.) interval linear programming, 376–377 IPWM, 380–382, 382t ISCCP, 383–385, 383t, 386t, 391t optimization approaches, 374 stochastic mathematical programming, 375 Agriculture sustainability, 251 agroecosystem biodiversity, 252–253 air and atmosphere, 252 soil, 251–252 quality assessment, 253–254 strategies for, 254 water, 252 Agro-ecosystem analysis, 349 Agroecosystem biodiversity, 252–253 Agronomic force, integrated farming systems, 172 Agronomic rate, 452 Air, sustainable agriculture, 252 Allelopathic crops, 21 Alluvial stratigraphy, 268 Almond orchards, navel orangeworm management, 483–484 Ambient air modern pesticides in extract cleanup, 66 extracting analytes from sorbents and filters, 65–66 sampling techniques, 64–65 AMD, see Acid Mine Drainage (AMD) Amelioration, 150 Ammonia (NH3), 246–247 Ammonium (NH4+), 246, 247 Amyelois transitella, 483 Anagrus epos, 499 Ancient lynchets, 544 Andisols, 564 Animal pests anthropogenic measures, 417 phenomenon of biotic (natural) resistance, 416–417 strategy to control, 415–416 Anionic PAM, 285 ANSWERS, see Areal Nonpoint Source Watershed Response Simulation (ANSWERS) Antagonistic microorganisms, 179 Antimicrobial chemicals, 179 APEX model, 614–615 Aphis fabae, 178 Applied population ecology, 349, 350 Aquatic weed management, 486–487 Areal Nonpoint Source Watershed Response Simulation (ANSWERS), 294 Aridity index map, 219 ASE, see Accelerated solvent extraction (ASE) Asia, soil degradation, 221–222 ASW, see Available soil water (ASW) Atmosphere, sustainable agriculture, 252 Atmospheric CO2, 166–168, 166t, 235

Index Australia, soil degradation, 228–230 Australian bush fly management, 484–485 Australian Pesticide Act (1999), 50 Automatic sampling equipment, 293 Available soil water (ASW), 206 AVHRR, see Advanced very high-resolution radiometer (AVHRR)

B Bacillus thuringiensis (Bt), 21, 161, 353, 491 Beauvaria bassiana, 492 Bench terraces, 297 formation, 543 Biocontrol agents, 160, 161 Biodiversity, 185 Bioenergy crops, 275 benefits of, 276, 277 C sequestration, 275–278 overview of, 275 at temporal scale, 276 Biological effects, soil erosion, 155–156 Biological nitrogen fixation (BNF), 425 Biomass burning, 5 Biomass fuels, 275 Biosolids, 451–452 Biotechnology to food crops, 14 impacts on pest management, 353 Biotic factors, 352 “Black Blizzards,” 226 The Brazilian Agency for Water Resources, 302 Buffer strips, 527 Bulking agents, 451 Bush fly Australian bush fly management, 484–485 Kwajalein Atoll, 486 origins and habitat, 485 “Business as usual” concept, 228

c Canadian Organic Products Regulation, 158 Capillary GC, 70 Carbon cycles, 243 in agricultural soils, 100–101 Carbon dioxide (CO2), 244 agricultural sources of, 167 atmospheric, 166–168 soil organic carbon emission estimation, 238, 239 mass balance approach, 236–239 Carbon gases soil quality, 243–244 carbon dioxide (CO2), 244 methane (CH4), 244 Carbon monoxide (CO), 244

629

Index Carbon pools, 165, 165t, 166, 166f Caribbean Environment Program, 228 Caribbean islands, soil degradation, 227, 228 Carpophilus pilosellus, 486 Cascade impactor, 65 Cation exchange capacity (CEC), 155, 199 Cation ratio of soil structural stability (CROSS), 86 Capacity factor, 113 CEC, see Cation exchange capacity (CEC) Central Andes agricultural production in, 259–260 droughts in El Niño Southern Oscillation, 259–263 (see also El Niño Southern Oscillation (ENSO)) precipitation variability, 260–261 Certified organic crops, 19 ceteris paribus, 125 CFC, see Critical flocculation concentration (CFC) Chemical deterioration map, 219 Chemical effects, soil erosion, 155 Chemical pesticides, 55, 55t, 56t Chemicals, Runoff, and Erosion from Agricultural Systems (CREAMS), 318 Chemical speciation, of phosphorus, 111–112 China, soil erosion in, 119–120, 120f, 120t Circulation of pesticides in air, 57–58, 57f in aquatic environment, 56, 57 in crops, 58 in soil, 58 Clay–cation interaction, 202–204 Clay flocculation, 283, 284 Clay–sodium bonding, 85 Clay-to-clay bonding, 85 Clean Water Act (CWA), 504–505 Climate change, 188 direct and indirect effects of, 31 Climatic erosivity, 122 Climatic (C) factor, wind erosion, 139 Closed ecosystem, 499 Coleomegilla maculata, 500 Community supported agriculture (CSA), 21 Compaction, 189–190 Competitive crops, 160 Compost, 285, 451 Conservation agricultural systems, 104, 104f, 108 Conservation agriculture, 11–12 Conservation practices, 276 Conservation prioritization, 303 Conservation Reserve Program (CRP), 277 Conservation structures, 300 Conservation tillage, 343 Contamination, of plant foods, 58 Contemporary lynchets, 544–545 Continuous osmotic phase, 84 Continuous simulation models, 318–319 Contouring, 296, 297

Contour ridges, 299 Controlled traffic farming, 12, 14 Conventional approach, integrated farming systems, 172, 173 Conventional farming vs. organic agriculture, 22 Conversion to sprinklers, 36 Copidosomopsis plethorica, 483, 484 Cover crops, 158, 159 Coversoil resources, 268–269, 268t thickness requirements, 269 Coversoiled acid spoils, 269 CREAMS, see Chemicals, Runoff, and Erosion from Agricultural Systems (CREAMS) Critical flocculation concentration (CFC), 283 Crop-livestock systems, 187 Cropping sector, energy in, 10, 13, 15 Crop rotation, 159, 161–162 plan, 21 Crop systems, simulation models of, 351–352 CROSS, see Cation ratio of soil structural stability (CROSS) CRP, see Conservation Reserve Program (CRP) Crust development, 269 CSA, see Community supported agriculture (CSA) C sequestration, 168–169, 224, 236 bioenergy crops, 275–278 Ctenopharyngodon idella, 487 Cultural services, grazing systems, 186 Curve number method, 46

D Dairy wastewater rainwater, 455 recycling through irrigation, 455 water budget development, 456–457 DDL, see Diffuse double layer (DDL) Decomposition, organic matter climate, 561–562 C : N ratio, 562–563 Decomposition processes, 244 Deep ripping, 94 Deforestation, 131–132, 132f, 190 Denitrification, 4 Depth-and-width integrated sampling, 293 Depth stratification, 104 Derivatization, of pesticides, 70 Desertification, 188 biophysical parameters, 588 definition, 587, 593–594 degree of desertification, assessment, 589 evaluation frameworks, 597–598 human–environmental frameworks, 599 landscape functioning, 587–588 lessons learned, 595–597 participatory approach, 598–599

630 Desertification (cont.) prevention and mitigation actions, 594–595 rehabilitation monitoring, 590, 595 reversing, 589–590 Desulfobacter, 516 Desulfotomaculum, 516 Desulfovibrio, 516 Detachment, 618 Detention storage, 27 Diffuse double layer (DDL), 202 Dikrella cruentata, 499 Dispersion process, 205, 205f, 207–208, 208t Dispersion ratio of clay (DRC), 200 Dispersive liquid-liquid microextraction (DLLME), 63 Diversion/graded terraces, 297, 298 DLLME, see Dispersive liquid-liquid microextraction (DLLME) DPSIR, see Driving force–pressure–state–impact– response (DPSIR) Drainage, of sodic soils, 95–96 DRC, see Dispersion ratio of clay (DRC) Driving force–pressure–state–impact–response (DPSIR), 230, 231, 231f Droughts in Central Andes, El Niño Southern Oscillation, 259–263 Dryland ecosystems, 188 Dryland/seepage salinity, 213 Dryland sodic soils, 214, 215 Dust storms, 138 Dynamic sampling, 64 Dysfunctional/desertified landscapes, 588

e Earth Science Data and Information System (ESDIS), 219 Earthworm, 344 ECD, see Electron capture detector (ECD) Ecological agriculture, 491 Ecological infrastructure management, 174 Econometric methods, 353 Economic pressure, 172 Ecosystem services by grazing systems, 184 cultural services, 186 provisioning services, 185 regulating services, 185 supporting services, 185–186 Electricity generation technologies, 14, 15, 15t Electron capture detector (ECD), 71 El Niño Southern Oscillation (ENSO), 259–261 classification of, 262t impact on crop yield, 261–263 Embodied energy, 13, 14 Empirical erosion model soil water erosion ABAG, 311

Index development of, 308–309 Modified Universal Soil Loss Equation, 310 overview of, 307–308 predictions, accuracy of, 312–314 Revised Universal Soil Loss Equation, 312 Soil Loss Estimation Model for Southern Africa, 310–311 Soiloss, 312 Universal Soil Loss Equation, 309–310 Empirically based model, 317 Encroachment, 188–189 Endangered Species Act (ESA), 504–505 Energy, in agriculture cropping sector, 10, 13, 15 electricity generation technologies, 14, 15, 15t energy audit, 9–10 farming systems, 10 conservation agriculture, 11–12 irrigation methods, 12–13 machinery operation, 10–11 overview of, 9 Energy audit, 9–10 Engineered terraces bench terraces, 545 gradient terraces, 546–547 ENSO, see El Niño Southern Oscillation (ENSO) Environmental force, integrated farming systems, 172 Environmental legislation, 301, 302 Environmental preservation, 254 Environmental Protection Agency (EPA), 50, 501, 502 Environmental Quality Indicators Program (EQIP), 22 Environmental sustainability, 251–253 Enzymes, 100 EPA, see Environmental Protection Agency (EPA) Ephemeral gullies, 281 EPIC, see Erosion Productivity Impact Calculator (EPIC) EQIP, see Environmental Quality Indicators Program (EQIP) Eroding clods, 295 Erosion, 125–128, 189 assessment, 608–609 control, 609 deforestation, 131–132, 132f effects on soil quality biological effects, 155–156 chemical effects, 155 overview of, 153 physical effects, 155 gravity-induced erosion, 607–608 impacts, 609 induced carbon dioxide, SOC emission estimation, 238, 239 mass balance approach, 236–239 irrigation-induced, 35 control methods, 36–37 unique characteristics, 35–36

631

Index overview of, 131 sedimentation rates, 132–134 snowmelt amelioration, 150 modeling, 150 overview of, 147–148, 147f precipitation, 148 runoff events, 150 snow and freezing conditions, 149–150 soil, 148–149 soil, 605 conservation methods, 609–610 water, 606 wind, 606 dynamics, 142, 142f global hot spots, 137–139, 137f particle entrainment, 142, 143, 143f processes of, 141 self-balancing concept, 144, 144f Erosion by water empirical erosion model development of, 308–309 Modified Universal Soil Loss Equation, 310 overview of, 307–308 predictions, accuracy of, 312–314 Revised Universal Soil Loss Equation, 312 Soil Loss Estimation Model for Southern Africa, 310–311 Soiloss, 312 Universal Soil Loss Equation, 309–310 future climate change, 30, 31 impacts of, 30 overview of, 25 problem of accelerated, 30 processes of, 25–28 spatial and temporal scale, 28–29, 28f, 29f Erosion control, 285 mulch tillage, 346–347 no-till, 344–345 ridge tillage, 345–346 soil erosion, 343 strip tillage, 347 tillage, 343 effect on soil properties, 343–344 Erosion Productivity Impact Calculator (EPIC), 310 Erosion sediment, 41 ESDIS, see Earth Science Data and Information System (ESDIS) ESP, see Exchangeable sodium percentage (ESP) ESR, see Exchangeable sodium ratio (ESR) EU-Com: “Strategy for soil protection 1995–2005,” 50 Europe, soil degradation in, 219–221 European soil erosion model (EUROSEM), 294, 319 European tillage-based systems, 10, 11 European Water Framework Directive, 49 EUROSEM, see European soil erosion model (EUROSEM)

EUROSEM erosion model, 618–619 EU standard, EEC No. 2092/91, 158 Eutrophication, 449–451 Event erosivity index (Re), 612–613 Event sediment yield (SYe), 613–615 Exchangeable sodium percentage (ESP), 85, 86, 199–201, 200f, 207, 211, 573 Exchangeable sodium ratio (ESR), 199 Extract cleanup modern pesticides, determination of in ambient air, 66 in fruit and vegetables, 67–68

F Farm energy calculators, 10 Farmer field schools (FFS), 492 Farming systems on energy, 10 conservation agriculture, 11–12 irrigation methods, 12–13 machinery operation, 10–11 Farm irrigation methods, 12–13 Federal Environmental Pesticide Control Act (FEPCA), 502 Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), 502 Fertilizer consumption, 6 Field border (FB), 400 Field erosion plots, 291 Filter strip (FS), 400 terraces, 526–527 Filth fly management, dairies and poultry, 487–488 Fingerprinting techniques, 294 Flame photometric detector (FPD), 71 Flood disasters, 121 Flooded soils, 5 Floodplain aggradation, 133 Flow-driven saltation (FDS), 622, 623 Flow surfaces, 295 Food insecurity, 189 Food Quality Protection Act (FQPA), 502 Forage crops, irrigation, 455–456 Forest soils, 131–132 FPD, see Flame photometric detector (FPD) Freshwater shortage, 122 Frost heave, 570 Fruit and vegetables modern pesticides in extract cleanup, 67–68 isolating techniques, 67 sample collection and preparation, 66–67 Functional-group diversity, 185 Functional/non-desertified landscapes, 587–588 Functional pools, 112 Functional-response diversity, 185 Fuzzy approach, 219

632

G Gas chromatography (GC), 70–72 GCMs, see General circulation models (GCMs) Gel chromatography (GPC), 68 General circulation models (GCMs), 122, 570 Genetically engineered organisms, 159 Genetically modified organisms (GMOs), 20 Genetic modifications, 353 Geographic information system (GIS), 353–354, 532 Geotextiles, 301 German Federal Immission Control Act (2002), 50 German Soil Protection Act (1999), 50 GHGs, see Greenhouse gas emissions (GHGs) GIS, see Geographic information system (GIS) The global assessment of soil degradation (GLASOD), 139 Global change, 119 soil erosion and, 119, 121, 122 in China and in the World, 119–120, 120t deteriorates global environment, 121–122 threatens to global food security, 121 Global climate change agricultural activities, 167–168, 167f greenhouse gas emissions, 167–168, 167f overview of, 165–167 world soils C sequestration, 168 for greenhouse effect mitigation, 168–169 Global environment, soil erosion deteriorates, 121–122 Global food security, soil erosion threat, 121 Global positioning systems (GPS), 529–530 Global warming, 252 GMOs, see Genetically modified organisms (GMOs) Goniozus emigratus, 483 Goniozus legneri, 483, 484 GPC, see Gel chromatography (GPC) Grasslands, 183, 185 Grass strips, 525 Gravity-induced erosion, 607–608 Grazing lands, 183, 186 Grazing systems challenges in, 187 climate change, 188 deforestation, 190 desertification, 188 greenhouse gas emissions, 190 nutrient loss, 189 poverty and food insecurity, 189 soil compaction, 189–190 woody encroachment, 188–189 ecosystem services by, 184 cultural services, 186 provisioning services, 185 regulating services, 185 supporting services, 185–186 knowledge gaps in, 190

Index management of, 186–187 overview of, 183–184 Greenhouse effect mitigation, 168–169, 169t Greenhouse gas emissions (GHGs), 12–15, 167–168, 190 Green marketing incentive force, integrated farming systems, 172 Green revolution technologies, 20, 301 Griffith University Erosion System Template (GUEST), 294, 319 Ground residue cover, 318 GUEST, see Griffith University Erosion System Template (GUEST) Gullies, 281 Gully erosion, 148, 148f, 290 Gully stabilization structures, 301, 302 Gypsum, 214, 283–284

H Hand tillage, 545 HC, see Hydraulic conductivity (HC) Heisenberg’s theory, 111 Herbicide-resistant crop, 353 Heterogeneous equilibrium, phosphorus, 116 Heterotrophs, 4 High-performance liquid chromatography (HPLC), 70 Histosols, 564 HMF samplers, 329–332 Holdridge diagram, 560 HPLC, see High-performance liquid chromatography (HPLC) Hydraulic conductivity (HC), 93, 94 Hydraulic properties, of sodic soils, 93–94 Hydrological models, water erosion, 46–47 Hydrologic properties, sodicity, 204–206 Hypoxia, 90

i ICM, see Integrated crop management (ICM) IFOAM, see International Federation of Organic Agriculture Movements (IFOAM) Impact force, 143 Impact threshold, 143 Indigenous technical knowledge (ITC), 231 Induced wind erosion, 138–139 Industrial desertification, 220 Inexact double-sided fuzzy chance-constrained programming (IDFCCP), 385–389 Infiltration excess overland flow, 27 Infiltration rate (IR), 93, 95 Inorganic fertilizer, 173 Inorganic mulches, 282, 285–286 Insecticide Act (1910), 502 Insects, compounds toxic release, 178–179 In situ acid minesoil remediation, 269–270

633

Index In situ sodic minesoil remediation, 269 In situ soil reclamation, 269 acid minesoil remediation, 269–270 sodic minesoil remediation, 269 Integrated crop management (ICM), 171, 174 Integrated farming systems conventional approach, 172, 173 development of, 172–173 driving forces of, 172 implementation of, 175 overview of, 171 principles of ecological infrastructure management, 174 integrated crop management, 174 integrated nutrient management, 173–174 minimum soil cultivation, 174 multifunctional crop rotation, 173 Integrated fly management, 486, 492, 493 Integrated nutrient management (INM), 173–174 biofertilizers, 425 components, 420–421 micronutrients, 425 mineral and synthetic fertilizers, 422–423 nitrogen, 423–424 nitrogen-fixing biofertilizers, 427, 427t nutrient losses, 429 phosphorus, 424 potassium, 424–425 soil fertility improvement, 430–431 toxic accumulation, 430 Integrated pest management (IPM), 171, 174, 177 need, 436–437 packages, 437 Integrated weed management biologically controlling, 440–443, 441t–442t other integrated approaches, 443–444 Intensity factor, 112 Intensive rotational grazing systems, 159 Intergovernmental Panel on Climate Change (IPCC), 3, 7, 236 Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES), 217 International Federation of Organic Agriculture Movements (IFOAM), 19, 157 Interval parameter water quality management model (IPWM), 380–382, 382t Interval-stochastic chance-constrained programming (ISCCP), 383–385, 383t, 391t IPBES, see Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES) IPCC, see Intergovernmental Panel on Climate Change (IPCC) IPM, see Integrated pest management (IPM) IR, see Infiltration rate (IR) Iron-oxidizing bacteria, 270

Irrigation farming on sodic soils drainage, 95–96 hydraulic properties, 93–94 irrigation practices, 95 nutrient status, 94 soil management, 94 water quality, 94–95 Irrigation-induced erosion, 35 control methods conversion to sprinklers, 36 site modification, 37 soil protection and tillage, 36 water properties, 37 unique characteristics, 35–36 Irrigation methods, in agriculture, 12–13 Irrigation technologies, 226 Isokinetic sampling, 293 ITC, see Indigenous technical knowledge (ITC)

K Knowledge gaps in grazing systems, 190

L Labidura riparia, 486 LADA, see Land Degradation Assessment in Drylands (LADA) Land cover map, 218, 219 Land degradation, 190, 218, 218t Land Degradation Assessment in Drylands (LADA), 219 Land degradation neutrality (LDN), 231, 594 Land intensification, 223 Landscape regrading, 267–268 LDN, see Land degradation neutrality (LDN) Leaching, 189 Legislation force, integrated farming systems, 172 Light detection and ranging (LIDAR), 531–532 Limburg Soil Erosion Model (LISEM), 294 Lime amendment, 282–283 Limit of detection (LOD), 67 Limonoid azadirachtin, 178 Liquid-liquid extraction (LLE), 63 Liquid manure, 463–464 Liquid manure injection system, 470 LISEM, see Limburg Soil Erosion Model (LISEM) Livestock production systems, 183, 185, 187, 189, 191 Living mulches, 285 Lixophaga diatraege, 492 LLE, see Liquid-liquid extraction (LLE) LOD, see Limit of detection (LOD) Lynchets ancient, 544 contemporary, 544–545

634

M Machinery operation, agricultural energy, 10–11 Manure, 449–451 flushing manure, 454 Manure management biosolids, 451–452 compost, 451 dairy drinking water requirements, 454 estimation, water use, 453 flushing manure, 454 washing cow and equipment, 454–455 waste water management, 455–456 water budget development, 456–457 phosphorus concentration diet modification impacts, 460–461 practices for environmental protection, 459 reducing manure P concentration, 459–460 poultry manure collection and handling, 463–464 earthen anaerobic lagoon/outdoor storage, 464, 467 high-rise layer facilities/indoor storage, 465, 466 storage of solid and liquid manure, 465 uses, 471 utilization, 467–471 MAR, see Mean annual rainfall (MAR) Mass balance approach, SOC, 236–239, 237f Mass balance differential equation, 317 Mass flux profiles, 144, 144f Mean annual rainfall (MAR), 126 Mean organic carbon content, 563, 564 Mechanical properties, sodicity, 206–207 Membrane extraction, 63, 64 Methane (CH4), 244 Micronesia, Australian bush fly management, 484–485 Millennium Ecosystem Assessment (MA), 593–594 Mineralization, 101, 106 Minimum soil cultivation, 174 Mobility, of phosphorus, 112 Modern pesticides, determination of, 58–60 in ambient air extract cleanup, 66 extracting analytes from sorbents and filters, 65–66 sampling techniques, 64–65 in fruit and vegetables extract cleanup, 67–68 isolating techniques, 67 sample collection and preparation, 66–67 in soil samples isolation and enrichment techniques, 69–70 purification of extract, 70 sample collection and preparation, 68, 69 in water samples collection and preparation, 62

Index isolation and enrichment techniques, 62–64, 63t, 64t Modified Fournier Index, 126 Modified Universal Soil Loss Equation (MUSLE), 310, 613–615 Modulus of rupture (MOR), 207 Monitoring process soil water erosion, 291 field plots, 291 mathematical models, 294 quantitative indicators, 295 sediment transport in watersheds, 292–294 visual indicators, 295 Monoculture, 498 MOR, see Modulus of rupture (MOR) Mortality composting, 451 M. sorbens, 485, 486 Mulches, 94, 158, 159 Mulch tillage, 346–347 Multifunctional crop rotation, 173 Musca domestica L., 484 MUSLE, see Modified Universal Soil Loss Equation (MUSLE)

n NASA-EOSDIS, see National Aeronautics and Space Administration-Earth Observing System Data and Information System (NASA-EOSDIS) National Aeronautics and Space Administration-Earth Observing System Data and Information System (NASA-EOSDIS), 219 National Landcare Organisation, 230 National Organic Standards Board (NOSB), 19 Natural parasiticide formulations, 22 Natural pesticides, 417 Natural wind erosion, 137–138 Navel orangeworm management, 483–484 Negligence, 507 Neonicotinoids, 59 Net photosynthetic activity, 102 Net primary productivity (NPP), 218, 219 New Zealand Landcare Trust, 230 Nino3.4, 260, 261 Nitric oxide (NO), 245 Nitrification, 4 Nitrogen gases cycling, 100–101, 101f losses from agricultural soils, 106–108 soil quality, 244–245 ammonia (NH3), 246–247 loop-within-a-loop, 244, 245, 245f nitric and nitrous oxides, 245–246 nitric oxide (NO), 245 Nitrogenous fertilizers, 168 Nitrogen phosphorus detector (NPD), 71, 71t

635

Index Nitrous oxide emission from agriculture, 3–4 biomass burning, 5 calculation, 6, 6t denitrification, 4 flooded soils, 5 nitrification, 4 fertilizer consumption, 6 overview of, 3 practices to decrease, 7 Nitrous oxides, 245–246 production, 6 N leaching, 108 Nonpoint source pollution, 581 North and Central America, soil degradation in, 226–228 NOSB, see National Organic Standards Board (NOSB) No-till, erosion control, 344–345 NPD, see Nitrogen phosphorus detector (NPD) NPP map, 219 NRCS National Handbook of Conservation Practices, 397–399 Nuisance, 508 Nutrient cycling, 185–186 Nutrient loss, 189 Nutrient management, 252 diet modification impacts, 460–461 practices for environmental protection, 459 reducing manure P concentration, 459–460 Nutrient status, of sodic soils, 94

o Oak Ridge National Laboratory Distributed Active Archive Center (ORNL DAAC), 219 Off-grid electricity, 14 OFPA, see Organic Foods Production Act (OFPA) On-farm energy, 11–13 Open ecosystem, 499 Open pit mining, 267 acid mine drainage, 270 alluvial stratigraphy, 268 coversoil resources, 268–269 thickness requirements, 269 landscape regrading, 267–268 in situ soil reclamation, 269 acid minesoil remediation, 269–270 sodic minesoil remediation, 269 steep slope reclamation, 270 Oreochromis hornorum, 486 Oreochromis mossambica, 486 Organic agriculture, 157 vs. conventional farming, 22 crop rotation, 21–22 definition of, 19 history of, 19–20

pest management in, 21–22 worldwide statistics, 20–21 Organic farmers, 20, 21, 160, 162 Organic farming animal pests anthropogenic measures, 417 phenomenon of biotic (natural) resistance, 416–417 strategy to control, 415–416 definition, 415 genetically engineered organisms, 159 overview of, 157 pest management practices, 159 diseases, 161–162 insects and other invertebrates, 161 weeds, 160 soil building, 158–159 standards, 157–158 Organic fertilizers, 226 Organic Foods Production Act (OFPA), 19, 158 Organic matter andisols, 564 decomposition climate, 561–562 C : N ratio, 562–563 histosols, 564 inputs placement in soil, 561 plant species composition, 561 quantity, 559–561 soil carbon density, 560 physical and chemical influences, 563 Organic mulches, 282, 285–286 Organic producers, 162, 163 Organic production, 157–158 Organic soil amendments, 177, 285 mechanisms antagonistic microorganisms, 179 compounds toxic release, 178–179 on plant health and weeds, 177 plant resistance, 178 in pest and disease control, 180 in 21st century, 180 Organic substrates quality, 105–107 spatial distribution of, 103–106 Organochlorine pesticides, 58–60, 58f, 60t “Organonitrogen pesticides,” 59, 60f ORNL DAAC, see Oak Ridge National Laboratory Distributed Active Archive Center (ORNL DAAC) Osmotic effect, 87–89

P PAM, see Polyacrylamide (PAM) Particle entrainment, 142, 143

636 PASs, see Passive air samplers (PASs) Passive air samplers (PASs), 64 Passive dosimeters, 63 Passive fractions, 106 Passive sampling, 64 Pasturelands, 184, 187 PCA, see Principal component analysis (PCA) Percent tree cover map, 219 Periodic leaching, 95, 96 Permafrost characteristics, 567–569 climatic change, 570 effects on groundwater, 570 ice content, 570 schematic distribution of Alaska, 568 soil freezing, 570 surface energy balance, 569–570 temperature profile, with annual variation, 569 Pesticide persistence, 44 Pesticides; see also Modern pesticides, determination of characteristics of, 54–55, 55t application, 55–56 chemical class and structure, 55, 55t toxicity, 56t circulation in air, 57–58, 57f in aquatic environment, 56, 57 in crops, 58 in soil, 58 dangers result from, 60–61 derivatization of, 70 determining methodologies for, 61–62 organochlorine, 58–60, 58f overview of, 53 poisoning symptoms of, 60 quantitative determination of, 70–72 Pesticide translocation control assessment of, 44–46 erosion and runoff control, 49–50 monitoring and modeling, 46–47 soil erosion, 41, 47–49 step-by-step analysis for, 47 water erosion, 42, 43f, 46–47 wind erosion, 42–44, 43f, 44f, 47 Pest management aquatic weed management, 486–487 biotechnology impacts, 353 bush fly Australian bush fly management, 484–485 Kwajalein Atoll, 486 origins and habitat, 485 components of, 350f crop systems, simulation models of, 351–352 diverse cropping systems, 492–493 ecological practices, 492 economical benefits, 493–494

Index environmental benefits, 493 filth fly, dairies and poultry, 487–488 modeling approaches, 350–351 navel orangeworm management, 483–484 negligence, 507 nuisance, 508 in organic agriculture, 21–22 overview of, 349–350 pesticide common law actions, 506–507 enforcement power, EPA, 503–504 federal regulations, 504–505 preemption, 506 registration, 502–503 tolerance in food, 504 plant demand-side pests, 352–353 populations, 497–498 regional distribution, 353–354 regulatory trends, 509 risks involved, adopting ecological strategies, 494 statutory laws and regulations, 502 strict liability, 508 supply-side pests, 352 trespass, 507 Pest management practices organic farming, 159 diseases, 161–162 insects and other invertebrates, 161 weeds, 160 Pests definition, 415 regulation of animal pests, 416–417 Pheromone traps, 161 Phosphorus, 111, 116 chemical speciation of, 111–112 mobility of, 112 spatial speciation of, 112–115 Phosphorus concentration, manure diet modification impacts, 460–461 practices for environmental protection, 459 reducing manure P concentration, 459–460 Photosynthesis process, 101 Physical effects, soil erosion, 155 The Physics of Blown Sand and Desert Dunes, 141 Physiological models, pest management, 351 Phytosarcophaga gressitti, 484 Pitting techniques, 270 Plant-available P, 112 Plant demand-side pests, 352–353 Plant pathogens, compounds toxic release, 178–179 Plant resistance, 178 Point sources of pollution, 581 Poisoning symptoms, of pesticides, 60 Polar pesticides, 59 Polyacrylamide (PAM), 284, 285 Polyculture, 498 Polyethylene dosimeters, 63

637

Index Ponding, 27 Potassium fixation, 198 Potential acid sulfate soils (potential acid SS) sulfidization biogeochemistry, 515 definition, 518 environments, 518 oxidation of organic carbon, 516 reactive iron, 517–518 saturated soils, 516 sulfate reducing bacteria, 516–517 sulfate reduction, 515–516 sulfuricization process of, 519–520 sulfide oxidation, 519 Potential C mineralization, 103 Potential evapotranspiration (PET), 562 Poultry manure management collection and handling, 463–464 storage of solid and liquid manure, 465 earthen anaerobic lagoon/outdoor storage, 464, 467 high-rise layer facilities/indoor storage, 465, 466, 466f uses, 471 utilization, 467–471 Precipitation, 125, 127–128, 128f, 148 Precipitation variability, 260–261 Precision agriculture GIS, 532 GPS, 529–530 high-speed computer processing systems, 530–531 LIDAR, 531–532 remote sensing, 531–532 Pre-rills, 295 Principal component analysis (PCA), 260–261 Process-based models, 317–318 runoff events, 150 Provisioning services, grazing systems, 185

Q Quantitative determination, of pesticides, 70–72 Quantitative indicators, 295, 303 Quantity factor/richness factor, 112 QuEChERS (quick, easy, cheap, effective, rugged, and safe), 68, 69f Quesungual agro-forestry method, 227

R Radioisotopic techniques, 332–333 Raindrop-induced saltation (RIS), 622 Rainfall erosivity, 125–127, 611 Rainfall erosivity index, 125 Rainfall intensity, 125 Rainfall kinetic energy, 616–617

Rainsplash detachment, 554, 556 redistribution, 25–27 transport, 556 Rangelands, 183–184, 186–187 Rapid wetting, 204 Readily available pool, 112 Rear delivery manure spreaders, 469 Regional pest management, 353–354 Regulating services, grazing systems, 185 Rehabilitation; see also Open pit mining of open pit mining, 267 alluvial stratigraphy, 268 coversoil resources, 268–269 landscape regrading, 267–268 in situ soil reclamation, 269–270 Remote sensing, 531–532 Renewable energy, 14 Reptation, 143 Resource concentration hypothesis, 498 Resource conservation, 254 Resource degradation poverty traps, 189 Retention/absorption terraces, 297, 298 Reversibly available pool, 112 Revised Universal Soil Loss Equation (RUSLE), 125, 126, 310, 312 Revised wind erosion equation (RWEQ), 47, 335 R-factor, 125, 126 Rhizobia, 4, 158 Ridge tillage, 345–346 Rill erosion, 606–607 Rills, 281, 290, 295 Riparian forest buffer (RFB), 400 Riparian strips, 527, 528 Rough surface elements, 145 Runoff, 189 Runoff control, 35, 49–50 Runoff events, snowmelt erosion, 150 RUSLE, see Revised Universal Soil Loss Equation (RUSLE) RUSLE model, 616 RWEQ, see Revised wind erosion equation (RWEQ)

S Sahel, 138, 139 Saline environment adaptation, 90–91 Saline–sodic soils, 87 Salinity stress, 87–89 Salinization, 221, 226 Salt accumulation, 211, 214 prevention and removal of, 90 Salt-affected soils agricultural production in, 90 accumulated salts removal, 90 saline environment adaptation, 90–91 salt accumulation prevention, 90

638 Salt-affected soils (cont.) categories of, 87, 88 classification, 573t clay–cation interaction, 202–204 crop exchangeable sodium tolerance, 576, 576t crop tolerance, 575–576, 576t definition, 573 distribution, 574 hypoxia and salinity effects, 90 management of, 207–208 overview of, 83, 197–199 physical property and behavior, 204 hydrologic, 204–206 mechanical, 206–207 salinization, 575 salt concentration measurement, 573 sodicity, 199–202 sodic soils, 198, 201, 208 global distribution of, 199t management of, 207–208 soil structural stability in, 85–86, 91 soil water dynamics and salinity stress, 87–89 soluble salts and salinity impact, 84–85 toxic elements in, 87 Saltation, 143 Sampling techniques modern pesticides, determination of in ambient air, 64–65 in fruit and vegetables, 66–67 in soil samples, 68, 69 in water samples, 62 SAR, see Sodium adsorption ratio (SAR) Sarotherodon species, 486, 487 Saturation excess overland flow, 26, 27 SBSE, see Stir bar sorptive extraction (SBSE) Scotia segetum, 417 SDGs, see Sustainable Development Goals (SDGs) SDR, see Sediment delivery ratio (SDR) Sea level change, 123 Sea surface temperature (SST), 259 Secondary salinity, 199 Sedimentation, 290 Sedimentation rates, 132–134 Sediment concentration, 127, 150 Sediment delivery ratio (SDR), 294 Sediment flux, 292 Sediment-laden runoff, 396 Sediment yield, 294 Self-balancing concept, 144 Settling basins, 527 SIC, see Soil inorganic carbon (SIC) Simulator for Water Resources in Rural Basins (SWRRB), 318 Site modification, 37 Slaking process, 85, 198, 204, 205, 207–208, 208t

Index SLEMSA, see Soil Loss Estimation Model for Southern Africa (SLEMSA) SLM, see Sustainable land management (SLM) Slower-acting ionic phase, 84 Slow fractions, 106 Snow and freezing conditions, 149–150 Snowmelt erosion amelioration, 150 modeling, 150 overview of, 147–148 precipitation, 148 runoff events, 150 snow and freezing conditions, 149–150 soil, 148–149 SOC, see Soil organic carbon (SOC) Socioeconomic interventions, 218 Sodicity measurement, 199–202 properties and behaviors of soils, 204 hydrologic, 204–206 mechanical, 206–207 Sodic soils, 198, 201, 208, 215 dryland management, 214 global distribution of, 199t irrigation farming on drainage, 95–96 hydraulic properties, 93–94 irrigation practices, 95 nutrient status, 94 soil management, 94 water quality, 94–95 management of, 207–208 overview of, 211 physical and chemical properties, 213 root zones of salt accumulation in, 211, 214 world distribution of, 212t yield obtained in, 211–213 Sodic soils, reclamation addition of organic matter, 539 ameliorant calcareous sodic soils, 537 gypsum, 537, 538 lime, 537 effects of plant growth, 537, 539 exchangeable Na ions, displacement of, 535–536 strategies optimal supply, ameliorants, 539 rate of supply, ameliorants, 540–541 supply of water, 539–540 without ameliorants, 539 water flow improvement, 535 Sodium adsorption ratio (SAR), 85, 86f, 199, 200, 269, 573 Soil holistic management approaches, 253

Index modern pesticides in isolation and enrichment techniques, 69–70 purification of extract, 70 sample collection and preparation, 68, 69 quality assessment, 253–254 snowmelt erosion, 148–149 sustainable agriculture, 251–252 Soil accumulation, 295 Soil acidity, 269–270 Soil adsorption, 44 Soil and Water Assessment Tool (SWAT), 294 Soil biological activity, 159 Soil building, 158–159 Soil carbon density, 560 Soil carbon sequestration, 474–478 Soil compaction, 189–190 Soil conditioner, 282 Soil conservation, 185–186 community-based approach, 412 landcare approach, 412 practices, 156, 296f vegetative and mechanical approaches, 409–411 Soil conservation practices, 296 Soil contamination, 220–221 Soil cover crops, 94 Soil cultivation, 174 Soil degradation, 217–218, 247 in Africa, 223–224, 223f, 224f on Asian lands, 221–222, 222f in Australia, 228–230 in Caribbean islands, 227, 228 in Europe, 219–221 in New Zealand, 229, 230 in North and Central America, 226–228 in South America, 224–226, 226f Soil erodibility, 397 Soil erosion, 119, 153–154, 154t, 343, 605 biological effects, 155–156 chemical effects, 155 control devices buffer strips, 527, 527f filter strip terraces, 526–527 grass strips, 525, 526f mass planting, vegetative cover, 524 riparian strips, 527, 528 settling basins, 527 vegetated waterways, 525–526 vegetative materials, 523–524 eroded soil organic carbon, 236 (see also Soil organic carbon (SOC)) factors, 619 and global change, 119, 121, 122 in China and in the World, 119–120, 120f, 120t deteriorates global environment, 121–122 threatens to global food security, 121 pesticide translocation control, 41, 47–49

639 physical effects, 155 productivity impacts, 409 scope and impact, 408 sediment composition and erodibility, 620–623 soil conservation community-based approach, 412 landcare approach, 412 vegetative and mechanical approaches, 409–411 by water, 317 (see also Soil water erosion) amendments, 282 continuous simulation models, 318–319 empirical erosion model, 307–308 (see also Empirical erosion model) empirically based model, 317 overview of, 281–282 process-based models, 317–318 Soil erosion by water, 30–31 Soil (K) factor, wind erosion, 139 Soil fauna, 101 Soil freezing, 570 Soil frost models, 150 Soil hydraulic conductivity, 200–204, 206 Soil inorganic carbon (SIC), 165 Soil Loss Estimation Model for Southern Africa (SLEMSA), 310–311 Soil loss tolerance, 295 Soil management, 94 Soil map, 218 Soil microaggregates, 85 Soil microbial activity, 101–102 organic substrates distribution, 103–106 soil temperature, 102f soil texture, 103 soil water content, 102f, 103 Soil microbial biomass, 105–106 Soil organic carbon (SOC), 165–166, 170, 235, 245, 275–277, 473 carbon dioxide (CO2) emission estimation, 238, 239, 239t mass balance approach, 236–239, 239t erosion, 236 for greenhouse effect mitigation, 168, 169, 169t hypothetical distribution of, 238 mineralization of, 167 world soil, 168 Soil organic matter (SOM), 106, 122, 473–478 Soil organisms, 155–156 SOILOSS, 312 Soil particle motion, 43 Soil Partnership, 217 Soil perturbation, 277 Soil protection, 36 Soil quality carbon gases, 243–244 carbon dioxide (CO2), 244 methane (CH4), 244

640 Soil quality (cont.) nitrogen gases, 244–245 ammonia (NH3), 246–247 loop-within-a-loop, 244, 245 nitric and nitrous oxides, 245–246 Soil quality degradation, 121 Soil quality indicators characteristics of soil and land, 357 concerns, categories of, 357–358 interpretative framework, 359 single indices, 359 spatial and time dimensions, 358 trends, 359 Soil saturation extract (EC), 573 Soil structural stability in salt-affected soils, 85–86, 91 Soil temperature, 102 Soil test P (STP) concentration, 459, 460 Soil texture, 103 Soil treatments, 179 Soil water content, 102f, 103 Soil water dynamics, 87–89 Soil water erosion, 289 amendments, 282 gypsum, 283–284, 284f inorganic mulches, 285–286 lime, 282–283 manure, compost, and organic sludge, 285 organic mulches, 285–286 synthetic polymer, 284–285 assessment, 290 controlling methods, 295 conservation structures, 300 contouring, 296, 297, 297f contour ridges and stone terraces, 299f environmental legislation, 301, 302 geotextiles, 301 gully stabilization structures, 301, 302 soil conservation practices, 296f terraces, 296–299 waterways, 299, 300 empirical erosion model ABAG, 311 development of, 308–309 Modified Universal Soil Loss Equation, 310 overview of, 307–308 predictions, accuracy of, 312–314 Revised Universal Soil Loss Equation, 312 Soil Loss Estimation Model for Southern Africa, 310–311 SOILOSS, 312 Universal Soil Loss Equation, 309–310 integrated environmental management, 303 monitoring process, 291 field plots, 291 mathematical models, 294 quantitative indicators, 295

Index sediment transport in watersheds, 292–294 visual indicators, 295 overview of, 281–282 problems of, 289–290 sedimentation, 290 sources of, 290 Solid manure, 463–464 Solid phase extraction (SPE), 63, 68 Solid phase microextraction (SPME), 63, 65 Soluble salts, 84–85 Solvent extraction, 67t SOM, see Soil organic matter (SOM) South America, soil degradation in, 224–226 Soxhlet extraction, 67 Sparingly available pool, 112 Spatial distribution, of organic substrates, 103–106 Spatial speciation, of phosphorus, 112–115 SPE, see Solid phase extraction (SPE) Species diversity, 498 Splash erosion, 25 SPME, see Solid phase microextraction (SPME) Sprinkler irrigation, 36 SSC, see Suspended sediment concentration (SSC) SST, see Sea surface temperature (SST) Steep slope reclamation, 270 Stir bar sorptive extraction (SBSE), 63 Stone terraces, 299 Strict liability, 508 Strip tillage, 347 Subsoil sodicity, 212–214 Summer precipitation, PCA, 260, 261 Supended sediment flux, 292 Supply-side pests, 352 Supporting services, grazing systems, 185–186 Surface drainage, 95 Surface irrigation systems, 2 Surface residue, 343, 344 Surface residue management, 343, 344, 346, 347 Surface runoff, 396 Surface soil, 198 Surface waterlogging, 206 Suspended sediment concentration (SSC), 292, 293 Sustainable agriculture, 83, 251 agroecosystem biodiversity, 252–253 air and atmosphere, 252 soil, 251–252 quality assessment, 253–254 strategies for, 254 water, 252 Sustainable Development Goals (SDGs), 259 Sustainable land management (SLM), 230, 231, 595 SWAT, see Soil and Water Assessment Tool (SWAT) SWG, see Switchgrass (SWG) Switchgrass (SWG), 276, 277 SWRRB, see Simulator for Water Resources in Rural Basins (SWRRB) Synanthedon exitiosa, 484

641

Index Synthetic nitrogen fertilizers, 158, 161 Synthetic polymer, 284–285

t Tachinaephagus zealandicus, 488 TCCs, see Total cation concentrations (TCCs) TEC, see Threshold electrolyte concentration (TEC) Terrace formation benefits, 547–548 engineered terraces bench terraces, 545 gradient terraces, 546–547 lynchet, 544–545 problems, 547–548 Terraces, 296–299 Thermionic specific detector (TSD), 71 Thiobacillus ferrooxidans, 519 Threshold electrolyte concentration (TEC), 93, 94 Tilapia zillii, 486, 487 Tillage, 36, 160, 343 effect on soil properties, 343–344 Tillage berms, 546, 547 Tillage erosion, 153 Tithonia diversifolia, 178 Total cation concentrations (TCCs), 203 Toxic elements, in salt-affected soils, 87 Toxicity, pesticides, 56t Transgenic crops, 353 Transient salinity, 213 Trespass, 507 Tri-trophic models, 349, 354 TSD, see Thermionic specific detector (TSD) Tunnel erosion, 290 Turbulence, 142 Tylenchulus semipenetrans, 178

U UNCCD, see United Nations Convention to Combat Desertification (UNCCD) United Nations Conference on Environment and Development (UNCED), 231 United Nations Convention to Combat Desertification (UNCCD), 231, 409, 593 United Nations Framework Convention on Climate Change, 252 United States Department of Agriculture (USDA), 19, 20 United States Department of Agriculture/ Natural Resources Conservation Service (USDA-NRCS), 218 United States Environmental Protection Agency (USEPA), 450–452 United States Salinity Laboratory (USSL), 201 Universal Index of Onchev, 125

Universal soil loss equation (USLE), 125, 235, 291, 294, 309–310, 314, 317, 319 rainfall kinetic energy, 616–617 two-staged mathematical approach, 611–612 variants, 612–616 Upland erosion, 281–282 USDA, see United States Department of Agriculture (USDA) USDA-Natural Resources Conservation Services (NRCS), 22 USDA-NRCS, see USDA-Natural Resources Conservation Services (NRCS) U.S. Department of Agriculture–Agricultural Research Service (USDA-ARS), 468 U.S. empirical erosion model, 309 U.S. Environment Protection Agency (EPA), 463 U.S. Food, Conservation, and Energy Act of 2008, 50 USLE, see Universal soil loss equation (USLE); Universal Soil Loss Equation (USLE) US mine operations, 268 U.S. organic industry, 20 USSL, see United States Salinity Laboratory (USSL)

V Vegetated waterways, 525–526 Vegetation cover, 133 wind erosion, 139 Vegetative buffers hydraulic resistance, 400–402 types, 397–400 Visual indicators, 295, 303 Volatilization, 108

W Water modern pesticides in collection and preparation, 62 isolation and enrichment techniques, 62–64 sustainable agriculture, 252 Water budget development, 456–457 Water erosion, 31, 235, 236, 606 control, 295 conservation structures, 300 contouring, 296, 297, 297f contour ridges and stone terraces, 299f environmental legislation, 301, 302 geotextiles, 301 gully stabilization structures, 301, 302 soil conservation practices, 296f terraces, 296–299 waterways, 299, 300 future climate change, 30, 31 hydrological models, 46–47, 47t impacts of, 30 overview of, 25

642 Water erosion (cont.) pesticide translocation control, 42, 43f problem of accelerated, 30 processes of, 25–28 by soil (see Soil water erosion) spatial and temporal scale, 28–29 Water erosion and vegetation controls increases infiltration, 396 reduced soil erodibility, 397 sediment trapping efficiency, 402–403 slower runoff, 395–396 vegetative buffers hydraulic resistance, 400–402 types, 397–400 Water Erosion Prediction Program (WEPP), 294, 318, 319, 617–618 Water erosion vulnerability map, 218 Water Productor Program, 302 Water properties, 37 Water quality, 94–95 Water solubility, 44 Water-soluble products, 51 Water use drinking water requirements, 454 estimation, 453 flushing manure, 455 washing cow, 454 milking equipment and parlor, 454 sprinkling and cooling, 455 Waterways, 299, 300 Weed control, 344–347 Weed management, 160 WEPP, see Water Erosion Prediction Program (WEPP) WEPS, see Wind erosion prediction system (WEPS) Wildlife habitat, 185 Wind-driven rain horizontal velocity, 553–554 impact angle, 554

Index impact frequency, 554 raindrop size distributions, 553 rainsplash detachment, 554, 556 transport, 556 silt loam soil, 555 Wind dynamics, 142 Wind erosion, 606 causes, 322–323 control, 336–337 dynamics, 142 equation, 335 global hot spots for, 137 induced erosion, 138–139 natural erosion, 137–138 implication, 323–325 models, 333–335 monitoring tools, 326–333 particle entrainment, 142, 143 pesticide translocation control, 42–44, 43f, 44f, 47 prediction system, 335–336 processes of, 141 regions affected, 325–326 self-balancing concept, 144 Wind Erosion Equation, 139 Wind erosion prediction system (WEPS), 47, 333–337 Wind erosion vulnerability map, 218 Winter erosion, 149 Winter hydrology, 147 Woody encroachment, 188–189 World Atlas of Desertification, 139 World soils; see also Global climate change C sequestration, 168 for greenhouse effect mitigation, 168–169 Worldwide statistics, organic agriculture, 20–21

Z Zabrus gibus, 417

Managing Water Resources and Hydrological Systems

Environmental Management Handbook, Second Edition Edited by Brian D. Fath and Sven E. Jørgensen

Volume 1 Managing Global Resources and Universal Processes Volume 2 Managing Biological and Ecological Systems Volume 3 Managing Soils and Terrestrial Systems Volume 4 Managing Water Resources and Hydrological Systems Volume 5 Managing Air Quality and Energy Systems Volume 6 Managing Human and Social Systems

Managing Water Resources and Hydrological Systems Second Edition

Edited by

Brian D. Fath and Sven E. Jørgensen Assistant to Editor

Megan Cole

Cover photo: Forsand, Norway, N. Fath

Second edition published 2021 by CRC Press 6000 Broken Sound Parkway NW, Suite 300, Boca Raton, FL 33487-2742 and by CRC Press 2 Park Square, Milton Park, Abingdon, Oxon, OX14 4RN © 2021 Taylor & Francis Group, LLC First edition published by CRC Press 2013 CRC Press is an imprint of Taylor & Francis Group, LLC Reasonable efforts have been made to publish reliable data and information, but the author and publisher cannot assume responsibility for the validity of all materials or the consequences of their use. The authors and publishers have attempted to trace the copyright holders of all material reproduced in this publication and apologize to copyright holders if permission to publish in this form has not been obtained. If any copyright material has not been acknowledged please write and let us know so we may rectify in any future reprint. Except as permitted under U.S. Copyright Law, no part of this book may be reprinted, reproduced, transmitted, or utilized in any form by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying, microfilming, and recording, or in any information storage or retrieval system, without written permission from the publishers. For permission to photocopy or use material electronically from this work, access www.copyright.com or contact the Copyright Clearance Center, Inc. (CCC), 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400. For works that are not available on CCC please contact [email protected] Trademark notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation without intent to infringe. ISBN: 978-1-138-34266-8 (hbk) ISBN: 978-1-003-04504-5 (ebk) Typeset in Minion by codeMantra

Contents Preface ....................................................................................................................... xi Editors ..................................................................................................................... xiii Contributors ............................................................................................................ xv

SECTION I APC: Anthropogenic Chemicals and Activities

1

Aquatic Communities: Pesticide Impacts ........................................................ 3

2

Coastal Water: Pollution .................................................................................. 17

3

Groundwater: Mining Pollution ..................................................................... 37

4

Groundwater: Nitrogen Fertilizer Contamination ........................................ 45

5

Groundwater: Pesticide Contamination ......................................................... 59

6

Lakes and Reservoirs: Pollution ..................................................................... 65

7

Mines: Acidic Drainage Water ......................................................................... 81

8

Rivers and Lakes: Acidification ...................................................................... 87

9

Rivers: Pollution . ............................................................................................ 105

10

Sea: Pollution . .................................................................................................. 111

David P. Kreutzweiser and Paul K. Sibley Piotr Szefer

Jeff Skousen and George Vance

Lloyd B. Owens and Douglas L. Karlen Roy F. Spalding

Subhankar Karmakar and O.M. Musthafa Wendy B. Gagliano and Jerry M. Bigham

Agniezka Gałuszka and Zdzistaw M. Migaszewski Bogdan Skwarzec Bogdan Skwarzec

v

vi

Contents

SECTION II COV: Comparative Overviews of Important Topics for Environmental Management

11

Rain Water: Harvesting ................................................................................. 123

12

Water Harvesting ........................................................................................... 129

13

Groundwater: Saltwater Intrusion ................................................................. 133

14

Irrigation Systems: Water Conservation ....................................................... 151

15

Irrigation: Erosion .......................................................................................... 155

16

Irrigation: River Flow Impact ........................................................................ 165

17

Irrigation: Saline Water ................................................................................. 171

18

Irrigation: Sewage Eff luent Use .................................................................... . 175

19

Irrigation: Soil Salinity .................................................................................. 179

20

Managing Water Resources and Hydrological Systems . ............................... 185

21

Runoff Water . ................................................................................................. 195

22

Salt Marsh Resilience and Vulnerability to Sea-Level Rise and Other Environmental Impacts . ............................................................... 215

K.F. Andrew Lo

Gary W. Frasier

Alexander H.-D. Cheng

Pai Wu, Javier Barragan, and Vince Bralts David L. Bjorneberg

Robert W. Hill and Ivan A. Walter B.A. Stewart B.A. Stewart

James D. Rhoades

Honglin Zhong, Zhuoran Liang, and Cheng Li Zaneta Polkowska

Daria Nikitina

23

The Evolution of Water Resources Management ......................................... 225

24

Wastewater and Water Utilities ..................................................................... 241

25

Wastewater: Municipal .................................................................................. 253

26 27

Francine van den Brandeler and Joyeeta Gupta Rudolf Marloth

Sven Erik Jørgensen

Water Quality and Quantity: Globalization ................................................. 265

Kristi Denise Caravella and Jocilyn Danise Martinez

Water: Cost . ................................................................................................... 277 Atif Kubursi and Matthew Agarwala

Contents

28

vii

Wetlands: Methane Emission ....................................................................... 293 Anna Ekberg and Tørben Røjle Christensen

SECTION III

CSS: Case Studies of Environmental Management

29

Alexandria Lake Maryut: Integrated Environmental Management ............. 301

30

Aral Sea Disaster ............................................................................................ 317

31

Chesapeake Bay ............................................................................................. 323

32

Giant Reed (Arundo donax): Streams and Water Resources ....................... 327

33

Inland Seas and Lakes: Central Asia Case Study .......................................... 337

34

Oil Pollution: The Baltic Sea ......................................................................... 349

35

Status of Groundwater Arsenic Contamination in the GMB Plain ............. 369

36

Yellow River ................................................................................................... 383

Lindsay Beevers Guy Fipps

Sean M. Smith

Gretchen C. Coffman Andrey G. Kostianoy Andrey G. Kostianoy

Abhijit Das, Antara Das, Meenakshi Mukherjee, Bhaskar Das, Subhas Chandra Mukherjee, Shyamapada Pati, Rathindra Nath Dutta, Quazi Quamruzzaman, Khitish Chandra Saha, Mohammad Mahmudur Rahman, Dipankar Chakraborti, and Tarit Roychowdhury Zixi Zhu, Ynuzhang Wang, and Yifei Zhu

SECTION IV DIA: Diagnostic Tools: Monitoring, Ecological Modeling, Ecological Indicators, and Ecological Services

37

Groundwater: Modeling ................................................................................ 389

38

Groundwater: Numerical Method Modeling ................................................ 399

39

Nitrogen (Nitrate Leaching) Index ............................................................... 407

40

Nitrogen (Nutrient) Trading Tool. ................................................................. 415

41

The Accounting Framework of Energy–Water Nexus in Socioeconomic Systems ................................................................................. 423

Jesus Carrera Jesus Carrera

Jorge A. Delgado Jorge A. Delgado

Saige Wang and Bin Chen

42

Water Quality: Modeling .............................................................................. 427 Richard Lowrance

viii

Contents

SECTION V ELE: Focuses on the Use of Legislation or Policy to Address Environmental Problems

43

Drainage: Hydrological Impacts Downstream ............................................. 433

44

Drainage: Soil Salinity Management ............................................................ 439

45

Lakes: Restoration ......................................................................................... 445

46

Wastewater Use in Agriculture: Policy Issues ............................................... 461

47

Water: Total Maximum Daily Load ............................................................... 481

48

Watershed Management: Remote Sensing and GIS ....................................... 491

49

Wetlands: Conservation Policy . .................................................................... 497

Mark Robinson and D.W. Rycroft Glenn J. Hoffman Anna Rabajczyk

Dennis Wichelns

Robin Kundis Craig

A.V. Shanwal and S.P. Singh Clayton Rubec

SECTION VI ENT: Environmental Management Using Environmental Technologies

50

Irrigation Systems: Subsurface Drip Design ................................................ 505

51

Recent Approaches to Robust Water Resources Management under Hydroclimatic Uncertainty . ........................................................................... 511

Carl R. Camp, Jr. and Freddie L. Lamm

J. Pablo Ortiz-Partida, Mahesh L. Makey, and Alejandra Virgen-Urcelay

52

Rivers: Restoration ......................................................................................... 519

53

Waste: Stabilization Ponds ............................................................................. 535

54

Wastewater Treatment Wetlands: Use in Arctic Regions 5-Year Update .... 543

55

Wastewater Treatment: Biological ................................................................. 561

56

Wastewater Treatment: Conventional Methods ........................................... 577

57

Water and Wastewater: Filters ...................................................................... 583

Anna Rabajczyk

Sven Erik Jørgensen

Colin N. Yates, Brent Wootton, and Stephen D. Murphy

Shaikh Ziauddin Ahammad, David W. Graham, and Jan Dolfing Sven Erik Jørgensen Sandeep Joshi

Contents

ix

58

Wetlands: Constructed Subsurface ............................................................... 603

59

Wetlands: Sedimentation and Ecological Engineering ................................. 617

60

Wetlands: Treatment System Use .................................................................. 621

Jan Vymazal

Timothy C. Granata and J.F. Martin Kyle R. Mankin

SECTION VII NEC: Natural Elements and Chemicals Found in Nature

61

Cyanobacteria: Eutrophic Freshwater Systems ............................................. 631

62

Estuaries ........................................................................................................ 635

63

Everglades ....................................................................................................... 651

64

Water Quality: Range and Pasture Land ...................................................... 655

65

Water: Drinking ............................................................................................. 661

66

Water: Surface ............................................................................................... 679

67

Wetlands ........................................................................................................ 685

Anja Gassner and Martin V. Frey Claude Amiard-Triquet

Kenneth L. Campbell, Rafael Munoz-Carpena, and Gregory Kiker Thomas L. Thurow

Marek Biziuk and Matgorzata Michalska Victor de Vlaming Ralph W. Tiner

SECTION VIII PRO: Basic Environmental Processes

68

Eutrophication . .............................................................................................. 693

69

Wastewater Use in Agriculture . ..................................................................... 701

70

Wetlands: Biodiversity .................................................................................. 709

71

Wetlands: Carbon Sequestration ................................................................... 715

Sven Erik Jørgensen

Manzoor Qadir, Pay Drechsel, and Liqa Raschid-Sally Jean-Claude Lefeuvre and Virginie Bouchard Virginie Bouchard and Matthew Cochran

Index ....................................................................................................................... 721

Preface Given the current state of the world as compiled in the massive Millennium Ecosystem Assessment Report, humans have changed ecosystems more rapidly and extensively during the past 50 years than in any other time in human history. These are unprecedented changes that need certain action. As a result, it is imperative that we have a good scientific understanding of how these systems function and good strategies on how to manage them. In a very practical way, this multivolume Environmental Management Handbook provides a comprehensive reference to demonstrate the key processes and provisions for enhancing environmental management. The experience, evidence, methods, and models relevant for studying environmental management are presented here in six stand-alone thematic volumes, as follows: VOLUME 1 – Managing Global Resources and Universal Processes VOLUME 2 – Managing Biological and Ecological Systems VOLUME 3 – Managing Soils and Terrestrial Systems VOLUME 4 – Managing Water Resources and Hydrological Systems VOLUME 5 – Managing Air Quality and Energy Systems VOLUME 6 – Managing Human and Social Systems In this manner, the handbook introduces in the first volume the general concepts and processes used in environmental management. The next four volumes deal with each of the four spheres of nature (biosphere, geosphere, hydrosphere, and atmosphere). The last volume ties the material together in its application to human and social systems. These are very important chapters for a wide spectrum of students and professionals to understand and implement environmental management. In particular, features include the following: • The first handbook that demonstrates the key processes and provisions for enhancing environmental management. • Addresses new and cutting-edge topics on ecosystem services, resilience, sustainability, food– energy–water nexus, socio-ecological systems, etc. • Provides an excellent basic knowledge on environmental systems, explains how these systems function, and gives strategies on how to manage them. • Written by an outstanding group of environmental experts. Since the handbook covers such a wide range of materials from basic processes, to tools, technologies, case studies, and legislative actions, each handbook entry is further classified into the following categories: APC: Anthropogenic chemicals: The chapters cover human-manufactured chemicals and their activities COV: Indicates that the chapters give comparative overviews of important topics for environmental management xi

xii

Preface

CSS: The chapters give a case study of a particular environmental management example DIA: Means that the chapters are about diagnostic tools: monitoring, ecological modeling, ecological indicators, and ecological services ELE: Focuses on the use of legislation or policy to address environmental problems ENT: Addresses environmental management using environmental technologies NEC: Natural elements and chemicals: The chapters cover basic elements and chemicals found in nature PRO: The chapters cover basic environmental processes. Volume 4, Managing Water Resources and Hydrological Systems, has extensive coverage in over 80 entries of water supply, water treatment, wetlands, lakes, and other natural water systems. New entries cover the evolution of water management, with application of optimization tools, and the innovative move toward integrating the energy and water nexus. Case studies include the Aral Sea, Chesapeake Bay, Baltic Sea, and Yellow River to name a few. Policy implications regarding wetland conservation, use of remote sensing and GIS, and agricultural water use are included. Brian D. Fath Brno, Czech Republic December 2019

Editors Brian D. Fath is Professor in the Department of Biological Sciences at Towson University (Maryland, USA) and Senior Research Scholar at the International Institute for Applied Systems Analysis (Laxenburg, Austria). He has published over 180 research papers, reports, and book chapters on environmental systems modeling, specifically in the areas of network analysis, urban metabolism, and sustainability. He has co-authored the books A New Ecology: Systems Perspective (2020), Foundations for Sustainability: A Coherent Framework of Life–Environment Relations (2019), and Flourishing within Limits to Growth: Following Nature’s Way (2015). He is also Editor-in-Chief for the journal Ecological Modelling and Co-Editor-in-Chief for Current Research in Environmental Sustainability. Dr. Fath was the 2016 recipient of the Prigogine Medal for outstanding work in systems ecology and twice a Fulbright Distinguished Chair (Parthenope University, Naples, Italy in 2012 and Masaryk University, Czech Republic in 2019). In addition, he has served as Secretary General of the International Society for Ecological Modelling, Co-Chair of the Ecosystem Dynamics Focus Research Group in the Community Surface Modeling Dynamics System, and member and past Chair of Baltimore County Commission on Environmental Quality. Sven E. Jørgensen (1934–2016) was Professor of environmental chemistry at Copenhagen University. He received a doctorate of engineering in environmental technology and a doctorate of science in ecological modeling. He was an honorable doctor of science at Coimbra University (Portugal) and at Dar es Salaam (Tanzania). He was Editor-in-Chief of Ecological Modelling from the journal inception in 1975 until 2009. He was Editor-in-Chief for the Encyclopedia of Environmental Management (2013) and Encyclopedia of Ecology (2008). In 2004, Dr. Jørgensen was awarded the Stockholm Water Prize and the Prigogine Medal. He was awarded the Einstein Professorship by the Chinese Academy of Sciences in 2005. In 2007, he received the Pascal Medal and was elected a member of the European Academy of Sciences. He has published over 350 papers, and has edited or written over 70 books. Dr. Jørgensen gave popular and well-received lectures and courses in ecological modeling, ecosystem theory, and ecological engineering worldwide.

xiii

Contributors Matthew Agarwala Bennett Institute for Public Policy University of Cambridge Cambridge, United Kingdom Shaikh Ziauddin Ahammad School of Civil Engineering and Geosciences Newcastle University Newcastle, United Kingdom Claude Amiard-Triquet French National Center for Scientific Research (CNRS) University of Nantes Nantes, France Javier Barragan Department of Agroforestry Engineering University of Lleida Lleida, Spain Lindsay Beevers Lecturer in Water Management School of the Built Environment Heriot Watt University Edinburgh, United Kingdom Jerry M. Bigham School of Environment and Natural Resources Ohio State University Columbus, Ohio Marek Biziuk Chemical Faculty Department of Analytical Chemistry Gdansk University of Technology Gdansk, Poland

David L. Bjorneberg Northwest Irrigation and Soil Research Laboratory Agricultural Research Service (USDA-ARS) U.S. Department of Agriculture Kimberly, Idaho Virginie Bouchard School of Natural Resources Ohio State University Columbus, Ohio Vince Bralts Agricultural and Biological Engineering Purdue University West Lafayette, Indiana Carl R. Camp, Jr. Agricultural Research Service (USDA-ARS) U.S. Department of Agriculture Florence, South Carolina (Retired) Kenneth L. Campbell Agricultural and Biological Engineering Department University of Florida Gainesville, Florida Kristi Denise Caravella Florida Atlantic University Boca Raton, Florida Jesus Carrera Technical University of Catalonia (UPC) Barcelona, Spain

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Contributors

Dipankar Chakraborti School of Environmental Studies Jadavpur University Calcutta, India

Bhaskar Das School of Environmental Studies Jadavpur University Kolkata, India

Bin Chen State Key Joint Laboratory of Environmental Simulation and Pollution Control School of Environment Beijing Normal University Beijing, China

Jorge A. Delgado Soil Management and Sugar Beet Research Unit Agricultural Research Service (USDA-ARS) U.S. Department of Agriculture Fort Collins, Colorado

Alexander H.-D. Cheng Department of Civil Engineering University of Mississippi Oxford, Mississippi

Jan Dolfing School of Civil Engineering and Geosciences Newcastle University Newcastle, United Kingdom

Tørben Røjle Christensen Climate Impacts Group Department of Ecology Lund University Lund, Sweden

Pay Drechsel International Water Management Institute (IWMI) Colombo, Sri Lanka

Matthew Cochran School of Natural Resources Ohio State University Columbus, Ohio Gretchen C. Coffman Department of Environmental Science University of San Francisco San Francisco, California Robin Kundis Craig Attorneys’ Title Professor of Law and Associate Dean for Environmental Programs Florida State University College of Law Tallahassee, Florida Abhijit Das Vijoygarh Jyotish Ray College University of Calcutta Kolkata, India Antara Das School of Environmental Studies Jadavpur University Kolkata, India

Rathindra Nath Dutta Department of Dermatology Institute of Post Graduate Medical Education and Research SSKM Hospital Kolkata, India Anna Ekberg Department of Ecology Lund University Lund, Sweden Guy Fipps Agricultural Engineering Department Texas A&M University College Station, Texas Gary W. Frasier U.S. Department of Agriculture (USDA) Fort Collins, Colorado Martin V. Frey Department of Soil Science University of Stellenbosch Matieland, South Africa

Contributors

Wendy B. Gagliano Clark State Community College Springfield, Ohio Agniezka Gałuszka Division of Geochemistry and the Environment Institute of Chemistry Jan Kochanowski University Kielce, Poland Anja Gassner Institute of Science and Technology University of Malaysia-Sabah Kota Kinabalu, Malaysia David W. Graham School of Civil Engineering and Geosciences Newcastle University Newcastle, United Kingdom Timothy C. Granata Department of Civil and Environmental Engineering and Geodetic Science Ohio State University Columbus, Ohio Joyeeta Gupta Department of Human Geography Planning and International Development Amsterdam Institute of Social Science Research University of Amsterdam Amsterdam, the Netherlands and IHE Institute for Water Education in Delft Delft, the Netherlands Robert W. Hill Biological and Irrigation Engineering Department Utah State University Logan, Utah Glenn J. Hoffman Biological Systems Engineering University of Nebraska–Lincoln Lincoln, Nebraska

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Sven Erik Jørgensen Section of Environmental Chemistry Copenhagen University Copenhagen, Denmark Sandeep Joshi Shrishti Eco-Research Unit (SERI) Pune, India Douglas L. Karlen U.S. Department of Agriculture (USDA) Ames, Iowa Subhankar Karmakar Center for Environmental Science and Engineering (CESE) Indian Institute of Technology Bombay Mumbai, India Gregory Kiker University of Florida Gainesville, Florida Andrey G. Kostianoy P.P. Shirshov Institute of Oceanology Russian Academy of Sciences Moscow, Russia David P. Kreutzweiser Canadian Forest Service Natural Resources Canada Sault Sainte Marie, Ontario, Canada Atif Kubursi Department of Economics McMaster University Hamilton, Ontario, Canada Freddie L. Lamm Northwest Research-Extension Center Kansas State University Colby, Kansas Jean-Claude Lefeuvre Laboratory of the Evolution of Natural and Modified Systems University of Rennes Rennes, France

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Cheng Li Guangdong Key Laboratory of Agricultural Environment Pollution Integrated Control Guangdong Institute of Eco-Environmental Science and Technology Guangzhou, China

Contributors

Zdzistaw M. Migaszewski Division of Geochemistry and the Environment Institute of Chemistry Jan Kochanowski University Kielce, Poland

Zhuoran Liang Hangzhou Meteorological Bureau Hangzhou, China

Meenakshi Mukherjee School of Environmental Studies Jadavpur University Kolkata, India

K.F. Andrew Lo Department of Natural Resources Chinese Culture University Taipei, Taiwan

Subhas Chandra Mukherjee Department of Neurology Medical College Kolkata, India

Richard Lowrance Agricultural Research Service (USDA-ARS) U.S. Department of Agriculture Tifton, Georgia

Rafael Munoz-Carpena University of Florida Gainesville, Florida

Mahesh L. Makey University of California Oakland, California

Stephen D. Murphy Faculty of Environment University of Waterloo Waterloo, Ontario, Canada

Kyle R. Mankin Department of Biological and Agricultural Engineering Kansas State University Manhattan, Kansas

O.M. Musthafa Center for Pollution Control and Environmental Engineering Pondicherry University Pondicherry, India

Rudolf Marloth San Diego State University San Diego, California

Daria Nikitina Department of Earth and Space Sciences West Chester University of Pennsylvania West Chester, Pennsylvania

J.F. Martin Department of Food, Agricultural, and Biological Engineering Ohio State University Columbus, Ohio Jocilyn Danise Martinez University of South Florida Tampa, Florida Matgorzata Michalska Institute of Maritime and Tropical Medicine Gdynia, Poland

J. Pablo Ortiz-Partida Union of Concerned Scientists Cambridge, Massachusetts Lloyd B. Owens U.S. Department of Agriculture (USDA) Coshocton, Ohio Shyamapada Pati Department of Obstetrics and Gynaecology Calcutta National Medical College Kolkata, India

Contributors

Zaneta Polkowska Gdansk University of Technology Gdansk, Poland Manzoor Qadir Institute for Water, Environment and Health (UNU-INWEH) Hamilton, Ontario, Canada Quazi Quamruzzaman Dhaka Community Hospital Dhaka, Bangladesh Anna Rabajczyk Independent Department of Environment Protection and Modeling Jan Kochanowski University of Humanities and Sciences Kielce, Poland Mohammad Mahmudur Rahman School of Environmental Studies Jadavpur University Kolkata, India Liqa Raschid-Sally International Water Management Institute (IWMI) Colombo, Sri Lanka James D. Rhoades Agricultural Salinity Consulting Riverside, California

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D.W. Rycroft Department of Civil and Environmental Engineering Southampton University Southampton, United Kingdom Khitish Chandra Saha School of Environmental Studies Jadavpur University Kolkata, India A.V. Shanwal Department of Soil Science Chaudhary Charan Singh Haryana Agricultural University Hisar, India Paul K. Sibley School of Environmental Sciences University of Guelph Guelph, Ontario, Canada S.P. Singh National Bureau of Soil Survey and Land Use Planning Indian Agricultural Research Institute New Delhi, India Jeff Skousen Division of Plant and Soil Sciences West Virginia University Morgantown, West Virginia

Mark Robinson Center for Ecology and Hydrology Wallingford, United Kingdom

Bogdan Skwarzec Faculty of Chemistry University of Gdansk Gdansk, Poland

Tarit Roychowdhury School of Environmental Studies Jadavpur University Calcutta, India

Sean M. Smith Ecosystem Restoration Center Maryland Department of Natural Resources Annapolis, Maryland

Clayton Rubec Center for Environmental Stewardship and Conservation Ottawa, Ontario, Canada

Roy F. Spalding Water Science Laboratory University of Nebraska–Lincoln Lincoln, Nebraska

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B.A. Stewart Dryland Agriculture Institute West Texas A&M University Canyon, Texas Piotr Szefer Department of Food Sciences Medical University of Gdansk Gdansk, Poland Thomas L. Thurow Department of Renewable Resources University of Wyoming Laramie, Wyoming Ralph W. Tiner National Wetlands Inventory Program U.S. Fish and Wildlife Service Hadley, Massachusetts Francine van den Brandeler Department of Human Geography Planning and International Development University of Amsterdam Amsterdam, the Netherlands George Vance Department of Renewable Resources University of Wyoming Laramie, Wyoming Alejandra Virgen-Urcelay University of British Columbia Vancouver, Canada Victor de Vlaming Aquatic Toxicology Laboratory University of California—Davis Davis, California Jan Vymazal Faculty of Environmental Sciences Department of Landscape Ecology Czech University of Life Sciences Prague, Czech Republic Ivan A. Walter Ivan’s Engineering, Inc. Denver, Colorado

Contributors

Saige Wang State Key Joint Laboratory of Environmental Simulation and Pollution Control School of Environment Beijing Normal University Beijing, China Ynuzhang Wang Academy of Yellow River Conservancy Science Zhengzhou, China Dennis Wichelns International Water Management Unit Colombo, Sri Lanka Brent Wootton Center for Alternative Wastewater Treatment Fleming College Lindsay, Ontario, Canada Pai Wu College of Tropical Agriculture and Human Resources University of Hawaii Honolulu, Hawaii Colin N. Yates Centre for Research and Innovation Fanshawe College London, Ontario, Canada Honglin Zhong Department of Geographical Sciences University of Maryland College Park, Maryland Yifei Zhu Gemune LLC Fremont, California Zixi Zhu Henan Institute of Meteorology Zhengzhou, China

APC: Anthropogenic Chemicals and Activities

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1 Aquatic Communities: Pesticide Impacts

David P. Kreutzweiser and Paul K. Sibley

Introduction ...................................................................................................... 3 Measuring Impacts on Aquatic Communities ............................................. 4 Assessing Risk of Pesticide Impacts on Aquatic Communities.................. 5 Some Examples of Pesticide Impacts on Aquatic Communities................ 7 Reducing Risk of Pesticide Impacts on Aquatic Communities .................. 9 Recent Advances and Outstanding Issues ..................................................... 9 Conclusions ...................................................................................................... 11 Refer erences ences ........................................................................................................ 12

Introduction A biotic community can be defined as an assemblage of plant or animal species utilizing common resources and cohabiting a specific area. Examples could include a fish community of a stream, an insect community of a forest pond, or a phytoplankton community of a lake. Interactions among species provide ecological linkages that connect food webs and energy pathways, and these interconnections provide a degree of stability, or balance, to the community. Community balance can be described as a state of dynamic equilibrium in which species and their population dynamics within a community remain relatively stable, subject to changes through natural adjustment processes. Toxic effects of pesticides can disrupt these processes and linkages and thereby cause community balance upsets. For example, this can occur when a pesticide has a direct impact on a certain species in a community and reduces its abundance while other unaffected species increase in abundance in response to the reduced competition for food resources or increased habitat availability. Some of the best examples of pesticide impacts on biological communities are found in freshwater studies. Freshwater aquatic communities are usually contained within distinct boundaries or systems, and this generates a high degree of connectivity among species, thereby increasing their susceptibility to pesticide-induced disturbances at the community level. We examine traditional and developing methods for measuring pesticide impacts on freshwater communities, with emphasis on recent improvements in risk assessment approaches and analyses, and provide some examples for illustration. We then describe some advances in impact mitigation strategies and discuss some ongoing issues pertaining to understanding, assessing, and preventing pesticide impacts including probabilistic risk assessment (PRA), population and ecological modeling, and pesticide interactions with multiple stressors. The integration of improved risk assessment and mitigation approaches and technologies together with information generated from the numerous impact studies available will provide a sound scientific basis for decisions around the use and regulation of pesticides in and near water bodies.

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Managing Water Resources and Hydrological Systems

Measuring Impacts on Aquatic Communities Changes in aquatic communities can be measured directly in water bodies by a number of quantitative and qualitative sampling methods. Descriptions of those methods can be found in any up-to-date text or handbook (e.g., Hauer and Lamberti[1]). Measurements can be in terms of community structure (species composition) or community function (a measurable ecosystem process attributable to a biotic community that causes a change in condition) and can include both direct and indirect effects.[2,3] Community structure is a measure of biodiversity in its most general sense, that is, the number of species or other taxonomic units and their relative abundances. Some community functions are referred to as environmental or ecosystem services. Examples include organic matter breakdown and nutrient cycling that is largely mediated by microbial communities, or water uptake, filtration, and flood control mediated by shoreline plant communities.[4] Both community structure (biodiversity) and function (ecosystem services) are being increasingly valued by society and global economies,[5,6] and therefore sustaining healthy aquatic communities will be an important driver of pesticide impact mitigation efforts. Detecting impacts of pesticides typically involves repeated sampling and a comparison of community attributes among contaminated and uncontaminated test units over time, or across a gradient of pesticide concentrations. The test units can range from petri dishes to natural ecosystems, with a trade-off between experimental control in small test units and environmental realism in field-level testing and whole ecosystems.[7] In an effort to incorporate both experimental control and environmental realism in pesticide impact testing, the use of microcosms or model ecosystems for measuring impacts on aquatic communities has increased over the past couple of decades.[8,9] Model ecosystems for community-level pesticide testing can be quite simple at lower-trophic levels such as with microbial communities (e.g., Widenfalk et al.[10]) but will necessarily be more complex for testing higher-order biological communities (e.g., Wojtaszek[11]). Regardless of the test units, an important consideration for measuring pesticide impacts will be an assessment of the duration of impact or rate of recovery. A rapid return to pre-pesticide or reference (nopesticide) community condition will reduce the longterm ecological consequences of the pesticide disturbance.[12] Traditional measures of community-level impacts have focused on structure and have usually been expressed in terms of single-variable indices such as species richness, diversity, or abundance. These indices are useful descriptors of community structure but suffer from the fact that they reduce complex community data to a single summary metric and may miss subtle or ecologically important changes in species composition across sites or times. Over the last couple of decades, ecotoxicologists have increasingly turned to multivariate statistical techniques for analyzing community response data.[13] A variety of multivariate statistical techniques and software are available and are usually considered superior for the analysis of community data because they retain and incorporate the spatial and temporal multidimensional nature of biological communities.[14] This includes various ordination techniques that can provide graphical representation of spatiotemporal patterns in community structure in which points that lie close together in the ordination plot represent communities of similar composition (richness, abundance), while communities with dissimilar species composition are plotted further apart. Figure 1 illustrates the use of an ordination plot generated by nonmetric multidimensional scaling for detecting differences among aquatic insect communities in four control and eight insecticide-treated streams. These data have been adjusted for illustrative purposes but are based on real invertebrate community responses to an insecticide in outdoor stream channels.[15] At both concentrations of the insecticide, the community structure of stream insects clearly shifted away from the natural community composition in control streams as depicted by the separation of treated streams (T1 and T2) from controls (C) in the ordination bi-plot. The plot also illustrates that the variability among treated streams (relative distance between points) was greater than that among control streams, that the lowconcentration streams (T1) and high-concentration streams (T2) tended to separate along axis 1, and that the T2 streams were further removed from controls than the T1 streams, indicating a differential response by the insect communities to the two test concentrations. Canonical correspondence analysis

Aquatic Communities: Pesticide Impacts

5

FIGURE 1 Ordination by nonmetric multidimensional scaling of aquatic insect communities in stream channels. Each point represents the community structure of control channels (C) and channels treated with a neem-based insecticide at a low (T1) and high (T2) concentration. Source: Adapted from Kreutzweiser et al.[15]

and redundancy analysis have also been commonly used to assess aquatic community responses to pesticide contamination.[16,17] A useful refinement of an ordination technique for detecting and interpreting pesticide impacts on aquatic communities is principal response curves (PRCs).[18] PRC is derived from redundancy analysis, and time-dependent responses in the treatments are expressed as deviations from the control or reference system allowing for clear visualization of pesticide effects.

Assessing Risk of Pesticide Impacts on Aquatic Communities The likelihood or risk of harmful effects on aquatic communities from exposure to pesticides will depend on the exposure concentration, bioavailability, exposure duration, rate of uptake, inherent species sensitivities, community composition, and other community attributes. All of these must be measured, estimated, modeled, or predicted to derive an assessment of risk to aquatic communities for any given pesticide. Formalized risk assessment frameworks and guidelines for pesticides have been developed in the United States,[19] the European Union,[20] Canada,[21] and elsewhere and can be consulted for detailed information on the various components of a risk assessment. In brief, pesticide risk assessments typically include the following phases: 1) defining the problem by determining the pesticide use patterns and developing conceptual models and hypotheses around how it is expected to behave, the anticipated exposure regimes, the kinds of organisms that are likely to be at risk, the community or entity that is to be protected, and the level of protection that will be acceptable; 2) developing the measurement endpoints for assessing risk of harm by establishing which response measurements are relevant and applicable, and how the measurements will be made; 3) outlining the risk assessment process by specifying the kinds of data to be used and how they will be derived including simulation modeling, empirical laboratory, microcosm or field testing, their appropriate spatial and temporal scales, and their statistical analyses; 4) applying the risk assessment by running models or collecting data, completing analyses, summarizing outputs, and providing risk estimates; 5) conducting risk communication and management by answering questions posed in the problem formulation, suggesting risk mitigation strategies if necessary, and communicating those to appropriate users; and 6) conducting follow-up monitoring to evaluate the success of mitigation strategies and to implement adaptive management to address deficiencies if or when necessary.[22,23] Traditionally, pesticide risk assessments have relied on standardized, single-species toxicity tests to predict effects on communities, the underlying assumption being that protecting the most sensitive

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Managing Water Resources and Hydrological Systems

species will protect whole communities. In this case, the selection and relevance of test species are critically important to a successful and meaningful risk assessment.[24] However, the accuracy and relevance of estimating the potential risk to aquatic communities can be greatly improved by consideration of specific species or community attributes. In particular, attribute information can improve the ecological relevance and predictive capabilities of conceptual models and the generation of hypotheses in the risk assessment process. Insofar as these attributes affect exposure, sensitivity, or both, they can increase or decrease risk beyond what could be determined from toxicity estimates or species sensitivity distributions alone. Behavioral attributes can elevate the risk of pesticide effects on species by increasing the likelihood of intercepting the stressor. For example, young-of-the-year bluefish (Pomatomus saltatrix) typically feed in estuaries during their early life stages where agricultural runoff can elevate concentrations of pesticides in food items. This feeding behavior can result in bioaccumulation and in adverse effects such as reduced migration, overwinter survival, and recruitment success in fish communities.[25] Incorporating this kind of information into conceptual models and risk hypotheses will generate more realistic risk assessments. In addition, behavioral attributes themselves can be relevant measurement endpoints if the pesticide mode of action indicates risk of sublethal behavioral effects at expected concentrations. For example, some pesticides have been shown to impair the ability to capture prey in fish[26] and the ability to avoid predators in zooplankton.[27] These types of adverse effects can disrupt trophic linkages and reduce survival or reproduction, thus impacting community balance. Inclusion of life history information into conceptual models and risk hypotheses can also refine and improve the risk assessment process. Life history strategies can influence a species susceptibility to a stressor through effects on a population’s resilience or ability to recover from disturbance.[28] Different species exposed to the same pesticide and experiencing similar levels of effect in terms of population declines do not necessarily recover at the same rates when recovery is dependent on reproduction or dispersion. Populations of organisms with short regeneration times (e.g., several generations per year) and/or high dispersal capacity have higher likelihood of recovery from pesticide-induced population declines than those with longer regeneration periods and limited dispersal capacity. These differential life history strategies and their influences on community response and recovery from pesticide effects have been demonstrated empirically (e.g., van den Brink et al.[29] and Kreutzweiser et al.[30]) and through population modeling.[31] These community balance upsets could not have been predicted from screening-level toxicity data or from species sensitivity data; thus, inclusion of life history information in conceptual models can improve risk hypotheses and direct the assessment to focus on species at higher risk owing to specific life history strategies. Life history attributes can also influence the risk of pesticide effects through differential life-stage sensitivity or susceptibility. Early life stages are often (but not exclusively) more sensitive to pesticides than later stages. An organism’s life stage can also influence its susceptibility to a pesticide by increasing or decreasing the likelihood of intercepting the stressor. If a contaminant is present in the environment at effective concentrations during a period in which the particular life stage of a species is present, then the risk to that species is increased. For some amphibians, aquatic (larval) stages could be at higher risk of direct and indirect effects of pesticides than their terrestrial (adult) life stages when their larval stage coincides with pesticide contamination of water bodies.[32] Thus, while a species sensitivity and geographical distribution may indicate potential risk, the life-stage information coupled with pesticide use pattern, timing, or fate information may indicate little likelihood of exposure to the pesticide and the risk assessment can be adjusted accordingly. Functional attributes may also be important for refining or improving pesticide risk assessments. Protection goals for populations and communities often include the safeguarding of critical biological processes or ecosystem function. Measuring ecosystem function integrates responses of component populations and can be a relevant measurement endpoint when species loss affects ecosystem function such as energy transfer and organic matter cycling.[33] However, most ecosystems are complex and it may not be clear which functional attributes are critical for sustaining ecological processes or the extent

Aquatic Communities: Pesticide Impacts

7

to which they can sustain changes in structural properties (e.g., population levels, diversity) without adversely affecting ecosystem function. Neither is it clear if functional endpoints are more or less sensitive than structural endpoints for detecting ecosystem disturbance. Some studies investigating the relationship between species diversity and ecosystem function have indicated that ecosystems can tolerate some species loss because of functional redundancy.[34] Functional redundancy is thought to occur when several species perform similar functions in ecosystems such that some may be eliminated with little or no effect on ecosystem processes. Others have suggested that redundant species are required to ensure ecosystem resilience to disturbance as a form of biological insurance, especially at large spatial scales.[35] Given these discrepancies, measurement endpoints based on functional attributes are not typically used in pesticide risk assessments because it is generally accepted that protection of community structure will protect ecosystem function. However, when specific functional attributes can be identified and are known or suspected to be at risk from a pesticide, they can be included in the data requirements for a risk assessment. An example would be the risk of adverse effects on leaf litter decomposition (a critical ecosystem function in forest soils and water bodies) posed by a systemic insecticide for control of woodboring insects in trees.[36] In that case, the protection goal was maintaining leaf litter decomposition, the community at risk was decomposer invertebrates feeding on leaves from insecticide-treated trees, and the selection of test species was directed to a specific functional group because of the unique route of exposure to decomposer organisms identified in the risk hypotheses.

Some Examples of Pesticide Impacts on Aquatic Communities A few examples will serve to illustrate how pesticides can cause disruptions to aquatic communities. DeNoyelles et al.[37] reviewed studies into pesticide impacts on aquatic communities and reported that herbicides like atrazine, hexazinone, and copper sulfate were directly toxic to most species of phytoplankton (waterborne algae). After herbicide applications, reductions in phytoplankton caused secondary reductions in herbivorous zooplankton, resulting from a depleted food source for the zooplankton. They further showed that direct adverse effects on phytoplankton can also cause disruptions to the bacterial-based energy pathways by reducing carbon flow from phytoplankton to bacteria, and ultimately to grazing protozoans and zooplankton. Boyle et al.[38] found that applications of the insecticide diflubenzuron to small ponds reduced populations of several aquatic invertebrate species. This in turn resulted in indirect effects on algae (increased productivity because of release from grazing pressure by the invertebrates) and on juvenile fish populations (reduced production because of limited invertebrate prey availability). George et al.[39] used a novel approach to predict effects of pesticide mixtures on zooplankton communities and then tested the predictions in outdoor microcosms. Responses among zooplankton populations within the community differed, depending on the pesticide mixture, and those differences appeared to reflect the relative susceptibilities among specific taxa within groups. Cladocerans declined but were less sensitive than copepods to a chlorpyrifosdominated mixture, while rotifers actually increased after application in response to release from competition or predation pressures. Kreutzweiser et al.[40] applied a neem-based insecticide to forest pond enclosures and measured effects on zooplankton community structure, respiration, and food web stability. Significant concentrationdependent reductions in numbers of adult copepods were observed, but immature copepods and cladocerans were unaffected (Figure 2). There was no evidence of recovery of adult copepods within the sampling season. During the period of maximal impact (about 4 to 9 weeks after the applications), total plankton community respiration was significantly reduced, and this contributed to significant concentration-dependent increases in dissolved oxygen and decreases in specific conductance. The reductions in adult copepods resulted in negative effects on zooplankton food web stability through elimination of a trophic link and reduced interactions and connectance. Van Wijngaarden et al.[41] evaluated the responses of aquatic communities in indoor microcosms to a suite of pesticides used for bulb crop protection. At pesticide concentrations equivalent to 5%

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Managing Water Resources and Hydrological Systems

FIGURE 2 Mean abundance (±1 SE, n = 5) of (a) adult copepods, (b) immature copepods, and (c) cladocerans in natural pond microcosms (controls) and microcosms treated at three different rates of a neem-based insecticide Source: Taken from Kreutzweiser et al.[40]

spray drift deposition, zooplankton taxa within communities showed significant changes relative to non-treated controls, reflecting taxon-specific sensitivities. Some copepods and rotifers in particular showed significant declines for at least 13 weeks, while many other rotifers and cladocerans were unaffected or increased weeks, while many other rotifers and cladocerans were unaffected or increased. Several macroinvertebrate taxa were negatively affected, and this contributed to significant declines in leaf litter decomposition among treated microcosms. The herbicide asulam was among the suite of pesticides, and it induced significant reduction of the macrophyte Elodea nuttallii. This in turn caused significant changes in water chemistry (decreases in dissolved oxygen and pH, increases in alkalinity and specific conductance) and increases in phytoplankton biomass from decreased competition for nutrients. Increased phytoplankton and reduced zooplankton predators combined to support higher abundance of less sensitive zooplankton taxa. The authors point out that most of these effects were not measurable at more realistic rates of spray drift deposition.

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9

Relyea and Hoverman[42] investigated impacts of the insecticide malathion on aquatic communities in microcosms designed to mimic a simple aquatic food web that can be found in ponds and wetlands. The insecticide generally reduced zooplankton abundance, and these reductions stimulated increases in phytoplankton, decreases in periphyton (attached algae), and decreases in growth of frog tadpoles. While invertebrate predator survival was not affected, amphibian prey survival increased with insecticide concentration, apparently the result of insecticide-induced impairment of predation success by the invertebrates. Overall, the study demonstrated that realistic concentrations of an insecticide can interact with natural predators to induce large changes in aquatic community balance.

Reducing Risk of Pesticide Impacts on Aquatic Communities For pesticides applied to crops and forests, exposure to aquatic communities can be minimized by the implementation of vegetated spray buffers or setbacks to intercept off-target spray drift and runoff.[43] Pesticide runoff can be further reduced by using formulations that are less prone to wash-off, leaching, and mobilization. Recent advances in spray drift reduction and improved spray guidance systems can also significantly reduce the off-target movement of pesticides to water bodies.[44] Examples include new technologies in map-based automated boom systems for row crops[45] and Geographical Information System (GIS)-based landscape analysis for predicting off-target pesticide movement.[46] The risk of adverse effects on aquatic communities may also be decreased by intentional selection and use of pesticides that are inherently safer to the environment. This would include so-called reduced-risk pesticides that are bioactive compounds usually with unique modes of action and derived from microbial, plant, or other natural sources. These are generally thought to be less persistent and toxic to non-target organisms than conventional synthetic pesticides.[47] Examples include the bacteria-derived insecticide Bt (Bacillus thuringiensis), the plant-derived insecticide neem, and the microbe-derived herbicide phosphinothricin. However, Thompson and Kreutzweiser[48] caution that it cannot be assumed that this group of pesticides is inherently safer or more environmentally acceptable than synthetic counterparts and that full environmental risk evaluations must be conducted to ensure their environmental safety. These types of technologies combined with the use of non-pesticide approaches to pest management form the basis of integrated pest management (IPM) strategies. IPM strategies are those in which the judicious use of pesticides is only one of several concurrent methods to control or manage losses from pest damage. This can include the use of natural enemies and parasites, biological control agents, insect growth regulators, confusion pheromones, sterile male releases, synchronizing with weather patterns known to diminish pest populations, and cultivation methods and crop varieties to improve conditions for natural enemies or degrade conditions for pest survival.[49] Increasing the use of IPM approaches can reduce reliance on pesticides and thus reduce the risk of pesticide impacts overall.

Recent Advances and Outstanding Issues Pesticide risk assessments and risk reductions have recently been advanced in terms of ecological realism and effectiveness through some developing methods and techniques. Traditional risk assessments have estimated hazards from pesticides by comparing the expected environmental concentration (often predicted from worst-case scenarios) to the toxic threshold for the most sensitive test species. When the expected concentration is higher than the toxicity threshold, the pesticide is considered to have potential for environmental effects. These so-called hazard or risk quotient approaches are still widely used in pesticide risk assessment and regulation, but more recently, PRA and probabilistic hazard assessment (PHA) approaches are being adopted. In these approaches, pesticide exposure levels and the likelihood of toxic effects are estimated from probability distributions based on all reliable data available.[50] In PRA, exposure and effects distributions are developed from modeling or measurements in laboratory, microcosm, or field studies and used to improve the accuracy and relevance of the estimated likelihood

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Managing Water Resources and Hydrological Systems

FIGURE 3 Schematic illustrating the principle of PRA (a) and PHA (b). PRA is based on a comparison of exposure and effects distributions using a predetermined criterion typically in the range of 5%-10% (shaded area and dashed lines in panel A) to determine the probability of exceeding the criterion (ellipse on y-axis); PHA is based on a comparison of an endpoint-derived sensitivity distribution within a test species to a threshold value such as a hazard quotient (dashed line in panel B).

of environmental effects compared to the traditional worst-case (hazard/risk quotient) approach (e.g., Solomon[51]). In PHA, a distribution approach is also used, except that the probability of hazard is estimated from distributions built on the relative sensitivity of interspecies endpoints rather than species sensitivity itself.[52] Figure 3 illustrates the principles of PRA (Figure 3a) and PHA (Figure 3b). Regardless of the approach, one important aspect of PRA that is ongoing is the development and use of uncertainty analysis to quantify variability and uncertainty in exposure and effects estimates. Characterizing and quantifying uncertainty will provide more meaningful risk assessments and improved decision making for minimizing potential risk of pesticide impacts in or near water.[53] Efforts at incorporating population or ecological modeling into pesticide risk assessments have also improved their accuracy and relevance for predicting, and therefore mitigating, risk of harm to aquatic communities.[54] The use of ecological models to incorporate a suite of factors including lethal and sublethal effects and their influences on the risks to organisms, populations, or communities can provide useful insights into receptor/pesticide interactions and can thereby improve risk assessments and direct mitigation measures. Population models that account for differential demographics and population

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11

growth rates within communities have been shown to provide more accurate assessments of potential pesticide impacts on populations and communities than what conventional lethal concentration estimates can provide.[55] Ecological and population modeling combined with pesticide exposure modeling and case-based reasoning (drawing on past experience or information from similar chemical exposures) can provide further refinements and improve risk assessment for aquatic communities.[56] Another recent advancement in ecological modeling to predict pesticide effects is the use of trait-based information such as organism morphology, life history, physiology, and feeding ecology in risk assessments.[57] This approach includes some of the functional attributes and concepts described above in the section on “Assessing Risk of Pesticide Impacts on Aquatic Communities” and has the advantage of formally expressing communities as combinations of functional traits rather than as groups of species, thereby yielding a more meaningful description of community structure and function. Taken together, these modeling approaches that incorporate probability distributions, toxicological sensitivities, population dynamics, ecological information, and functional trait attributes can be integrated into improved risk assessments that will inform mitigation and prevention strategies for pesticide use.[58] Two additional issues that present challenges to pesticide risk assessment and mitigation are pesticide mixtures and the combined or cumulative effects of multiple stressors on pesticide impacts. Pesticides frequently occur as mixtures in aquatic systems, particularly in agricultural regions, and methods to assess and/or predict pesticide mixture toxicity under laboratory conditions have been relatively well developed. However, there are still large uncertainties associated with the prediction of pesticide mixture toxicity, and additional studies are needed to evaluate the performance of mixture models when evaluating community-level endpoints and toxicity thresholds over long-term exposures.[59] Secondly, whereas most pesticide assessment data are derived from tests or experiments under controlled or semicontrolled environmental conditions, pesticides in natural environments may interact with a number of other natural or human-caused stressors that can substantially alter the likelihood and magnitude of pesticide impacts.[60] Other stressors could include overarching effects of climate change that can influence water temperature and quality; land use activities that result in chemical, sediment, and nutrient pollution of waterways; and biotic interactions with invasive species in aquatic communities. A number of studies have examined the combined effects of a pesticide with other stressors, but they have usually been single stressor effects tested at the si ngle-species level. Examples of studies that examined combined effects include pesticide interactions with water temperature,[61] pH,[62] dissolved organic matter,[63] UV radiation,[64] predators,[65] competitors,[66] food availability,[67] elevated sediments,[68] and other chemical stressors.[69] However, potential multiple stressors and their interactions with pesticides can be myriad and testing or extrapolating to community-level impacts is onerous at best. Sorting out and mitigating pesticide impacts from among these multiple stressors continues to be a challenge, and the suggestion by Laskowski et al.[70] to include studies of toxicant interactions with a range of environmental conditions in risk assessments seems warranted.

Conclusions Because of the high degree of connectivity among species in an aquatic community, pesticides pose a risk of harm to the community stability or balance. The community structure can be altered by direct effects, indirect effects, or both, and this can cause disruptions to the interactions and linkages among species and to their ecological function. This risk of harm will depend on exposure concentration, bioavailability, exposure duration, rate of uptake, species sensitivities, community composition, and other community attributes. Recent advances in pesticide risk assessment for aquatic communities have improved the ecological relevance and predictive capabilities for determining, and thus mitigating, potential harmful impacts. Pesticide impacts on aquatic communities can be minimized by the use of improved application technologies to reduce application rates and to decrease off-target movement to water bodies. Potential impacts can be further minimized through the selection and use of

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pesticides that are demonstrated to be inherently safer to the environment and through the application of IPM strategies. Given the preponderance of pesticide impact studies in freshwater aquatic ecosystems, the improved risk assessment frameworks and regulatory requirements for pesticide evaluations, and the recent advances in mitigation technologies, many decisions around the use of pesticides can be made on a sound scientific basis rather than on misinformed perceptions or politically driven agendas. Integrated, science-based pest management strategies including the prudent use of appropriate pesticides will contribute to ensuring the sustainability of aquatic communities in areas subjected to pest management programs.

References 1. Hauer, F.R.; Lamberti, G.A. Methods in Stream Ecology, 2nd Ed.; Elsevier: Amsterdam, 2006. 2. Fleeger, J.W.; Carman, K.R.; Nisbet, R.M. Indirect effects of contaminants in aquatic ecosystems. Sci. Tot. Environ. 2003, 317, 207–233. 3. Rohr, J.R.; Crumrine, P.W. Effects of an herbicide and an insecticide on pond community structure and processes. Ecol. Appl. 2005, 15, 1135–1147. 4. Daily, G.C. Introduction: What are ecosystem services? In Nature’s Services: Societal Dependence on Natural Ecosystems; Daily, G.D., Ed.; Island Press: Washington, DC, 1997; 1–10. 5. Anderson, J.; Gomez, W.C.; McCarney, G.; Adamowicz, W.; Chalifour, N.; Weber, M.; Elgie, S.; Howlett, M. Ecosystem Service Valuation, Market-based Instruments and Sustainable Forest Management: A Primer; Sustainable Forest Management Network: Edmonton, Alberta, 2010. 6. Bayon, R.; Jenkins, M. 2010. The business of biodiversity. Nature 2010, 466, 184–185. 7. Sibley, P.K.; Chappel, M.J.; George, T.K.; Solomon, K.R.; Liber, K. Integrating effects of stressors across levels of biological organization: Examples using organophosphorus insecticide mixtures in field-level exposures. J. Aquat. Eco-syst. Stress Recovery 2000, 7, 117–130. 8. Campbell, P.J.; Arnold, D.J.S.; Brock, T.C.M.; Grandy, N.J.; Heger, W.; Heimbach, F.; Maund, S.J.; Streloke, M. Guidance Document on Higher-Tier Aquatic Risk Assessment for Pesticides; SETAC Europe: Brussels, Belgium, 1999. 9. Kennedy, J.H.; LaPoint, T.W.; Balci, P.; Stanley, J.K.; Johnson, Z.B. Model aquatic ecosystems in ecotoxicological research: Considerations of design, implementation, and analysis. In Handbook of Ecotoxicology, 2nd Ed.; Hoffman, D.J.; Rattner, B.A.; Burton, G.A., Jr.; Cairns, J., Jr., Eds.; Lewis Publishers: Boca Raton, Florida, 2003; 45–74. 10. Widenfalk, A.; Svensson, J.M.; Goedkoop, W. Effects of the pesticides captan, deltamethrin, isoproturon, and pirimi-carb on the microbial community of a freshwater sediment. Environ. Toxicol. Chem. 2004, 23, 1920–1927. 11. Wojtaszek, B.F.; Buscarini, T.M.; Chartrand, D.T.; Stephenson, G.R.; Thompson, D.G. Effect of Release® herbicide on morality, avoidance response, and growth of amphibian larvae in two forest wetlands. Environ. Toxicol. Chem. 2005, 24, 2533–2544. 12. Barnthouse, L.W. Quantifying population recovery rates for ecological risk assessment. Environ. Toxicol. Chem. 2004, 23, 500–508. 13. Maund, S.; Chapman, P.; Kedwards, T.; Tattersfield, L.; Matthiessen, P.; Warwick, R.; Smith, E. Application of multivariate statistics to ecotoxicological field studies. Environ. Toxicol. Chem. 1999, 18, 111–112. 14. Clarke, K. R.; Warwick, R.M. Change in Marine Communities: An Approach to Statistical Analysis and Interpretation, 2nd Ed.; PRIMER-E: Plymouth, U.K., 2001. 15. Kreutzweiser, D.P.; Capell, S.S.; Scarr, T.A. Community-level responses by stream insects to neem products containing azadirachtin. Environ. Toxicol. Chem. 2000, 19, 855–861. 16. Frieberg, N.; Lindstrom, M.; Kronvang, B.; Larsen, S.E. Macroinvertebrate/sediment relationships along a pesticide gradient in Danish streams. Hydrobiologia 2003, 494, 103–110.

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37. deNoyelles, F., Jr.; Dewey, S.L.; Huggins, D.G.; Kettle, W.D. Aquatic mesocosms in ecological effects testing: Detecting direct and indirect effects of pesticides. In Aquatic Mesocosm Studies in Ecological Risk Assessment; Graney, R.L., Kennedy, J.H., Rodgers, J.H., Jr., Eds.; Lewis Publishers: Boca Raton, 1994; 577–603. 38. Boyle, T.P.; Fairchild, J.F.; Robinson-Wilson, E.F.; Haver-land, P.S.; Lebo, J.A. Ecological restructuring in experimental aquatic mesocosms due to the application of diflubenzuron. Environ. Toxicol. Chem. 1996, 15, 1806–1814. 39. George, T.K.; Liber, K.; Solomon, K.R.; Sibley, P.K. Assessment of the probabilistic ecological risk assessment-toxic equivalent combination approach for evaluating pesticide mixture toxicity to zooplankton in outdoor microcosms. Archives of Environmental Contamination and Toxicology 2003, 45, 453–461. 40. Kreutzweiser, D.P.; Sutton, T.M.; Back, R.C.; Pangle, K.L.; Thompson, D.G. Some ecological implications of a neem (azadirachtin) insecticide disturbance to zooplankton communities in forest pond enclosures. Aquat. Toxicol. 2004, 67, 239–254. 41. van Wijngaarden, R.P.A.; Cuppen, J.G.M.; Arts, G.H.P.; Crum, S.J.H.; van den Hoorn, M.W.; van den Brink, P.J.; Brock, T.C.M. Aquatic risk assessment of a realistic exposure to pesticides used in bulb crops: a microcosm study. Environ. Toxicol. Chem. 2004, 23, 1479–1498. 42. Relyea, R.A.; Hoverman, J.T. Interactive effects of predators and a pesticide on aquatic communities. Oikos 2008, 117, 1647–1658. 43. Zhang, X.; Liu, X.; Zhang, M.; Dahlgren, R.A.; Eitzel, M. A review of vegetated buffers and a metaanalysis of their mitigation efficacy in reducing nonpoint source pollution. J. Environ. Qual. 2010, 39, 76–84. 44. van de Zande, J.C.; Porskamp, H.A.; Michielsen, J.M.; Holterman, H.J.; Juijsmans, J.M. Classification of spray applications for driftability to protect surface water. Aspects Appl. Biol. 2000, 57, 57–64. 45. Luck, J.D.; Zandonadi, R.S.; Luck, B.D.; Shearer, S.A. Reducing pesticide over-application with map-based automatic boom section control on agricultural sprayers. Trans. Am. Soc. Agric. Biol. Eng. 2010, 53, 685–690. 46. Pfleeger, T.G.; Olszyk, D.; Burdick, C.A.; King, G.; Kern, J.; Fletcher, J. Using a geographical information system to identify areas with potential of off-target pesticide exposure. Environ. Toxicol. Chem. 2006, 25, 2250–2259. 47. PMRA. Regulatory Directive DIR2002–02: The PMRA Initiative for Reduced Risk Pesticides; Pest Management Regulatory Agency, Health Canada Information Services: Ottawa, Ontario; 2002, availabe at http://www.hc-sc.gc.ca/pmra-arla/english/pdf/dir/dir2002–02-e.pdf. (accessed September 2010). 48. Thompson, D.G.; Kreutzweiser, D.P. A review of the environmental fate and effects of natural “reduced-risk” pesticides in Canada. In Crop Protection Products for Organic Agriculture: Environmental, Health, and Efficacy Assessment, ACS Symposium Series 947; Felsot, A.S., Racke, K. D., Eds.; American Chemical Society: Washington, DC, 2007; 245–274. 49. van Emden, H. Integrated pest management. In Encyclopedia of Pest Management; Pimentel, D., Ed.; Marcel Dekker Inc.: New York, 2002; 413–415. 50. Solomon, K.R.; Takacs, P. Probabilistic ecological risk assessment using species sensitivity distributions. In Species Sensitivity Distributions in Ecotoxicology; Posthuma, L., Suter, G.W., Traas, T.P., Eds.; Lewis Publishers: Boca Raton, Florida, 2002; 285–314. 51. Solomon, K.R.; Baker, D.B.; Richards, R.P.; Dixon, K.R.; Klaine, S.J.; La Point, T.W.; Kendall, R.J.; Weisskopf, C.P.; Giddings, J.M.; Giesy, J.P.; Hall, L.W.; Williams, W.M. Ecological risk assessment of atrazine in North American surface waters. Environ. Toxicol. Chem. 1996, 15, 31–76. 52. Hanson, M.L.; Solomon, K.R. New technique for estimating thresholds of toxicity in ecological risk assessment. Environ. Sci. Technol. 2002, 36, 3257–3264.

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53. Warren-Hicks, W.J.; Hart, A., Eds. Application of Uncertainty Analysis to Ecological Risks of Pesticides; CRC Press: Boca Raton, Florida, 2010. 54. Thorbek, P.; Forbes, V.E.; Heimbach, F.; Hommen, U.; Thulke, H.; van den Brink, P.; Wogram, J.; Grimm, V. Ecological Models for Regulatory Risk Assessments of Pesticides; SETAC Press: Pensacola, Florida, 2010. 55. Stark, J.D.; Banks, J.E. Population-level effects of pesticides and other toxicants on arthropods. Annu. Rev. Entomol. 2003, 48, 505–519. 56. van den Brink, P.J.; Roelsma, J.; van Nes, E.H.; Scheffer, M.; Brock, T.C.M. PERPEST model: A case-based reasoning approach to predict ecological risks of pesticides. Environ. Toxicol. Chem. 2002, 21, 2500–2506. 57. Baird, D.J.; Baker, C.J.O.; Brua, R.B.; Hajibabaei, M.; McNicol, K.; Pascoe, T.J.; de Zwart, D. Towards a knowledge infrastructure for traits-based ecological risk assessment. Integr. Environ. Assess. Manage. 2010, online DOI 10.1002/ieam.129 (accessed September 2010). 58. van den Brink, P.J. Ecological risk assessment: From bookkeeping to chemical stress ecology. Environ. Sci. Technol. 2008, 42, 8999–9004. 59. Beldon, J.B.; Gilliom, R.J.; Lydy, M.J. How well can we predict the toxicity of pesticide mixtures to aquatic life? Integr. Environ. Assess. Manage. 2007, 3, 364–372. 60. Heugens, E.H.W.; Hendricks, A.J.; Dekker, T.; van Straalen, N.M.; Admiraal, W. A review of the effects of multiple stressors on aquatic organisms and analysis of uncertainty factors for use in risk assessment. Crit. Rev. Toxicol. 2001, 31, 247–285. 61. Lydy, M.J.; Lohner, T.W.; Fisher, S.W. Influence of pH, temperature and sediment type on the toxicity, accumulation and degradation of parathion in aquatic systems. Aquat. Toxicol. 1990, 17, 27–44. 62. Howe, G.E.; Marking, L.L.; Bills, T.D.; Rach, J.J.; Mayer, F.L. Jr. Effects of water temperature and pH on toxicity of terbufos, trichlorfon, 4-nitrophenol and 2,4-dinitrophenol to the amphipod Gammarus pseudolimnaeus and rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 1994, 13, 51–66. 63. Yang, W.; Spurlock, F.; Liu, W.; Gan, J. Effects of dissolved organic matter on permethrin bioavailability to Daphnia species. J. Agric. Food Chem. 2006, 54, 3967–3972. 64. Puglis, H.J.; Boone, M.D. Effects of technical-grade active ingredient vs. commercial formulation of seven pesticides in the presence or absence of UV radiation on survival of green frog tadpoles. Arch. Environ. Contam. Toxicol. 2010, online DOI 10.1007/s00244–010–9528-z (accessed September 2010). 65. Sandland, G.J.; Carmosini, N. Combined effects of a herbicide (atrazine) and predation on the life history of a pond snail, Physa gyrina. Environ. Toxicol. Chem. 2006, 25, 2216–2220. 66. Davidson, C.; Knapp, R.A. Multiple stressors and amphibian declines: Dual impacts of pesticides and fish on yellowlegged frogs. Ecol. Appl. 2007, 17, 587–597. 67. Barry, M.J.; Logan, D.C.; Ahokas, J.T.; Holdway, D.A. Effect of algal food concentration on toxicity of tow agricultural pesticides to Daphnia carinata. Ecotoxicol. Environ. Saf. 1995, 32, 273–279. 68. Wu, Q.; Riise, G.; Pflugmacher, S.; Greulich, K.; Steinberg, C.E.W. Combined effects of the fungicide propiconazole and agricultural runoff sediments on the aquatic bryophyte Vesicularia dubyana. Environ. Toxicol. Chem. 2005, 24, 2285–2290. 69. Boone, M.D.; Bridges, C.M.; Fairchild, J.F.; Little, E.E. Multiple sublethal chemicals negatively affect tadpoles of the green frog, Rana clamitans. Environ. Toxicol. Chem. 2005, 24, 1267–1280. 70. Laskowski, R.; Bednarska, A.J.; Kramarz, P.E.; Loureiro, S.; Schell, V.; Kudlek, J.; Holmstrup, M. Interactions between toxic chemicals and natural environmental factors— A meta-analysis and case studies. Sci. Tot. Environ. 2010, 408, 3763–3774.

2 Coastal Water: Pollution Introduction ..................................................................................................... 17 Heavy Metals and Metalloids ......................................................................... 18 Radionuclides...................................................................................................20 Organic Compounds ...................................................................................... 22 Biological Pollution ......................................................................................... 24 Eutrophication and Algal Bloom • Invasive Species

Piotr Szefer

Fertilizers and Pesticides ................................................................................ 26 Sewage Effluents .............................................................................................. 26 Oils .................................................................................................................... 26 Marine Debris and Plastics ............................................................................ 27 Noise Pollution ................................................................................................ 27 Light Pollution ................................................................................................. 28 Conclusion ....................................................................................................... 28 References......................................................................................................... 29

Introduction The anthropogenic activity of man in coastal regions and even in areas located far inland is responsible for generating a huge amount of pollutants that are transported to marine ecosystems directly or by means of coastal watersheds, rivers, and precipitation from air. Therefore, water pollution is a key global problem that has threatened marine organisms, including edible ones, and marine life in general. There are two types of water pollutants, i.e., point source and nonpoint source. The point source type is attributable to harmful contaminants released directly to the aquatic environment while nonpoint source delivers pollutants indirectly to the site of their approach. The former one is a single, well-localized source, e.g., directly discharging sewage or industrial waste to the sea, whereas in the latter, the source of pollution is not well defined. Examples of such nonpoint source are agricultural runoff and windblown debris. Nonpoint sources are considered to be much more difficult to control and regulate as compared to point source pollutants. The following are the classifications of other sources of pollution in coastal waters: • Discharge of sewage and industrial waste • Exploration and exploitation of the seabed • Accidental pollution by oil and other pollutants from the land via air and other routes Among these sources of pollutants dominate those connected with the discharge of municipal sewage and industrial wastes into coastal or estuarine regions, especially in the case of their inadequate treatment to remove persistent and harmful compounds. However, natural (and not anthropogenic) phenomena (e.g., volcanoes, storms, algae blooms, earthquakes, and geysers) could also be responsible 17

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for polluting aquatic systems. Their influence causes crucial changes in the ecological status of aquatic ecosystems. The following are factors that determine the severity of pollutants:[1] • Chemical structure • Concentration • Persistence Independent on their sources in the water, pollutants may be classified as those for which the environment has some or little/no absorptive capacity. They are named “stock pollutants” (e.g., persistent synthetic chemicals, non-biodegradable plastics, and heavy metals).[2] Most marine pollutants have land origin. They are often transported via rivers from agricultural sources and also via atmospheric trajectory. A lot of pollutants may be taken up by various compartments (biotic and abiotic) of aquatic environments; some of them could be biomagnified along the successive members of the food chain. A good example of having such ability to biomagnify is mercury. Such biomagnification could have negative effect on the quality of the water and hence on the health of the plants, animals, and humans whose lives depend on the quality of aquatic environments. It should be emphasized that coastal areas are generally damaged from pollution, resulting in considerable impact on commercial coastal and marine fisheries. The pollution problem is very complex because of its interactions, interconnectedness, and uncertainty.[3–5] Pollutants, independent of their origin (e.g., air, water, land), enter the ocean, whether earlier or later.[3] Spatial distribution patterns of contamination concentrations exhibit a trend of their increase during transition from the south to the northern part of all oceans, i.e., to areas neighboring with both industrial centers and concentration of main pollution sources.[6] The following are considered major pollutants:[6–11] fertilizers, pesticides, and agrochemicals; domestic and municipal wastes and sewage sludge; oil and ship pollution; trace elements; radionuclides; organic compounds; plastics; sediments; eutrophication and algal bloom; biological pollution; noise pollution; and light.

Heavy Metals and Metalloids In contrast to organic pollutants, e.g., Polichlorinated Biphenyls (PCBs), heavy metals occur as natural elements of particular abiotic and biotic components of continental and aquatic ecosystems. They are present at a natural background level in rocks, soils, sediments, water, and biota. Human industrial and agricultural activities result in the elevation of this natural level to sometimes significantly higher values. Typical metal concentrations are generally observed in open waters of marine ecosystems, although these remote regions can be affected by elevated levels of trace elements of anthropogenic or volcanic origin. For instance, the atmosphere affects the oceans and continental matter facilitating metal fluxes between these two compartments. Therefore, the atmosphere is a very important component as it makes it possible to transport metals that are natural in origin into distances far from their sources, e.g., from areas closest to forest fires as well as windblown dust, vegetation, and sea aerosols.[12,13] These sources are responsible for contributing metals to the lower troposphere, and therefore, their transport is associated to local and regional wind patterns, in contrast to specific sources such as volcano eruption, which can be responsible for injecting particulate metals not only into the troposphere but also into the stratosphere. In the latter case, particulate metals can be transported long distances under the appropriate circumstances.[13] Another example of long-distance transport of metals from their sources is dust carried from the Sahara Desert resulting in deposition of Fe, Mn, Al, and trace elements across the Mediterranean Sea, Atlantic Ocean, and Caribbean Sea.[13,14] Therefore, wind-driven dust transports particulate metals far offshore, in contrast to riverine flux carrying greater pole of mineral components from continental material to the coastal waters. These metals are promptly deposited to bottom

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sediments or taken up by biota, especially by phytoplankton in the surface waters, and transferred next to the food chain, recycles or settled to the bottom.[13] The mass of metal of anthropogenic origin emitted to ecosystems is now equal to or greater than the mass introduced to the natural cycle on a global scale.[13] Some metals, e.g., Pb, Hg, and Cd, owing to their great toxicity, pose a high health risk; therefore, great attention has been paid to estimate their inputs to marine ecosystems, particularly to coastal waters. For instance, the largest masses of Pb are emitted to the atmosphere during processing of the metal (smelters) or from combustion-related sources like motor vehicles. Lead from motor vehicle exhaust was identified not only in the atmosphere but also in remote surface waters as well as in remote terrestrial areas. A ban on Pb usage in vehicle fuel has resulted in the effective reduction of metal inputs to the ecosystem since the 1970s. The decline in atmospheric Pb detected over a time scale of 10 years (from 1979 to 1989) because of the reduction of leaded gasoline in western European countries should be reflected in decreasing Pb levels in surficial water and biota. In fact, the temporal negative trend for cod in the Baltic Sea seems to support this argumentation.[15] According to Mason et al.,[16] preindustrial fluxes and reservoirs of Hg pose one-third of its fluxes in the civilization era. It is suggested that modern emission of Hg to the atmosphere increased considerably, even 4 to 5 times, due to the human activities. The extremely elevated levels of Hg are frequently associated with Hg mining. As has been reported, ca. 300 metric tons of dissolved Cd annually enter the oceans from rivers while ca. 400 to 700 metric tons of dissolved and particulate Cd are annually deposited to the oceans from the atmosphere.[17,18] It is estimated that human activities have contributed to increased Cd inputs to the ocean by 60% in the 1980s. It is also found that the higher proportion of land deposition of Cd is associated to the rapid removal of this metal from the atmosphere near inputs of air pollution. A substantial pole of Cd transported by rivers is deposited in estuaries and continental margins of oceans. It is found that increased concentrations of Cd occur locally, reflecting its mosaic contamination, especially near mining and industrial point sources—not managed. A significant fraction of Zn entering the oceans is derived from atmospheric deposition.[18] Soils and sediments are main natural reservoirs of Zn. Zinc, like Cd, is not distributed evenly across the Earth’s surface, since its increased concentrations occur locally in the vicinity of increased inputs, i.e., specific points of source inputs. The concentrations of many trace elements, e.g., Pb, Cd, Hg, Cu, Zn, Se, and As in coastal and estuarine waters, especially in highly industrialized areas, are generally significantly greater than those in open oceanic waters.[13,18–20] The waters of harbors and marinas around the world contain variable concentrations of tributyltin (TBT),[13] but its extremely high levels may be characteristic for marinas in southwest England.[21] Human industrial and agricultural activities affect inputs of several metals to reservoir/reservoirs and hence increase their concentrations, even sometimes very dramatically, above natural background levels. There are numerous examples of worldwide events leading to serious contamination of coastal waters by heavy metals and metalloids.[15] Therefore, relationships between man and ecosystem health have been explored, especially in relation to perturbed ecosystems. This includes the pollution status of coastal regions harmed by some catastrophes, large-scale pollution, environmental accidents and episodes, etc. High risk groups consume extremely high quantities of trace metals present in specific assortment of seafood or offal and it concerns seaside populations. Marine fish and shellfish may be the dominant dietary sources of Hg for local populations.[22–25] A notable example of aquatic pollution by a toxic metal is the Minamata incident, commencing in 1953 and resulting in fish, shellfish, and bird mortalities in waters of the partially landlocked Minamata Bay.[15] Dogs, pigs, and especially cats were also victims of this incident. By the end of 1974, 107 of 798 officially verified patients had died. According to Tomiyasu et al.,[26] the sediments from the Minamata Bay contained levels of Hg that highly exceeded its background level. Among other incidents resulting in the release of Hg compounds to the environment, the most significant ones happened in the 1960s and early 1970s in Sweden, Canada and the United States, northern Iraq, Guatemala, Pakistan, and Ghana.[27–31] MeHg in aquatic ecosystems, especially

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those that bioaccumulated in fish, is a major public health problem all over the world.[32] Its levels in the hair of fishermen represent the critical group for dietary exposure. For instance, the concentrations of Hg (total and MeHg) in the hair of fishermen from Kuwait were 2 times higher than the “normal” level according to the World Health Organization.[33] Biomass burning in tropical forests also seems to have contributed significantly to the Hg input to the atmosphere. Approximately 31% of the Hg concentrations were associated with the vegetation fire component.[34] It is postulated (based on long-range air mass trajectory analyses) that Hg occurs in the Amazon basin over two main routes: to the South Atlantic and to the Tropical Pacific, over the Andes.[34] Global emission flux estimates exhibited that biomass burning could be major contributor of heavy metals and black carbon to the atmosphere.[15] It is estimated that savannah and tropical forest biomass burning could emit huge amounts of Cu, Zn, and black carbon to the atmosphere, corresponding to 2%, 3%, and 12%, respectively, of the global level of these elements.[35] The toxic effects of TBT were first indicated towards the end of the 1970s in Arcachon Bay, France, as the “TBT problem.”[36,37] The release of TBT (from antifouling paints) to the area resulted in shell abnormalities and reduced growth and settlement in oysters, Crassostrea gigas, cultured in the vicinity of marinas. In much polluted water, oyster production was severely affected by the absence of reproduction, resulting in a strong decline in the marketable value of the remaining stock.[37] Imposex, i.e., the development of male sexual characteristics in female marine mesogastropoda and neogastropoda caused by TBT pollution, is a widespread phenomenon concerning several coastal species and, more recently, offshore species as well.[23,38,39] Subsequent regulations in 1990 that prohibit the use of TBT-based antifoulants on vessels less than 25 m in length have been highly effective in reducing TBT levels in coastal waters. However, larger vessels have continued the release of TBT, and major harbors remained pollution hot spots.[40] The Organotin Antifouling Paint Control Act restricted in the United States the use of TBT paints to vessels greater than 25 m in length.[41] The voluntary stoppage of TBT production in January 2001 by major U.S. and European manufacturers resulted in the decline of its presence in marine biota, but TBT paint is still being used in most Asian countries. The International Maritime Organization (IMO) imposed an international ban for the use of organotin compounds in antifouling treatments on ships longer than 25 m. The target is to prohibit their application starting 2003 and to require the removal of TBT from ships’ hulls by January 1, 2008.[41,42] The extensive flooding, especially occurred in river area of former or operating metalliferous mining can be responsible for wide-spreading of heavy metals and metalloids far distance from pollution source. An example of such environmental events is the flooding of the Severn catchment (United Kingdom) in January 1998.[43]

Radionuclides Physicochemical aspects and applications of radioactivity in the environment were extensively presented in a book by Valcovic.[8] There are numerous papers reporting on problems resulting from radionuclide pollution and their sources in different ecosystems.[15,44] One of the first low- level emissions of radioactivity took place in the Hanford reactors (Columbia River, Washington, United States), which released radionuclides (mainly 60Co, 51Cr, and 65Zn) to its environs from 1940 to 1971.[28] The nuclear reactors in Cumbria (northwest England) have also been responsible for discharging quantities of radioisotopes, i.e., 144Ce, 137Cs, 95Nb, 106Ru, and 95Zr, to the marine environment. Although these emissions have been diminished recently, discharges from nuclear power stations such as Sellafield (formerly named Windscale) could still be identified, even in distances far away from their source.[45,46] Significant quantities of artificial radionuclides (137Cs, 134Cs, 90Sr, 99Tc) have been transported to the North Atlantic and Arctic from Sellafield, together with measurable amounts of Pu and Am.[46–48] The nuclear reprocessing plant at La Hague in France emitted 137Cs and 239+240Pu to the environment, although this plant mainly supplies 129I and 125Sb.[28,48,49] Besides the expected emission of radionuclides from nuclear and

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reprocessing facilities, significant quantities of radioisotopes contaminate aquatic and terrestrial environments from either nuclear weapons testing or nuclear reactor accidents.[15] For instance, the thermonuclear detonation that took place in 1954 at Bikini Atoll resulted in the contamination of a large area of the Marshall Islands. A number of atmospheric tests (520 in total) were mostly carried out in the Northern Hemisphere, including eight underwater tests, with a total yield of 542 Mt. Moreover, there have been a total of 1352 underground tests with a total yield of 90 Mt.[50] A number of nuclear incidents were concernedly noted, including those affecting the crew of the Japanese fishing vessel “Fukuru Maru”.[28] Plutonium released from the Kyshtym accident in the Urals has been much probably detected in deep basins of the Arctic Ocean.[51] In 1968, an aircraft from the U.S. Strategic Air Command crashed near the Thule Airbase in NW Greenland, releasing to the marine  environment ca. 1 TBq 239+240Pu.[52] As a consequence, marine sediments as well as benthic organisms, i.e., bivalves, shrimps, and sea stars, have been contaminated by Pu, although their levels rapidly decreased.[53] A number of American and Russian nuclear submarines have been lost in the world’s oceans. For instance, the Soviet Komsomolets submarine sank at a depth of 1700 m at Bear Island in the eastern part of the Norwegian Sea. The estimated radioactivity in the wreck was 2.8 PBq 90Sr and 3 PBq 137Cs.[54] Some nuclear powered satellites can incidentally be sources of radioactivity. They can burn up in the upper atmosphere, resulting in the contamination of the ocean. For instance, such an accident happened in 1964 when a SNAP-9A nuclear power generator containing 0.6 PBq 238Pu aboard a U.S. satellite re-entered the atmosphere in the Southern Hemisphere. The estimated 238Pu/239+240Pu ratio in this region was higher than that in the ocean water from the Northern Hemisphere.[55,56] Sea dumping was carried out since the late 1940s to mid-1960s mainly by the United States in the Atlantic Ocean and Pacific Ocean as well as by the United Kingdom in the Northeast Atlantic Ocean.[56] In 1967, an international operation was initiated by the former European Nuclear Energy Agency that contributed to the deposition of ca. 0.3 PBq solid waste at a depth of 5 km in the eastern Atlantic Ocean. Other international operations were continued until 1982 when ca. 0.7 PBq α activity, 42 PBq β activity, and 15 PBq tritium activity have been dumped in the North Atlantic.[57] It has been assessed that the radiological impact of the NEA (former European Nuclear Energy Agency) dumping activities resulted in some releases of Pu from the dumped waste.[15] This source would be responsible for only a part of the total body burden radioactivity in local benthic organisms, e.g., sea cucumbers; the remainder has been attributed to fallout.[58] According to Consortium for Risk Evaluation with Stakeholder Participation (CRESP) evaluation, the individual dose of a critical group consuming seafood such as molluscs from the Antarctic Ocean was estimated to be 0.1 μSv yr–1, in effect labeling 239Pu and 241Am as critical radionuclides. The indefinite collective dose to the world’s population coming from sea dumping was estimated at 40,000 manSv with predominance of 14C and 239Pu.[56,58] U.S. weapons production facilities account for a large fraction of radiocaesium discharges during the 1950s.[15] A striking incident occurred at Chernobyl in the former USSR where an explosion of a reactor core of the nuclear plant took place in April 1986. The Baltic countries and a large part of central and western Europe have been contaminated principally by 131I, 134Cs, and 137Cs.[28,59] It is found that a significant part of the activity fell over the European marginal seas from which the Baltic Sea was the most affected by contamination.[56,60] It has been mainly responsible for additional inflow of the radioactive contaminants to the Northeast Atlantic Ocean.[56] Due to the Chernobyl accident, significant levels of 137Cs were also found in the Black Sea. The outflow from this Sea has been the major source of additional 137Cs in the Mediterranean Sea.[56] In the summer of 1987, the Chernobyl-derived 137Cs was also detected in surficial waters of the Greenland Sea, Norwegian Sea, and Barents Sea as well as in the west coast of Norway and the Faroe Islands. According to Aarkrog,[56] the total Chernobyl 137Cs input to the world’s oceans was relatively significantly smaller than that estimated for nuclear weapons fallout because of the tropospheric nature of this accident that has contaminated the surrounding European continental areas.[15] After the 2011 Tōhoku earthquake and tsunami, the radiation effects from the Fukushima Daiichi nuclear disaster resulted in the release of radioactive isotopes from the crippled Fukushima Daiichi

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Nuclear Power Plant. The total amount of 131I and 137Cs released into the atmosphere has been estimated to exceed 10% of the emissions from the Chernobyl disaster. Large amounts of radioactive isotopes have also been released into the Pacific Ocean.[61]

Organic Compounds The high lipophilicity of many persistent organic pollutants (POPs) enhances their bioconcentration/ biomagnification, resulting in potential health hazards on predators at higher trophic levels, including humans. These xenobiotics occur widely in coastal waters and oceans from the Arctic to the Antarctic and from intertidal to abyssal. It should be emphasized that most of these compounds exist at a very low concentration level, and hence, their threat to marine biota is still not well recognized. However, it is well known that exposure to extremely low levels of halogenated hydrocarbons, e.g., PCBs, Dichlorodiphenyltrichloroethane (DDT), and TBT, may disrupt the normal metabolism of sex hormones in fish, birds, and marine mammals. Moreover, sublethal effects of these organic chemicals over long-term exposure may result in serious damage to marine populations since some of these POPs may impair reproduction functions of organisms while others may show carcinogenic, mutagenic, or teratogenic activity.[6] Some of the effects of these compounds have been reported by Goldberg.[62] For instance, very low levels of TBT (as endocrine disruptor) cause a significant disruption in sex hormone metabolism, resulting in the malformation of oviducts and suppression of oogenesis in female whelks, e.g., Nucella lapillus.[63] As a consequence, sex imbalance leads to species decline if not species extinction in some field populations.[64] Butyltins may be responsible for mass mortality events of bottlenose dolphins in Florida through suppression of the immune system.[65] Trace environmental levels of other compounds like chlorinated hydrocarbons, organophosphates, and diethylstilbestrol may be responsible for significant endocrine disruption and reproductive failure in different groups of animals, i.e., marine invertebrates, fish, birds, reptiles, and mammals.[6] For instance, high levels of DDT, PCBs, and organochlorines in the Baltic Sea significantly reduced the hatching rates of the fish-eating whitetailed eagle (Haliaeetus albicilla) in the 1960s and the 1970s.[66] Another example of the toxic impact of POPs is organochlorine contamination in different cetacean species dependent upon their diet, sex, age, and behavior. Many of these compounds, as endocrine disruptors, reduce reproduction and/or suppress immune function. DDT and PCBs are known as compounds affecting steroid reproductive hormones and can increase mammalian vulnerability to bacterial and viral diseases. Jepson et al.[67] reported a statistically significant relationship between elevated PCB level and infectious disease mortality of harbor porpoises (Phocoena phocoena). The assessment and monitoring of existing and emerging chemicals in the European marine and coastal environment have been overviewed based on numerous, most recent worldwide references.[5] From this report, the extensive range of chemicals that are capable of disrupting the endocrine systems of animals can be categorized into the following: environmental estrogens (e.g., bisphenol A, methoxychlor, octylphenol, and nonylphenol), environmental anti-estrogens (e.g., dioxin, endosulfan, and tamoxifen), environmental anti-androgens [e.g., dichlorodiphenyldichloroethylene (DDE), procymidone, and vinclozolin], chemicals that reduce steroid hormone levels (e.g., fenarimol and ketoconazole), chemicals that affect reproduction primarily through effects on the central nervous system (e.g., dithiocarbamate pesticides, and methanol), and chemicals with multiple mechanisms of endocrine action (e.g., phthalates and TBT). There is a high level of international concern regarding developmental and reproductive impacts on marine organisms from exposure to endocrine- disrupting chemicals. This is the case for “new” substances such as alkylphenols; there is also renewed interest for some “old” organochlorines such as DDT and its metabolites. Brominated flame retardants (BFRs), particularly the brominated diphenyl ethers (BDEs) and hexabromocyclododecane (HBCD), have been detected in the European marine environment. It has been reported that the input of BDEs into the Baltic Sea through atmospheric deposition now exceeds that of PCBs by almost a factor of 40. BDEs are found in fish from various geographic regions. This resulted from the long-range atmospheric transport and deposition

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of these substances.[5] HBCD was detected in liver and blubber samples from harbor seals and harbor porpoises from the Wadden Sea and the North Sea. It is found that environmental concentrations of these BFRs in Japan and South China increased significantly during the last decades. PBDE levels in marine mammals and sediments from Japan, after showing peak concentrations in the 1990s, appear to have leveled off in recent years. Furthermore, in recent years, HBCD concentrations in marine mammals from Japanese waters appear to exceed those of PBDEs, presumably reflecting the increasing use of HBCDs over PBDEs. Pentabromotoluene (PBT) and Decarbomodiphenyl (DBDPE), for example, have been found in Arctic samples remote from sources of contamination. It is an indication of their potential for long-range atmospheric transport, showing a tendency for accumulation in top predators. Polymeric BFRs may be a source of emerging brominated organic compounds to the environment. Medium- and short-chain chlorinated paraffins (SC- CPs) are ubiquitous in the environment and tend to behave in a similar way to POPs. They have been found in water as well as in fish and marine mammals.[5] Perfluorinated compounds (PFCs), namely, perfluorooctane sulfonate (PFOS), have been detected in marine mammals.[5] They are globally distributed anthropogenic contaminants. PFCs, such as PFOS, have been industrially manufactured for more than 50 years and their production and use have increased considerably since the early 1980s. The main producer of PFOS voluntarily ceased its production in 2002. Furthermore, the large-scale use of PFOS has been restricted. PFOS has been used in many industrial applications such as fire-fighting foams and consumer applications such as surface coatings for carpets, furniture, and paper. PFCs are released into the environment during the production and use of products containing these compounds. About 350 polyfluorinated compounds of different chemical structures are known.[5] The most widely known are PFOS (C8F17SO3) and perfluorooctanoic acid (PFOA; C8F15O2), which are chemically stable and thus may be persistent (substance dependent). PFCs do not accumulate in lipid but instead accumulate in the liver, gallbladder, and blood, where they bind to proteins. PFCs have been detected worldwide, including the Arctic Ocean and Antarctic Ocean, in almost all matrices of the environment. High concentrations of PFCs have been found in marine mammals.[5] A screening project in Greenland and the Faroe Islands indicated high biomagnification of PFCs, with elevated concentrations in polar bear liver. A time trend study (1983–2003) showed increasing concentrations for all PFCs for ringed seals from East Greenland. In the United Kingdom, a study on stranded and by-catch harbor porpoise liver (1992 and 2003) found PFOS at up to 2420 pg kg-1 wet weight. There is a decreasing trend going from south to north.[5] Antifouling paint booster biocides were recently introduced as alternatives to organotin compounds in antifouling products.[5] These replacement products are generally based on copper metal oxides and organic biocides. Commonly used biocides in today’s antifouling paints are as follows: Irgarol 1051, diuron, Sea-Nine 211, dichlofluanid, chlorothalonil, zinc pyrithione, TCMS (2,3,3,6-tetra- chloro-4methylsulfonyl) pyridine, TCMTB [2-(thiocyanomethylthio) benzothiazole], and zineb. It has been reported that the presence of these biocides in coastal environments around the world is a result of their increased use (notably in Australia, the Caribbean, Europe, Japan, Singapore, and the United States). For example, Irgarol 1051, the Irgarol 1051 degradation product GS26575, diuron, and three diuron degradation products [1-(3-chlorophenyl)-3,1-di- methylurea (CPDU), 1-(3,4-dichlorophenyl)3-methylu- rea (DCPMU), and 1-(3,4-dichlorophenyl)urea (DCPU)] were all detected in marine surface waters and some sediments in the United Kingdom. Risk assessments indicate that the predicted levels of chlorothalonil, Sea-Nine 211, and dichlofluanid, in contrast to Irgarol 1051, in marinas represent a risk to marine invertebrates. Finally, non-eroding silicone-based coatings can effectively reduce fouling of ship hulls and are an alternative to biocidal and heavy- metal-based antifouling paints. Although polydimethylsiloxanes (PDMSs) are unable to bioaccumulate in marine organisms and their soluble fractions have low toxicity to marine biota, undissolved silicone oil films or droplets can cause physical– mechanical effects such as trapping and suffocation of organisms.[5] Human and veterinary pharmaceuticals are designed to have a specific mode of action, affecting the activity of, e.g., an enzyme, ion channel, receptor, or transporter protein.[5] Clotrimazole, dextropropoxyphene, erythromycin, ibuprofen, propranolol, tamoxifen, and trimethoprim were detected

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in U.K. coastal waters and in U.K. estuaries. Concentrations of some pharmaceutical compounds are effectively reduced during their passage through a tertiary wastewater treatment works, while others are sufficiently persistent to end up in estuaries and coastal waters.[5] Compared with mammalian and freshwater organisms, there is a lack of experimental data on the impacts of pharmaceuticals in marine and estuarine species. However, there is experimental evidence that selected pharmaceuticals have the potential to cause sublethal effects in a variety of organisms. It has been concluded that antibiotic substances in marine ecosystems can pose a potential threat to bacterial diversity, nutrient recycling, and removal of other chemical pollutants. Although data on the occurrence of pharmaceuticals and antibiotics in the marine environment are becoming more available, the true extent of the potential risks posed by this group of contaminants cannot, at present, be assessed, mainly due the lack of effect data.[5] Several studies showed that among personal care products (PCPs), synthetic musks (nitromusks, polycyclic musks, and macrocyclic musks) are widespread in marine and freshwater environments and bioaccumulate in fish and invertebrates.[5] There were identified products such as benzotriazole organic UV filters, namely, UV-320 [2-(3,5- di-i-butyl-2-hydroxyphenyl)benzotriazole], UV-326 [2-(3- i-butyl-2-hydroxy-5-ethylphenyl)-5-chlorobenzotriazole], UV-327 [2,4-di-t-butyl-6-(5-chloro2H-benzotriazol-2-yl) phenol], and UV-328 [2-(2H-benzotriazol-2yl)-4,6-di-t- pentylphenol]. Their relatively high concentrations were found in marine organisms collected from waters of western Japan. There are indications that marine mammals and seabirds accumulate UV-326, UV-328, and UV-327. Benzotriazole UV filters were also detected in surface sediments from this area. The results suggest a significant bioaccumulation of UV filters through the marine food webs and a strong adsorption to sediments. Although a full risk assessment of some of these has been performed (e.g., musks), for most PCPs, there is little data on their occurrence and their effects in the marine environment.[5]

Biological Pollution Eutrophication and Algal Bloom Nutrient loadings in coastal waters cause direct responses such as changes in chlorophyll, primary production, macro- and microalgal biomass, sedimentation of organic matter, altered nutrient ratios, and harmful algal blooms. The indirect responses of nutrient loadings are responsible for changes in benthos biomass, benthos community structure, benthic macrophytes, habitat quality, water transparency, sediment organic matter, sediment biogeochemistry, dissolved oxygen, mortality of aquatic organisms, food web structure, etc. Moreover, increase in phytoplankton biomass and attributing decrease in transparency and light intensity limit growth of submerged vascular plants.[6,68] Generally speaking, eutrophication leads to major changes in qualitative and quantitative species composition, structure, and function of marine communities over large areas. As for phytoplankton communities, such changes are connected with an increase in biomass and productivity.[69] For instance, a general shift from diatoms to dinoflagellates, as well as dominance of small-size nanoplankton (microflagellates, coccoids), has been reported. Similar trends were observed in the case of zooplankton communities, indicating replacement of herbivorous copepods by small- size zooplankton.[70,71] Some examples of consequences of eutrophication have been reported based on worldwide references.[15] The harmful deoxygenation of water giving rise to fish kills was producing nutrient-derived large mats of macroalgae in the Peel-Harvey Estuary, Western Australia.[72] Similar events took place in the northern Adriatic Sea where diatom blooming in summer resulted in the production of mucilage, affecting tourism in northeastern Italy and reducing fish catch.[28,73,74] Insufficient water exchange and increasing production of organic matter during this century caused depletion of O2 in all deep waters of the Baltic Proper.[15] It resulted in devastating consequences for marine biota, leading to the replacement of O2 by H2S in these bottom waters.[75] Although eutrophication generally leads to an increase in fish

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productivity, it can also cause negative environmental changes in fish populations. Fish such as cod and plaice are threatened by O2 depletion in Baltic deep basins, causing decreasing fish catch in Koge Bay in the Sound.[75] The blooms of blue-green algae as well as Nodularia produce a toxic peptide hepatoxin under particular conditions, which can pass through the food web, affecting top consumers, e.g., man. The toxin is responsible for the degeneration of liver cells, promoting tumors and causing death from hepatic hemorrhage.[75] Paralytic shell poisoning (PSP) and/or ciguatera has/have been identified predominantly in the subtropical and tropical zones such as Australia[76–80] and especially in other Indo-Pacific regions, e.g., India, Thailand, Indonesia, Philippines, and Papua New Guinea.[81,82] Principal toxic dinoflagellate species, i.e., Pyrodinium bahamense var. compressa, killed many fish and shellfish from these regions.[15] The consumption of seafood in the Indo-Pacific area posed considerable public health problems.[28] The significant PSP incidences also took place in temperate zones. For instance, in May 1968, a poisoning episode affected 78 persons inhabiting Britain after consumption of soft tissue of the blue mussel Mytilus edulis.[83] Another dinoflagellate-poisoning event again happened in northeast England in the summer of 1990, possibly attributed to a specific combination of elevated nutrient inputs from rivers and exceptionally warm weather conditions, which could be favorable for algae growing.[28] It has been reported that anthropogenically derived atmospheric N deposition to the North Atlantic Ocean was strictly responsible for harmful algal bloom expansion.[84] This event concerned especially the Eastern Gulf of Mexico, U.S. Atlantic coastal waters, the North Sea, and the Baltic Sea.[84–95] Expanding blooms of the noxious dinoflagellate Alexandrium tamarense have been observed along the Northeast U.S. Atlantic coastline.[84,92] There are numerous examples of specific harmful algal bloom expansions in coastal and off-shore waters in case of significant atmospheric deposition of N, e.g., in the North Sea, Adriatic Sea, Western Mediterranean Sea, and Baltic Sea.[84,96] Great attention has been paid to toxic hypoxia- inducing dinoflagellate blooms in the North Sea and the Western Baltic.[84] In the summer of 1991, a very extensive bloom of Nodularia spumigena in the open Baltic Sea and along the southern and southeastern Swedish coasts was observed. Dogs’ mortalities caused by toxic Nodularia blooms have been observed in Denmark, Gotland, and the Swedish coastal waters.[15] In other Baltic areas, horses, cows, sheep, pigs, cats, birds, and fish also suffered from this event. Nodularia blooms have caused human health problems such as stomach complaints, headaches, eczema, and eye inflammation.[75] In the Skagerrak and Kattegat, harmful algal bloom expansion of toxic algae species such as Prorocentrum, Dinophysis, Dichtyocha, Prymnesium, and Chrysochromulina has taken place.[88] The recent blooms mostly killed pelagic organisms and the phyto- and zoobenthic organisms.

Invasive Species The impacts of introduction and invasion of species throughout the world have recently been identified. There are an increasing number of reports that document this phenomenon taking place in coastal, estuarine, and marine waters.[6] For instance, the Chinese mitten crab (Eriocheir sinensis), as invasive species, now inhabits coastal regions in northwestern Europe, and it has caused damage to flood defense walls by burrowing, affecting local community structure. Worldwide fish species introduction is connected with various consequences.[97] It has been pointed out that many aquaculture species are recently genetically modified. Such modified populations are frequently released and mixed with the natural populations and are breeding with them. It causes biological pollution from a molecular level to community and ecosystem levels. An example of such events is the flooding in Central Europe that caused the release of hybrid and modified fish like sturgeon (Acipenser spp.) from aquaculture installations.[98] The local populations of fish are generally not resistant to the pathogenic organisms carried by the introduced species and vice versa. Therefore, deliberate genetic selection and breeding for a long time may have numerous consequences in the aquaculture unit itself as well as the loss of the natural stock for numerous species in a global scale.[6,98,99]

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Fertilizers and Pesticides Agricultural activity as an important pollution source has contributed to significant enrichment of nutrients (mainly ammonium ion and nitrates) in coastal marine waters. It is found[100] that wastes, manures, and sludges provide soils with significantly more hazardous substances as compared to fertilizers for achieving the equivalent plant nutrient content. The worldwide use of fertilizers, including organic fertilizers like manure, is huge. In the case of intensively monocultivated areas, a relatively small number of pesticides have been widely used in spite of their variety.[6] The large mass of pesticide residues is accumulated in the environment since they are not rapidly degradable. The total global DDT production from the 1940s to 2004 was estimated as ca. 4.5 Mt.[101] Duursma and Marchand[102] estimated the world production of DDT to be ca. 2.8 Mt, of which 25% is assumed to be released to the ocean. According to Shahidul Islam and Tanaka,[6] the total emission of DDT through agricultural applications amounts to 1030 kt between 1947 and 2000. Organochlorine pesticides (OCPs) originating mostly in temperate and warmer areas of the world can be transported to coastal waters and even via atmospheric long-range transport and ocean currents to the Arctic. Owing to their bioaccumulative abilities (as lipophilic compounds) and biomagnification along the sequential trophic levels of the food chain, pesticides are classified as one of the most destructive agents for marine organisms. As a consequence, their very high levels can be observed among top predators, including man. Their toxic effects to marine organisms are often complex because they may be associated with the combination of exposure to pesticides and other POPs with environmental stresses such as eutrophication and pathogens.[6]

Sewage Effluents Sewage effluents contain industrial, municipal, and domestic wastes; animal remains; etc. The huge amounts of these effluents generated in big cities are transported by drainage systems into rivers or other aquatic systems, e.g., coastal waters. It is estimated that the annual production of sewage amounts to ca. 1.8 × 108 m3 for a population of 800,000. This load is equivalent to an annual release of 3.6 × 103 tons of organic matter.[6] Sewages pose significant effects on coastal marine ecosystems because they contain POPs (heavy metals/trace elements, organic pollutants) as well as viral, bacterial, and protozoan pathogens and organic substances subjected to bacterial decay. In case of such bacterial activity, the content of oxygen in water is reduced, resulting in the destruction of proteins and other nitrogenous compounds. Releasing hydrogen sulfide and ammonia exhibits toxic activity to marine biota, even at low levels. As for pathogens, domestic sewage released to coastal waters contains such harmful pathogens as Salmonella spp., Escherichia coli, Streptococcus sp., Staphylococcus aureus, Pseudomonas aeruginosa, the fungi Candida, and viruses such as enterovirus, hepatitis, poliomyelitis, influenza, and herpes.[6] Different bacteria and viruses can be transferred to some representatives of marine fauna, e.g., marine mammals.

Oils The recently observed increase in tanker operations and oil use as well as marine tanker catastrophes has been responsible for the presence of excessively large amounts of oil spillage in coastal and marine ecosystems. It is estimated that ca. 2.7 million tons of oil pollution enter the ocean each year. The tanker accidents between 1967 and 2007 released ca. 4.5 million tons of oil to seawater. Notable examples of ecological catastrophes are the huge spill from a drilling platform in Gulf of Mexico (Mexico) in 1979 and the Deepwater Horizon drilling rig explosion in the Gulf of Mexico (United States) (April 20 to July 15, 2010), resulting in massive amounts of oil in the gulf. Another similar example took place during the Persian Gulf War in 1991, where ca. 2 million tons of oil was spilled, resulting to the death of many species of marine biota.[7,103,104] Therefore, oil pollution poses serious adverse effects on aquatic environment and marine organisms represented different trophic levels from primary producers to the top predators.[6]

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Although aerial and flying birds (e.g., gulls, gannets) are not seriously exposed to oil toxicity, birds that spend most of their time in contact with oil on the water surface (e.g., ducks, auks, divers, penguins) are at greater risk of oil toxicity. According to Smith,[105] the annual release of hydrocarbon can range from 0.6 to 1 million tons. Coastal refineries can be an important source of oil pollution since millions of gallons of crude oil and its fractions are processed and stored there. During their operation, pollutants are continuously released by way of leakages, spills, etc.

Marine Debris and Plastics Marine debris, especially plastics, is one of the most pervasive pollution problems. Nets, food wrappers, bottles, resin pellets, etc., have serious impacts on humans and marine biota. Medical and personal hygiene debris can enter coastal water through direct sewage outflows, posing a serious threat to human health and safety. Contact with water contaminated with these pollutants and pathogens (e.g., E. coli) can result in infectious hepatitis, diarrhea, bacillary dysentery, skin rashes, typhoid, and cholera.[106] There are numerous reviews devoted to an important topic such as pollution by marine debris.[106–110] Entanglement in marine debris such as nets, fishing line, ropes, etc., can hamper an organism’s mobility, prevent it from eating, inflict wounds, and cause suffocation or drowning. It was estimated that 136 marine species have been involved in entanglement incidents, including some species of seabirds, marine mammals, and sea turtles.[111] The decline in the population of the northern sea lion (Eumetopias jubatus), endangered Hawaiian monk seal (Monachus schauinslandi), and northern fur seal has been explained by entanglement of young specimens in lost or discarded nets and packing bands.[112] Abandoned fishing gear, e.g., fishing net, can contribute to catching and killing marine animals. This process called ghost fishing or ghost net can kill a huge number of commercial species.[108] An example of another serious pollution problem is ingestion of debris by marine animals. Plastic pellets and plastic shopping bags can be swallowed and lodged in animals’ throats and digestive tracts, causing some animals to stop eating and slowly starve to death.[106] According to the U.S. Marine Mammal Commission,[111] ingestion incidents concerned 111 species of seabirds, 26 species of marine mammals, and 6 species of turtles. For instance, plastic cups were found in the gut of some species of fish from British coastal waters; the ingested cups were eventually responsible for their deaths.[112] Even Antarctic and sub-Antarctic seabirds, e.g., Wilson’s storm-petrel (Oceanites oceanicus) and white-faced storm-petrel (Pelagodroma marina), are at risk for this ingestion hazard.[112–115] It is reported that the proportion of plastic debris among litter increases with distance from source because it is transported more easily as compared to a denser material like glass or metal and because it lasts longer than other low-density materials (paper). Floating plastic articles (material less dense than water, e.g., polyamide, polyterephthalate, polyvinyl chloride) pose a global problem because they can contaminate even the most remote islands.[107,116] Drift plastics can increase the range of some marine organisms or introduce unwanted and aggressive alien taxa species into an environment. It could be risky to littoral, intertidal zones, and the shoreline.[112,117] There is also potential danger to marine ecosystems from the accumulation of plastic debris (material more dense than water) on the seafloor. Such bottom accumulation of plastic can inhibit the gas exchange between overlying waters and the pore water. This process can result in hypoxia or anoxia in the benthic fauna, altering the makeup of life on the sea bottom.[6] Another threat is connected with potential entanglement and ingestion hazards for pelagic and benthic animals.[62,112,118] Plastic can adsorb and concentrate some pollutants in coastal waters, including PCBs, DDE, nonylphenyl, and phenanthrene. It has been reported that these sorbed POPs could subsequently be released if the plastics are ingested.[109,110] For instance, PCBs in tissues of great shearwaters (Puffinus gravis) were derived from ingested plastic debris.[119]

Noise Pollution In recent years, the marine biota has been affected by noise pollution. Natural sources of underwater noise may be physical and biological in character. Physical sources include wind, waves, rainfall,

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thunder and lighting, earthquake-generated seismic energy, and the movement of ice. Biological sources include marine mammal vocalizations and sounds produced by fish and invertebrates.[120,121] Anthropogenic sound sources can be grouped into six categories, namely, shipping, seismic surveying, sonars, explosions, industrial activity, and miscellaneous.[122] Vessel traffic significantly contributes to underwater noise, mainly at low frequencies. Commercial shipping vessels generate noise mainly in areas confined to ports, harbors, and shipping lanes.[122] In contrast to wide geographic distribution of shipping industry, the oil and gas industry activities have taken place along continental margins in specific worldwide areas. Such resources exploration activities have been typically observed in shallow waters less than 200 m in depth. Other activities, in spite of their geographically widespread range, are also confined to near-shore coastal regions, namely, pile driving, dredging, operation of land- and ocean-based wind power turbines, power plant operations, and typical harbor and shipyard activities.[120] Offshore wind turbines may have significantly contributed noise to the underwater ecosystem bearing in mind that the relatively recent growth in offshore wind development has increased. It has been suggested that marine mammals may be indirectly affected by noise from offshore wind turbines, e.g., prey fish avoiding the sound source as well as the masking of marine mammals’ mating and communication calls. On the other hand, a number of mass stranding of marine mammals, especially whales, found on worldwide beaches may be associated with the use of concurrent military sonar.[120] Another example of noise pollution affecting marine animals is continued exposure to anthropogenic noise pressure in vital sea turtle habitats, resulting in potential impact on its behavior and ecology. Brown shrimp exposed to higher pressure levels of noise in experimental area exhibited increased aggression, higher mortality rates, and significant reduction in their food uptake, growth, and reproduction. Sound exhibits measurable damage to sensory cells in the ears of fish.[123]

Light Pollution A remarkable recent interest concerns the introduction of light to the coastal zone and nearshore environment. It is estimated that at least 3351 cities in the coastal zones all over the world are illuminated. It is expected that artificial light will be continuously intensified not only by population growth but also by dramatically increasing the number of locations of high-intensity artificial light. According to the United Nations World Tourism Organization (UN- WTO), there were ca. 900 million international tourist arrivals all over the world.[9] Tourist visits to beaches cause light pollution along the coastline since tracking the movement of population over time by research using satellite imagery showed that wherever human population density increases, the use of artificial light at night also increases. Living organisms are mostly sensitive to changes in the quality and intensity of natural light in the ecosystem. For instance, for algae and seaweeds, photosynthetic activity is highly dependent on available light, i.e., different cycles in natural light intensity and quality.[9] Light pollution takes place when biota is exposed to artificial light, especially in coastal areas, resulting in damaging effects on marine species in seas. The behavior, reproduction, and survival of marine invertebrates, amphibians, fish, and birds have been influenced by artificial lights. Light pollution disrupts the migration patterns of nocturnal birds and can result in hatchling sea turtles to head inland, away from the sea, which could be eaten by predators or run over by cars.[124] Ecological effects of light pollution concern disruption of predator–prey relationship. For instance, artificial light disturbs natural vertical migrations of zooplankton in the water column in accordance with the day-night cycle when natural light helps to reduce their predation by fish and other animals.[125]

Conclusion The anthropogenic activity of man in coastal regions and even in offshore areas is responsible for emission of a huge amount of pollutants that are transported to marine ecosystems directly or by means of coastal watersheds, rivers, and precipitation from air. A lot of pollutants may be taken up by various compartments, i.e., biotic and abiotic, of aquatic environments and some of them could be biomagnified

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along the successive members of the food chain. Therefore, water pollution could have a negative effect on the quality of the water and hence on the health of the plants, animals, and humans whose lives depend on the quality of aquatic environments. Coastal areas are generally damaged from pollution, resulting in considerable impact on commercial coastal and marine fisheries. There are numerous examples of worldwide events leading to serious contamination of coastal waters by persistent pollutants. Therefore, these areas have been extensively explored, especially in relation to perturbed ecosystems by heavy metals, radionuclides, POPs, oils, etc. Elevated levels of nutrients in coastal waters resulted in eutrophication and proliferation of toxic algal blooms. The recently observed increase in tanker operations and oil use as well as marine tanker catastrophes has been responsible for the presence of excessively large amounts of oil spillage in coastal and marine ecosystems. Marine debris, especially plastics, is one of the most pervasive pollution problems. Marine pollutants are generally present in increased concentrations in the enclosed seas and coastal areas than in the open seawaters. Spatial distribution patterns of contamination concentrations exhibit a trend of their increase during transition from the south to the northern part of all oceans, i.e., in areas neighboring with industrial centers and concentration of main pollution sources.

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3 Groundwater: Mining Pollution

Jeff Skousen and George Vance

Introduction .................................................................................................... 37 Groundwater Resources ................................................................................ 38 Groundwater Contaminants ......................................................................... 39 Groundwater Analysis ................................................................................... 41 Strategies for Remediating Contaminated Groundwaters ........................ 42 References ........................................................................................................ 43

Introduction Mining activities can impact the quantity, quality, and usability of groundwater supplies. Underground mining for coal by longwall or room and pillar mining methods often interrupts and depletes groundwater, and can also alter its quality. Surface mining can enhance the introduction of surface water with dissolved solids into groundwater systems through fractures or other conduits. The type and nature of the mining activity, the disturbed geologic strata, and alteration of surface and subsurface materials will determine how groundwater supplies will be impacted. As waters contact and interact with disturbed geologic materials, constituents such as salts, metals, trace elements, and organic compounds become mobilized [1,2]. The dissolved substances can leach into deep aquifers and cause groundwater quality impacts [3]. In addition to concerns due to naturally occurring contaminants from disturbance activities, mining operations may also contribute to groundwater pollution from leaking underground storage tanks, improper disposal of lubricants and solvents, and contaminant spills. Blasting and hydraulic fracking activities can provide additional connection to surface water inputs, and underground injection of wastes can also occur during these operations [4]. In the United States, the Clean Water Act (CWA) and its subsequent amendments establish the authority for all water pollution control actions at the federal level [5] and regulate discharges into surface streams, wetlands, and oceans. Mining operations must acquire National Pollutant Discharge Elimination System (NPDES) permits for discharges to surface waters. Groundwater quality in the United States is regulated by the Safe Drinking Water Act (SDWA), which was originally enacted in 1974 and amended in 1996. The SDWA was passed to protect drinking water supplies by requiring discharges into groundwaters to meet the use standard or the ambient condition, whichever is of higher quality [6]. This is done by legislating maximum contaminant levels (MCLs) above which waters are considered unsafe for human consumption. The Office of Water within the Environmental Protection Agency provides guidance, specifies scientific methods and data collection requirements, and performs oversight for entities that supply drinking water including groundwater. Examples of some water contaminants with specified MCLs associated with mining activities are listed in Table 1 [7]. Because mining activities can result in poor-quality groundwaters, enforcement of regulations is needed to minimize and/or eliminate potential problems. The Surface Mining Control and Reclamation 37

38

Managing Water Resources and Hydrological Systems TABLE 1 Selected Contaminants in Drinking Waters That May Be Influenced by Mining Activities [7] Contaminant

MCL (mg/L)

MCLG

Inorganics Arsenic Cadmium Chromium Copper Cyanide Fluoride Lead Mercury Nitrate (NO3-N) Selenium Sulfate

0.010 0.005 0.1 LV 0.2 4.0 LV 0.002 10 0.05 500

0 0.005 0.1 1.3 0.2 4.0 0 0.002 10 0.05 500

Radionuclides Radium Uranium

5 pCi/L 30 ug/L

0 0

Organics Benzene Carbon tetrachloride Pentachlorophenol Toluene Xylenes

0.005 0.005 0.001 1 10

0 0 0 1 10

Microbiological Total coliforms Viruses

LV LV

0 0

MCL, Maximum contaminant levels permissible for a contaminant in water that is delivered to any user of a public water system; MCLG, Maximum contaminant level goals of a drinking water contaminant that is protective of adverse human health effects and which allows for an adequate margin of safety; LV, Lowest value that can be achieved using the best available technology.

Act (SMCRA) of 1977 identifies policies and practices for mining and reclamation to minimize water quality impacts [8]. SMCRA requires that specific actions be taken to protect the quantity and quality of both on- and off-site groundwaters. All mines are required to meet either state or federal groundwater guidelines, which are generally related to priority pollutant standards described in the CWA and SDWA.

Groundwater Resources Groundwater resources are the world’s third largest source of water behind oceans (97%) and glaciers (2%), and represent 0.6% of the earth’s water content [9]. Approximately 53% of the US population uses groundwater as a drinking water source, but this percentage increases to almost 97% for rural households. In areas of low rainfall, weathering and translocation of dissolved constituents are relatively slow compared to high rainfall areas. For example, only 12% of precipitation will recharge underground water supplies in a dry coal mining area like Gillette, Wyoming, while almost 47% of precipitation was available for recharge in coal mining areas of Tennessee [10]. Transport of contaminants from surface

Groundwater: Mining Pollution TABLE 2

39

Important Hydrogeological Characteristics of a Site That Determine Groundwater Quantity and Quality Geological

Type of water-bearing unit or aquifer (rock type, overburden). Thickness and areal extent of water-bearing units and aquifers. Type of porosity (primary, such as intergranular pore space, or secondary, such as bedrock discontinuities, e.g., fracture or solution cavities). Presence or absence of impermeable units or confining layers. Depths to water tables; thickness of vadose zone. Permeability and connectivity to other voids or conduits. Hydraulic Hydraulic properties of water-bearing unit or aquifer (hydraulic conductivity, transmissivity, storability, permeability, dispersivity). Pressure conditions (confined, unconfined, leaky confined). Groundwater flow directions (hydraulic gradients, both horizontal and vertical), volumes (specific discharge), rate (average linear velocity). Recharge and discharge areas. Groundwater or surface water interactions; areas of groundwater discharge to surface water or vice versa. Seasonal variations of groundwater conditions. Groundwater Use Existing or potential underground sources of drinking water. Existing or near-site use of groundwater.

and subsurface environments to groundwaters is generally accelerated as the amount of percolating water increases. Infiltrating water moves through the vadose zone (unsaturated region) into groundwater zones (saturated region). The upper boundary of the groundwater system (e.g., water table) fluctuates depending on the amount of water received or removed from the groundwater zone. Groundwater movement is a function of hydraulic gradients and hydraulic conductivities, which represent the combined forces with which water moves as a function of gravitational, osmotic, and pressure forces and the permeability of geologic strata. Groundwater moves faster in coarse-textured materials and where hydraulic gradients are high. Aquifers are groundwater systems that have sufficient porosity and permeability to supply enough water for specific purposes. For an aquifer to be useful, it must be able to store, transmit, and yield sufficient amounts of good-quality water. Important hydrogeological characteristics of a site that determine groundwater quantity and quality are listed in Table 2.

Groundwater Contaminants Several types of substances affect groundwater quality [1,11]. Water contaminants include inorganic, organic, and biological materials. Some have a direct impact on water quality, while others indirectly cause physical, chemical, or biological changes that make the water unsuitable for its designated use. Substances that degrade groundwaters include nutrients, salts, heavy metals, trace elements, and organic chemicals, as well as contaminants such as radionuclides, carcinogens, pathogens, and petroleum wastes (Table 3, [12]). Several types of organic chemicals entering groundwaters are less dense than water and tend to move to and along the surface of the water table. Changes can also occur in groundwaters due to temperature fluctuations and odors. Some groundwaters near coal seams contain natural organic substances (such as dissolved methane gas) and synthetic organic chemicals. Methane gas can be extracted from coal beds where underground and surface mining operations are projected,

40 TABLE 3

Managing Water Resources and Hydrological Systems Different Classes of Groundwater Pollutants and Their Causes [12]

Water Pollutant Class Inorganic chemicals Organic chemicals Infectious agents Radioactive substances

Contributions Toxic metals and acidic substances from mining operations and various industrial wastes Petroleum products, pesticides, and materials from organic wastes industrial operations Bacteria and viruses Waste materials from mining and processing of radioactive substances or from improper disposal of radioactive isotopes

and this extraction can alter methane gas concentrations in groundwaters [10]. Organic contamination may also result from leaking gas tanks, oil spills, or runoff from equipment-serving areas. In these cases, the source of the contamination must be identified and removed. Gasoline, diesel, or oil-soaked areas should be immediately excavated and disposed of by approved methods. The chemistry of groundwaters and potential levels of naturally occurring contaminants are related to (1) groundwater hydrologic conditions, (2) mineralogy of the mined and locally impacted geological materials, (3) mining operations (e.g., extent of disturbed materials and its exposure to atmospheric conditions), and (4) time. Movement of metal contaminants in groundwaters varies depending on the chemical of concern. Solubility considerations include metals such as cobalt, copper, nickel, and zinc being more mobile than silver and lead, and gold and tin being even less mobile [1]. As conditions such as pH, redox, and ionic strength change over time, dissolved constituents in groundwaters may decrease due to adsorption, precipitation, and chemical speciation reactions and transformations. Acid mine drainage (AMD) is the most prevalent groundwater quality concern at inactive and abandoned surface and underground mine sites. If geologic strata containing reduced S minerals (e.g., pyrite (FeS2)) are exposed to weathering conditions, such as when pyritic overburden materials are brought to the surface during mining activities and then reburied, high concentrations of sulfuric acid (H 2SO4) can develop and form acid waters with pH levels below 2 [2]. Neutralization of some or all of the acidity produced during the oxidation of reduced S compounds can occur when carbonate minerals in proximity to the acid-producing materials dissolve [3]. Neutralization can also occur when silicate minerals dissolve, but sometimes high levels of potentially toxic metals such as Al, Cu, Cd, Fe, Mn, Ni, Pb, and Zn may be released. For example, mining of coal in the Toms Run area of northwestern Pennsylvania resulted in groundwater contamination by AMD containing high concentrations of Fe and sulfate (SO4) that leached into the underlying aquifer through joints, fractures, and abandoned oil and gas wells. The Gwennap Mining District in the United Kingdom contained numerous mines that operated over several centuries to extract various mineral resources. One of these mines, the Wheal Jane metal mine in Cornwall, extracted ores that included cassiterite (Sn-containing mineral), chalcopyrite (Cu), pyrite (Fe), wolframite (tungsten, W), arsenopyrite (arsenic, As), in addition to smaller deposits of Ag, galena (Pb), and other minerals. After closure in the early 1990s, extensive voids remaining in the Wheal Jane mine that contained oxidized and weathered minerals were flooded. Initial groundwater quality was poor with a pH of 2.9 and a total metal concentration of 5000 mg/L, which contained elevated levels of Fe, Zn, Cu, and Cd. Water quality worsened with depth, and at 180 m, the groundwater had a pH of 2.5 and a metal concentration of 7000 mg/L. Treatment of discharge waters originating from the mine involves an expensive process that will continue long term to preserve environmental quality in surface and groundwaters in the region. A similar situation occurred when a Zn mine in southwestern France was closed. In this case after flooding, discharge mine waters had a solution pH near neutral, but the water still contained high concentrations of Zn, Cd, Mn, Fe, and SO4. Within the Coeur D’ Alene District of Idaho at the Bunker Hill Superfund site, groundwater samples were found to contain high concentrations of Zn, Pb, and Cd [13]. The contamination originated from the leaching of old mine tailings deposited on a sand and gravel aquifer. When settling ponds were

Groundwater: Mining Pollution

41

constructed to catch the runoff from the tailings, water from the ponds infiltrated into the aquifer and caused an increase in metal concentration in the local groundwater system [14]. Gold mining operations have used cyanide as a leaching agent to solubilize Au from ores, which often contain arsenopyrite (As, Fe, and S) and pyrite [1]. Unfortunately, cyanide, in addition to being toxic on its own, is a powerful nonselective solvent that solubilizes numerous substances that can be environmental contaminants. These ore waste materials are often stored in tailing ponds and, depending on the local geology and climate, the cyanide present in the tailings can exist as free cyanide (CN−, HCN); inorganic compounds containing cyanide (NaCN, HgCN2); metal-cyanide complexes with Cu, Fe, Ni, and Zn; and/or the compound CNS. Because cyanide species are mobile and persistent under certain conditions, a large potential exists for trace element and cyanide migration into groundwaters. For example, a tailings dam failure resulted in cyanide contamination of groundwater at a gold mining operation in British Columbia, Canada [1]. Arsenic and uranium (U) contamination has resulted from extensive mining and smelting of ores containing various metals (Ag, Au, Co, Ni, Pb, and Zn) and/or nonmetals (As, P, and U). Arseniccontaminated groundwaters have been a source of surface recharge and drinking water supplies. At one site, a nearby river had As levels 7 and 13 times greater than the recommended national and local drinking water standards, respectively [1]. Arsenic is known as a carcinogen and has been the contributing cause of death to humans in several parts of the world that rely on As-contaminated drinking water [11]. Waters from dewatering a U mine in New Mexico had elevated levels of U and radium (Ra) activities as well as high concentrations of dissolved Mo and Se, which were detected in stream water 140 km downstream from the mine.

Groundwater Analysis Both the remediation and prevention of groundwater contamination by nutrients, salts, heavy metals, trace elements, organic chemicals (natural and synthetic), pathogens, and other contaminants require the evaluation of the composition and concentration of these constituents either in situ or in groundwater samples [2,10]. Monitoring may require the analysis of physical properties, inorganic and organic chemical compositions, and/or microorganisms according to well-established protocols for sampling, storage, and analysis [15]. For example, if groundwater will be used for human or animal consumption, the most appropriate tests would be nitrate-nitrogen (NO3-N), trace metals, pathogens, and organic chemicals. Several common constituents measured in groundwaters are listed in Table 4. However, other tests can be conducted on waters including tests for hardness, electrical conductivity (EC), chlorine, radioactivity, water toxicity, and odors [16]. Recommendations based on interpretation of the groundwater test results should be related to the ultimate use of the water [2]. The interpretation and recommendation processes may be as simple as determining that a drinking water well exceeds the established MCLs for NO3-N and recommending the

TABLE 4

Groundwater Quality Parameters and Constituents Measured in Some Testing Programs [16]

Physical Conductivity Salinity Sodicity Dissolved solids Temperature Odors

Metals and Trace Elements

Nonmetallic Constituents

Al, Ag, As, Ca, Cd, Cr, Cu, Fe, Mg, Mn, Na, Ni, Pb, Se, Sr, Zn

pH, acidity, alkalinity, dissolved oxygen, carbon dioxide, bicarbonate, B, Cl, CN, F, I, ammonium, nitrite, nitrate, P, Si, sulfate

Organic Chemicals

Microbiological Parameters

Methane Oil and grease Organic acids Volatile acids Organic C Pesticides Phenols Surfactants

Fecal coliforms Bacteria Viruses

42

Managing Water Resources and Hydrological Systems

well should not be used as a drinking water source or that a purification system be installed. However, interpretations of most groundwater analyses can be quite complicated and require additional information for proper interpretation. If a contaminant exceeds an acceptable concentration, all potential sources contributing to the pollution and pathways by which the contaminant moves must be identified. In many cases, multiple groundwater contaminants are present at different concentrations. Because the interpretation of water analyses is a complex process, recommendations should be based on a complete evaluation of the water’s physical, chemical, and biological properties. Integrating water analyses into predictive models that can assess the effects of mining activities on water quality is needed in the long term to determine the most effective means to preserve and restore water quality.

Strategies for Remediating Contaminated Groundwaters Mine sites that have been contaminated generally contain mixtures of inorganic and/or organic constituents, so it is important to understand these multi-component systems in order to develop remediation strategies. Therefore, a proper remediation program must consider identification, assessment, and correction of the problem [17,18]. Identification of a potential problem site requires that the past history of the area and activities that took place are known, or when a water analysis indicates a site has been contaminated. Assessment addresses questions such as (1) what is the problem, (2) where and to what extent is the problem, and (3) who and what is affected by the problem. Afterward, a remediation action plan must be developed that will address the specific problems identified. A remediation action program may require that substrata materials (e.g., backfill) and groundwater be treated. If remedial action is considered necessary, then three general options are available: (1) containment, (2) in situ treatment, or (3) pump-and-treat method (Figure 1). The method(s) used for the containment of contaminants are beneficial for restricting contaminant transport and dispersal. Of the remediation techniques, in situ treatment measures are the most appealing because they generally do less surface damage, require a minimal amount of facilities, reduce the potential for human exposure to contaminants, and when effective, reduce or remove the contaminant so that the groundwater can be utilized again [18]. In situ remediation can be achieved by physical, chemical, and/or biological techniques. Biological in situ techniques used for groundwater bioremediation can either rely on the indigenous (native) microorganisms to degrade organic contaminants or on amending the groundwater environment with specialized microorganisms (bioaugmentation). The pump-and-treat method, however, is one of the more commonly used processes for remediating contaminated groundwaters [17]. With the pump-and-treat methods, the contaminated waters are pumped to the surface where one of the

FIGURE 1

Remedial options to consider if cleanup of contaminated groundwater is deemed necessary.

Groundwater: Mining Pollution

43

many treatment processes can be utilized. A major consideration in the pump-and-treat technology is the placement of wells, which is dependent on the contaminant and site characteristics (see Table 2). Extraction wells are used to pump the contaminated water to the surface where it can be treated and re-injected or discharged. Injection wells can be used to re-inject the treated water, water containing nutrients and other substances that increase the chances for chemical alteration or microbial degradation of the contaminants, or materials for enhanced oil recovery. Treatment techniques can be grouped into three categories, namely, physical, chemical, and biological methods [2,18]. Physical methods include several techniques. Adsorption methods physically adsorb or trap contaminants on various types of resins. Separation treatments include physically separating contaminants by forcing water through semipermeable membranes (e.g., reverse osmosis). Flotation, or density separation, is commonly used to separate low-density organic chemicals from groundwaters. Air and steam stripping can remove volatile organic chemicals. Isolation utilizes barriers placed above, below, or around sites to restrict movement of the contaminant. Containment systems should have a permeability of 10−7 cm/s or less. Chemical methods are also numerous. Chemical treatment involves addition of chemical agent(s) in an injection system to neutralize, immobilize, and/or chemically modify contaminants. Extraction (leaching) of contaminants uses one of the several different aqueous extracting agents such as an acid, base, detergent, or organic solvent miscible in water. Oxidation and reduction of groundwater contaminants are commonly done using air, oxygen, ozone, chlorine, hypochlorite, and hydrogen peroxide. Ionic and nonionic exchange resins can adsorb contaminants, thus reducing their leaching potential. Biological methods for contaminant remediation are less extensive than physical and chemical techniques. Land treatment is an effective method for treating groundwaters by applying the contaminated waters to lands using surface, overland flow, or subsurface irrigation. Activated sludge and aerated surface impoundments are used to precipitate or degrade contaminants present in water and include both aerobic and anaerobic processes. Biodegradation is one of the several biological-mediated processes that transform contaminants, and it utilizes vegetation and microorganisms.

References 1. Ripley, E.A., R.E. Redmann, and A.A. Crowder. 1996. Environmental Effects of Mining. St. Lucie Press, Delray Beach, FL. 356 pp. 2. Pierzynski, G.M., J.T. Sims, and G.F. Vance. 2005. Soils and Environmental Quality. 3rd Edition. CRC Press, Inc., Boca Raton, FL. 584 pp. 3. Prokop, G., P. Younger, and K. Roehl. 2003. Groundwater management in mining areas. Proceedings of the 2nd IMAGE TRAIN Advanced Study Course. Available at http://www.umweltbundesamt.at/ fileadmin/site/publikationen/CP035.pdf. 4. U.S. Environmental Protection Agency. 2016. EPA’s study of hydraulic fracturing for oil and gas and its potential impact on drinking water resources. Available at https://www.epa.gov/hfstudy. 5. U.S. Congress. 1977. Clean Water Act. Public Law 95-217. Available at https://www.epa.gov/ laws-regulations/summary-clean-water-act. 6. U.S. Congress. 1996. Safe Drinking Water Act and Amendments. Public Law 104-182. Available at https://www.epa.gov/laws-regulations/summary-safe-drinking-water-act. 7. U.S. Environmental Protection Agency. 2002. Potential impacts of hydraulic fracturing of coalbed methane wells on underground sources of drinking water. Available at http://permanent.access. gpo.gov/lps21800/www.epa.gov/safewater/uic/cbmstudy.html. 8. U.S. Congress. 1977. Surface Mining Control and Reclamation Act. Public Law 95-87. Available at https://www.osmre.gov/lrg/docs/SMCRA.pdf. 9. National Groundwater Association (NGWA). 2019. Information on earth’s water. Available at https://www.ngwa.org/what-is-groundwater/About-groundwater/information-on-earths-water.

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10. National Academy of Sciences. 1990. Surface Coal Mining Effects on Ground Water Recharge. Committee on Ground Water Recharge in Surface-Mined Areas, Water Science and Technology Board, National Research Council. 170 pp. Available at http://www.nap.edu/catalog. php?record_id=1527. 11. Manahan, S.E. 2017. Environmental Chemistry. 10th Edition. Lewis Publishers, Chelsea, MI. 12. Nielsen, D.M. 2005. Practical Handbook of Groundwater Monitoring. 2nd Edition. Lewis Publishers, Boca Raton, FL. 13. Todd, D.K., D.E.O. McNulty. 1975. Polluted Groundwater. Water Information Center, Inc., Huntington, NY. 14. National Research Council. 2005. Superfund and Mining Megasites: Lessons learned from the Coeur D’ Alene River Basin. The National Academies Press, Washington, DC. https://doi. org/10.17226/11359. 15. Rice, E.W., R.B. Baird, and A.D. Eaton (Eds). 2017. Standard Methods for the Examination of Water and Wastewater. 23rd Edition. American Public Health Association, Washington DC. 16. U.S. Environmental Protection Agency. 2009. Drinking Water Contaminants. Available at https:// www.epa.gov/sdwa/drinking-water-regulations-and-contaminants 17. Interstate Technology & Regulatory Council. 2005. Overview of Groundwater Remediation Technologies for MTBE and TBA. Available at https://www.itrcweb.org/GuidanceDocuments/ MTBE-1.pdf. 18. Hyman, M.H. 1999. Groundwater and Soil Remediation, pp. 684–712. In: Meyers, R.A., (ed.), Encyclopedia of Environmental Pollution and Cleanup. Volume 1. John Wiley & Sons, Inc., New York, NY.

4 Groundwater: Nitrogen Fertilizer Contamination Introduction .................................................................................................... 45 Why N in Groundwater Is a Problem ..........................................................46 Human Health Impacts • Environmental Impacts

Agricultural Practices Contributing to Groundwater NO3 ...................... 47 Row Crops • Grasslands/Turf • Containerized Horticultural Crops

Practices That Can Mitigate NO3 Leaching from Agriculture.................. 50

Lloyd B. Owens and Douglas L. Karlen

Testing, Timing, Rates of Application, and Nitrification Inhibitors • Winter Cover Crops, Diversified Crop Rotations, and Reduced Tillage • Use of Alternate Grassland/Turf Management

Conclusions ..................................................................................................... 52 References ........................................................................................................ 53

Introduction Groundwater is widely used for domestic and public water supplies, particularly in rural areas.[1] When it resurfaces, groundwater also becomes surface water that ultimately flows through streams and lakes to the oceans. Generally, groundwater is quite pure unless it contains natural contaminants such as high iron, sulfur, or possibly an intrusion of saltwater if the aquifer is located close to an ocean or estuary. Such contaminants usually result from the geological formations through which the groundwater source is flowing or residing in. Another groundwater contaminant, nitrate-nitrogen (NO3-N), can originate naturally but is most often associated with areas where human intervention and the management practices being used have significantly increased the amount of NO3-N that is available for leaching as water moves through soil and the underlying geology into groundwater aquifers. Why is NO3-N in groundwater a problem? When water with high nitrate (NO3) concentrations is consumed by humans, it can cause several adverse health problems. One of the most common is known as “blue baby syndrome” or methemoglobinemia, an illness that arises when an infant’s blood is unable to carry enough oxygen to body cells and tissue. Consumption of high-NO3 water has also been associated with increased levels of nitrosamine and some types of cancer. High levels of NO3-N in streams, lakes, and oceans can stimulate excess growth of plants and bacteria, which upon death and decay subsequently deplete much of the oxygen in water. This causes fish kills and “dead zones,” such as the hypoxic area in the Gulf of Mexico. Elevated levels of NO3-N in groundwater have been reported in many parts of the world.[2,3] Although this can occur naturally, leaching of N fertilizer is often a major factor responsible for high NO3-N concentrations in groundwater. This occurs because nitrogen (N) is an essential plant nutrient, and to achieve optimum crop yields, the amount of N that becomes available through natural cycling must be supplemented with additional N from either inorganic (i.e., fertilizer) or organic (i.e., manure or 45

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Managing Water Resources and Hydrological Systems

legume) sources. When the total amount of N provided through natural and supplemental sources exceeds that removed by harvested portions of the crop, excess N can accumulate in the soil profile and become available for leaching into subsurface drainage lines or directly into groundwater aquifers. The challenge of providing adequate plant-available N without creating an excess that is available for leaching has been studied for decades in attempts to determine the economically optimum N level (EON). Such studies have been conducted not only for agronomic crops (e.g., maize, wheat, cotton, and rice) but also for turf grasses and containerized horticultural crops, which usually receive frequent and high N fertilizer rates. Despite these efforts to identify EON rates, there are some situations when there will still be a positive yield response to N fertilizer at rates exceeding the amount of N that the plants will utilize. When this occurs, excess NO3-N accumulates in the soil profile and becomes available for leaching, thus resulting in NO3-N concentrations that exceed the maximum contaminant levels (MCLs) for drinking water (i.e., 10 mg L−1). Many approaches can be used to reduce the NO3 leaching potential. This includes reducing overall N fertilizer inputs; applying N in split applications, coinciding with the time when plants can use the N most efficiently; using slow-release N fertilizers; and growing cover crops to capture part of the residual N accumulating in the soil profile before it can be leached to the groundwater. This entry focuses on groundwater and how it is impacted by NO3-N leaching. Several factors contributing to leaching are addressed and supported by accompanying references, but space does not permit this to be a comprehensive literature review. Our primary emphasis is given to alternate management practices for row crops, grasslands/turf, and containerized horticultural crops that can be used to reduce potential N loss to groundwater resources.

Why N in Groundwater Is a Problem Human Health Impacts Groundwater is a major source of water for human consumption. When contaminated, the N is usually in the form of NO3, which can pose major human health concerns at high levels, especially for infants. The link between high NO3 in polluted water and serious blood changes in infants was first reported in 1945. From 1947 to 1950, 139 cases of methemoglobinemia were reported, including 14 deaths in Minnesota alone. In response to this documented threat to human health, a MCL standard has been set stating that NO3 in excess of 45 mg L−1 (10 mg L−1 NO3-N) is considered hazardous to human health.[4] Even though the number of reported cases of methemoglobinemia has greatly decreased, recent studies indicate possible adverse impacts on human health at NO3 levels below MCL.[5–7] Occurrences of bladder and ovarian cancer[8] and non-Hodgkin’s lymphoma[9] have been linked to people with longterm exposure to public water supplies with NO3-N concentrations of 2 to 4 mg L−1.

Environmental Impacts As part of the hydrologic cycle, a portion of what is classified as groundwater, especially shallow groundwater, resurfaces to feed streams, rivers, reservoirs, and eventually, estuaries and oceans. Nutrients and pollutants in the groundwater are thus transported in surface waters as well. A major challenge associated with groundwater contamination is that its flow is generally very slow. It can take years or decades for water to move through an aquifer from recharge areas to discharge areas and for land management changes/chemical applications to be reflected in “downstream” groundwater quality.[1,10,11] Green et al.[12] concluded that current fertilizer management practices in the United States will likely affect regional groundwater quality for decades to centuries. Over the past 40 years, there has been an eightfold increase in the use of synthetic N fertilizers.[13] This has led to increased NO3-N contamination in both groundwater and surface water bodies.[2,14,15] High NO3-N levels can cause excess plant and bacterial growth in aquatic systems. Subsequently, the

Groundwater: Nitrogen Fertilizer Contamination

47

decay of this organic matter can deplete much of the oxygen in the water, causing fish kills and dead zones to appear. Phosphorus receives much of the attention with regard to eutrophication in fresh waters because it often is the limiting nutrient, but as water systems become more brackish, N often becomes the limiting factor.[16] Well-known examples of this situation include hypoxia in the Gulf of Mexico,[17] Danish coastal waters,[18] the western Indian continental shelf,[19] and the Changjiang Estuary in the East China Sea.[20] Hypoxia occurs when the concentration of dissolved oxygen is less than 2 mg L−1. Nitrogen contributions to the Gulf of Mexico, and other large bodies of water, come from multiple sources, including surface runoff, subsurface drainage, and resurfacing of groundwater. There are several sources of N, e.g., fertilizer, animal manures, septic tanks, atmospheric deposition, land application of treated wastewater and biosolids, and mineralization of organic matter that can contribute to NO3-N contamination. For example, in a 960 km2 basin in Florida, it was estimated that fertilizer applied to cropland, lawns, and pine stands contributed 51% of the annual N load to groundwater in the basin.[21]

Agricultural Practices Contributing to Groundwater NO3 Row Crops High levels of NO3-N in subsurface drainage from row crops, especially corn (Zea mays L.), are well documented. Nitrate-nitrogen concentrations in tile drainage from silt loam soils in Iowa that were fertilized for either continuous corn or corn grown in rotation with soybean [Glycine max (L.) Merr.] were reported to exceed 10 mg L−1 more than two decades ago.[22–24] Similar findings have been reported for clay loam soil in Minnesota,[25–27] silty clay loam in Illinois,[28] silt loam in Indiana,[29] silt loam and silty clay loam in Ohio,[30] and clay over silty clay loam and fine sand over clay in Ontario, Canada.[31] Analyses of subsurface water collected with monolith lysimeters[32,33] and ceramic porous-cup samplers[34,35] are in agreement with these findings. Nitrate-nitrogen concentrations in tile drainage have been studied frequently because of the widespread use of artificial drainage throughout the U.S. Corn and Soybean Belt and the relative ease of collecting water samples. The majority of NO3-N moves in the subsurface water during the late autumn, winter, and early spring recharge period[28,29,32] in this region. There are several factors influencing the amount of N exported via tile drainage, including timing, rate, source, and area receiving N fertilization.[28] Variable weather patterns, especially rainfall amounts, have a major influence on the amount of drainage that occurs. Increasing subsurface drain spacing can decrease NO3-N losses,[25,29] although the NO3-N concentration in the drainage water may change very little.[29] Furthermore, even though the increased spacing may reduce NO3-N losses through the drainage water, it probably increases NO3-N losses in seepage below the drains. Model simulation studies show that reducing N fertilization rates will have much greater impact for reducing NO3-N losses than changing tile drain spacing or depth.[25] There are many management factors contributing to the leaching loss of NO3-N. Nitrogen is an essential plant nutrient that exhibits easily recognizable visual symptoms (e.g., yellow or light green plants, slow growth and development, decreased yield) when plant-available supplies are low. Also, until recently, N fertilizer was relatively inexpensive, and therefore, to reduce the risk of encountering a deficiency, it was often applied at rates in excess of crop need to ensure that inadequate N would not limit crop yield. Another factor is the difficulty in synchronizing N application with crop need because of the narrow window of time that producers may have if weather conditions are not optimum. As a result, it has been shown conclusively that there is a direct relationship between NO3-N loss by leaching and N application rates that exceed crop needs.[34,36,37] Soil NO3-N in excess of current plant needs can result from overapplication of N fertilizer or manure or from not accurately accounting for residual N from previous years, mineralization of organic N, or decomposition of legume crops. The latter is especially true following a dry year because if water limits crop growth, the plant will not utilize as much N as when a “normal” amount of water is available.[27] When this occurs, there is an increased amount of residual N to begin the next cropping season. Nitrogen management is therefore very difficult, and even

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when EON practices are used, it is not unusual to find NO3-N concentrations in subsurface water that exceed the 10 mg L−1 MCL standard for drinking water.[34–36] This leads to the conclusion that optimum corn production will likely result in elevated NO3-N concentrations in groundwater.[36] Stated in another manner, to achieve NO3-N concentrations in groundwater that are less than the MCL, N fertilizer rates will need to be below the level associated with normal crop production recommendations.[37] Nitrogen fertilizer management is also difficult within irrigated production systems, as evidenced by high NO3-N concentrations found in subsurface water beneath sprinkler-irrigated crops in Spain[38] and Nebraska.[39] High NO3-N concentrations also occurred in North Dakota[40] with sprinkler-irrigated corn and intermittent soybean and potato (Solanum tuberosum L.). The different crops did not directly affect NO3-N concentrations. The most important factor leading to increased NO3-N moving through the root zone and into groundwater was the amount of fall residual NO3-N in the soil profile. Floodirrigated wheat in Arizona[41] also produced high NO3-N concentrations in groundwater, and even with best management practices (BMPs) for flood irrigation, NO3-N concentrations in groundwater in excess of the MCL can be expected. Animal manures are generally land applied for both disposal and their fertilizer nutrient value. This is especially true for organic production systems. High rates of manure application can result in high groundwater NO3-N concentrations,[42,43] but even at low rates, manure management can be very difficult. For example, when liquid swine manure was applied at rates to meet the N requirements for either continuous corn or the corn phase of a corn/soybean rotation in Iowa, NO3-N concentrations in tile flow from manure and urea ammonium NO3 treatments consistently exceeded the MCL for drinking water. Furthermore, NO3-N concentrations from the manure treatment exceeded those from the fertilizer treatments.[42] In another Iowa study, swine manure applications to both corn and soybean in a corn/soybean rotation were compared with application to only the corn phase of the rotation.[44] Average flow-weighted NO3-N concentrations and leaching losses to subsurface drainage water were more than 50% greater when manure was applied to both corn and soybean. This response was credited to greater total N application when manure was applied every year compared with every other year. The studies also identified some of the challenges associated with using manure as a nutrient source. These include variability in manure composition and the importance of knowing the composition at the time of application.[43] Because of these factors, knowing the quantity of manure-applied N is much more difficult than for mineral fertilizer. Therefore, manure is often applied at a target rate to meet crop P requirements, with additional N subsequently being applied using another source. Although leaching of NO3-N dominates N loss to surface and groundwater, movement of dissolved organic N (DON) has also been recognized for more than 100 years.[45] Numerous studies have shown that DON leaching from forest ecosystems can be substantial, but DON leaching from agricultural soil has received little attention. Significant amounts of DON leaching were reported for agricultural soils in England,[46] and for cropped soils in Germany, DON leaching accounted for 6–21% of the total N flux.[47] In Ohio, DON accounted for 32–37% of the total N leaching from corn/soybean rotations,[48] with the primary source of DON in agricultural soils being crop residues and soil organic matter.[45] Just as for mineral N dynamics,[46] several factors, including leaching, mineralization, immobilization, and plant uptake, also affect DON leaching.

Grasslands/Turf Because of the animal component, NO3-N leaching in grazed grassland is very complex,[49,50] and even on highly fertilized pastures, much of the N loss has been attributed to excreta.[51,52] Considerable NO3-N leaching can occur from the feces and urine in a management-intensive grazing system[53] and can result in a greater impact on water quality than moderate N fertilization.[54] A New Zealand study showed that when subsurface drainage occurred immediately following intensive grazing, drainage water had higher total N concentrations as a result of total organic N and NH4-N from direct drainage or preferential flow of cattle urine.[55] However, those elevated concentrations were short lived and contributed a relatively

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small amount to the total N loss. Studies in England,[56,57] the Netherlands,[58] and the eastern United States[59,60] have shown that NO3-N concentrations in subsurface water are often greater than the MCL when >100 kg N ha−1 was applied annually to grazed grasslands. In grazed bahiagrass (Paspalum notatum Fluegge) pastures in Florida that were fertilized with 76.5 kg N ha−1 (the recommended rate), NO3-N leaching was not harmful to water quality.[61] Other processes, including accumulation of fertilizer N during drought or release of N from decaying plant material following tillage or chemically killing the sod in preparation for reseeding, may affect N leaching from pastures more than the loss from small areas affected by urine.[51] Once again, management is the key as demonstrated by a field experiment in the Netherlands showing that cattle slurry can replace some or most of the mineral N fertilizer for cut grassland on wet sandy soils without increasing NO3-N leaching if the slurry is applied during the growing season and at rates that do not exceed crop uptake.[62] This was not the case on dry sandy soil, perhaps because denitrification in the wet soil minimized the amount of N available for leaching. Nevertheless, long-term leaching risk may increase as soil organic N increases unless mineral N applications are decreased accordingly.[61] In some nongrazed systems, NO3-N leaching from highly fertilized systems has been reported to be very low, e.g., 29 kg N ha−1 lost from ryegrass (Latium perenne L.) receiving 420 kg N ha−1.[63] Fertilized turf, whether for home lawns or golf courses, is another N source with potential environmental issues. Annual applications up to 244 kg N ha−1 to turfgrass on sandy loam soils in Rhode Island did not appear to pose a threat to drinking water aquifers,[64] but overwatering did increase N loadings to bays and estuaries in coastal areas. A literature review by the Horsley Witten Group[65] concluded that a 20% loading rate to groundwater was an adequate estimate for modeling N leaching in the Cape Cod, Massachusetts, area. The excess N movement was most prevalent with late-summer N applications. Several studies have shown that late-fall applications of N fertilizer to turf result in higher N leaching rates during the late fall and winter.[66–69] Some late-season (October through December) N fertilization recommendations are for the purpose of maintaining/improving turf color, but increased N leaching from such applications indicated they should be managed to achieve acceptable water quality and not to maximize turf color.[67,70] Slow-release formulations of N fertilizer have also shown great reductions in the rate of N leaching.[66,69] The differences were even greater during a high-precipitation year than during a low-or normal-precipitation year,[69] especially with high autumn precipitation.[68] Irrigation management is critical, as evidenced by studies showing that increasing irrigation from 70% to 140% replacement of daily pan evaporation increased N leaching by almost 400%[71] and that flushing porous golf greens with high rates of irrigation can leach elevated rates of N.[72] Grass species is an important factor influencing N leaching, with less occurring beneath grasses with greater aboveground biomass and deeper root systems, e.g., bentgrass species (Agrostis spp.). High-quality grass could also reduce fertilizer N leaching under sand-based putting greens.[73] Turfgrass is most vulnerable for NO3-N leaching during establishment. Even though this period represents a small period of the average turf’s life,[65] the soil disturbance, limited root biomass and N uptake, and tendency of turf managers to overfertilize during this period cause the NO3-N leaching losses to exceed the MCL.[74,75] A mature turf may have actual N requirements below recommended levels. Research in Michigan on a 10-year-old Kentucky bluegrass (Poa pratensis L.) compared annual applications of 245 kg N ha−1 (49 kg N ha−1 per application) and 98 kg N ha−1 (24.5 kg N ha−1 application).[76] With resultant NO3-N leachate concentrations often greater than 20 mg L−1 and below 5 mg L−1, respectively, the conclusion was that applying high rates of water-soluble N to mature turfgrass should be avoided to minimize the potential for NO3-N leaching. Returning clippings to the turfgrass ecosystem reduced the N fertilizer requirement by 25–60%, with the reduction increasing as the time after establishment increased from 10 to more than 50 years.[77] Grass sod has the capacity to use large amounts of N, with 85–90% of fertilizer N being retained in the turf-soil ecosystem.[78] Roots and thatch can represent a large N pool that becomes available for mineralization and subsequent leaching if disturbed.[79] Reseeding and sod establishment within 2 mo of “turfdeath” can stabilize this N pool.[78] High rates of NO3-N leaching can occur at very high N fertilizer

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rates, e.g., 450 kg N ha−1 per year. Although most NO3-N leaching occurred during autumn and winter, it was the accumulation of all N fertilizer applications and not just the autumn application[68] that determined the actual loss. Excess soil NO3 in the fall is the driving force that causes leaching, regardless of N source or time of application. Therefore, high rates of N application to turf should be avoided in the fall because it can result in high NO3-N leaching rates. A significant portion of the N leached can be in the organic form.[71] A survey of several golf courses across the United States indicated that NO3-N concentrations above the MCL occurred in only 4% of the samples,[80] with most of these being attributed to prior agricultural land use. Pollution of groundwater by NO3-N leaching from N-fertilized turf should be minimal with good management, which includes consideration of soil texture, N source, rate and timing, and irrigation/rainfall.[81]

Containerized Horticultural Crops Although the acreage for containerized horticultural crops is small compared with row crops or grasslands, the production intensity is great, and “hot spots” of potential NO3-N leaching can develop. Assuming 80,000 pots ha−1 for a typical foliage plant nursery and using a soluble granular fertilizer, over 650 kg N ha−1 could be lost annually through leaching.[82] During a 10-week greenhouse study of potted flowers, average NO3-N in the leachate ranged from 250 mg NL−1 to 450 mg NL−1.[83] Irrigation and fertilization are critical management components of this intensive industry. Trickle irrigation has been shown to move less water and leach less N than overhead irrigation,[84] but precipitation could nullify this difference.[85] Increasing the irrigation rate, e.g., from 1 to 2 cm day−1, increased the amount of water lost and N leached[86] even though N concentrations were decreased. The use of controlled-release fertilizers (CRFs) is one practice that can significantly reduce leaching losses.[82,87] Even with the use of CRF, NO3-N concentrations can be high in leachate[88]; if the use of CRF is combined with large irrigation volumes, NO3-N can move into the soil profile beneath containers.[89] Water management is crucial, as illustrated by one experiment that compared a low leaching fraction (0.0 to 0.2 of the irrigation water) with a high leaching fraction (0.4 to 0.6). The lower leaching fraction reduced irrigation volume, effluent volume, and NO3-N in the effluent by 44%, 63%, and 66%, respectively.[90] However, these gains in efficiency resulted in a 10% loss in total plant growth (shoots and roots). Thus, establishing an acceptable balance among the level of plant growth, water and nutrient use efficiencies, and the potential environmental impacts becomes a management decision. Although beyond the scope of this entry, vegetable crops also have a high N demand but low apparent N recovery, as illustrated by sweet peppers,[91] which can leave large amounts of N in the soil and residues at harvest.

Practices That Can Mitigate NO3 Leaching from Agriculture Nitrate contamination of groundwater is not caused by any single factor, because non-agricultural and agricultural practices contribute to the problem. Fertilizer management decisions (i.e., rates, formulations, timing, etc.) are an agricultural contributing factor, but so are tillage, crop selection, soil organic matter levels, and drainage.[92] Temperature and precipitation patterns also combine with these factors. Likewise, various strategies and approaches are needed to reduce NO3-N loss to groundwater. This includes using appropriate N fertilizer rates, proper timing, soil testing and plant monitoring, nitrification inhibitors, cover crops, diversified crop rotations, and reduced tillage[92] as well as various combinations of these and other management practices.

Testing, Timing, Rates of Application, and Nitrification Inhibitors Preplant N tests or pre-side-dress N tests[93] can assess soil N levels from cover crops and help provide adequate N credits for legume or fertilizer carryover from prior crops. For example, in Iowa, use of the late-spring NO3 test reduced fertilizer N application[94] and resulted in up to a 30% decrease in NO3-N in

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discharge waters compared with using traditional fall application. Testing plant tissue for N concentration can determine in-season crop N status and be used to guide supplemental N fertilizer applications. Calibrated chlorophyll meter readings have been correlated to plant N status[95–97] and used to increase N fertilizer efficiency and reduce leaching.[96] Splitting N applications also has the potential to reduce the total application rate and to apply the fertilizer N when it is needed by the crop. However, one disadvantage of applying N fertilizer after initial crop growth is that weather conditions may create a very small window of opportunity for side-dress N applications. Nitrification inhibitors used with ammoniacal N sources can slow the rate of oxidation to NO3-N and thus decrease the amount of NO3-N available for leaching,[98] especially with fall N applications. In a Minnesota study, NO3-N losses in subsurface drainage from a corn/soybean rotation were reduced 14% by spring N application compared with late-fall anhydrous ammonia application and 10% by late-fall N application using the nitrification inhibitor nitrapyrin.[99] The use of nitrapyrin with spring-applied N showed no further reduction of N losses. However, even with these improved practices, it may be necessary to reduce the N fertilization rate to below the EON level to achieve NO3-N concentrations in groundwater that are below the MCL[37] for drinking water.

Winter Cover Crops, Diversified Crop Rotations, and Reduced Tillage Winter cover crops have been shown to be an effective strategy for reducing NO3-N leaching.[100–103] A variety of crops including annual grasses, cereals, and legumes have been used as cover crops with varying degrees of success—often depending upon the specific soil and climatic pattern of cropping sequence being used. Winter rye (Secale cereale L.) is the most common cover crop used to reduce NO3-N leaching following corn or corn-soybean rotations in the United States.[100,101,104–106] Winter cover crops have also significantly reduced NO3-N leaching for broccoli (Brassica oleracea L.) crops[107] and potato (S. tuberosum L.)–based rotations[108] in the United States. For potato, spring wheat (Triticum aestivum L.), sugar beet (Beta vulgaris L.), and oat (Avena sativa L.) rotation in the Netherlands, adding cover crops helped to decrease NO3- N concentrations in leachate to near or below the European Union standard (11.3 mg L−1).[109] Care needs to be exercised with long-term cover crops because if they are disturbed, some of the accumulated N may become mineralized and actually increase NO3-N leaching.[110] Even in short-term situations, the efficacy of cover crops to reduce NO3-N leaching is relatively low when considering the entire crop succession, and N saved by the cover crop generally does not increase N utilization by the next crop.[111] Deep-rooted cover crops may help capture N leached to deeper soil layers,[112] and rapid establishment of a deeprooting system is one factor influencing the efficacy of cover crops for reducing NO3-N leaching.[113] Although sometimes cover crops are considered to be a BMP for reducing NO3-N leaching,[114] this is not always true. For example, on the Delmarva Peninsula in the mid-Atlantic United States, a rye winter cover crop following corn did not reduce NO3-N leaching, presumably because the existing corn crop did not allow the rye to be seeded early enough in the autumn. Similarly, owing to the restricted time for cover crop growth in a Wisconsin study, winter rye did not utilize significant amounts of fertilizer N from the previous crop residues or soil.[115] However, in temperate regions with mild winters that favor long growing seasons and winter N mineralization, cover crops have been demonstrated as a valuable tool to reduce NO3-N leaching.[116] Diversifying cropping systems to include perennial crops is another strategy that will probably reduce NO3-N leaching.[117] Compared with annual crops, perennials have an extended period of growth and therefore greater N utilization. However, having an adequate market for satisfactory economic return on such crops can be problematic. Fortunately, the emerging biofuel industry provides a potential market that may make such cropping systems more profitable. Reduced tillage is a BMP for reducing soil erosion, but it is not usually an option for N management. Tillage promotes N mineralization from soil organic matter, which decreases soil quality and makes more N available for oxidation and subsequent leaching. Although water movement through the soil profile tends to be greater with no-till than conventional or limited tillage, NO3-N concentrations are

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usually lower with no-till. Therefore, the amount of NO3-N leaching can be either similar or less with no-till.[118] Overall, differences are small, and it is often concluded that tillage has less impact on NO3-N leaching than factors such as crop rotation.[92,117,119]

Use of Alternate Grassland/Turf Management Demonstrated management options to reduce NO3-N leaching from grasslands include using grass–legume mixtures instead of highly fertilized grass[52,120]; ensuring total N input does not exceed 100 kg ha−1[59,60]; coordinating fertilizer application rate and timing with other N sources (e.g., manure applications); avoiding excessive N application rates[121]; using irrigation, especially during dry periods, to encourage N uptake[52]; and integrating forage cutting and grazing for optimum management throughout the year. Concentrations of NO3-N in groundwater associated with high-fertility, highstocking-density grazing systems can also be reduced by continuous grazing or haying if external N inputs are reduced or eliminated.[122] In areas where NO3-N contamination from turf is a concern, late-summer N fertilizer applications should be reduced, and the amount of irrigation water should be limited.[64] Other management practices for reducing N leaching from turf include using slow-release fertilizers,[66,69] avoiding excess irrigation, and using grass species with greater aboveground biomass and deeper root systems.[73] Because the vulnerability for leaching during turf establishment, irrigation and fertilizer rates should be limited until the turf root system is well developed. Throughout the midwestern United States, excess soil NO3-N in the fall is the driving force responsible for NO3-N leaching. Therefore, using any or all of the management practices mentioned above should help avoid this buildup and thus help reduce NO3-N leaching.

Conclusions High NO3-N concentrations in drinking water can pose human health problems, and when found in surface water bodies, they can create many environmental problems. Several factors have contributed to high groundwater NO3-N concentrations, including the increased use of relatively inexpensive N fertilizer to minimize the yield risk associated with encountering nutrient deficiencies. The unintended consequence is that when N is applied at rates exceeding levels to which crops respond, the potential for excess N to accumulate in the soil and leach is increased. Thus, EON rates may often produce NO3-N levels in groundwater that exceed MCLs. This has been measured with grain crops, grasslands, and turf, which receive high rates of fertilizer N, and nursery container crops. There are several management practices that can help reduce excess N application. These include splitting applications of N so that it is most available when plants need it; using preplant, pre-sidedress, and plant tissue N tests to determine appropriate rates of application; applying slow-release fertilizer sources so that the N is available as plants need it instead of all at once; improving timeliness of irrigation to help plants take up N during drought periods and to avoid excess irrigation, which will reduce leaching; avoiding late-season fertilizer applications, especially common for turf; and using cover crops to capture excess N that remains in the soil profile after growth of the crop so that NO3-N leaching is reduced. Research challenges to improve N management and reduce NO3-N leaching are multidisciplinary. Although precision agriculture is a term generally applied to using soil tests, yield results, and Geographic Information system (GIS) mapping to match fertilizer rates with yield responses in a field, it also has the potential to reduce NO3-N leaching by encouraging the use of variable N rates instead of a constant rate for an entire field. Traditionally, the sole focus for agriculture has been on crop yield, but to simultaneously address groundwater quality, effects of climate change, and other emerging factors, interactions with other crops must also be included. Collectively, these practices can improve overall N use efficiency and have both economic and environmental benefits.

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43. Karlen, D.L.; Cambardella, C.A.; Kanwar, R.S. Challenges of managing liquid swine manure. Appl. Eng. Agric. 2004, 20 (5), 693–699. 44. Bakhsh, A.; Kanwar, R.S.; Baker, J.L.; Sawyer, J.; Malarino, A. Annual swine manure applications to soybean under corn-soybean rotation. Trans. Am. Soc. Agric. Biol. Eng. 2009, 52 (3), 751–757. 45. Van Kessel, C.; Clough, T.; van Groenigen, J.W. Dissolved organic nitrogen: An overlooked pathway of nitrogen loss from agricultural systems? J. Environ. Qual. 2009, 38 (2), 393–401. 46. Murphy, D.V.; Macdonald, A.J.; Stockdale, E.A.; Goulding, K.W.T.; Fortune, S.; Gaunt, J.L.; Poulton, P.R.; Wakefield, J.A.; Webster, C.P.; Wilmer, W.S. Soluble organic nitrogen in agricultural soils. Biol. Fertil. Soils 2000, 30 (5–6), 373–387. 47. Siemens, J.; Kaupenjohann, M. Contribution of dissolved organic nitrogen to N leaching from four German agricultural soils. J. Plant Nutr. Soil Sci. 2002, 165 (6), 675–681. 48. Shuster, W.D.; Shipitalo, M.J.; Subler, S.; Aref, S.; McCoy,E.L. Earthworm additions affect leachate production and nitrogen losses in typical Midwestern agroecosystems. J. Environ. Qual. 2003, 32 (2), 2132–2139. 49. Ball, P.R.; Ryden, J.C. Nitrogen relationships in intensively managed temperate grasslands. Plant Soil 1984, 76 (1–3), 23–33. 50. Ryden, J.E.; Ball, P.R.; Garwood, E.A. Nitrate leaching from grassland. Nature 1984, 311 (5981), 50–53. 51. Cuttle, S.P.; Scurlock, R.V.; Davies, B.M.S. A 6-year comparison of nitrate leaching from grass/clover and N fertilized grass pastures grazed by sheep. J. Agric. Sci., Cambridge 1998, 131 (1), 39–50. 52. Whitehead, D.E. Leaching of nitrogen from soils. In Grassland Nitrogen; CAB International: Wallingford, 1995; 129–151. 53. Stout, W.L.; Fales, S.A.; Muller, L.D.; Schnabel, R.R.; Priddy, W.E.; Elwinger, G.F. Nitrate leaching from cattle urine and feces in northeast USA. Soil Sci. Soc. Am. J. 1997, 61 (6), 1787–1794. 54. Decau, M.L.; Simon, J.C.; Jacquet, A. Nitrate leaching under grassland as affected by mineral nitrogen fertilization and cattle urine. J. Environ. Qual. 2004, 33 (2), 637–644. 55. Houlbrooke, D.; Hanly, J.; Horne, D.; Hedley, M. Nitrogen losses in artificial drainage and surface runoff from pasture following grazing by dairy cattle. In SuperSoil 2004, Proceedings of the 3rd Australian New Zealand Soil Conference, Sydney, Australia, Dec 5–9, 2004; The Regional Institute Ltd., http://www.regional.org.au/au/asssi/, 8 pp. 56. Haigh, R.A.; White, R.E. Nitrate leaching from a small, underdrained, grassland, clay catchment. Soil Use Manage. 1986, 2 (2), 65–70. 57. Roberts, G. Nitrogen inputs and outputs in a small agricultural catchment in the eastern part of the United Kingdom. Soil Use Manage. 1987, 3 (4), 148–154. 58. Steenvoorden, J.H.A.M.; Fonck, H.; Oosterom, H.P. Losses of nitrogen from intensive grassland systems by leaching and surface runoff. In Nitrogen Fluxes in Intensive Grassland Systems; van der Meer, H.G., Ryden, J.E., Ennik, G.E., Eds.; Martinus Nijhoff Publ.: Dordrecht, The Netherlands, 1986; 85–97. 59. Owens, L.B.; Van Keuren, R.W.; Edwards, W.M. Nitrogen loss from a high-fertility, rotational pasture program. J. Environ. Qual. 1983, 12 (3), 346–350. 60. Owens, L.B.; Edwards, W.M.; Van Keuren, R.W. Nitrate levels in shallow groundwater under pastures receiving ammonium nitrate or slow-release nitrogen fertilizer. J. Environ. Qual. 1992, 21 (4), 607–613. 61. Sigua, G.C.; Hubbard, R.K.; Coleman, S.W.; Williams, M. Nitrogen in soils, plants, surface water and shallow groundwater in a bahiagrass pasture of Southern Florida, USA. Nutr. Cycling Agroecosyst. 2010, 86 (2), 175–187. 62. Schroder, J.J.; Assinck, F.B.T.; Uenk, D.; Velthof, G.L. Nitrate leaching from cut grassland as affected by the substitution of slurry with nitrogen mineral fertilizer on two soil types. Grass Forage Sci. 2010, 65 (1), 49–57. 63. Garwood, E.A.; Ryden, J.E. Nitrate loss through leaching and surface runoff from grassland: Effects of water supply, soil type and management. In Nitrogen Fluxes in Intensive Grassland Systems; van der Meer, H.G., Ryden, J.C., Ennik, G.E., Eds.; Martinus Nijhoff Publ.: Dordrecht, Netherlands, 1986; 99–113.

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64. Morton, T.G.; Gold, A.J.; Sullivan, W.M. Influence of overwatering and fertilization on nitrogen losses from home lawns. J. Environ. Qual. 1988, 17 (1), 124–130. 65. Horsley Witten Group. Evaluation of Turfgrass Nitrogen Fertilizer Leaching Rates in Soils on Cape Cod, Massachusetts; MA Department of Environmental Protection: Hyannis, MA, 2009; 33 pp. 66. Guillard, K.; Kopp, K.L. Nitrogen fertilizer form and associated nitrate leaching from cool-season lawn turf. J. Environ. Qual. 2004, 33 (5), 1822–1827. 67. Mangiafico, S.S.; Guillard, K. Fall fertilization timing effects on nitrate leaching and turfgrass color and growth. J. Environ. Qual. 2006, 35 (1), 163–171. 68. Roy, J.W.; Parkin, G.W.; Wagner-Riddle, E. Timing of nitrate leaching from turfgrass after multiple fertilizer applications. Water Qual. Res. J. Canada 2000, 35 (4), 735–752. 69. Petrovic, A.M. Nitrogen source and timing impact on nitrate leaching from turf. In I International Conference on Turfgrass Management and Science for Sports Fields, Athens, Greece, June 2–7, 2003. ISHS Acta Hortic. 661, 427–432. 70. Mangiafico, S.S.; Guillard, K. Nitrate leaching from Kentucky bluegrass soil columns predicted with anion exchange membranes. Soil Sci. Soc. Am. J. 2007, 71 (1), 219–224. 71. Barton, L.; Wan, G.G.Y.; Colmer, T.D. Turfgrass (Cynodon dactylon L.) sod production on sandy soils: II. Effects of irrigation and fertilizer regimes on N leaching. Plant Soil 2006, 284 (1–2), 147–164. 72. Shuman, L.M. Normal and flush irrigation effects on nitrogen leaching from simulated golf greens in the greenhouse. Commun. Soil Sci. Plant Anal. 2006, 37 (3–4), 605–619. 73. Pare, K.; Chantigny, M.H.; Carey, K.; Johnston, W.J.; Dionne, J. Nitrogen uptake and leaching under annual bluegrass ecotypes and bentgrass species: A lysimeter experiment. Crop Sci. 2006, 46 (2), 847–853. 74. Geron, E.A.; Danneberger, T.K.; Traina, S.J.; Logan, T.J.; Street, J.R. The effects of establishment methods and fertilization practices on nitrate leaching from turfgrass. J. Environ. Qual. 1993, 22 (1), 119–125. 75. Easton, A.M.; Petrovic, A.M. Fertilizer source effect on ground and surface water quality in drainage from turfgrass. J. Environ. Qual. 2004, 33 (2), 645–655. 76. Frank, K.W.; O’Reilly, K.M.; Crum, J.R.; Calhoun, R.N. The fate of nitrogen applied to a mature Kentucky bluegrass turf. Crop Sci. 2006, 46 (1), 209–215. 77. Qian, Y.L.; Bandaranayake, W.; Parton, W.J.; Mecham, B.; Harivandi, M.A.; Mosier, A.R. Longterm effects of clipping and nitrogen management in turfgrass on soil organic carbon and nitrogen dynamics: The CENTURY model simulation. J. Environ. Qual. 2003, 32 (5), 1694–1700. 78. Bushoven, J.T.; Jiang, Z.; Ford, H.J.; Sawyer, E.D.; Hull, R.J.; Amador, J.A. Stabilization of soil nitrate by reseeding with ryegrass following sudden turf death. J. Environ. Qual. 2000, 29 (5), 1657–1661. 79. Jiang, Z.; Bushoven, J.T.; Ford, H.J.; Sawyer, E.D.; Amador, J.A.; Hull, R.J. Mobility of soil nitrogen and microbial responses following the sudden death of established turf. J. Environ. Qual. 2000, 29 (5), 1625–1631. 80. Cohen, S.; Svrjcek, A.; Durborow, T.; Barnes, N.L. Water quality impacts by golf courses. J. Environ. Qual. 1999, 28 (3), 798–809. 81. Petrovic, A.M. The fate of nitrogenous fertilizers applied to turfgrass. J. Environ. Qual. 1990, 19 (1), 1–14. 82. Broschat, T.K. Nitrate, phosphorus, and potassium leaching from container-grown plants fertilized by several methods. HortScience 1995, 30 (1), 74–77. 83. McAvoy, RJ. Nitrate nitrogen movement through the soil profile beneath a containerized greenhouse crop irrigated with two leaching fractions and two wetting agent levels. J. Am. Soc. Hort. Sci. 1994, 119 (3), 446–451. 84. Rathier, T.M.; Frink, C.R. Nitrate in runoff water from container grown juniper and Alberta spruce under different irrigation and N fertilization regimes. J. Environ. Hort. 1989, 7 (1), 32–35.

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85. Colangelo, D.J.; Brand, M.H. Nitrate leaching beneath a containerized nursery crop receiving trickle or overhead irrigation. J. Environ. Qual. 2001, 30 (5), 1564–1574. 86. Million, J.; Yeager, T.; Albano, J. Consequences of excessive overhead irrigation on runoff during container production of sweet viburnum. J. Environ. Hort. 2007, 25 (3), 117–125. 87. Million, J.; Yeager, T.; Albano, J. Effects of container spacing practice and fertilizer placement on runoff from overhead-irrigated sweet viburnum. J. Environ. Hort. 2007, 25 (2), 61–72. 88. Newman, J.; Albano, J.; Merhaut, D.; Blythe, E. Nutrient release from controlled-release fertilizers in neutral-pH substrate in an outdoor environment: I. Leachate electrical conductivity, pH, and nitrogen, phosphorous, and potassium concentrations. HortScience 2006, 41 (7), 1674–1682. 89. Colangelo, D.J.; Brand, M.K. Effect of split fertilizer application and irrigation volume on nitratenitrogen concentration in container growing area soil. J. Environ. Hort. 1997, 15 (4), 205–210. 90. Tyler, H.H.; Warren, S.L.; Bilderback, T.E. Reduced leaching fractions improve irrigation use efficiency and nutrient efficacy. J. Environ. Hort. 1996, 14 (4), 199–204. 91. Tei, F.; Benincasa, P.; Guiducci, M. Nitrogen fertilisation of lettuce, processing tomato and sweet pepper: Yield, nitrogen uptake and the risk of nitrate leaching. Welles- bourne, Warwick, United Kingdom. December 1999. Acta Hort. 1999, 506, 61–67. 92. Dinnes, D.L.; Karlen, D.L.; Jaynes, D.B.; Kaspar, T.C.; Hatfield, J.L.; Colvin, T.S.; Cambardella, C.A. Nitrogen management strategies to reduce nitrate leaching in tile-drained Midwestern soils. Agron. J. 2002, 94 (1), 153–171. 93. Magdorff, F.R.; Ross, D.; Amadon, J. A soil test for nitrogen availability to corn. Soil Sci. Soc. Am. J. 1984, 48 (6), 1301–1304. 94. Jaynes, D.B.; Dinnes, D.L.; Meek, D.W.; Karlen, D.L.; Cambardella, C.A.; Colvin, T.S. Using the Late Spring Nitrate Test to reduce nitrate loss within a watershed. J. Environ. Qual. 2004, 33 (2), 669–677. 95. Schepers, J.S.; Francis, D.D.; Vigil, M.; Below, F.E. Comparison of corn leaf nitrogen concentration and chlorophyll meter readings. Commun. Soil Sci. Plant Anal. 1992, 23 (17–20), 2173–2187. 96. Varvel, G.E.; Schepers, J.S.; Francis, D.D. Ability for in-season correction of nitrogen deficiency in corn using chlorophyll meters. Soil Sci. Soc. Am. J. 1997, 61 (4), 1233–1239. 97. Hawkins, J.A.; Sawyer, J.E.; Barker, D.W.; Lundvall, J.P. Using relative chlorophyll meter values to determine nitrogen application rates for corn. Agron. J. 2007, 99 (4), 1034–1040. 98. Owens, L.B. Nitrate leaching losses from monolith lysimeters as influenced by nitrapyrin. J. Environ. Qual. 1987, 16 (1), 34–38. 99. Randall, G.W.; Vetsch, J.A. Nitrate losses in subsurface drainage from a corn-soybean rotation as affected by fall and spring application of nitrogen and nitrapyrin. J. Environ. Qual. 2005, 34 (2), 590–597. 100. McCracken, D.V.; Smith, M.S.; Grove, J.H.; MacKown, E.T.; Blevins, R.L. Nitrate leaching as influenced by cover cropping and nitrogen source. Soil Sci. Soc. Am. J. 1994, 58 (5), 1476–1483. 101. Rasse, D.P.; Ritchie, J.T.; Peterson, W.R.; Wei, J.; Smucker, A.J. Rye cover crop and nitrogen fertilization effects on nitrate leaching in inbred maize fields. J. Environ. Qual. 2000, 29 (1), 298–304. 102. Zhou, X.; MacKenzie, A.F.; Madramootoo, C.A.; Kaluli, J.W.; Smith, D.L. Management practices to conserve soil nitrate in maize production systems. J. Environ. Qual. 1997, 26 (5), 1369–1374. 103. Ball-Coelho, B.R.; Roy, R.C. Overseeding rye into corn reduces NO3 leaching and increases yield. Can. J. Soil Sci. 1997, 77, 443–451. 104. Kaspar, T.C.; Jaynes, D.B.; Parkin, T.B.; Moorman, T.B. Rye cover crop and gamagrass strip effects on NO3 concentration and load in tile drainage. J. Environ. Qual. 2007, 36 (5), 1503–1511. 105. Li, L.; Malone, R.W.; Ma, L.; Kaspar, T.C.; Jaynes, D.B.; Saseendran, S.A.; Thorp, K.R.; Yu, Q.; Ahuja, L.R. Winter cover crop effects on nitrate leaching in subsurface drainage as simulated by RZWQM-DSSAT. Trans. ASABE 2008, 51 (5), 1575–1583. 106. Strock, J.S.; Porter, P.M.; Russelle, M.P. Cover cropping to reduce nitrate loss through subsurface drainage in the northern U.S. corn belt. J. Environ. Qual. 2004, 33 (3), 1010–1016.

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107. Wyland, L.J.; Jackson, L.E.; Chaney, W.E.; Klonsky, K.; Koike, S.T.; Kimple, B. Winter cover crops in a vegetable cropping system on nitrate leaching, soil water, crop yield, pests and management costs. Agric. Ecosyst. Environ. 1996, 59 (1–2), 1–17. 108. Weinert, T.L.; Pan, W.L.; Moneymaker, M.R.; Santo, G.S.; Stevens, R.G. Nitrogen recycling by nonleguminous winter cover crops to reduce leaching in potato rotations. Agron. J. 2002, 94 (2), 365–372. 109. Vos, J.; van der Putten, P.E.L. Nutrient cycling in a cropping system with potato, spring wheat, sugar beet, oats and nitrogen catch crops. II. Effect of catch crops on nitrate leaching in autumn and winter. Nutr. Cycling Agroecosyst. 2004, 70 (1), 23–31. 110. Hansen, E.M.; Djurhuus, J.; Kristensen, K. Nitrate leaching as affected by introduction or discontinuation of cover crop use. J. Environ. Qual. 2000, 29 (4), 1110–1116. 111. Herrera, J.M.; Liedgens, M. Leaching and utilization of nitrogen during a spring wheat catch crop succession. J. Environ. Qual. 2009, 38 (4), 1410–1419. 112. Kristensen, H.L.; Thorup-Kristensen, K. Root growth and nitrate uptake of three different catch crops in deep soil layers. Soil Sci. Soc. Am. J. 2004, 68 (2), 529–537. 113. Herrera, J.M.; Feil, B.; Stamp, P.; Liedgens, M. Root growth and nitrate-nitrogen leaching of catch crops following spring wheat. J. Environ. Qual. 2010, 39 (3), 845–854. 114. Ritter, W.F.; Scarborough, R.W.; Christie, A.E.M. Winter cover crops as a best management practice for reducing nitrogen leaching. J. Contam. Hydrol. 1998, 34 (1), 1–15. 115. Bundy, L.G.; Andraski, T.W. Recovery of fertilizer nitrogen in crop residues and cover crops on an irrigated sandy soil. Soil Sci. Soc. Am. J. 2005, 69 (3), 640–648. 116. Hooker, K.V.; Coxon, C.E.; Hackett, R.; Kirwan, L.E.; O’Keeffe, E.; Richards, K.G. Evaluation of cover crop and reduced cultivation for reducing nitrate leaching in Ireland. J. Environ. Qual. 2008, 37 (1), 138–145. 117. Randall, G.W.; Mulla, D.J. Nitrate nitrogen in surface waters as influenced by climatic conditions and agricultural practices. J. Environ. Qual. 2001, 30 (2), 337–344. 118. Kanwar, R.S.; Baker, J.L.; Baker, D.G. Tillage and split N-fertilization effects on subsurface drainage water quality and crop yields. Trans. ASAE 1988, 31 (2), 453–461. 119. Zhu, Y.; Fox, R.H.; Toth, J.D. Tillage effects on nitrate leaching measured by pan and wick lysimeters. Soil Sci. Soc. Am. J. 2003, 67 (5), 1517–1523. 120. Owens, L.B.; Edwards, W.M.; Van Keuren, R.W. Groundwater nitrate levels under fertilized grass and grass-legume pastures. J. Environ. Qual. 1994, 23 (4), 752–758. 121. Jarvis, S.C. Progress in studies of nitrate leaching from grassland soils. Soil Use Manage. 2000, 16 (s1), 152–156. 122. Owens, L.B.; Bonta, J.V. Reduction of nitrate leaching with haying or grazing and omission of nitrogen fertilizer. J. Environ. Qual. 2004, 33 (4), 1230–1237.

5 Groundwater: Pesticide Contamination Introduction .................................................................................................... 59 Pesticide Use.................................................................................................... 59 Associated Pesticide Behavior in Soils and Water......................................60 Insecticides • Fungicides and Fumigants • Herbicides • Groundwater Contamination

Management of Point Sources of Groundwater Contamination ............. 62 Management of Non-Point Sources of Groundwater Contamination .... 62 Irrigation Management

Roy F. Spalding

Future Research ..............................................................................................63 References ........................................................................................................ 63

Introduction Trace concentrations of most of the commonly used pesticides have been confirmed in groundwaters of the United States. Since groundwater is the source of 53% of the potable water, the more toxic pesticides and their transformation products are a concern from the standpoint of human health. Others are a risk to the environment in areas where contaminated groundwater enters surface water. Through toxicological testing, the USEPA has established Maximum Contaminant Levels (MCLs) or lifetime Health Advisory Levels (HALs) for several pesticides (Table 1). The EPA also has a separate list of unregulated compounds, including newly registered pesticides and their transformation products, such as acetochlor and alachlor ESA, that are presently being evaluated or being considered for toxicological evaluation. Based on the results of the EPAs National Pesticide Assessment,[2] 10.4% of 94,600 community systems contained detectable concentrations of at least one pesticide. Evaluation of these results led to an estimated 0.6% of rural domestic wells containing one or more pesticides above the MCL.

Pesticide Use In the United States about 80% of pesticide usage is in agriculture. The remainder is used by industry, homeowners, and gardeners. About 500 million pounds of herbicide, 180 million pounds of insecticide, and 70 million pounds of fungicide were applied for agricultural purposes in 1993.[3] Several maps of the United States delineate usage patterns of several pesticides.[4] The majority of the triazine and amide herbicides are applied to fields in the north central corn belt states of Michigan, Wisconsin, Minnesota, Nebraska, Iowa, Illinois, Indiana, and Ohio. Commonly used organophosphorus insecticides are more heavily applied to fields in California and along the southeastern seaboard than in the northern corn belt. Carbamate and thiocarbamate pesticides are heavily used in potato growing areas of northern 59

60 TABLE 1

Managing Water Resources and Hydrological Systems U.S. Maximum Contaminant Levels for Drinking Water

Organic Chemical Name MCL (mg/L) Organic Chemical Name MCL (mg/L) Organic Chemical Name MCL (mg/L) 2,4,5-TP (Silvex) 2,4-D Alachlor Aldicarb Aldicarb sulfone Aldicarb sulfoxide Atrazine Carbofuran Carbon tetrachloride

0.05 0.07 0.002 0.007 0.007 0.004 0.003 0.04 0.005

Chlordane Dalapon Dinoseb Diquat Endothall Endrin Ethylene dibromide Glyphosate

0.002 0.2 0.007 0.02 0.1 0.002 0.00005 0.7

Heptachlor Heptachlor Epoxide Lindane Methoxychlor Oxamyl (Vydate) Picloram Simazine Toxaphene

0.0004 0.0002 0.0002 0.04 0.2 0.5 0.004 0.003

Source: U.S. Environmental Protection Agency.[1]

Maine, Idaho, the Delmarva Peninsula, and vegetable fields of California and the southeastern coastal states. Fungicide use is concentrated in high humidity and irrigated areas of the coastal states and to some extent along the Great Lakes and Mississippi River Valley. The fumigants carbon tetrachloride and ethylene dibromide (EDB) were used heavily in the past at grain storage elevators throughout the Midwest and elsewhere in the United States.

Associated Pesticide Behavior in Soils and Water Although pesticide use is a dominant factor in groundwater contamination, leaching variability among pesticides exhibiting similar behaviors is striking and explains why several heavily used pesticides seldom if ever are detected in groundwater. In general, pesticides within a class have similar chemical characteristics upon which soil leaching predictions can be made based on persistence, solubility, and mobility. Pesticide class relationships with soils and water transport described in the following text are detailed in Weber.[5] Individual frequencies of groundwater pesticide detection, in parenthesis next to commonly used products, are calculated from the Pesticide Groundwater Data Base (PGWDB)[4] and the National Water Quality Assessment (NAWQA) database.[6] High frequencies of detection identify those pesticides with a disposition to leach.

Insecticides Chlorinated hydrocarbons are one of the oldest chemical classes of insecticides. Some of the bestknown compounds include aldrin, dieldrin, DDE, DDT, endrin, and toxaphene. Although banned since the 1960s, their extremely persistent nature precludes their detection in very trace quantities in groundwater of the upper Midwest. On the other hand, heavily used organophosphates like malathion, methylparathion, disulfoton, and others have been extensively surveyed during several groundwater monitoring studies and have not been detected. The organophosphate insecticides, parathion (not reported (NR), Ecological Footprint) may use their available biocapacity to satisfy their own domestic consumptions or export ecological resources to other nations. This is generally the case for low-income countries that use only a fraction of their locally available resources, like some African and Latin America countries.[13,15,16] Conversely, countries with a biocapacity deficit (where Ecological Footprint > biocapacity) must rely on biocapacity from outside their own borders or draw down their own natural budget. Many high-income countries (like the United States, Canada, and some western European countries such as Italy, the United Kingdom, and France) have footprints several times larger than their domestic biocapacity.[13,15,16] Such a biocapacity deficit is becoming an increasing economic risk for countries, particularly in a world of growing global overshoot. It highlights a country’s dependence on additional external goods and ecological services, which are provided through one or more of the three following mechanisms:[10,13] 1) the biocapacity trade deficit, which consists of net import of biocapacity from other regions of the world; 2) the biocapacity deficit due to depletion, due to an overuse of local resources; and 3) the demand on biocapacity due to occupation of global commons, such as emissions of greenhouse gases into the global atmosphere (rather than domestic absorption) or fishing in international waters.[17] The analysis of biocapacity and Ecological Footprint trends reveals how human consumption is changing over time. At the global level, the latest data released[13,16] show that humanity is currently operating in a state of overshoot: in other words, demand for natural resources exceeds the regenerative capacity of existing natural capital by at least 50% according to calculations of Global Footprint Network.[15] Furthermore, the gap between Ecological Footprint and biocapacity globally has been continuously increasing since the mid-1970s (Figure 1).

FIGURE 1

Humanity’s Ecological Footprint, 1961–2007.

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Managing Human and Social Systems

In 2007, humanity’s total Ecological Footprint worldwide was 18 billion gha; with world population at 6.7 billion people, the average person’s footprint was 2.7 gha.[13,16] However, there was only 11.9 billion gha of biocapacity available that year, equivalent to 1.8 gha per person. This overshoot of approximately 50% means that in 2007, humanity used the equivalent of 1.5 earths to support its consumption. In other words, it would have taken the earth approximately a year and a half to regenerate the resources used by humanity in that year. The largest Ecological Footprint component was the carbon footprint. This has increased by 35% since 1961 and currently accounts for more than half of the global Ecological Footprint.[13,16] Even if the earth has a high resilience, prolonged biocapacity deficit is not possible since vital ecosystems and nonrenewable stocks would be depleted due to insuperable ecological and thermodynamic constraints. Also, it is not obvious that high-input agriculture can maintain its yields in the long run, particularly in the face of soil loss and potential phosphate limitations. It has become an urgent task to reduce our consumption levels back within the limits of our ecological budget.[14] The growing global trends, however, hide significant regional variation (Figure 2). Both demand on and supply of biocapacity are unevenly distributed across the world. Ecological Footprint and biocapacity values can therefore be used to develop new criteria for distinguishing among world nations. For instance, an alternative approach can be used to look at countries, based on their “biocapacity balances” (Figure 3), which helps to identify where resources are located and who uses what and to what extent. While in 1961, approximately 80% of the world population was living in countries characterized by a biocapacity remainder, in 2007, most of the world population was living in countries running a biocapacity deficit situation. The total Ecological Footprint demanded by a country is strongly related to GDP[18,19] and changes accordingly, in both its extent and its composition, among high-, middle-, and low-income countries (Figure 4).[13] Generally, high-income countries have per-capita Ecological Footprint values nearly 3 times higher than the world average, the majority of which is from the carbon footprint (approximately 65% of the total value). Conversely, middle- and low-income countries have average Ecological Footprint values that are equal to and lower than the world average, respectively. These countries are frequently characterized by transition economies, in which the carbon footprint component, although increased over the last decades, still constitutes less than 50% of the overall demand (Figure 5).

FIGURE 2

Ecological footprint (EF) and biocapacity (BC) by income level and country. The unequal distribu-

tion of human demand on bioproductive lands was also investigated by White.[20]

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FIGURE 3 Biocapacity reminder/deficit status for world countries in 1961 and 2007. Green nations represent countries where the local biocapacity is greater than their residents’ footprint (biocapacity reminder countries); red represents countries where the footprint is greater than local biocapacity (biocapacity deficit countries).

Ecological Footprint of a Product The Ecological Footprint of a final or intermediate product is defined as the total amount of resources and waste assimilation capacity required in each of the phases required to produce, use, and/or dispose of that product.[8] The lifecycle boundaries can be flexible and changed according to the aim and scope of the analysis i.e., from cradle to gate (production to distribution) or cradle to grave (production to destruction). The Ecological Footprint is evaluated by considering all the direct and indirect inputs that are associated with the analyzed system for its entire life cycle. Each of these inputs is converted in terms

Managing Human and Social Systems

308

FIGURE 4 Variations on Ecological Footprint (EF), biocapacity (BC), and population for the world and lowincome, middle-income, and high-income countries, indexed to 1961.

of the global hectares needed to support their production The EFp is expressed in units of global hectare years (gha yr), not just global hectares.[8] As the Ecological Footprint is strictly related to the production system, the way a product is produced should be clear and identifiable. The functional unit for the analyzed system should be also defined as well as the temporal and spatial boundaries. There are two widely used approaches for calculating the Ecological Footprint of a product, both standards compliant: process-based life-cycle assessment (P-LCA) and environmentally extended inputoutput life-cycle assessment.[8] Process-based life-cycle assessment has the advantage of a large amount of detail, as individual product types and even brands can be analyzed, with the general disadvantage of lacking complete upstream coverage of the production chain (e.g., truncation error). Extended inputoutput life-cycle assessment has the advantage of full upstream coverage but the disadvantage of generality, as input-output tables typically do not disaggregate down to the level of individual product types (e.g., homogeneity assumption). Following the P-LCA, Ecological Footprint of product (EFP) is given by the sum of the footprints of each input consumed and disposed within the life cycle of the production process as reported below. 6

EFP =

n

∑∑ YT j =1 i =1

i

N ,i

* YFj * EQFj

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FIGURE 5 Per-capita Ecological Footprint for the world and high-, middle-, and low-income countries by land type, 1961–2007.

6

=

n

∑∑ YT j =1 i =1

i

W ,i

* EQFj

(5)

where the variables are as follows: • i refers to the n-input needed • j refers to the six different land-use types (cropland, grazing land, fishing grounds, forest area, built-up land, and carbon uptake land) • YFj is the YF of the jth land type • EQFj is the EQF of the jth land type

Toward a Multi-Indicator Approach Building on the premise that no single indicator is able to provide a full sustainability diagnosis and indicators should rather be used and interpreted jointly (i.e., the joint use of more than one indicator), this section reports some of the most interesting applications. For instance, the HDI[21] can be used together with the Ecological Footprint to provide important  insights on whether a high level of consumption is necessary for a high level of human development.[15] The HDI is a composite indicator used to rank countries by level of “human development” and then of well-being. It is a comparative measure of life expectancy, literacy, education, and standards of living for countries worldwide. The relationship between Ecological Footprint and HDI has two different categories (Figure 6).

310

FIGURE 6

Managing Human and Social Systems

Human development index vs. Ecological Footprint, 2007.

While countries with a low level of development (HDI 4 gha per capita) report small variations in life satisfaction (Figure 8).

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FIGURE 7 The comparison among temporal trends of GDP vs. ISEW (left scale) and Ecological Footprint (EF) vs. biocapacity (BC) (right scale) for Sweden, United States, and Germany.

In general, people with high consumption and income levels are more satisfied with life than people with lower consumption and income levels despite the differing levels and patterns of consumption that are necessary to obtain the same level of satisfaction. On the contrary, people with low consumption and income levels are less satisfied, but the same unit of consumption produces different perceptions of satisfaction (life satisfaction values are more unpaired).

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Life satisfaction (LS) vs. Ecological Footprint (EF) for 130 world countries.

Biocapacity and Ecological Footprint temporal trends have recently been used to develop alternative interpretations of the geopolitical context of nations around the world.[25,26] Based on their development paths, Ecological Footprint and biocapacity trends for nations around the world can be grouped into four main dynamic typologies: parallel, scissor, wedge, and descent.[25] Each typology corresponds to a particular environmental situation, the implications of which could have extreme relevance for environmental management, economic, and social prosperity as well as the development of sustainability policies. In particular, the role of biocapacity is highlighted in maintaining healthy economies, in offering an acceptable quality of life, and as an essential asset to ensure national competitiveness.[27] Finally, under the recent EU-funded One Planet Economy Network Europe (OPEN:EU) project, three indicators have been identified as useful and complementary in assessing environmental issues— ecological, carbon, and water footprints—and therefore grouped together to form a suite of indicators called “footprint family.”[7] Although not yet comprehensive, this suite provides a quantifiable platform for discussions regarding the limits to biotic resource and freshwater consumption and greenhouse gas emissions, as well as how to address the sustainability of natural capital use across the globe, thus enabling decision makers to more easily understand the environmental consequences of economic activities.

New Insight in Footprint Theory: Toward a ThreeDimensional Ecological Footprint Geography The presence of global overshoot proves that the current human economy partially relies on natural capital depletion rather than just on sustainable flow consumption.[4] Considering natural capital and its limits, differentiating between these two components is fundamental for environmental planning and management. Recently, a variant of the classical Ecological Footprint model has been proposed, where a distinction between depletion of natural capital stocks and use of natural capital flows is operated.[28] The Ecological Footprint was redesigned as a three-dimensional model (3DEF) with two relevant components, called size (EFsize) and depth (EFdepth), related to the two different uses of natural capital. The EFsize regards the appropriation of the so-called “income” of natural capital, i.e., the consumption of flows of resource yearly regenerated by natural cycles. It is the spatial component expressed in global hectares and plotted in the (x, y) plane. By definition, its value ranges from zero to biocapacity. On the other side, the EFdepth regards the depletion of natural capital, which is the use of stocks of resources requiring a regeneration time longer than the flows. It is the intensity component plotted on

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the z-axis. In particular, the EFdepth component arises when overshoot is present and expresses the number of years it would theoretically take to regenerate the natural capital used in one year. By definition, its value ranges from 1 (the reference value called “natural depth”) upward (additional depth). It should be remarked that the two approaches, classical and three-dimensional, are simply two different ways of representing the same footprint values. The 3DEF originates from the fact that flow and stock are technically incommensurable and cannot be summed up because the former is consumed each year and regenerated the following year, whereas the latter represents the irreversible erosion of natural stocks that add up from year to year into an accumulated “environmental debt.” Footprint size and depth have been characterized by opposite trends:[17] 1) EFsize grew continuously until mid-1970s, when it reached the asymptote (i.e., the earth biocapacity), and has remained constant ever since; and 2) EFdepth has remained equal to the natural depth until the appearance of overshoot, and it has subsequently been growing. Recently, the 3DEF model has been theoretically applied to national case studies with the aim of enhancing the significance and potential usefulness of the Ecological Footprint in tracking relevant issues in the sustainability debate, such as the differentiation between resource stocks and flows.[17] Moreover, EFsize can be used as a proxy to highlight the existing intragenerational (in)equity in the appropriation of resources and ecological services by the residents of different nations. At the same time, EFdepth enables the relationships between current and future generations to be examined. Although several questions remain to be addressed, the implementation of both a multilateral trade framework and the 3DEF model in the National Footprint Accounts could form the basis for a new Ecological Footprint-based geography able to differentiate pressures on flows and stocks and identify the spatial/geographical location of such pressures.[4]

Applications of Ecological Footprint Territorial Systems Examples of national Ecological Footprint studies can be found in the literature,[29,42] although the most comprehensive set of national Ecological Footprint assessments is represented by Global Footprint Network’s National Footprint Accounts.[16] As of today, more than 35 nations have engaged with the Global Footprint Network directly, 17 nations have completed reviews of the Ecological Footprint, and a few have formally adopted it. Wales has adopted the Ecological Footprint as its headline indicator for sustainability. The Swiss government has incorporated the footprint into its national sustainable development plan. Japan includes the footprint as a measure in its environmental plan, the United Arab Emirates is using the Ecological Footprint as a tool to recommend and assist in the development of longterm science-based policies,[43] and Ecuador has set official footprint reduction targets in its 2009–2013 National Development Plan. There are several other countries that are currently collaborating with Global Footprint Network.[44] Among nongovernmental organizations, WWF (World Wide Fund for Nature) International, one of the world’s most influential conservation organizations, uses the Ecological Footprint in its communication and policy work for advancing conservation and sustainability. WWF has recently established a target of bringing humanity out of overshoot by 2050 and is actively pursuing this goal through its One Planet programs. Numerous applications have also been performed at various subnational scales.[45–53] Currently, there are two suggested methodologies for subnational Ecological Footprint evaluations: component (or bottom-up) and compound (or top-down) approaches. The component method starts from specific individual consumption and waste production data to then calculate the total Ecological Footprint. While the method is detailed and flexible, several problems, including double counting, the lack of detailed data, and specific conversion factors, make this approach less acceptable. The compound approach evaluates

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the subnational Ecological Footprint by scaling the national Ecological Footprint value according to differences in consumption and life style. This is the widest used and more complete method.[54,8]

Products and Services Despite its diffusion and popularity, product Ecological Footprint applications are still scarce, though a few interesting case studies exist. Studies on cultivation of tomatoes,[55] conventional vs. organic wine farming,[56] nectarine production,[57] shrimp and tilapia aquaculture,[58] marine aquaculture of reef fish,[59] and fisheries products[60] have been performed to highlight appropriation of natural capital, efficiency of natural resource use, and environmental pressure. Evaluations of the environmental impact of farms[61] and dairy production[62] as well as assessment of economic and ecological carrying capacity of crops[63] have been proposed via the combined use of the Ecological Footprint with other methods, such as life-cycle assessment, emergy analysis and economic cost, and return estimation. In the context of product, Ecological Footprint has been used as a basis for the elaboration of the model of the double pyramid, which flanked the food pyramid with the environmental pyramid.[64] This model highlights that, in general, modern lifestyles produce a growing impact on the planet; it tries to promote eco-sustainable life and eating styles. In the context of industrial processes, the potential of using EFA as an environmental indicator for the textile sector has been considered, although the contribution of wastes other than carbon dioxide should be included in the footprint methodology.[65] The EFA has also been applied to the tourism sector. Starting from some pioneering studies,[66,67] the EFA has received attention as a key environmental indicator of sustainable tourism,[68] and several studies have been published at different scales.[69–72] Over time, several tools and software have been developed for Ecological Footprint assessment for both territorial systems and products, although only a few are standard compliant. Most of them are freely downloadable from Web sites such as the footprint calculator (available at http://www.footprintnetwork.org/en/index.php/GFN/page/personal_footprint/). The REAP (Resource and Energy Analysis Programme) software (available at http://www.sei.se/reap) is a scenario based, integrated resource– environment modeling tool developed by the Stockholm Environment Institute to help local authorities in the U.K. make decisions about how to reduce their Ecological Footprint. The Footprinter software (available at http://www.foot-printer.com), developed by Best Foot Forward, is comprehensive and powerful analytical software for carbon and Ecological Footprint assessment for products and organizations. It is based on the use of EcoIndex database.

Role of Business The Ecological Footprint has been used to evaluate the environmental pressure of production processes; this type of investigation is becoming increasingly important to integrate sustainability issues (as natural capital consumption) into industrial and business decision-making processes. In a recent World Business Council for Sustainable Development project (WBCSD), several companies and industries came together to assess their role in helping to shape a future sustainable society for mankind.[73] With the help of Global Footprint Network, the consequences, in Ecological Footprint and biocapacity terms, of the hypothetical scenarios (up to the year 2050) envisioned by WBCSD were estimated. Results from this study showed that humanity will likely require the equivalent of 2.3 planets’ worth of resources upon following a business-as-usual scenario or, conversely, 1.1 planets’ worth of resources by implementing all the envisioned actions and activities, such as 50% reduction in CO2 emissions compared with 2005 levels; enhanced forest productivity using better management techniques and an extension of their acreage between 2030 and 2050; increased crop productivity (+2% each year over past trends) due to technological advances and the diffusion of best practices; and changes to the average global nutritional regime, in terms of both diet and calorie content.[74]

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Weakness and Limitations of EFA Despite its popularity, EFA, like most indicators, is not exempt from criticisms regarding philosophical as well as methodological issues. In this section, a brief summary of the main weaknesses and limitations is offered. For further details on this topic see, for example, Best et al.,[75] Fiala,[76] Kitzes et al.,[77] and van der Bergh and Verbruggen.[78] The first criticism is that the EFA cannot be fully defined as a measure of sustainability. The EFA research question is limited to identify the extent to which humanity is consuming bioproductive land compared with the available biocapacity.[21] We believe consuming resources within the capacity of the planet is a first necessary although not sufficient criterion for sustainability; as such, in order to depict a comprehensive picture of the system analyzed, it is strongly recommended to combine the Ecological Footprint with other complementary indicators (environmental as well as social and economic). The use of a spatial unit makes some impacts difficult to determine.[75] Ecological Footprint is not able to directly account for all resources that cannot be referred to in spatial terms. This is the case for the depletion of nonrenewable deposits, such as metals, minerals, or fossil fuel reserves. For processes of extraction and refining, only the CO2 emissions related to these processes are accounted for. The use of fossil fuels is evaluated in an indirect way, considering the amount of forestland required for the absorption of the CO2 that is emitted. To date, carbon dioxide is the only greenhouse gas accounted for, and its associated footprint relies on the assumption that all emissions are absorbed only by forests and the oceans, neglecting carbon uptake by other biomes. Other missing elements in EFA are freshwater consumption and soil erosion, even if the latter could be accounted for, at least theoretically. A possible way to include the overexploitation phenomena possible in agricultural land as well as in other land types into the classical Ecological Footprint framework was proposed by Bastianoni et al.[79] As EFA is unable to show unsustainable practices and their consequences, when agricultural EFA is performed, there can be some misunderstanding and misinterpretation of the results. Sometimes, it seems that EFA encourages more intensive farming, as this increases agricultural intensities, resulting in a higher biocapacity. The Ecological Footprint shows pressures that could lead to degradation of natural capital (e.g., reduced quality of land or reduced biodiversity) but does not predict this degradation. Furthermore, multifunctional land-use patterns are not considered, in order to prevent double counting.[79] Each hectare is counted only once, even though it might provide multiple services. Counting them multiple times would produce an overestimation on Ecological Footprint. Finally, EFA also is not able to capture the impacts due to the release of long-life toxic materials (e.g., pollution in terms of waste generation, toxicity, eutrophication, etc.), for which no regenerative capacity exists.

Conclusion This entry offers a comprehensive insight to the EFA. Based on its simple logic and unit of measure, the Ecological Footprint has become a very popular sustainability indicator preserving scientific rigor on an ecological and thermodynamic basis. By tracking a wide range of human activities, the Ecological Footprint is able to monitor the combined impact of anthropogenic pressures that are more typically evaluated independently and can thus be used to understand, from multiple angles, the environmental consequences of human activities. The main strength of this methodology is its ability to explain, in simple terms, the concept of ecological limits, thus helping to safeguard the long-term capacity of the biosphere to support mankind and understand how resource issues are linked with economic and social issues. One of the positive characteristics of this methodology is its ability to make visible aspects that are traditionally invisible for conventional economic analyses. For instance, the choice of an area as a unit of measure reflects the fact that many basic ecosystem services and ecological resources are provided

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by surfaces where photosynthesis takes place. Unfortunately, these surfaces are limited by physical and planetary constraints: reporting results in terms of an area helps to better communicate the existence of physical limits to the growth of human economies. Currently, the Ecological Footprint is a robust method widely used to give a measure of the (un)sustainability of consumption patterns at different scales as well as to establish the natural capital requirement of products, services, and activities. Due to the growing number of applications, Global Footprint Network has released the Ecological Footprint Standards[8] in order to enhance the consistency and quality of footprint assessments. This standard contains a list of mandatory requirements for standards compliance. The document also suggests the best way to present the results avoiding distortion and misinterpretations. The National Footprint Accounts are updated and improved every year, and the annual release of the newest version ensures that the method is more robust, reliable, and detailed than previous versions, though some shortcomings still exist and remain to be addressed. The Ecological Footprint Standards are also periodically updated. The information derived from Ecological Footprint assessments could be included in the environmental management and future planning of territories to promote more competitive lifestyles, resource-efficient strategy, and a more effective management of our ecological assets.

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76. Fiala, N. Measuring sustainability: Why the Ecological Footprint is bad economics and bad environmental science. Ecol. Econ. 2008, 67 (4), 519–525. 77. Kitzes, J.; Moran, D.; Galli, A.; Wada, Y.; Wackernagel, M. Interpretation and application of the Ecological Footprint: A reply to Fiala. Ecol. Econ. 2009, 68 (4), 929–930. 78. van den Bergh, J.C.J.M.; Verbruggen, H. Spatial sustainability, trade and indicators: An evaluation of the ‘Ecological Footprint’. Ecol. Econ. 1999, 29, 61–72. 79. Bastianoni, S.; Niccolucci, V.; Pulselli, R.M.; Marchettini N. Indicator and indicandum: “Sustainable way” vs. “prevailing conditions” in Ecological Footprint definition. Ecol. Indic. 2012, 16, 47–50.

28 Environmental Legislation: Asia Introduction .................................................................................................. 321 Law on Biodiversity Conservation

Recent Development in International Law and Its Potential Consequences on National Legislations .............................................. 324 Development of Biodiversity Laws at National Level .............................. 325 Challenges to Effective Implementation and Enforcement of Biodiversity Law ...................................................................................... 330 Regional Cooperation

Wanpen Wirojanagud

Law on E-Waste Management .....................................................................331 Transboundary Movements of Hazardous Waste .................................... 334 Challenges to Effective Implementation and Enforcement of E-Waste Management Law .................................................................... 335 Law on Environmental Assessment ........................................................... 335 Conclusion .................................................................................................... 339 Acknowledgments ........................................................................................ 339 References ......................................................................................................340 Bibliography ..................................................................................................342

Introduction The 21st century has been dubbed “The Asian Century.” The region’s economic growth is unprecedented, with the rise of China and India as the next economic superpowers.[1] China is the second largest economy and is predicted to take over after the United States as the largest economy in the world by 2020. India is one of the fastest-growing economies and is expected to become the world’s third largest economy in the near future.[2–3] More than 58% of the world’s population live in Asia and the Pacific region, and Asia is categorized both as the factory of the world as well as the booming market with tremendous growth potential. As a result, natural resources, which are the capital for development of both economic and social sectors, have been remarkably exploited, with a consequent substantial increase in environmental pollution. Without proper environmental management, the ecosystem continues to suffer, resulting in loss of biodiversity, depletion of ecosystem services, desertification, loss of fertile land, atmospheric pollution, aquatic and marine pollution, etc. The consequences of environmental degradation have become more and more acute and chronic over the years. Moreover, natural disasters are more frequent and more severe. As indicated in the Asia-Pacific Disaster 2010 Report of the United Nations (UN), “while the region generated one quarter of the world’s GDP, it accounted for a staggering 85 per cent of deaths and 38 per cent of global economic losses due to natural disasters over the last three decades.” The report concludes that Asia is the most disaster-prone region in the world.[4] The change in climate patterns has also become more severe. 321

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Environmental status in Southeast Asia, as summarized in Table 1, evidently indicates that over the last decades pressure mostly from anthropogenic activities (rapid population growth, urbanization, economic growth, and consumptive lifestyle) has had a substantial impact on the plentiful natural resources and the environment. Natural resources depletion results in impact on agricultural TABLE 1

Key Environmental Issues and Causes in ASEAN

Country Brunei Cambodia

Indonesia

Laos

Malaysia

Myanmar

Philippines

Singapore Thailand

Vietnam

Shared Issues Seasonal smoke and haze; solid wastes

Key Causes

Transboundary pollution from land and forest fires; inadequate waste management facilities and practices Soil erosion; sedimentation; water pollution; Unmanaged waste and effluent discharge into Tonle Sap deforestation; loss of biodiversity; and Lake; destruction of mangrove wetlands through extensive threats to natural fisheries industrial and aquaculture development Deforestation; loss of biodiversity; water Deficiencies in urban infrastructure—unmanaged industrial pollution; air pollution in urban areas; wastes and municipal effluents and waste; vehicle national and transboundary seasonal congestion and emissions; extensive land clearance and smoke and haze; land degradation; forest fires for pulp wood and oil palm production; pollution of Malacca Straits extensive and unmanaged mining activities; national and transboundary industrial pollution; tourist developments in coastal regions beyond carrying capacity Deforestation; loss of biodiversity; soil Land clearance; shifting cultivation; inadequate water supply erosion; limited access to potable water; and sanitation infrastructure water-borne diseases Urban air pollution; water pollution; Vehicle congestion and emissions; deficiencies in urban deforestation; loss of biodiversity; loss of infrastructure industrial and municipal effluents; extensive mangrove habitats; national and land clearance and forest fires for pulp wood and oil palm transboundary smoke/haze production; unmanaged coastal developments; tourist developments in coastal regions beyond existing carrying capacity Deforestation; loss of biodiversity; urban air Land clearance; excessive mineral extraction; vehicle pollution; soil erosion; water congestion and emissions; deficiencies in urban contamination and water-borne diseases infrastructure—unmanaged industrial and municipal effluents Deforestation in watershed areas; loss of Illegal forest cutting; land clearance; rapid urbanization and biodiversity; soil erosion; air and water deficiencies in urban infrastructure—unmanaged industrial pollution in Manila leading to waterborne and municipal effluents, inadequate water supply and diseases; pollution of coastal mangrove sanitation; tourist developments in coastal regions beyond habitats; natural disasters (earthquakes, existing carrying capacity floods) Industrial pollution; water shortages; waste Seasonal smoke/haze; limited natural fresh water resources; disposal problems limited land available for waste disposal Deforestation; loss of biodiversity; land Sporadic development and destruction of watersheds; degradation and soil erosion; shortage of unmanaged aquaculture; tourist growth exceeding growth water resources in dry season and flooding in carrying capacity; deficiencies in urban and rural in rainy season; conflict of water users; infrastructure; freshwater resources polluted by domestic/ coastal degradation and loss of mangrove industrial wastes, sewage, and contaminated runoff habitat; urban air pollution; pollution from solid waste, hazardous materials and hazardous waste Deforestation and soil degradation; loss of Land clearance for industry; forest clearance and chronic biodiversity; loss of mangrove habitat; impact of Agent Orange; extensive aquaculture and water pollution and threats to marine life; overfishing; growing urbanization and infrastructure groundwater contamination; limited deficiencies; inadequate water supply and sanitation potable water supply; natural disasters (particularly in Hanoi and Ho Chi Minh City) (e.g., floods)

Source: Data from Nguyen.[5]

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productivity; increased frequency of disasters such as floods, landslides, and soil erosion[5]; depletion of aquifers; deterioration of diversity; increased loads from municipal, industrial, and hazardous wastes; atmospheric pollution; and marine and aquatic pollution.[6] Hence, the key policy question facing Asian governments is how to reconcile economic development and environmental protection. In other words, how to make a transition from the “grow first, clean later” approach to the policy of “sustainable development”—a holistic integration of economic, social, and environmental dimensions.[7] More importantly, the main driving forces in the development of environmental management tools in Asia are international agreements, the occurrence of environmental disasters, and non-harmonized legal framework. Asian countries have shown increasing interest in environmental management by which environmental legislation has been used as one of the important tools in this complex task of integrating environmental protection and economic development. Environmental legislation is among the most determined elements in environmental management toward sustainable development.[8] “Legislation is an important element of the institutional framework for environmental management. The role of legislation is to implement and enforce policy and to provide effective administrative and regulatory mechanisms.”[9] Relevantly, development and implementation of environmental law may involve interaction with legislation and administrative practices and institutions. Thus, environmental laws are undeniably increasingly important in Asia. Environmental law is defined as a body of state and federal statutes intended to protect the environment, wildlife, land, and beauty; prevent pollution and overcutting of forests; save endangered species; conserve water; develop and follow general plans; and prevent damaging practices.[9] The particular law gives individuals and groups the right to bring legal actions or seek court orders to enforce the protection, or demand revisions of private and public activity that may have detrimental effects on the environment. In terms of the evolution of environmental legislation, two different types of statutes can be distinguished, referred as to the “first generation” and the “second generation” of environmental laws. Although there is no total agreement about how to characterize each generation, generally, the first generation of environmental legislation refers to “command-and-control statutes and regulations administered with technology-based standards and enforced by rule-of-law litigation.”[10] Then arrived the “second generation” in the late 1980s with the new concept of “sustainable development” aimed at reconciling environmental protection and economic development.[11] The role of the people, particularly local communities, is highlighted, and compliance incentives and market-based mechanisms have been developed to encourage compliance and provide flexibility.[12] In the Asia-Pacific region, environmental legislation began to emerge in the 1970s, the same period as the growing global interest in environmental protection. Examples of the early framework laws and regulations for environmental protection are presented in Table 2. Noticeably, there is no lack of environmental legislation in Asia. However, there is a wide gap between law and practice. The countries have to put an emphasis on ways and means to increase the effectiveness of the implementation and enforcement of environmental law. This entry aims to provide a comparative overview of environmental law and legislation in Asia. It should be noted, however, that the development of environmental legislation in Asia has been uneven and reflects the equally uneven socioeconomic development in the region. In some cases, differences may not be significant in written laws, but lie in the effectiveness of their implementation and enforcement. Due to a wide range of environmental issues and the number of countries in Asia, it is impossible to cover all of them. The entry spotlights four themes—biodiversity conservation, electronic waste (e-waste) management, environmental assessment, and climate change—because of their significance in the Asian context. The choice of countries as case studies aims to demonstrate the uneven development of environmental legislation. Thus, Japanese law serves as a reference of a well-developed body of environmental law, with the secondgeneration of legislation. China, Korea, Malaysia, Indonesia, Singapore, Hong Kong, Cambodia, Thailand, Laos, and Vietnam are at various stages of development ranging from the first-generation statutes, to development of second-generation environmental laws, and implementation and enforcement.

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324 TABLE 2

Early Laws and Regulations on Environment in the Region

Country Cambodia China Hong Kong Indonesia Japan Korea Lao PDR Mongolia Philippines Singapore Thailand Vietnam

Law or Regulation

Year

Environmental Protection and Natural Resource Management Law Environmental Protection Law Water Pollution Control Ordinance Environmental Management Act No. 4 Cabinet Directive Environmental Preservation Act Lao PDR Constitution Environmental Protection Law Environmental Policy Presidential Decree No. 1151 Environmental (Public Health) Act Enhancement and Conservation of the National Environmental Quality Act Environmental Protection Law

1996 1978 1980 1982 1972 1977 1991 1996 1977 1969 1992 1994

Source: Data from World Bank.[42]

Law on Biodiversity Conservation Asia-Pacific is one of the richest regions in terms of biodiversity and 60% of the world’s species are found in this region. As of 2008, the Asian and Pacific regions had the highest number of threatened species in any of the world’s regions, almost one-third of all threatened plants and more than one-third of all threatened animal species.[13] The first generation of biodiversity law consisted mainly in establishing protected areas where human intervention is curtailed or prohibited. However, this exclusionary approach has caused notable socioeconomic impacts, in particular to local communities who used to live in the areas before the establishment of national parks. Therefore, more inclusive approaches have been proposed, such as community-based conservation (comanagement of a protected area by the government and local communities) and payments-based conservation (individuals are paid for their activities to conserve biodiversity). Biodiversity conservation capability is thus a function of a sound legislation in both international and national levels.

Recent Development in International Law and Its Potential Consequences on National Legislations The international legal instruments, together with agreements that initiate the legislative framework of individual countries for biodiversity conservation, include The Convention on Biological Diversity (CBD) that entered into force on December 29, 1993, the Convention on Wetlands of International Importance Especially as Waterfowl Habitat (Ramsar 1971), the Convention Concerning the Protection of the World Cultural and Natural Heritage (Paris 1972), the Convention on International Trade in Endangered Species of Wild Fauna and Flora (Washington 1973), the Convention on the Conservation of Migratory Species of Wild Animals (Bonn 1979), etc.[14] However, an effective implementation for the international legal agreements is largely dependent on the actions taken by individual countries. Consequently, the most important of all is to establish the legal systems for biodiversity conservation at state leve1. Very recently, on February 2, 2011, the Nagoya Protocol on Access to Genetic Resources and the Fair and Equitable Sharing of Benefits Arising from their Utilization was opened for signature by Parties to the Convention of Biological Diversity. Many Asian countries are in the process of amending their legislation to implement the provisions under this new protocol. As most Asian countries are resourceproviding countries, the Nagoya Protocol will allow better control of their genetic resources and ensure fair financial compensation among various stakeholders, including local communities.[15]

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As for forestry, the new market-based mechanism of “Reducing Emission from Deforestation and Degradation (REDD),” which is under discussion in the UN Climate Change negotiations, should provide additional impetus to the improvement of forest law in Asian developing countries, particularly in the areas of monitoring, reporting, and data collection. Cambodia, Indonesia, Papua New Guinea, the Philippines, and Vietnam are among the first countries to receive support from the UN REDD Program.[16]

Development of Biodiversity Laws at National Level At present, the biodiversity conservative laws of Asian countries like Japan, Thailand, Indonesia, and Laos contain more or less elements of public participation that characterize the second generation of legislations. Except for Japan, however, communities’ rights are still very limited and difficult to enforce. The payment-based conservation approach has generally only been implemented through small-scale projects and not adopted by Asian legislators yet. China. China covers an enormous land area of 9,600,000 km2, including complex and varied geomorphology, climate, and natural conditions. That creates a country rich in ecosystems, which can be categorized into five types, namely forest, grassland, desert, inland wetland and other freshwater ecosystems, and ocean and coast. Due to a massive and distinct diversity of flora and fauna, China is regarded as one of the most important biodiversity countries. It is ranked among the top 10 nations in the world diversity of its mammal, bird, amphibian, and plant species. China has been considered to have one of the most important stocks of genetic diversity in the world. It is very important to protect and conserve this biodiversity for the national and international heritage. China has promulgated a series of laws and regulations. The main domestic laws are Forest Law (1984), Grassland Law (1985), Fishery Law (1986), and Wild Animal Conservation Law (1988). Examples of regulations related to biodiversity include Reproduction and Conservation of Aquatic Resources (1979), Regulation on Forest and Wild Animal Nature Reserves Management (1985), Regulation on Forest Fire Prevention and Control (1988), Regulation on Seed Management (1989), Regulation on Conservation of Terrestrial Wild Animals (1992), Regulation on Nature Reserves (1994), etc. Regarding the enforcement of the statutes, great progress for in situ and ex situ biodiversity conservation has been achieved. There are, however, still some gaps in the legislation. Based on the current status of conservation legislation in China and in accordance with the Convention on Biological Diversity, more attention should be paid to the conservation legislation for genetic resources, wild plant species, and various natural ecosystems. Korea. Ecosystems in Korea comprise forest, mountain, freshwater, coastal and marine, and agriculture ecosystems. The total forest area covers 6.394 million hectares, estimated as about 64% of the country’s land area. Forests are mainly coniferous, deciduous, and mixed forests. The variety of habitats creates a rich biodiversity of plants, animals, and other living organisms (fungi, protista, prokaryotes, etc.). Some species are considered to be extinct, such as the tiger and Siberian leopard, fox, wolf, and sitka deer. The decline of biodiversity in Korea is associated with its economic development. The main threats to biodiversity include overexploitation of land and biological resources, and environmental pollution. Under the guiding principles of the Framework Act on Environmental Policy 1990 and the Constitution, the Natural Environment Conservation Act 1991 administered by the Ministry of Environment is Korea’s basic law for biodiversity and nature protection. It defines categories of protected areas and provides for species and habitat protection. The law serves as a common framework for nature conservation and strengthens the provisions of other nature laws administered by government agencies. Several government agencies share the responsibility of conservation and sustainable use of biodiversity, in accordance with various laws. The Ministry of Environment is responsible for general biodiversity conservation under the Law of Natural Environment Conservation, Law of Wildlife Protection and Hunting, Law of Wetland Conservation, Law of Natural Parks, and Law of Ecosystem Conservation for Uninhabited Islands. The Ministry of Environment is also responsible for preventing

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inappropriate uses of natural resources through the Environmental Impact Assessment (EIA) process by the Law of Environmental Impact Assessment. The Forestry Administration, part of the Ministry of Agriculture and Forests, which manages forests under the Law of Forests. Japan. The country is 3000 km long in the north–south direction, with a vertical range from coasts to mountains, with thousands of islands, and a geological history of intermittent connection to and separation from the continent, and various disturbances such as eruption of volcanoes, flooding of precipitous rivers, and typhoons. Those geological characteristics, together with four definite seasons due to the monsoon climate, create diverse habitats. It is such rich biodiversity that makes Japan one of the 34 biodiversity hotspots identified worldwide. A biodiversity hotspot refers to a region that is originally rich in biological diversity and endemic species but is now exposed to a serious threat of loss of such diversity. Besides the original local geohistorical and natural conditions, the tradition of wetpaddy rice agriculture and the rural lifestyle, which rely on a secondary natural environment known as “satochi-satoyama,” or simply “satoyama,” as well as the way the land has been used for agricultural purpose, have also contributed to the area’s biodiversity richness. As in other countries, biodiversity loss in Japan is due to high economic growth with industrial development, and also natural disasters. It can, however, be said that the Japanese law on biodiversity is regarded as one of the best laws in the field. It goes beyond the protected area–based approach by mainstreaming biodiversity conservation into the daily life of the people. The law defines not only the responsibilities of national government and local governments but also those of businesses, citizens, and private bodies. The elaboration of the National Biodiversity Strategy and regional biodiversity strategies are mandatory. The 4th National Biodiversity Strategy, which was adopted in 2010, sets a long-term goal of 100 years, mid-and short-term targets for 2020 and 2050, and indicates about 720 measures with 35 targets.[17] The results of the implementation have to be reported to the Diet every year in the Annual Report on the State of Biodiversity. The law puts emphasis on preventive and adaptive approaches, including land use planning, research and technology development, EIA, and prevention of global warming. The role of the public is also highlighted with mandatory public consultation before formulation of policies and support of voluntary activities by businesses and citizens for the conservation of biodiversity. Other relevant legislations are the Nature Conservation Law, Natural Parks Law, Law for the Promotion of Nature Restoration, Law for the Promotion of Biodiversity Conservation Activities, Law for the Conservation of Endangered Species of Wild Fauna and Flora, Wildlife Protection and Proper Hunting Law, Invasive Alien Species Act, and Law Concerning the Conservation and Sustainable Use of Biological Diversity through Regulations on the Use of Living Modified Organisms. Malaysia. Malaysia has been identified as one of the world’s mega-diversity areas with extremely rich biodiversity. Covering much of the country are the tropical forests, the oldest and most biologically diverse ecosystem on Earth. With ratification of the CBD, Malaysia is working toward incorporating into its national policies and planning a set of commitments under the treaty as well as setting the goal to become a world leader in conservation, research, and sustainable utilization of tropical biodiversity by 2020. To accomplish the ratification and goal, Malaysia has enacted a spectrum of legislation aimed at protecting biodiversity. Examples of law relevant to biodiversity conservation are as follows: Environment Quality Act 1974, Fisheries Act 1985, Pesticides Act 1974, Plant Quarantine Act 1976, Protection of Wildlife Act 1972, National Parks Act 1980, National Forestry Act 1984, Parks Enactment 1984, Forest Enactment 1992, Fauna Conservation Ordinance 1963, etc.[18,19] Malaysia has enacted a number of laws and regulations to protect the nation’s environment but, while adequate, there was no single overarching statute (or policy?) that relates to biodiversity conservation and management until the just approved National Policy on Biodiversity. Much of the present legislation is sector based. As stated above, the attainment of biodiversity conservation is significantly dependent on implementation and enforcement of the legislation. The effectiveness of legislation can be accomplished by the dedication of government agencies as well as public participation for accountability.[18,19] Indonesia. Indonesia is a rich and diversified archipelagic nation. With the topographical characteristics of approximately 13,500 islands and extensive reef system, Indonesia has a wide range of natural

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habitats, with a wealth of fauna and flora, corals, fish, and other reefs; thus, it is recognized as a major world center for biodiversity. However, that substantial biodiversity is decreasing in the country owing to illegal clearing and deforestation, including large-scale burning in oil palm plantations and smallscale slash and burn for shifting farming, as well as illegal logging and trade in timber. Indonesia has one of the world’s worst deforestation rates. In addition, illegal poaching, trade in protected species, and illegal and unsustainable fishing have threatened the country’s biodiversity.[20,21] Regarding the domestic legal framework, as the result of 1972 UN Stockholm Conference on the Environment, Indonesia promptly established of the Office of the State Minister for the Environment and enacted the Environmental Management Act (EMA) No. 4 of 1982 replaced in 1997 by the EMANo. 23 of 1997. This act and its implementing regulations are set in the broader context of the state policies passed every 5 years by the People’s Consultative Assembly, which have since the early 1970s progressively entrenched the concepts of sustainable development and natural resources management. The EMA must also be read in the context of other natural resources management acts, such as the Forestry Act No. 41 of 1999, and the Fisheries Act No. 31 of 2004, and associated regulations and decrees transferred to the local governments’ agencies. Laws and regulations related to biodiversity conservation are exampled as follows: Act on the Conservation of Biological Resources and their Ecosystems (Act No. 5 of 1990), Decree No. 1 of the Minister of Agriculture on the Conservation of the Riches of the Fish Resources of Indonesia, and Decree of the Ministry of Forestry No. 424/ KPTS-VI/1994 on the Guidelines on Crocodile Management in Indonesia, Fisheries Law (No. 9 of 1985). Similarly to other countries, law enforcement is still not an effective function. Indonesia has no specialized environmental law courts. Environmental cases are heard by the general and administrative courts as well as the Supreme Court on appeal.[21] Furthermore, a multitude of biodiversity laws in Indonesia tend to be conflicting and uncoordinated. Also, in the forestry sector, the power to manage forests is shared between the central government and regional governments with no clear division of powers and responsibilities. It is believed that the Regional Autonomy Law, which entered into force in 1999, has been abusively used by local authorities to issue their own logging concessions, thus resulting in massive deforestation and forest fires.[22] The Indonesian case shows that a legislation that transfers power from center to periphery has to be carefully drafted to ensure that the power is put in the hands of local communities whose livelihood depends on the forests and biodiversity, and not in the hands of local officials known for corruption and nepotism. Some main legislations of Indonesia relating to biodiversity are the Conservation of Biodiversity and Ecosystems Law (1990) and the Basic Forestry Law (2000). These laws include provisions on participatory forestry planning, people’s economic empowerment, partial transfer of authority to regional governments, and community-based forest monitoring. Philippines. The Philippines is a tropical archipelago of 7100 islands located off the southeast coast of mainland Asia. It occupies a land area of 299,400 km2, and territorial waters covering around 2,200,000 km2, that create precious terrestrial and aquatic ecosystems and habitat types. As a consequence of rapid loss of biodiversity, as well as widespread destruction of the country’s environment, a strong effort has been put into biodiversity conservation, including the formation of the multisector Philippine Council for Sustainable Development (1992), ratification of the Convention on Biological Diversity (1993), and preparation of the Philippine Biodiversity Assessment Report and the National Biodiversity Strategy and Action Plan (NBSAP) (1995–1997). The NBSAP proposes a wide range of strategies and actions, including information generation, in situ and ex situ conservation, legislative and policy development, institutional capability building, information, education and communication, and strengthened international cooperation.[23] Regarding ecosystem and habitat conservation in the Philippines, it was innovated through the National Integrated Protected Areas System Act of 1992, a landmark piece of legislation that provides the framework for a decentralized, community-based reserve management strategy. The legislation relevant to biodiversity conservation include the following: Act No. 2590 (1916), An Act for the Protection of Game and Fish; RA 7586 National Integrated Protected Areas System Act of 1992; RA  7900,

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High-Value Crops Development Act of 1995; PD 1433, Plant Quarantine Decree of 1978; Proc. No. 926, Establishing Subic Watershed Forest Reserve; DAO 20, s 1996, Implementing Rules and Regulations on the Prospecting of Biological and Genetic Resources; DAO 24, s 1991, Shift in Logging from the Old Growth (Virgin) Forests to the Second Growth (Residual) Forests; DAO 20, s 1996, Implementing Rules and Regulations on the Prospecting of Biological and Genetic Resources; etc.[24] Singapore. The Republic of Singapore, situated off the southern tip of the Malay Peninsula, comprises one major and more than 50 adjacent islands, with a total area of 648 km 2. The main island is separated from Malaysia by the narrow Johor Strait on the north, and from Indonesia’s Riau Archipelago by the Singapore Strait on the south. It is a small country with an urbanized character. Forest land and coastal areas have been cleared to provide land for residential and commercial sites and other developments, resulting in a substantial depletion of flora/fauna and deterioration of natural habitats. At present, the environmental policy states that 5% of the land area should be set aside for nature reserves, national parks, catchment areas, bird sanctuaries, and gardens. Some of the environmental laws related to biodiversity conservation are Fisheries Act 1966, Wild Animals and Birds Act 198, Parks and Trees Act (for parks not gazetted as national parks) 1985, Endangered Species Act 1989, National Parks Act 1990, Animals and Birds Act (Revised) 2002.[25] To accomplish the vision of Singapore Today, which is to be “A Garden City, A Haven for Biodiversity,” the Singapore Green Plan (SGP) 2012 was established to provide the direction for protected area management in the next decade. One of the SGP objectives is to ensure the quality of the living environment, including the enhancement of the country’s environmental heritage. SGP 2012 also recaps the state’s commitment to maintain the 5% of land set aside for nature areas, and provides the direction.[26] Hong Kong Special Administration Region. Like Singapore, Hong Kong SAR (or Hong Kong in short) experienced very rapid growth and development, to become urbanized and industrial in character. This would have jeopardized the biodiversity of the area, which is recognized as part of the natural heritage of Hong Kong. The most common causes of biodiversity loss in Hong Kong now are habitat destruction associated with infrastructure development, population growth, hillside fires, unsustainable exploitation of wild species, introduction of alien species, pollution, and global environmental change. Possible relevant factors contributing to such causes are lack of a comprehensive legal framework to protect areas of high conservation value and poor enforcement of existing conservation, environmental, and planning laws.[27] Environmental protection in Hong Kong is constituted by 16 ordinances. Nonetheless, overall environmental laws in Hong Kong are still short of an unambiguous conservation objective. Only the 1995 Marine Parks Ordinance (Cap. 476) and the 1997 Protection of the Harbour Ordinance (Cap. 531) enclose evidently expressed conservation principles. Other legislations protecting flora and fauna are focused on conserving particular species; however, they fail to address the values and principles that lie beneath these objectives. Regarding the Country Parks Ordinance (Cap. 208) originally enacted in 1976, a revision is needed, “To provide for the designation, control and management of country parks and special areas for the purposes of conservation of biological diversity, countryside recreation and education.” Additionally, the Environmental Impact Assessment Ordinance (EIAO) is the newest piece of legislation that seeks to protect the environment. The EIAO requires certain designated projects (generally major infrastructure projects) to undertake an EIA before they can be granted an environmental permit for development to proceed. A Technical Memorandum contains guidance on the criteria and guidelines to use for ecological, fisheries, landscape, and visual impact assessment.[27] As well as domestic laws, Hong Kong is obligated to protect its natural and cultural heritage by international treaties, including the 1973 Convention on International Trade in Endangered Species of Wild Fauna and Flora, the 1979 Convention on the Conservation of Migratory Species of Wild Animals, the 1971 (Ramsar) Convention on Wetlands of International Importance, and the 1972 Convention for the Protection of World Cultural and Natural Heritage. Seemingly, application of the 1992 Convention on Biodiversity in Hong Kong remains limited as there is no sign that the treaty will be formally applied to accomplish the CBD’s biodiversity objectives after the government’s endorsement.[27,28]

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Thailand. Thailand has a total land area of 513,115 km2, lying in a hot and humid climatic zone in the middle of Southeast Asia. Much of Thailand is situated in the Mekong River basin, and it is one of the Greater Mekong Subregion (GMS) countries. Thailand also has an extensive coastline. With such a location, it is enriched with biodiversity associated with terrestrial and aquatic (both freshwater and marine) ecosystems. It also covers agricultural ecosystems (about one-fifth of the country), which include biodiversity components of rice, farm crops, and livestock. Similarly to other countries, the major threat to biodiversity is human disturbance through overexploitation of natural resources/ habitat, illegal logging and trading of animals, overhunting of wildlife, deforestation, urban expansion and pollution, etc. Such disturbances cause an adverse reduction of bio-diversity.[29] Biodiversity in Thailand is safeguarded by a number of laws and regulations. Some of the important ones are the National Park Act 1961, National Forest Reserve Act 1964, Wild Animal Reservation and Protection Act 1992, Plant Quarantine Act of 1964 and Plant Quarantine Act (second issue) 1994, Animal Species Maintenance Act 1966, Importing and Exporting of Goods Act 1979, Enhancement and Conservation of National Environmental Quality Act 1992, and Plant Varieties Protection Act 1999.[30] In addition, Section 66 Paragraph 1 of the Thai Constitution of 2007 guarantees the right of an individual and communities to participate in the preservation and exploitation of natural resources and biological diversity. However, out of 35 pieces of legislation relating to natural resources, biodiversity and environment, only 4 provides for public participation, namely, the National Promotion and Conservation of Environment Quality Act, B.E. 2535 (1992); the Private Irrigation Act, B.E. 2482 (1939); the Plant Varieties Protection Act, B.E. 2542 (1999); and the Protection and Promotion of the Thai Traditional Medicine Act, B.E. 2542 (1999). Moreover, the rights given to individuals are still very limited, and the authorities retain most of the control in natural resources management. Lao People’s Democratic Republic. Lao PDR, situated in the Mekong River basin, is counted as one of the GMS countries. For Southeast Asia, Lao PDR is one of the countries with a large proportion of land covered with undisturbed forest. It covers a land area of 236,800 km2, where the topography is largely mountainous, with elevations above 180 m typically characterized by steep terrain, narrow river valleys, and low agricultural potential. Lao PDR is rich with natural resources such as forestry, minerals, and hydroelectric power. Due to the country’s still abundant forestry resources, it thus generates a prosperous biodiversity.[31,32] Similarly to other GMS countries, Lao PDR has enacted laws for biodiversity conservation, but the enforcement is ineffective. Illegal logging and wildlife hunting have been frequently found. Mining and hydropower developments might also cause loss of biodiversity.[31,32] In Lao PDR, biodiversity conservation is provided for in the Forest Law of 1996, a 1993 decree that designated the first national biodiversity conservation areas, logging ban, decree on (PM Decree No. 67/PM, 1991) protecting trees by logging, decision on adoption of (PM Decree No. 66/PM, 1991) forest conservation, and seed material import regulation (quarantine regulation) that controls importation of plant material.[32] Cambodia. Cambodia is a tropical country in mainland Southeast Asia with a territory of 181,035 km2. It is adjacent to the Gulf of Thailand on the south and shares borders with Thailand (west and north), Laos (north), and Vietnam (east). Cambodia is in the Mekong River basin. The country is dominated by the Mekong River, known as the Tonle Thom or “great river,” and the Tonle Sap or “fresh water lake.” The Mekong River basin is one of the most biodiverse regions in the world. As situated in the Mekong River Basin, Cambodia has an abundance of terrestrial and aquatic ecosystems that are significant habitat for plants and aquatic organism, fishes in particular. Such ecosystems are invaluable resources for economic development and human well-being.[33,34] To protect and sustain the country’s biodiversity, Cambodia enacted the framework for the Law on Environmental Protection and Natural Resources Management in 1996. Subsequently, the Ministry of Environment was created in 1998, which manages natural resources along with the Ministry of Water Resources and Meteorology and the Ministry of Land Use Management, Urbanization, and Construction. Cambodia has continued to enact more environmental and conservation laws. Examples of the relevant laws are as follows: Environmental Protection and Natural Resources Law 1996, Law on Commune

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Administration (part of the decentralization process) 2001, Land Law 2001, Forestry Law 2002, and Wildlife Law 2002. Important subdecrees include Subdecree on Concession Management (moratorium on logging and log transport), Subdecree on Community Forestry, Subdecree on Environmental Impact Assessment, Subdecree on Industrial Agricultural Concessions, Subdecree on Social Concessions, and Royal Decree on Protected Areas 1993. Cambodia is party to a number of important international conventions of which those stated here are relevant to biodiversity conservation. The Ramsar Convention on Wetlands has been ratified, and Cambodia has identified three wetland sites for recognition: BoengChhmar in the Tonle Sap floodplain, KohKapik on the coast, and a portion of the middle Mekong river north of Stung Treng. In addition, the Tonle Sap is recognized as a UNESCO Biosphere Reserve. Cambodia ratified the Convention on Biological Diversity in 1995. In 1997, the government prepared a biodiversity prospectus and in 2002 completed a National Biodiversity Strategy and Action plan.[19,33] Vietnam. Vietnam is situated at the eastern side of a peninsula that protrudes into the Eastern Sea, which is a bay of the Pacific Ocean. Within the Mekong River Basin, Vietnam is at the most downstream length of the Mekong River. Due to the country occupied with land, river, and sea, Vietnam is enriched with a variety of ecosystems, including tropical rainforests and monsoon savannah, marine life, and mountainous subalpine scrubland. Additionally, a specific feature of Vietnam is its length of more than a thousand miles from north to south, with a width of only 30 miles from east to west at its narrowest point, thus generating an abundance of natural resource along the coast and sea, and a richness of biodiversity. The threats tobiodiversity are mainly due to transformation of forest and wetland areas to other uses, infrastructure construction, urbanization, industrialization, and environmental pollution. Vietnam’s environmental law is based on its constitution. The Law on Environmental Protection was initially promulgated in 1993; subsequently, Vietnam has enacted a variety of laws and decrees on conservation issues. These laws affect, directly or indirectly, the conservation of biodiversity. Examples of the direct laws are the Decree on the Conservation and Development of Wetlands, the Decree on Protection of the Environment (which details rare and precious flora and fauna), and a related decree that determines methods for regulating their protection and management. The indirect laws include decrees regulating wastewater, controls on businesses creating environmental damage, the 2003 Land Law, etc. Regarding the protected areas, Vietnam has two laws concerning the establishment and management of protected areas. The two statutes are the Law on Forest Protection and Development of December 2004 (revised from the 1991 original), and the Law on Biodiversity that became effective starting July 2009. The Law on Forest Protection and Development provides the guiding principles for the development and use of special-use forests, while the Law on Biodiversity focuses on protected area concerns such as categorization and decentralization of protected area management.

Challenges to Effective Implementation and Enforcement of Biodiversity Law Evidently, the countries presented here have enough laws and regulations to protect their environment; however, much of the present legislation is sector based. Moreover, the achievement of biodiversity conservation is not a function of the number of laws and regulations but of the implementation and enforcement of such legislation. The effectiveness of legislation can be accomplished by the dedication of government agencies as well as public participation for accountability.[19] As partly demonstrated through the case studies mentioned above, challenges facing Asian governments in the implementation and enforcement of biodiversity law lie in corruption, weak institutional capacity, lack of reliable data and budget, high demand for alternative land use, and traditional beliefs. With regard to weak institutional capacity, in most Asian countries, forestry staffs are not adequately trained and are underpaid. To address this shortcoming, governments such as China and Vietnam have raised the pay and living conditions for forest personnel.[19] Inter-agency coordination is another serious

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institutional problem. There are usually many agencies involved in forest management with overlapping areas of responsibilities or, on the contrary, areas where no agency is in charge. Economic development has put pressure on forest protection. The construction of hydropower dams and roads, and agricultural activities have often encroached on protected lands. In an effort to reconcile forest protection and economic development, Lao PDR has piloted hydropower levies that support the management of protected areas affected.[34] Forests and wildlife statistics are often out of date, inaccurate, and incomplete. The remoteness of the forest areas and the reluctance of officials to report forest crimes (as they may be seen as a sign of failed forest management on their part) are among factors that contribute to the lack of reliable data and information sharing among relevant agencies. The use of satellite imagery has helped improve the data collection system to a certain extent. Traditional beliefs may sometimes come in the way of forest and wildlife protection. In many Asian countries such as China, Vietnam, Thailand, Laos, and Cambodia, the consumption of rare wildlife, such as bear’s paw and monkey brains, is seen as status symbols or medicines.

Regional Cooperation Regional cooperation has played an important role in promoting sustainable forest management. Examples of regional initiatives are the East Asia Forest Law Enforcement and Governance process (EAFLEG) and the Asia Forest Partnership (AFP). Both the EAFLEG and the AFP aim to bring together various stakeholders in forest management, including governments, non-governmental organizations (NGOs), and the private sector. They serve as an informal forum for information sharing, dialogue, and joint action. Within the Southeast Asian Nations Association (ASEAN), the ASEAN Senior Officials on Forestry (ASOF) was entrusted with the task of policy coordination and decision making on regional cooperation in the forest sector.

Law on E-Waste Management The quantity of e-wastes in Asia has exploded in recent years with the exponential growth in the use of electronic equipment (computers, mobile phones, televisions, refrigerators, etc.) coupled with the consumers’ behavior of regularly replacing their devices in order to stay up-to-date with the latest technology. E-wastes often end up in incinerators and/or landfills. Toxic substances such as mercury and lead that are commonly used in electronic products can contaminate the environment, including land, water, and air. E-waste is commonly characterized as hazardous waste. As of July 2008, among 46 countries in the Asia-Pacific region, there are 32 countries that have ratified the Basel Convention on the Control of Transboundary Movements of Hazardous Wastes and Their Disposal and some of them have also ratified the Ban amendment.[35] However, there are some countries in Asia that still lack regulations for controlling hazardous wastes including e-waste. An effective regulatory framework on the management of such wastes is urgently needed. A new legal concept created to deal with e-waste management is the Producer Extended Responsibility (PER) where producers are held liable for the costs of managing their products to end of life of the product. In this way, producers are encouraged to design environmentally friendly products to reduce disposal costs. Only a few countries, such as Japan, Korea, and Taiwan, have a well-established legislation on e-waste management based on the concept of PER and many years of experience in implementation. Some Asian countries such as China have recently enacted a law on the matter; however, the effectiveness of the implementation remains to be seen. Other countries have already passed similar laws, but the laws have yet to enter into force (Thailand and India) or are still at the drafting stage. An adaptation period is usually granted for businesses before the entry into force of the law. Other actors in a product’s life cycle such as distributors, repair and customer service providers, consumers, and recyclers may also be required to bear some of the treatment and disposal costs. E-waste discussed herein includes used and waste electrical and electronic equipment (UEEE/WEEE).

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China. The Ministry of Environment and Protection is responsible for regulating and controlling e-waste management. At the end of February 2006, China promulgated the law entitled “Administration on the Control of Pollution Caused by Electronic Information Products,” which is simply called as China RoHS in the industry.[36] Import of WEEE has taken place since 2002 under the existing laws, including Law of the People’s Republic of China on Prevention of Environmental Pollution Caused by Solid Waste, Interim Provisions on Administration of Environmental Protection on Import of Waste and its supplementary provisions, List of Wastes Prohibited against Import (Notice No. 25, 2002), and Catalogue of Solid Waste Forbidden to Import in China (Announcement No. 11, 2008). Regarding UEEE, it is allowed except used TVs; UEEE requires 3C certification. Applicable laws for UEEE are as follows: Administrative Method on Inspection and Supervision of Imported Used Mechanical and Electrical Products, Measures for Administration of Import of Specified Used Mechanical and Electrical Products (Order No. 5, 2008), and Catalogue of Import of Specified Used Mechanical and Electrical Products (Announcement No. 37, 2008). A new law on the management of e-waste has established higher standards for recycling processes and allows only certified recyclers to engage in the e-waste recycling business. To support recyclers to improve their equipment and facilities, a centralized mandatory fund has been established with contributions from domestic producers and sellers of imported electronic devices. The law also places responsibility on manufacturers, distributors, repair and consumer service providers, and recycling companiesto collect and responsibly handle e-waste. One shortcoming of the law, however, is that the scope of their responsibility and the penalizing measures for non-compliance remains vague.[35] Hong Kong, China. The regulation of e-waste management is under the responsibility of the Environment Protection Department. Hong Kong has begun its waste import and export control through the Waste Disposal Ordinance in 1996. For the purpose of import, WEEE and UEEE (that is classified as WEEE) have been controlled through a permit system in accordance with guidelines on “import and export of hazardous waste including electrical and electronic appliances containing hazardous constituents or components.”[35] Republic of Korea. The Resource Recirculation Policy Division, Ministry of Environment, is responsible for regulating and controlling e-waste management. Specific regulations applicable to UEEE and WEEE do not currently exist; however, in general, e-waste management falls under the “Waste Control Act” of December 1986 and later amendments. Import control of WEEE is performed through application for a license from the Ministry of Environment in accordance with the Act on the Control of TransBoundary Movement of Hazardous Wastes and Their Disposal (Basel Convention) and Act on Resource Recycling of Electrical and Electronic Equipment and Vehicles. Import of UEEE is allowed; no specific law is applicable.[35] Japan. The Ministry of Environment is responsible for regulating e-waste management. Since 2001, Japan has enforced the Fundamental Law for Establishing a Sound Material-Cycle Society to promote comprehensively and systematically the policies for realizing a Sound Material-Cycle Society, providing an umbrella framework for the relevant waste management laws of the country. Regarding the import and export control of WEEE, the Law for the Control of Export, Import, and Others of Specified Hazardous Wastes and Other Wastes was entered into force in 1993. This law stipulates the necessary import/export procedures of hazardous waste to comply with the requirements of the Basel Convention. The Waste Management and Public Cleansing Law of 1970 was amended in 1993 to regulate import and export of waste. None of these laws is applicable for the management of UEEE. With the specific purpose to complement to the Basic Act on Establishing a Sound Material-Cycle Society (2000), the following two specific e-waste recycling legislations have been established: the Law for the Promotion of Effective Utilization of Resources (LPUR), and the Law for Recycling of Specified Home Appliances (LRHA) (1991). The LPUR applies to used computers and small-sized secondary batteries and encourages manufacturers’ voluntary efforts to take part in collection and recycling. Recycling costs are borne by both manufacturers and consumers. The LRHA is a stricter regulation. It covers television sets, refrigerators, washing machines, and air conditioners. Manufacturers and retailers have an obligation to take back

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used products and recycle them. Consumers are required to pay for the cost of transportation and recycling. The LRHA also sets up a procedure that allows tracking a product from the beginning until the end of its life cycle.[35] Malaysia. The Department of Environment, Ministry of Natural Resources and Environment, is responsible for regulating e-waste management. Under the Environmental Quality Act, 1974, several regulations for the control of scheduled wastes (hazardous wastes) management in Malaysia have been enacted. The principal regulation on e-waste management is the Environmental Quality (Schedule Waste) Regulation of 2005, enforced by Department of Environment, and in which specific categories of e-waste are defined and coded. The Guidelines for the Classification of Used Electrical and Electronic Equipment entered into force in January 2008, which prohibits the import of WEEE and export for the purpose of disposal. Waste generators are allowed to export waste for recycling, recovery, or treatment provided prior written consent are obtained from the importing state. The Ministry of Local Government and Housing has the jurisdiction over households and business entities/institutions and has enacted the Solid Waste Management and Public Cleansing Act of 2007. The Royal Malaysian Customs enforces transboundary movements of hazardous waste under the Customs Act 1967, Customs (Prohibition of Import) Order 2008, and Customs (Prohibition of Export) Order 2008.[35] Indonesia. The Environmental Impact Management Agency, Ministry of Environment, is responsible for regulating e-waste management; however, neither specific criteria on e-waste nor specific regulations on e-waste management have been established. The existing laws have been employed for WEEE and UEEE. WEEE is only allowed for export, but prohibited for import to Indonesia under Act No. 23 of 1997 on Environmental Management, Articles 20 and 21, Presidential Decree No. 61/1993 Basel Convention Ratification, and Ministerial Decree No. 231/ MPP/Kp/07/1997 Regarding Import Procedure of Waste. Import of UEEE and e-waste for direct consumption by consumers is prohibited under Decree No. 756/MPP/ Kep/12/2003 on Import of Non-New Capital Goods and Decree No. 610/MPP/Kep/10/2004 Regarding Amendment of No. 756/MPP/Kep/12/2003.[35] Philippines. The Department of Environment and Natural Resources (DENR) is responsible for regulating e-waste management. Import of WEEE and UEEE requires permit. Laws applicable to WEEE are the Toxic Substances and Hazardous and Nuclear Wastes Control Act of 1990 (Republic Act No. 6969), DENR Administrative Order 200–6 (Implementing Rules and Regulations for RA 6969), DENR Administrative Order 1994-28 (Interim Guidelines for the Importation of Recyclable Materials Containing Hazardous Substances), DENR Administrative Order 1997-28 (Amending Annex A of DAO 1994-28), and DENR Administrative Order 2004–27 (Amending Annex A of DAO 1994-28). The law applicable to UEEE is DENR Administrative Order 1994-28 (Interim Guidelines for the Importation of Recyclable Materials Containing Hazardous Substances). DAO-94-28 allows the import of electronic assemblies and scrap with the condition that residuals from recycling of materials that contain hazardous substances without any acceptable method of disposal in the Philippines must be shipped back.[35] Singapore. Export, import, or transit waste requires a permit from the Pollution Control Department of Singapore in accordance with the Hazardous Waste (Control of Export, Import, and Transit) Act. Import/export of UEEE are allowed if there are documents to support that the appliances for import/ export are in working condition and suitable for reuse. Export of UEEE that are not suitable for reuse is prohibited. Import of UEEE for the purpose of dismantling and re-export of the dismantled components are prohibited.[35] Thailand. The Ministry of Natural Resources and Environment and the Ministry of Industry are the administrative authorities for hazardous waste and e-waste management. Both UEEE and WEEE are controlled under the Hazardous Substance Act B.E. 2535 (AD 1992) in Thailand. UEEE can be imported only under a subordinate law for import control of UEEE. Import of UEEE is allowed only for reuse, repair/refurnish back to its original purposes, disassembly and recycle/recovery under certain conditions. Thirty-two UEEE items require import permits from the Ministry of Industry. WEEE can

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be imported and exported under a subordinate law and following Basel Convention procedures. The e-waste management act will only enter into force in 2014 and will seek to regulate the entire life cycle of an electronic product. At the beginning of its life cycle, importers and producers will be taxed, and the tax money collected will be used for e-waste management. At the end of the product’s life cycle, the law will promote construction of an integrated waste management facility at production sites and other areas throughout the country.[35] Lao People’s Democratic Republic. Lao PDR does not have legislation discretely mentioning e-wastes; however, e-waste is considered hazardous waste by defining that hazardous and toxic wastes include batteries, old paint cans, aerosol, and other refuse. Such wastes are mixed with municipal solid wastes that are disposed at landfills. Accordingly, there are no specific laws or regulations directed to e-wastes. For solid waste disposal itself, a decree on waste management is planned in connection with the finalization of the revision of the Environmental Protection Law of 1999. Nonetheless, some laws have implications for solid waste management in Lao PDR, for example, the Environmental Protection Law (Article 22), the Decree on Implementing the Environmental Protection Law (Article 9.4), and in addition some of the provinces and the Capital City of Vientiane have issued specific regulations on urban environmental management including solid waste management. Thus, Lao PDR is encountering solid waste problems due to several reasons, as follows: inadequate legal framework, ambiguous institutional responsibilities and lack allocation of responsibilities on solid waste management to specific institutions, insufficient budget allocation to carry out functions in accordance with the law, etc.[37] Cambodia. At present, Cambodia does not produce EEE products. Regulation on e-wastes is the responsibility of the Ministry of Environment. Specific law and/or regulation to properly manage, recycle, and dispose EEW does not yet exist. Cambodia, however, uses the following existing laws for WEEE management: Subdecree on Solid Waste Management and the Inter-Ministerial Declaration on SWM in Cities and Provinces. Import of WEEE to Cambodia is banned, while import of both new EEE and UEEE is allowed for domestic consumption. Export of household waste and hazardous waste from Cambodia requires approval from the Ministry of Environment, export license from the Ministry of Trade, and permit from the import country (BCRC China, 2009). Vietnam. Regulating e-waste management is the responsibility of Hazardous Waste Management Division, Waste Management and Environment Promotion Agency, Vietnam Environment Administration, Ministry of Natural Resources and Environment. In January 2006, Vietnam promulgated the Implementation Rules for the Law on Trade (No. 12/2006/ND CP), which bans import of waste materials (both WEEE and UEEE), toxic chemical substances, and second-hand commodities, including electronic, cooling, and home appliances. Other applicable laws are the Regulation of Management of Hazardous Waste (Decision No. 155/1999/QD-TTg), and Decision No. 23/2006/QD-BTNMT on the List of Hazardous Waste. In Circular No. 12/2006/TT-BTNMT, export of hazardous waste shall follow Basel Convention procedures.[35]

Transboundary Movements of Hazardous Waste Since the entry into force of the Basel Convention on the Control of Transboundary Movements of Hazardous Wastes in 1992, most Asian countries have acceded to the Convention and enacted national laws banning import of waste into their territories. However, despite the legal prohibition, China and India continue to be the world’s largest waste dumping yards. One of the reasons is that there are loopholes in the laws that provide for some exceptions to the ban. For example, in India, the law allows for imports of secondhand computers and laptops if they are intended for donations to educational institutions or NGOs. Due to several arrests for e-waste smuggling recently, the government is considering a complete ban on e-waste.[38] In China, some recyclable wastes such as poly-silicon and artificial fiber continue to be allowed into the country. To prevent the smuggling of prohibited waste, all solid waste imports are required to undergo electronic inspections.[39]

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Challenges to Effective Implementation and Enforcement of E-Waste Management Law Barriers to effective implementation and enforcement of e-waste management laws, like other pollutioncontrol laws, are as follows: 1) lack of specific legal frameworks; 2) complexity of institutional arrangements; and 3) lack of technological, financial, and human resources. Lack of specific legal frameworks. As the above-mentioned case studies show, many Asian countries have yet to enact a specific e-waste management law. Complexity of institutional arrangements. Agencies involved in pollution control are often numerous with overlapping areas of responsibilities as well as gray areas where the institutional responsibilities are unclear or lacking. Lack of coordination is not only an issue between ministries but also between central governments and local authorities, as well as between the administration and the industrial sector. Lack of technological, financial, and human resources. E-waste management requires considerable investments in planning, staff training, purchasing technology, and building new facilities. The laws need to provide funding or economic/fiscal incentives to manufacturers and local authorities in order to allow them to comply with environmental standards. Or else, many would prefer to pay a little extra money to officials who would turn a blind eye on their polluting practices. How to raise sufficient resources for pollution control remains a big challenge for Asian governments. Thus far, taxation is the most commonly adopted method in resource raising (property taxes, sewerage charges, and vehicle taxes, for instance). However, taxation is never popular. International assistance by means of finance, technology transfer, and capacity building is also essential to help improve the capacity of developing countries in addressing pollution problems.[40]

Law on Environmental Assessment Environmental assessment considered herein includes EIA and strategic environmental assessment (SEA). EIA and SEA are related methods with the purpose to prevent, mitigate, and compensate adverse environmental impacts that may be caused by a proposed activity. EIA focuses on projects such as construction of a dam, industry, mining, etc., whereas SEA is applied to policies, plans, programs, and macro projects (in some countries) such as a transportation development plan, energy development plan, international airport project in Hong Kong, etc. EIA regulations were established in most Asian countries in the 1980s and 1990s. In terms of legislation, they are varied ranging from none (Myanmar), to very recent and not widely applied legislation (Laos and Cambodia), to moderately to highly applied legislation (Thailand, Malaysia, the Philippines, Indonesia), and to extensively applied EIA regulation within a broader planning framework (Japan, Hong Kong, South Korea, China).[41,42] Accordingly, EIA has been practiced throughout Asia with varying degrees of rigor and effectiveness. However, SEA is a relatively new method and, thus far, few Asian countries have incorporated SEA into their legislation.[42] In 2006, the World Bank released a report called “Environmental Impact Assessment Regulations and Strategic Environmental Assessment Requirements: Practices and Lessons Learned in East and Southeast Asia.” The report divides countries in the region into three categories: 1) Hong Kong SAR, Japan, and Korea, which have the most advanced EIA/SEA legislations and effective implementation; 2) China, the Philippines, Indonesia, and Thailand, which have less strict EIA/SEA legislations and have encountered difficulties in their implementation; and 3) Vietnam, Mongolia, Lao PDR, and Cambodia, which have recently established EIA/SEA legislations and are at an early stage of implementation. Some countries are exampled for discussion herein with the content is mostly drawn from such mentioned report. China. The EIA system has been in place since 1979, which was the year that China enacted the Environmental Protection Law. This law contained broad elements requiring EIA, particularly for construction projects. Later in 1986, the first legal document on EIA in China was issued by the

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Ordinance of Environmental Protection of Construction Projection (1986). A series of regulations on construction projects were issued, including Environmental Protection Procedures for Construction Project (SEPA, 1990), Regulation of Environmental Protection of Construction Projects (State Council No. 253, 1998), and Environmental Management Catalogue for Construction Projects (SEPA, 1999). To broaden environmental assessment, the current EIA law has been modified and extended to cover plans, and has become SEA inclusive. A new law on EIA was approved by the National People’s Congress in 2002, and has functioned since September 1, 2003. The new EIA law incorporates the concept of SEA for plans and programs, but not for policies. Subsequently, the EIA law covers two large areas: plans and construction projects. Plans are divided into two categories: 1) plans for land use, regional, watershed, and offshore development; and 2) “specific plans,” which include agriculture, industry, livestock breeding, forestry, natural resources, cities, energy, transportation, tourism, etc. However, enforcement in legislation, public participation, and capacity building should be undertaken for applications of policy-and plan-based SEA.[42] Hong Kong SAR. To solve environmental problems encountered in the country, the Hong Kong government has put forth effort to mitigate, control, and prevent such problems. Environmental assessment is considered an important tool in preventing environmental pollution. It is applied not only to individual projects but also to strategic policies and proposal that facilitate the country moving toward sustainability. The EIA process has been applied to projects since 1986, to plans since 1988, and to strategies and policies since 1992. Hong Kong’s EIAO was enacted in 1997 in order to formalize the 15 years experience with EIA, environmental monitoring, and auditing, and came into force in April 1, 1998. The EIA in Hong Kong is considered SEA inclusive. With a successful application of EIA/SEA tools with proven records in legal provision, technical capacity, training and implementation, Hong Kong has become one of the most transparent EIA systems in the world.[42] Japan. The EIA concept was introduced in Japan in the 1960s and implemented through various administrative guidelines, sector legislation (such as the Public Water Area Reclamation Law), and ordinances and guidelines issued by local authorities. The unified law called “the Environmental Impact Assessment Law” was finally adopted in 1997 and took effect in 1999. The law adopts a listing method by scale to identify projects for which environmental impact statement (EIS) is required.[42] Legal requirements that make the Japanese EIA system more strict and comprehensive than those of many other countries in the region are, for instance, EIA requirements for small-scale projects with potential adverse impacts on the environment and emphasis on public participation. Therefore, public opinion is requested at both the scoping stage and the EIA conduction stage, and a period of 100 days is provided for public hearings and information display before submission of EIA report. Although Japan has yet to make SEA into law, there is a strong political will at both the national and regional level to integrate SEA in the policy-making process. The Ministry of Environment and other ministries have adopted SEA guidelines such as the Ministry of Environment’s preliminary guideline on SEA in the formulation of municipal waste management plans. Local governments are also active in implementing SEA. Republic of Korea. The development of the EIA system was initiated in 1977 through the Environmental Preservation Act, and put into effect by the legislation of “Regulations on the Preparation of EIA” enacted in February 1981. After the Environmental Administration was upgraded to the ministerial level in 1990, the previous Environmental Preservation Act was divided into a number of separate laws. One of those is the Basic Environmental Policy Act enacted in August 1990. The Environmental Impact Assessment Act was enacted as a separate law on June 11, 1993, and was put into effect on December 12, 1993. To increase efficiency of the system, the EIA Act was further revised in 1997, and became the EIA law of 1997. The current EIA system is considered as SEA exclusive. Nonetheless, an SEA type of system was applied in the late 1990s known as the Prior Environmental Review System (PERS), which is mainly implemented for various plans and programs. The current PERS has been amended as an SEA type in general, but not to cover policy level. Thus, EIA in Korea includes two types, the PERS conducted at a planning stage and EIA carried out at the project-development stage. In this system, a decision on

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whether to execute a development project will be made at the planning stage, taking into account environmental concerns.[42] Malaysia. The Environmental Quality Act (EQA) was enacted in 1974 as the major federal environmental statute. It was not until 1987 that EIA procedures were introduced under the EQA as a control preventative mechanism. The EIA is well established in Malaysia under the responsibility of the federal government.[43] The situation is currently changing by delegation of EIA powers to the state level. The states of Sarawak and Sabah have adopted independent impact assessment procedures for natural resource management, and it is possible that other states may follow. However, the EIA at state level has faced some problems due to insufficiency of skilled staffs, low institutional capacity, and an absence of effective monitoring of mitigation measures.[43] Therefore, there is a need to strengthen the state capability on EIA implementation. EIA in Malaysia is considered SEA exclusive. However, there is now evidence of an up-and-coming commitment to SEA in the country. Government objectives in environmental protection and management are moving forward, although the regulatory framework to achieve these objectives is not, as yet, fully developed.[44] Some major infrastructure projects, such as roads and power facility development, could be subject to SEA procedures in the future. Integration of SEA into policy, plan, and programs is necessary to secure a more environmentally sustainable development in the country. Indonesia. The development of the EIA system was initiated by Government Regulation No. 29 (1986) in compilation with the provisions of Article 15 of the former Environmental Management Act No. 4/1982. Later, Government Regulation No. 51 (1993) concerning EIA imposed significant revisions to the assessment system. Currently, Regulation No. 27/1999 is a revision of EIA regulation No. 51/1993. The new regulation is expected to be improved and provide a more democratic basis. The EIA system is the responsibility of the Environmental Impact Management Agency. The EIA in Indonesia is project based and SEA exclusive. However, the government has realized the importance of SEA in the decision making process, but its application is not compulsory. The Ministry of Environmental published a book on Strategic Environmental Assessment that provides the fundamentals, procedures, and benefits of applying SEA in the policy, plan, and program process.[42] Philippines. Originally, the EIS system, which is equivalent to the EIA system, was conceived in Philippines Environmental Policy (P, D, No 1151). The actual establishment of the EIA system began with Presidential Decree (P.D.) No. 1586 in 1978. After issuance of some respective decrees, the EIS was adopted in the document DAO No. 30 of 2003 Implementing Rules and Regulations (IRR) for the Philippine Environmental Impact Statement (EIS) System, which was issued in 2003. The EIS system is the responsibility of the DENR. The EIS system is well established in the Philippines, including a legal mandate, administration, procedure, and guidelines. It is regarded as extremely comprehensive and perhaps entails the most stringent requirements in the whole Southeast Asia region.[45] The current EIA system in the Philippines is still project based and SEA exclusive. However, some SEA initiatives have been undertaken. In DAO 30/2003, it was stated that, “The EMB shall study the potential application of EIA to policy-based undertakings as a further step toward integrating and streamlining the EIS system” (Article II, Section 7). The SEA covering policy and plan are being considered to be contained in a new EIA Act.[46] Singapore. Implementation of EIA in Singapore has been operated through the Environmental Pollution Control Act 2000 and the Land Planning process. Regarding EIA through Pollution Control, an EIA may be required for particular projects specified by the Ministry of Environment and Water Resource that have potential to cause pollution affecting public health; for example, petrochemical works, gasworks, and refuse–incineration plants; foreign investment projects using or storing large quantities of hazardous substances; etc. This is likely a project-based EIA. By EIA through Land Planning, Singapore established a document called the Concept Plan, which broadly outlines land-use policies in the country, of which the policies are translated into detailed proposals for local areas called “Development Guide Plans” (DGPs). The basic environmental concerns are considered in the DGPs. This is an SEA-like approach for spatial planning. With the provision of relatively effective laws and

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efficient centralized planning mechanisms, the lack of an EIA law does not appear to obstruct environmental management endeavors. Not only the comprehensive planning and effective pollution control mechanisms but also, more importantly, a stringent enforcement system makes it possible for Singapore to move toward sustainable development.[42] Thailand. The EIA system was established in Thailand through the Improvement and Conservation of the National Environmental Act (1975), followed by the Enhancement and Conservation of Environmental Quality Act (1992). Although the EIA system is well established in Thailand with a good number of qualified personnel, the law remains vague on many issues and public hearing is optional for some projects. The recent EIA procedures for projects identified as having potential adverse impacts to natural resources, environment, and health legally require public involvement through a strict and specific process. EIA for such listed projects is named as Environmental Health Impact Assessment.[47,48] Regarding SEA implementation by law, Thailand has not made SEA mandatory yet. However, SEA has been performed for some projects specified by line agencies for certain areas of interest, such as SEA for Economic Zone at Border Territory at Chiang Rai Province by the Office of Natural Resource and Environment Policy and Planning (2005).[49] SEA legislation is now under development. At this moment, it can be stated that EIA in Thailand is project based and SEA exclusive. Lao PDR. Lao PDR enacted the regulations on EIA back in 2000, and these regulations were revised and upgraded to decree with the Prime Minister Decree on Environmental Impact Assessment No. 112 of 2010. The Ministry of Natural Resource and Environment is responsible for administrating the EIA system, approving EIAs, and for issuing Environmental Compliance Certificates Related legislation on social impacts, including the Decree on the Compensation and Resettlement of Development Projects, 192/PM of 2005, and the Regulations for Implementing Decree 192/PM on Compensation and Resettlement of People Affected by Development Projects. Since the EIA system of Lao PDR is quite recent, it sets relatively high legal standards and requirements, particularly focusing on social aspects. EIA reports in Lao PDR are called Environmental and Social Impact Assessment. Despite a relatively good legal basis, many barriers to effective implementation persist in Lao PDR, particularly the lack of qualified professionals in EIA. Thus, many capacity building programs for EIA personnel have been funded by international organizations and donor countries most prominently the long-term development cooperation with Sweden (Strengthening Environmental Management Phase I and II) and the assistance from the Government of Finland (Environmental Management Support Programme). At this time, EIA is legally implemented as project based and SEA exclusive; however, through the above-mentioned support, the Ministry of Natural Resources and Environment is revising the Environmental Protection Law and developing a decree and guidelines on SEA. This is combined with comprehensive capacity building and case studies on SEA.[42] Cambodia. In 1996, the National Assembly of Cambodia enacted the Law on Environmental Protection and Natural Resource Management (EPNRM) as a framework law governing environmental protection and natural resources management. The law requires the Royal Government to prepare national and regional environmental plans. In addition, there are subdecrees concerning a wide range of environmental issues, including EIAs, pollution prevention and control, public participation, and access to information (SIDA). Cambodia has subsequently established an EIA system under the EPNRM law through the EIA Subdecree on Environmental Impact Assessment issued in 1999, which mandates general requirements, procedures, and responsibilities. The subdecree instructs the Ministry of Environment to formulate implementing rules and guidelines.[50] This EIA system covers only projects and is SEA exclusive. Vietnam. EIA was first mentioned in the Law on Environmental Protection (LEP) 1994 Article 18, which stipulates that organizations and individuals must submit EIA reports to be appraised by the state management agency for environmental protection. The Government Decree on Providing Guidance for the Implementation of the Law on Environmental Protection (Government Decree No. 175/CP, 1994) is an important legal document on EIA in Vietnam.

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The implementation of the EIA system is the responsibility of the Ministry of Natural Resources and Environment. The current EIA system in Vietnam is basically consistent with international practice.[51] SEA has already been adopted conceptually in the Vietnamese legislative framework, for example, in the LEP, GD 175/CP, and Circular No. 490/TT-BKHCNMT, where “EIA not only must be carried out at project level, but also for master plans for development of regions, sectors, provinces, cities and industrial zones.” There are several cases of applying SEA in Vietnam in recent years. As plans are covered by the EIA system, the system is conceptually SEA inclusive. The government is considering accommodating SEA in the new environmental legislation.

Conclusion Development of environmental law in the region began about four decades ago. However, it has been uneven depending on the economic development of each country. According to the level of development of environmental legislation and effectiveness of the implementation, Asian countries may be divided into three categories: 1) countries with advanced economy, comprehensive environmental legislations, capable and well-coordinated institutional framework, and relatively effective implementation, such as Japan, Korea, Hong Kong, and Singapore; 2) countries with a developing economy, relatively wellestablished legislations but ineffective implementation due to problems such as institutional complexity and lack of qualified personnel, such as China, Malaysia, Indonesia, the Philippines, and Thailand; 3) countries that have recently emerged from conflict and are rebuilding, and therefore are at an early stage of developing environmental legislations and institutions, such as Laos, Cambodia, and Vietnam. Environmental legislation in Asia is incorporating more and more elements of the “second generation of environmental legislation,” which are public participation, compliance incentives, and marketbased mechanism. However, regarding the role of the people, in many cases, the law recognizes rights of local communities in natural resource management but does not provide for procedure or institutional framework that would enable the effective exercise of such rights. In addition, environmental legislations in Asia are struggling to keep pace with the rise of the emerging environmental problems and the level of environmental degradation. The problem of e-waste is one good example. For environmental assessment including EIA and SEA, all the countries presented herein have enacted with Environmental Protection Law that include EIA. However, SEA implemented by law is limited. The EIA systems of most countries by law are project based and SEA exclusive. Only some countries such as China, Hong Kong, and Vietnam have environmental assessment system by law for EIA with SEA inclusion. To ensure effective implementation, a good environmental legislation must be complemented with a capable and coordinated institutional framework, qualified personnel, and an adequate budgetary allocation. Regional and international cooperation are of utmost importance in the development and implementation of environmental legislations in Asia, particularly in the areas of capacity building, technical cooperation, and funding.

Acknowledgments This entry could not have been completed without the great help of Miss JinjutaManotham, Second Secretary, Ministry of Foreign Affairs, Thailand, for searching some meaningful environmental legislation-related topics; Mr. Brian J D’Arcy, Environmental Expert, Scotland, U.K.; Mr. Peter Gammelgaard Jensen, Chief Technical Advisor, Grontmij, Denmark, for editing the manuscript; and Miss SriratSuwannakom, my research assistant, for formatting the document. I would like to thank the Research Center for Environmental and Hazardous Substance Management, KhonKaen University, National Excellence Center for Environmental and Hazardous Substance Management for supporting me on the relevant research and academic services. I would also like to thank the Encyclopedia of Environmental Management, Taylor & Francis Group, for giving me the opportunity to write this entry.

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29 Environmental Policy Introduction .................................................................................................. 347 Multidisciplinary Approach to Developing Environmental Policy .......348 Impacts...........................................................................................................349 Air • Water • Land Use

Burden of Proof ............................................................................................ 350 Economic Framework...................................................................................351 Cost-Benefit Analysis (CBA) .......................................................................351 Cost-Effective Analysis

Policy Instruments ....................................................................................... 354 Tax on the Polluting Good • Tax on Emissions • Tradable Emissions Permits • Tighten Liability Rules • Emission Reduction Mandates (Quantity-Based Command-and-Control) • Mandate a Specific Control Technology (Technology-Based Command-and-Control)

Other Policy Issues ....................................................................................... 355 Intrinsic or Nonuse Benefits • Redistributive Effects • Sustainable Development • Renewable Energy Resources

Sanford V. Berg

Conclusions ................................................................................................... 356 References ..................................................................................................... 357

Introduction When economic activity leads to pollution and over-use of common property resources, government intervention can improve social welfare. Pollution involves a market failure in which damages caused by a producer or consumer are imposed on third parties. These damages can involve personal health, the physical deterioration of buildings, and foregone options for the future. Of course, if transaction costs are low, those that are causing pollution damage can be taken to court if the liability rules are clear. Destruction of common property resources such as losing unique ecological habitats, endangering particular species, or destroying valued scenic vistas is another form of market failure affecting the environment. Because there may not be clear property rights to such elements of the environment, these common property resources can be overutilized. Given the lack of well-defined property rights, government enacts environmental laws to address these market failures. However, identifying and quantifying the damage caused by pollution sources or inflicted upon sensitive ecosystems can be difficult, making any determination of the benefits and costs a contentious exercise. Consequently, choosing policies that define the extent of the environmental protection can be both contentious and problematic. The next section describes the multidisciplinary inputs that are incorporated into environmental policy analysis, selection, and implementation. Other topics addressed here include policy impacts, the burden of proof, economic evaluation, and the strengths and limitations of policy options.

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Multidisciplinary Approach to Developing Environmental Policy Environmental economics is the study of how economic and environmental issues interact. Issues addressed by environmental economists include but are not limited to evaluating ways to reduce pollution, analyzing the trade-offs between using renewable and nonrenewable resources, or estimating monetary values for ecosystems or habitat. While no single field of study contains all the insights needed to develop and implement sound environmental policies, the focus here will be on economics because it provides a system for incorporating many perspectives and it is the framework by which environmental policy is designed and evaluated. Depending on the burden of proof, the resulting policies might be excessively stringent (costly relative to their benefits) or inadequate for the protection and preservation of environmental features that affect human health and welfare and have intrinsic value. We know from materials balance that human activity does not create matter but only changes its form, concentration, and location, thus there is a need for physical sciences such as chemistry, physics, and biology to help inform environmental policy. While all societies affect natural systems, the scale of potential impacts has grown with economic development. There is evidence that as incomes rise, citizens are willing to devote relatively more resources to controlling environmental impacts. Moreover, many citizens would like to see much more attention given to reducing current damages and limiting the risks for future harm, hence an understanding of societal and political dynamics is also important for informing environmental policy. The development and implementation of sound environmental policy draws upon information and procedures from many fields of study. Here, economics is utilized as the framework for integrating the concepts, measurements, and values required for the steps: 1. Determine appropriate regulatory objectives (through citizen participation in political processes and community consensus-building) 2. Balance those objectives to determine regulatory priorities 3. Identify and legislate oversight responsibilities for environmental agencies 4. Develop (a) mechanisms for monitoring environmental impacts (such as ambient air and water quality) and (b) methodologies for integrating new scientific understandings of environmental impacts into the policy prioritization process 5. Define the appropriate targets for different types of pollutants and the protection of biodiversity 6. Determine (and then apply) the appropriate policies for meeting objectives 7. Analyze environmental indictors on a regular basis, checking for noncompliance 8. Evaluate the impacts, recognizing potential biases in the measures and the ways impacts are valued 9. Establish an effective process for monitoring and reviewing the framework, including the penalties and sanctions applied when there is noncompliance These steps require input from a number of disciplines that shape the way we see things. Although technical training allows analysts to delve deeply into subjects in a consistent manner, awareness of other disciplines’ perspectives can be important for constructive environmental policy-making, including: • Engineers look to technology for solutions to environmental problems. They are able to incorporate new (often expensive) control technologies into energy extraction, production, consumption (energy efficiency), and pollutant disposal and storage (as with nuclear waste). • Meteorologists and hydrologists analyze pollution transport in air and water systems. They have a deep understanding of the impacts of discharges under different conditions. In conjunction with demographers and epidemiologists, they can estimate the doses received by different population groups. • Medical scientists and toxicologists analyze the dose response relationships for citizen health, conducting exposure and risk assessments.

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• Ecologists study the impacts of pollutants on the local and global environment, assess the value of ecosystem services, and track invasive species and biodiversity. Climate scientists help assess the causes and consequences of changes in local and global temperatures and other weather patterns. • Materials scientists look at damages caused by air and water pollution. The associated impacts include cleaning and painting buildings, treatment costs, and shorter life spans for affected equipment. • Political scientists focus on issues of power, legitimacy, social cohesion, and the roles of different stakeholder groups in influencing environmental policies. Consensus is critical because ultimately, in a democratic system, there needs to be widespread agreement on the desired outcomes if the system is to avoid instability. • Economists emphasize the importance of efficiency in resource allocation. They apply benefit-cost analysis and tend to depend on price signals to provide incentives for the adoption of appropriate control technologies and conservation measures. • Planners deal with land-use and zoning issues, given population growth projections. Planners integrate legal constraints with historical experience, bringing topological, aesthetic, and geographical elements to the analysis. • Archeologists and anthropologists provide insights on the impacts of dams, mines, and their related economic activities on unique historical sites, local populations, and indigenous groups. Such impacts create difficult valuation issues.[1,2] • Lawyers spotlight the institutions of policy implementation. For example, rules and regulations attempt to pay significant attention to procedural fairness. Due process contributes to the legitimacy of outcomes. If different parties perceive that there is no transparency and no opportunity for participation, environmental policy will be perceived as unreasonable and the laws will either be changed or they will be disobeyed in a variety of ways. • Environmentalists advocate sustainability and environmental equity. The by-products of energy production affect public health and have environmental outcomes. Those impacts have economic value, but often that value is nonmonetary or difficult to quantify. For example, generation and transmission siting decisions incorporate impacts on biodiversity and sustainability. • Ethicists help society understand personal values and notions of stewardship. Humans have a clear responsibility to leave future generations with a legacy of sound institutions and a clean environment, though the best means to this end are often not obvious. Thus, physical, biological, and social scientists attempt to uncover patterns and identify lessons to help us improve policy. Given the complexity of environmental issues, most environmental problems are managed, not solved.

impacts Energy production and consumption impact people and the environment in a number of ways. For example, activities can damage ecosystems in the extraction phase (oil drilling or coal mining) or involve cross-media emissions in the consumption phase that can lead to further ecosystem damage. Emissions can be from a single point or a mobile source. In addition, they can be continuous or intermittent (with exposure and impacts depending on wind and other weather conditions or the presence of other chemicals). The transport mechanism can be complicated and involve multiple jurisdictions (as with SO2 and NOx—emissions lead to “acid rain” or ozone problems in downwind areas).

Air Issues range from local concentrations of particulate matter in the atmosphere to concerns over anthropogenic climate change. Consequences for health, ecosystems, agriculture, coastal settlements, species survival, and other impacts make atmospheric change a serious policy issue. For example, long-range

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transport means pollutants cross national boundaries and require coordination. Other pollutants— such as greenhouse gas emissions of CO2—require coordination not due to transport, but because the effects are global in nature regardless of where emissions occur.

Water Effects of contaminants vary in surface waters and groundwater. The United States has primary standards to protect public health (with maximum contamination levels [MCLs] for toxic wastes). Secondary standards and associated MCLs are meant to protect public welfare (for example, ensuring that the taste, odor, and appearance of groundwater do not result in persons discontinuing water use). Other environmental issues include species loss and dealing with nonindigenous, invasive species.

Land Use Siting is an issue for electricity generators, transmission lines, and distribution systems (other aspects of land use include urban sprawl and availability of land for agriculture. The focus here is on the environmental impact of energy systems. For example, social investments in mass transit affect emissions from mobile sources (autos). However, environmental policy addresses many other issues, such as the use of pesticides and fertilizers by agriculture or deforestation). The problem of not in my back yard (NIMBY) is universal: we like the convenience of electricity but do not want its production or transport to affect our own property. Surface coalmines are an eyesore, but restoration can be costly. Hydroelectric dams can affect fisheries, flood unique canyons (causing a loss of scenic vistas), damage ecosystems (as in the Amazon), or displace human populations (as with China’s Three Gorges Project). Solar collection stations and wind generators require space and have impacts on aesthetics. For some, viewing large windmills along the crest of a lovely mountain range is an eyesore. For others, the same scene is a symbol of hope. Environmental policy-makers must be aware of the relationship between changes in impacts in one medium and changes in impacts in other media. For example, reducing airborne emissions of mercury will also lead to reduced mercury concentrations in rivers and lakes. However, it may also be the case that reducing ozone precursors from auto emissions by using methyl tert-butyl ether (MTBE) leads to increasing harm to bodies of water as the MTBE precipitates out in rain. Finally, there may be policy and impact trade-offs that must be evaluated with reducing CO2 emissions through a greater use of nuclear energy. The policy reduces greenhouse gases but raises issues and associated risks of waste storage and protection. In all cases, the links between different environmental media and different environmental policy must be understood for society to properly evaluate the trade-offs.

Burden of Proof Because environmental issues tend to be complex, delays in responding to citizen concerns and new scientific information can lead to negative impacts or a local crisis. What is more problematic: erring on the side of environmental protection or erring on the side of development? When science is unclear or when studies yield conflicting outcomes, the issue of burden of proof arises. Two types of errors are possible. In a Type I error, a hypothesis is rejected when it is in fact true (e.g., deciding that a pollutant causes no health damages when in fact it does). Rejecting the hypothesis of a health link would lead to more emissions (and citizen exposure) than otherwise would be the case. A Type II error occurs when the decision maker fails to reject a hypothesis that is in fact false (e.g., not rejecting the hypothesis that low doses of a pollutant have no damaging side effects for certain types of citizens, such as asthmatics, who are viewed as potentially sensitive to a particular pollutant). If in fact at low doses the pollutant does not have negative health impacts, environmental regulators might have imposed standards that induced costly compliance strategies that were based on the Type II error.

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Dose-response models that do not reject linear functions when the actual relationships are non-linear would fall into this category. Both types of errors have costs. However, the political implications may depend on the type of error, leading decision-makers to prefer making errors that are difficult to detect. Thus, it can be argued that environmental regulators will tend to avoid making Type I errors. When evidence accumulates and shows conclusively that a pollutant has health impacts, those responsible for environmental policy do not want to be blamed for acting too slowly. Furthermore, citizens might prefer excessive caution (labeled a “precautionary bias”). On the other hand, Type II errors can result in regulators imposing high abatement costs onto polluters (and those purchasing associated products) in a manner that is not cost effective. A related issue is whether or not the environmental impact is irreversible. If it is not reversible, a case can be made that the burden of proof should be assigned to those who assert that relatively higher levels of pollution are not problematic. On the other hand, if abatement costs are systematically underestimated and the benefits of pollution reduction are overestimated, it is possible to devote excessive resources to limiting environmental impacts.

Economic Framework Economists are aware that it is difficult to place monetary values on many impacts of pollution but argue that environmental amenities must be balanced against other valued goods and services.[3] Some view economists as overemphasizing the efficacy of market incentives to the exclusion of other instruments. However, because economics offers a consistent framework for integrating insights from other fields, it will be described here.

Cost-Benefit Analysis (CBA) The most fundamental economic analysis looks at how pollution impacts (reflected in “external costs”) cause excessive consumption of polluting goods in the absence of government intervention. These external costs are the negative spillover effects of production or consumption for which no compensation is paid (e.g., a polluted stream that damages the health of those living along the stream). Producers consider the environment to be a free input; hence they only minimize private costs. If these external costs are added to the private costs (reflected in the supply curve), this is the total social cost. Figure 1 shows how a competitive product market yields an equilibrium price ($4) and quantity (80 units per week). However, in the absence of public intervention, the price only reflects the private costs of production, not damages imposed on others (amounting to $2 when the 80th unit is produced, but this is assumed to be less if only 65 units of the good are produced. The external costs are higher at

FIGURE 1

Private costs and external costs.

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higher levels of output presumably because damages rise dramatically when there are very high concentrations of the pollutant in the atmosphere). Determining the extent of those damages requires some valuation metric. For now, let us assume that the analysts “got it right” in estimating both benefits and costs. This is a strong assumption because environmental services are notoriously hard to price. This problem can limit the ultimate effectiveness of CBA because the abatement costs tend to be short-term and quantifiable, but the benefits (avoided damages) are often long-term and difficult to quantify. For now, consider the impacts of environmental regulation within the CBA framework. Regulation requires pollution abatement activity, raising production costs but reducing the pollution and associated damages (as shown in Figure 2). The imposition of environmental regulation raises production costs (shifting the supply curve up) and reduces equilibrium consumption of the polluting good (from 80 to 75 per week) because the price has risen (from $4.00 to $4.40). In addition, external costs are reduced (so the sum of private and external costs is now $5 when 75 units of the good are produced). Emissions are reduced (though this particular figure only indicates the reduction in damages, not the precise reduction in emissions). The next question is how much pollution abatement makes economic sense, since control costs rise rapidly as emissions are cut back towards zero. Continuing with our illustrative example, Default three depicts the total benefits of abatement and the total cost of abatement. The latter depends on the abatement technology and input prices and the interdependencies among production processes (for retrofitting control technologies). It is relatively easy to compute abatement costs from engineering cost studies, although predicting future control costs is not easy because innovations will create new control technologies. The benefits from abatement (or the reduction of pollution damages—the cost of pollution) depend on the size of the affected population, incomes (indicating an ability to pay for environmental amenities), and citizen preferences (reflecting a willingness to pay). The benefits can be very difficult to estimate. Consider, for example, the health benefits of reduced particulates in the atmosphere, habitat values, and citizen valuations of maintaining a habitat for a particular species. Physical benefits can be found from dose-response studies. Various survey methods and market proxies for computing willingness to pay to avoid experiencing the impacts of pollution have methodological problems. However, if the dollar metric is to be used for determining the benefits of environmental improvements, techniques can at least establish rough benchmarks (as shown in Figure 3) (Some argue strongly against the use of CBA.[4]). The total benefits and costs to the marginal benefits and costs can be related because, for economists, the issue is not zero emissions versus unlimited emissions. In the former situation, if 80 units of the good results in 80 tn of emissions, then zero emissions reduced would be characterized as having no abatement costs (but also, no benefits from abatement). When the total benefits equal the total cost of abatement (at 65 tn of emissions reduced per week in Figure 3), the last reductions in emissions were very

FIGURE 2

Reducing external costs.

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costly and the additional benefits were fairly small. Zero abatement activity would also be inefficient in this example because the marginal damages are very high and marginal abatement costs are quite low. Economics tries to determine the optimal amount of emissions. In the hypothetical example, the “optimal” quantity of reduced emissions is about 25, where the marginal abatement cost is just equal to the marginal benefits of $2. These are depicted in Figure 4. This outcome means that there are still 55 tn of emissions per week. If the estimated benefits and costs of pollution abatement are correct in this illustration, economic efficiency would be violated if additional resources were devoted to abatement activity. For example, if 50 tn of emissions were reduced (so only 30 tn of the pollutant are released), the marginal benefit would be about $1, but the marginal cost would be greater than $5. From the standpoint of economic efficiency, those resources could be used to create greater value in other activities. Of course, the difficulty of obtaining a common dollar metric for all the impacts of different pollutants means that benefit-cost analysis must incorporate a range of values. The range could be incorporated in Figure 4 as a band around the marginal benefit curve indicating one standard deviation from the calculated values. A conservative approach would recognize that the marginal benefit function could be above that depicted in Figure 4, which would lead to optimal emission reduction of more than 25 tn per week (improving the ambient air quality). Further complicating the analysis are production and exposure interdependencies. For example, the marginal cost of abatement associated with one type of emission may depend on the level of treatment (or abatement) for another contaminant. A joint optimization problem results, with the basic principles

FIGURE 3

Total benefits and total cost of abatement.

FIGURE 4

Marginal benefits and marginal cost of abatement.

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unchanged. Many investments in abatement equipment have this characteristic: once one set of contaminants is being reduced from a discharge flow and the cost of dealing with additional contaminants can be relatively low. For example, in the case of water discharges, if iron or manganese is removed via the precipitation method, total dissolved solids (TDS) is reduced and there may be an improvement in water clarity. Interdependencies can also arise on the benefit side when the dose-response relationship for a particular contaminant is influenced by the presence of other contaminants. Again, in the case of secondary groundwater standards, perceptions of odor and color will be affected by whether or not they occur in combination. Such considerations must be factored into the analysis when comparing the benefits and costs of different treatment options.

Cost-Effective Analysis Instead of trying to estimate the dollar benefits of saving a human life (or reducing the incidence of asthmatic attacks), one can compare the number of lives saved per dollar spent in abatement activity across programs. Thus, cost-effective analysis involves finding the least-cost method of achieving a given economic or social objective such as saving lives or retaining unique ecological settings. No dollar value (or explicit measure of avoided damages) is placed on that objective.[5] One advantage of this approach is that the focus is on minimizing the cost of meeting the (politically determined) target. It promotes consistency across a range of programs that might be designed to address a particular problem, whether that involves health impacts or a loss of habitat. Cost-effective analysis facilitates comparisons across programs, leading to reallocations of resources devoted to meeting such targets as new information is gathered over time.

Policy Instruments Political systems have passed legislation and created agencies to apply laws to improve environmental performance. For example, in the United States, the Water Pollution Control Act of 1956 and the Clean Air Act of 1963 and subsequent amendments to both pieces of legislation have focused on achieving ambient standards. The U.S. Environmental Protection Agency is responsible for implementing these laws, and in other nations agencies have also been established to reduce emissions and improve environmental outcomes. A number of policy options can lead to emission reductions.[6,7] These instruments have different economic efficiency implications. In addition, some of these approaches are difficult to implement (due to being information-intensive), some are not cost effective (in that other approaches achieve the same outcome at lower cost), and the distributional implications can differ across these approaches (tax burdens differ or some groups obtain valuable assets).

Tax on the Polluting Good An excise tax could be imposed on the good, cutting back consumption to 65 units per week (Figure 1). Of course, the problem is not with the product but with the emissions associated with its production. Thus, this option does not provide incentives for developing new technologies that reduce abatement costs—it represents a static approach to the problem because it does not promote technological innovation.

Tax on Emissions A penalty or charge for each ton of emissions would lead suppliers to cut back on emissions—to the extent that the abatement is less expensive than the tax. Thus, in Figure 4, a tax of $2/tn would lead to the optimal reduction of pollutants. In addition, it provides incentives for innovation in the control

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technology industry. Firms will seek ways to reduce abatement costs, thus reducing their pollution taxes. This strategy is likely to be opposed by polluters who will be passing the taxes on to customers (where the ultimate incidence depends on supply and demand elasticities in the product market).

Tradable Emissions Permits The same result (and incentive) is obtained if “allowances” of 25 tn are allocated to polluting firms, limiting emissions (the situation is not completely identical—a tax has certain costs to firms but yields uncertain overall abatement because regulators will not have precise estimates of abatement costs; the allowances have certainty in terms of overall abatement but uncertain cost. Of course, with monitoring, the tax can be varied over time to achieve a desired ambient condition). This approach provides an incentive for those with low abatement costs to reduce emissions and sell their permits (allowances) to others whose abatement costs would be very high. This places entrants at a disadvantage because incumbent firms are “given” these valuable allowances. The SO2 regime in the United States has this feature. Of course, the initial allocations raise political issues (because permits represent wealth). In establishing a tradable permit regime, an environmental agency must determine the allowed level of emissions (here, 25 tn) and whether additional constraints might be applied to local areas with particular circumstances. In addition, the energy sector regulator has to make decisions regarding the treatment of cost savings from the regime. For example, savings might be passed on to consumers or retained by firms. The latter situation provides an incentive for utilities to participate in the emissions trade markets. A sharing plan can also be utilized so customers benefit as well.

Tighten Liability Rules An alternative approach would utilize a court-based system, where fees would be assessed against those responsible for damaging the health of others, for reducing the economic value of assets, or for reducing the amenity values of ecosystems. Of course, this approach requires a well-specified set of property rights and clear causal links between specific emitters and affected parties. The transaction costs of such a system (resources devoted to negotiations and legal activity) could be prohibitive for many types of pollutants.

Emission Reduction Mandates (Quantity-Based Command-and-Control) Although equal percentage cutbacks sound “fair,” this strategy is not cost-effective because abatement costs can differ widely across pollution sources. If there are scale economies to emission reductions, it would be most efficient to have a few firms reduce emissions. The least-cost way to achieve a given overall reduction in emissions will involve differential cutbacks from different firms.

Mandate a Specific Control Technology (TechnologyBased Command-and-Control) This “command and control” strategy is not cost-effective because production conditions and retrofitting production processes differ across firms (based on the age of the plant and other factors). However, this policy option has been utilized in a number of situations as a “technology-forcing” strategy.

Other Policy Issues The above instruments have been utilized in different circumstances. Additional issues include intrinsic benefits, income distribution, sustainability, and renewable resources.

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Intrinsic or Nonuse Benefits Some people take a more expansive view of environmental amenities as they attempt to separate economic values from inherent values. However, this might be partly accounted for in terms of the perceived benefits to future generations. Intrinsic benefits from environmental programs include option values, existence values, and bequest values.[8] The first value represents a form of insurance so future access to a potential resource is not eliminated due to current consumption. The rationale behind option values is closely related to the “margin for error” argument noted earlier. Existence values reflect a willingness to pay for the knowledge that the amount of contaminant in the environment does not exceed particular levels or that a particular species (or level of biodiversity) is retained. The resource or ecological system is available for others. The bequest values can be interpreted as the willingness to pay for preserving a resource (or a geographic site) for future generations.

Redistributive Effects It is important to note that citizens being harmed by emissions are not necessarily the same as those who are consuming the polluting good (such as electricity). Even if a particular program has positive net benefits, some parties are likely to be losers. They are seldom compensated and left better off, raising concerns about the distributional consequences of alternative policies. Furthermore, those harmed may have lower incomes (and thus, a lower willingness to pay to avoid damages due to the lower ability to pay). This point underscores the role of fairness as a factor that might outweigh efficiency considerations in some circumstances. Some agencies have been forbidden to use CBA on the grounds that the numbers are too speculative and that social concerns should be given priority. Intergenerational concerns can be interpreted as reflecting redistributive considerations.

Sustainable Development Some of the issues associated with energy involve the use of nonrenewable resources (irreversibility). Some citizens argue that sustainability requires development that can be supported by the environment into the future. These people wish to ensure that resources are not depleted or permanently damaged. However, since sustainability depends on technology and innovations change resource constraints, defining the term with precision is quite difficult.

Renewable Energy Resources Generating electricity without fossil fuels (e.g., hydro, wind, solar, biomass) is sometimes referred to as using green options. Green options are often limited in the amount (and reliability) of energy produced in a given time period. Utility applications for renewable resources include bulk electricity generation, on-site electricity generation, distributed electricity generation, and non-grid-connected generation. A number of regulatory commissions have required utilities to meet renewable portfolio standards. Such strategies reduce dependence on a particular energy source (to reduce the region’s vulnerability to supply disruptions or rapid price run-ups). In addition, such requirements imply that managers are not making the most efficient investments in long-lived assets. Also, note that demand reduction through energy-efficient technologies is a substitute for energy, whatever the source.

Conclusions The three main trends in environmental regulation in recent years have been shifting from commandand-control regulation towards a greater use of economic instruments (such as emissions trading), seeking more complete information on the monetary value of environmental costs and benefits, and a tendency for addressing environmental objectives in international meetings, as with the Kyoto Protocol.[9]

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The interactions between economic and environmental regulation raise important policy issues. If energy sector regulation and environmental regulation remain separate, some means of harmonization may be necessary to promote improved performance. Collaboration would involve clarifying the economic duties of the environmental regulator and the environmental duties of the economic regulator. To avoid regulatory competition, agencies sometimes establish task forces or other mechanisms for identifying and resolving issues that might arise between jurisdictional boundaries (across states or between state and federal authorities). Such cooperation can serve to clarify the division of responsibilities and identify regulatory instruments that will most effectively meet economic and social objectives. In summary, policy-makers respond to domestic political pressures by devising institutions and instruments to address pollution and environmental sustainability.[10] Although no single field of study contains all the tools necessary for sound policy formulation, economics does provide a comprehensive framework for evaluating the strengths and limitations of alternative policy options. Because of the pressures brought to bear by powerful stakeholders, adopted policies and mechanisms are not necessarily cost minimizing. The resulting inefficiencies may partly be due to considerations of fairness, which places constraints on whether, when, how, and where environmental impacts are addressed. As emphasized in this survey, citizens want to be good stewards of the land. We appreciate the adage: “The land was not given to us by our parents; it is on loan to us from our children.” How to be good stewards— through the development and implementation of sound environmental policies—has no simple answer given the complexity of the issues that need to be addressed.

References 1. Maler, K.-G.; Vincent, J.R. In Handbook of Environmental Economics: Environmental Degradation and Institutional Responses; North-Holland: Amsterdam, 2003; Vol. 1. 2. Maler, K.-G.; Vincent, J.R. Handbook of Environmental Economics: Valuing Environmental Changes; North-Holland: Amsterdam, 2005; Vol. 2. 3. Viscusi, W.K.; Vernon, J.M.; Harrington, J.E., Jr. Economics of Regulation and Antitrust; MIT Press: Cambridge, MA, 2000. 4. Ackerman, F.; Heinzerling, L. Priceless: On Knowing the Price of Everything and the Value of Nothing; New Press: New York, 2004. 5. Freeman, A.M., III The Measurement of Environmental and Resource Values: Theory and Methods, 2nd Ed. Resources for the Future: Washington, DC, 2003. 6. Portney, P.; Stavins, R. Public Policies for Environmental Protection; Resources for the Future: Washington, DC, 2000. 7. Vig, N.; Kraft, M.E. Environmental Policy and Politics: New Directions for the Twenty-first Century, 5th Ed.; Congressional Quarterly Press: Washington, DC, 2003. 8. Krutilla, J. Conservation reconsidered. Am. Econ. Rev. 1967, 57 (4), 777–786. 9. Busch, P.-O.; Jörgens, H.; Tews, K. The global diffusion of regulatory instruments: the making of a new international environmental regime. In The Annals of the American Academy of Political and Social Science, Sage Publications: Thousand Oaks, CA, 2005, Vol. 598, 146–167. 10. Kraft, M.E. Environmental Policy and Politics, 3rd Ed.; Pearson/Longman: New York, 2004.

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30 Environmental Policy: Innovations Introduction .................................................................................................. 359 Environmental Policy Innovations............................................................. 359 The Nature of EPIs: Conflict to Collaboration ......................................... 361 Resources, Needs, Politics, and Other Determinants .............................. 363 Need/Problem Severity • Institutional Factors: The Importance of Commitment and Capacity • Interest Group Support • Regional Diffusion

Alka Sapat

Conclusion ....................................................................................................366 Acknowledgments ........................................................................................ 366 References .....................................................................................................366

Introduction Over the past three decades, American environmental policy management has often entailed the search for new solutions and innovative ways of managing environmental problems. These solutions have taken the form of environmental policy innovations (EPIs), which have frequently been pioneered and adopted by local and state entities, at times with the help of citizen groups and non-profit organizations. They have become an important means of managing environmental problems, particularly for intractable, multimedia, multijurisdictional environmental issues. While policy innovations have been studied in a number of areas at the state level,[1–19] less attention has been paid to policy innovations in environmental management.[20–24] The focus of this entry is to shed light on these environmental policy initiatives, termed here as EPIs. To provide an understanding of these policy initiatives, I begin by defining and discussing EPIs and supply some understandings of the concept. Next, I discuss the nature and type of a large majority of EPIs; in particular, I focus on their reliance and inclusion of collaborative forms of environmental management and provide some examples. The fourth section of the entry analyzes some of the main factors affecting the adoption of EPIs, including determinants such as institutional commitment, resources, the severity of the problem, and the role of interest groups. The effect of neighboring entities in spurring innovation adoption and diffusion is also considered in this section. The entry ends with a conclusion that summarizes the main points of the entry, provides some considerations for further research, and discusses the policy implications of adopting environmental policy initiatives as a strategy for managing the environment.

Environmental Policy Innovations EPIs here are defined in this entry as government-sponsored initiatives to protect the environment. This is based in part on the original definition of a policy innovation originally espoused by Walker[1] who examined the adoption of new policies by states, which he termed as “innovations.” Walker’s work also 359

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differentiated between states that were “leaders” and those that are deemed to be “laggards,” along with examining the diffusion or the spread of these innovations from leading states to other states. Thus, he analyzed the amount of time that elapsed between the invention of an innovative program and a state’s adoption of it. Based on 88 different state programs, he concluded that larger, wealthier, and more industrialized states adopt new programs more rapidly than do small, less-developed states. He also found strong regional relationships among states that were thought to reflect conditions of emulation and competition among proximate states. Walker’s research lead to further investigation into state policy innovations. Gray[3] undertook a more detailed analysis and rejected Walker’s contention that leading states lead no matter what the policy area; in a study of education, welfare, and civil rights policies, Gray found that innovativeness was not a pervasive factor. She concluded that based on the issue and the era, different states would comprise the leading, “middle-adopting,” and lagging clusters. During the next two decades, there were numerous other studies of state innovation that were published in a variety of policy areas. These included studies of innovation in juvenile corrections,[12] consumer affairs,[13] technology,[14] energy,[15] tort law,[16] judicial administration,[17] and human services.[18] These studies explored various determinants of innovativeness in different policy areas. However, while they expanded the scope of policy areas subject to innovation analysis, they did not lead to major advances in the conceptualization of state innovations or in the empirical approaches to its investigation. In the 1990s, however, significant advances were made in the field of state innovations research. Research on state innovation was tested with sophisticated empirical methods in the area of state lottery adoptions[4] and state taxation.[5] In both these studies, the authors extended theories of innovation to different policy areas and made a substantial methodological contribution to the literature by refining the empirical analysis of state innovation with the employment of event history analysis (EHA), a pooled cross-sectional time-series technique. Berry and Berry[4,5] also analyzed the importance of internal determinant explanations (which posit that the factors causing a state government to innovate are the political, economic, and social characteristics of a state) and regional diffusion explanations (which point toward the role of policy adoptions by neighboring states in prompting a state to adopt). An interesting twist on the diffusion aspect of state innovations was presented in a study of state living will laws.[19] In this study, Glick and Hays argued that as policy innovations diffuse across states, a process of “policy reinvention” occurs.[19] Policy innovations, in their view, are not adopted in their original form; rather, they are changed or selectively adapted as innovations by other states. The contribution of this study lay in refinements that added to earlier understandings of innovation diffusion. Another variation on state innovation research was the study undertaken by Berry.[6] In this study, Frances Stokes Berry[6] extended the original research undertaken by Berry and Berry[4,5] to state innovations in strategic planning by various regulatory agencies. This extension was interesting as it was one of the first studies to empirically evaluate an innovation adopted by state bureaucracies as opposed to legislatures. This research was followed by more sophisticated understandings of the catalysts and actors behind the adoption of innovations; in his analysis, Mintrom[7,8] focused specifically on the importance of policy entrepreneurs at the state level. Using both secondary data sources and primary data gathered from a 50-state survey of state education officials in an EHA, Mintrom examined both the consideration of school choice and its adoption by states across time. He focuses in particular on the role of policy entrepreneurs by developing a theory of policy entrepreneurship and testing that theory with respect to school choice. Using EHA, he concluded that policy entrepreneurs act as catalysts in the policy innovation process by promoting ideas for dynamic policy change, persuading legislators, and networking with others in government. In the area of environmental policy, state policy innovations has been the subject of research by Rabe,[20–23] Sapat,[24] and others. For instance, in a comparison of Canada and the United States, Rabe[21] found that despite the relatively decentralized Canadian institutional framework and more centralized and “bureaucratic” framework in the United States, the latter had been much more active and effective in devising innovative policy approaches. Comparing four U.S. states and four Canadian provinces,

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he concluded that within those sample jurisdictions, the United States was clearly ahead of Canada in cross-media permit integration, pollution prevention, disclosure of information on toxic substances, and achieving greater refinement and use of environmental outcome indicators. Similarly, in later research analyzing state experimentation in the realm of climate change, Rabe[22] analyzed the factors prompting successful state initiatives in reducing greenhouse gases. He finds that some states did succeed in cutting through traditional partisan divides and that agent- based entrepreneurs helped to develop policy ideas and form viable coalitions. State EPIs adopted by administrative agencies (rather than by legislatures) have also been the subject of past research. In a study of state policy innovations adopted by agencies in the areas of hazardous waste and groundwater contamination, Sapat[24] found that state agencies are more likely to adopt innovations to deal with problems created by hazardous waste contamination than for groundwater contamination. Further, she finds that state environmental managers are not directly influenced by interest groups and that the inclusion of all stakeholders is likely to lead to greater support for new policy initiatives. EPIs as defined here are drawn from and are based on this past research. An EPI as defined here can be one that is initiated by an agency or government institution (such as a state legislature, county government, or city) or it can be one that is borrowed from another government entity/ institution. Thus, EPIs can be adopted by legislatures or administrative agencies. Also, as defined in prior research by Sapat,[24] EPIs are defined as initiatives that protect the environment. The emphasis on the positive aspects of EPIs is important, because initiatives can have both positive and negative benefits. The term EPI as used in this entry is positive, i.e., it refers to laws, regulations, and policies that are beneficial to the environment. In other words, the goals of these EPIs are to regulate pollution problems and protect the environment. At this point, a caveat is necessary. Clearly, policy choices can be subjectively interpreted; they may be viewed as being positive for the environment by some and as bad for the environment by others. This is particularly true within the realm of environmental politics.[25–27] For instance, the use of economic incentives, such as fees, pollution taxes, or offsets, is viewed as being beneficial by some and negative by others.[25] Even environmental organizations are divided on this issue. For instance, large environmental interest groups such as the Environmental Defense Fund, which utilize the services of economists and scientists, view economic incentives as being beneficial for the environment.[28] However, a number of smaller environmental interest groups are more suspicious about the environmental benefits of adopting market-based incentives.[29] Moreover, the problem of clearly defining an environmentally “positive” policy solution is also compounded by the fact that for some environmental issues, even scientific experts cannot come to an agreement on what the best technical solution is to a pollution problem.[30] Hence, there is extensive debate about the benefits of particular environmental policies, because of conflicting perspectives regarding policy outcomes and because of incomplete and inconclusive technical knowledge of solutions to environmental problems. To overcome some of these problems, the EPIs discussed here are those that are deemed as being beneficial for the environment by the United States Environmental Protection Agency (USEPA).[31,32] These assessments are arguably less subjective than those carried out by environmental or industry interest groups.

The Nature of EPIs: Conflict to Collaboration Following the definition given above, EPIs can be categorized in various ways. For instance, they may be categorized according to their function; the USEPA provides a portfolio for instance, in which various environmental innovations are characterized by the core agency function that they address, such as helping in improving service delivery, enhancing regulatory outcomes, supporting superior environmental performance, designing targeted geographic solutions, etc. While it is useful to categorize environmental innovations thus, other ways of classifying environmental innovations may provide a broader and deeper understanding of these initiatives.

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In this entry, I contend that rather than categorizing EPIs purely by function, it is also useful to classify such innovations by the underlying regulatory and managerial premises of such policy innovations. More specifically, it is instructive to understand the extent to which such innovations rely on collaborative as compared to non-cooperative, more conflict-based approaches to environmental management. This perspective would be useful because there has been growing evidence in both practice and theory about the usefulness of collaborative approaches to managing the environment and natural resources. Reasons for the emergence of these collaborative approaches have been many and varied, but the primary reasons for the growth in collaborative forms of environmental management have been ascribed to the problems stemming from command and control measures.[33–39] These traditional forms of regulation have often resulted in failure to achieve key goals and/or protracted conflict. Traditional forms of environmental regulation that rely on bureaucratic, adversarial, and often technology-based regulatory approaches are also seen as being far too rigid and, more importantly, ineffective, in that they fail to address multimedia and multijurisdictional environmental hazards, particularly those stemming from non-point pollution.[33,36] Furthermore, traditional hierarchical approaches do not allow for more democratic forms of participation and often rely on unrealistic models of individual and administrative rationality.[37,40] To overcome such problems, a search for alternative solutions to managing environmental problems began to evolve and grow. For instance, a study by John in 1994[40] found that states and local governments were increasingly turning away from “top–down command-and-control” regulation to a bottom–up style, which he termed “civic environmentalism.” Civic environmentalism relies primarily on the use of non-regulatory tools, which encompass market-based incentives to regulation. Civic environmentalism or collaborative environmental approaches or similar interchangeable terms that can be used to describe a more cooperative, less adversarial form of environmental management are to be found in a number of EPIs that are undertaken by environmental agencies at different levels of government. Indeed, some EPIs were initially pioneered and adopted because of the failure of conflict-based approaches to manage environmental problems. For instance, environmental permitting processes engender a high level of stakeholder frustration and conflict due to permitting backlogs, long lead times, costs, and uncertainty. To improve these processes and reduce conflict and backlog times, some state and local agencies adopted EPIs to shift away from media-specific permitting for individual facilities and also to improve internal agency permitting processes. For instance, Massachusetts shifted away from a facility permitting approach to a multimedia, sectorbased regulatory approach, targeting sectors with large numbers of small sources, as an alternative to facility-specific state permits with industry-wide environmental performance standards and annual self-certifications of compliance.[31,41] Similarly, other EPIs take a collaborative approach by helping to enhance partnerships to resolve environmental problems that cannot be effectively solved without the participation and collaboration of multiple actors. Leveraging such partnerships enables environmental agencies to solve complex problems by harnessing energies and resources of stakeholders to achieve mutually desirable outcomes. An example of an EPI that does this is the effort undertaken in South Carolina to reduce neighborhood contamination. This initiative enlists the support of numerous community organizations and local businesses to provide education and outreach to reduce community exposure to lead and other hazardous substances as part of the Charleston–North Charleston Community-Based Environmental Partnership (CBEP).[42] Other examples of collaboration-based EPIs[43] are provided in Table 1. To summarize, EPIs are initiatives adopted to protect the environment and a number of them have been adopted at the local, state, and federal level by both legislative and administrative agencies. These EPIs are characterized by more collaborative approaches to environmental management. To understand EPIs further, we can try to ascertain which factors are likely to affect the adoption of these innovations. It is to this issue that I turn to in the next section.

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Examples of EPIs

Neighborhood Contamination Reduction—South Carolina Enlists numerous community organizations and local businesses to support education and outreach to reduce community exposure to lead and other hazardous substances as part of the Charleston–North Charleston Community- Based Environmental Partnership (CBEP). (http://www.epa.gov/Region4) Northeast Ohio Initiatives—Ohio Responds to regional economic, sprawl, ecosystem and infrastructure challenges through a 15-county, communitybased approach. (http://www.epa.gov/glnpo/lakeerie/leneohio.htm) Urban Environmental Program—Boston, Massachusetts Adopts a community-based approach including the city of Boston, Massachusetts DEP, and community organizations to improve the quality of life in urban settings by targeting issues such as asthma and indoor air quality, lead poisoning, vacant lots and green spaces, and pollution prevention. (http://www.epa.gov/boston/eco/uep/boston/) Pollution Complaint Response—Indiana Coordinates an agency-wide, multimedia response to citizen inquiries and complaints using Web-based information, enabling the agency to reduce costs and increase public trust. (http://www.in.gov/idem/5274.htm) Ford Good Neighbor Dialogue—Illinois Brings together stakeholders, academics, and agency representatives in a collaborative process to periodically discuss a large manufacturing facility’s environmental management and performance. (http://www.delta-institute.org/) Community Environmental Awareness Project—Michigan Develops an approach to improve the way environmental information is presented and made available to the public; the goal of the CEAP is to improve the public’s access to and understanding of how major industries are performing under environmental laws and regulations. (http://www.deq.state.mi.us/ceap)

Resources, Needs, Politics, and Other Determinants The factors that lead to the adoption of EPIs can be many and varied. Prior research on innovation adoption in other areas, as well as on EPI adoption is useful here. The following factors are some of the most influential determinants of innovation adoption as seen in extant research.

Need/Problem Severity Keeping in with the adage that “necessity is the mother of invention,” researchers have traditionally regarded problem severity as a significant influence on the adoption of innovations.[1,3,44,45] For innovations adopted at the state and local level, the expectation is that states and local governments rather than the federal government are likely to understand local problems more clearly and are hence likely to be responsive to the needs and problems present in their jurisdictions.[27,45] For instance, in a study of state innovation, ranging from state testing on teacher competency and rail passenger service to state regulation on sodomy, Nice[45] found that the problem environment was prominent in explaining five of eight policy innovations analyzed in this study. For EPIs, the severity of the pollution problem itself has been regarded as one of the most important reasons for innovation adoption. To deal with intractable environmental problems, several entities (local, state, non-profit, and even private institutions) have come up with new ways to solve old problems.[20,31]

Institutional Factors: The Importance of Commitment and Capacity While necessity can often drive innovation, institutional commitment and capacity have also been found to be important for the adoption of policy innovations.[24,40,46] The theoretical origins of the role of institutional actors may be found in ideas related to roles of institutional elites and theories of institutionalism and neo-institutionalism. Elites: The importance of elites in the policy-making process is theoretically supported by the elite perspective of policy analysis. Elite theorists argue that power is concentrated in the hands of elites who use

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the resources of their respective organizations to manage and impose order on society.[47] Societal stability, according to this perspective, rests not on a common political culture and a set of values, but on a forced consensus created and reinforced by the elite. Pluralist politics can coexist with the governmental elite; however, key decisions regarding policy are made by elites. Popular and electoral politics are, for the most part, mainly symbolic and concerned with middle-level policy issues, according to the elite perspective.[47] The elite theory of the policy process is close to the neopluralist view: neo-pluralists, such as Lindblom,[48] challenged the pluralist notion that power was diffuse and argued instead about the privileged position of business. Similarly, others such as E.E. Schattschneider pointed out, “the flaw in the pluralist heaven is that the heavenly chorus sings with a strong upper-class accent.”[49] Institutionalism and Neo-institutionalism: The motivation of institutional actors and elites is also stressed in theories that emphasize the importance of institutions in policy and governance.[50–54] Institutional theories hold that government actors can act independently of interest group pressures and other factors. According to this view, government actors are not just simply “pawns” of various interest groups; rather, the perceptions and attitudes of these actors shape the way they process information and affect independently the choices they make.[52,54] Institutional theories also recognize that informational constraints and computational limitations of political actors prevent actors from making purely “rational decisions” that are independent of the actor’s subjective representation of the decision problem.[52,55] Thus, the attitudes and ideological views of institutional actors can influence their choices in innovation adoption. Based on the theories discussed above, it is likely that pro-environmental policy actors will likely push for adoption of EPIs. Related to that issue is the importance of resources. Adoptions of EPIs require both the motivation and the commitment of key institutional actors in terms of time, money, and other resources. Typically for most policy adoptions, resources are required. For this reason, wealthier and larger states and local governments are more likely to be innovation adopters.[11,56] Resource provision by third- party institutions, such as non-profits, or by higher levels of government, such as by the federal government to state and local governments, can also help spur innovation adoption and diffusion. For instance, the USEPA provides State Innovation Grants to spur innovation adoption; these grants are typically provided to states with projects in three areas: environmental results programs, environmental management systems and permitting, and environmental leadership programs such as EPA’s Performance Track.[31] However, while institutional resources may be required for innovation adoption, some EPIs are adopted to economize on resources or to raise revenues. For instance, EPIs in pesticide regulation adopted by certain states helped raise fees, adding to revenue sources.[24] The role of resources or the importance of economic factors can also be played out in other ways. That is, objectives of economic development by states could deter their willingness to adopt strict environmental regulations.[57–60] For instance, states with a higher percentage of economic activity relying on manufacturing or industrial activities with negative environmental consequences are less likely to favor the adoption of innovative laws and programs that would negatively affect revenue-generating industries. This raises the issue then of interest group support and its importance for EPI adoption.

Interest Group Support Interest groups have always played an important role in the American policy process, due in part to constitutional provisions for pluralism. Theoretically, the pluralist perspective of policy making argues that public policy is the product of democratic participation by individuals who are represented by organizations such as interest groups. Theories of interest group influence on legislators and bureaucrats range from narrow views of interest group influence[61–63] to those that posit interest groups as not exercising any more influence over regulatory policy than other actors or bureaucrats.[64,65] Pluralist theorists, such as TrumanZS and Dahl[67] argued that group exchange (pluralism) was the dominant (and desired) method of political decision making. In their view, this was possible because the decentralization of American institutions and the relative openness of the political system guaranteed competing groups access and some degree of power. Pluralism thus presents a view of the political

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system in which multiple centers of power compete to shape policy. Power is diffuse, fluid, multifaceted, and dispersed among a number of groups. Political participation is largely a goal- oriented activity in which citizens take part in order to obtain some benefit from the government.[68] Pluralist group theory was questioned by a number of scholars, mainly neo-pluralists, who argued that dominant groups and interests had the ability to control the political agenda, preventing pluralist discourse over a full range of policy options.[47,48,69] Group theory was also evaluated in part in analyses of regulatory policies.[62,70,71] A number of scholars argued that regulatory policies emerged from a political equilibrium produced by coalitions of regulated industries and their customer groups, leading to “agency capture.”[62,72,73] This narrow interest group perspective is based on the assumption that certain groups of people, who are organized and powerful in terms of possessing economic resources, will have the capability to dominate policy at the subnational level. While this view is compatible with the Madisonian perspective of private parochial interests modifying policy outcomes, it has been modified considerably to account for characteristics of interest groups such as size and density that could affect their capabilities to exert influence over state regulations.[74] Moreover, the motivations and actions of other political actors, bureaucrats, and other interest groups themselves have also been found to be important in influencing policy outcomes.[75] Further, the narrow view of interest groups adduced by scholars such as Stigler[62] and Posner[73] have been more successful in explaining economic regulation as opposed to social regulation, such as consumer protection and regulation of the environment. The traditional “iron triangle” or “agency capture” theory does not adequately account for the emergence and power of newly organized and motivated consumer, environmental, and other “public interest” groups, which seek social benefits. With respect to environmental policy, past studies have found environmental interest groups to be important in influencing policy outcomes. For instance, Ringquist[76] and Hird[27] analyzed the influence of environmental interest groups that countered the regulatory priorities advocated by industry groups. Ringquist[76] found environmental interest groups to have a significant influence on state water quality regulation in a group influence model of state policy influence. With regard to EPIs, the presence and activities of environmental interest groups are likely to be important as well. However, industry interest group opposition to EPIs may be tempered by the nature of these policy innovations. As discussed above, a number of EPIs rely on collaborative nonadversarial approaches to solve environmental problems. Past research has also shown that some EPIs provide new ways to work with such industries to improve compliance with regulations and do not burden them with any additional regulatory costs.[24] Interest group opposition may also be mitigated by public support for certain EPIs, which provide broader benefits than they do costs to the public at large.

Regional Diffusion Previous empirical studies on the diffusion of innovations have found that external influences by other neighboring entities are important. For instance, adoption of innovations by neighboring states can affect a state’s adoption of a policy innovation.[1,4,11,56,77] Policy adoptions by neighboring states can decrease the information costs regarding the possible consequences, including the electoral consequences, of the adoption of a policy.[4] State officials can view the relative success/failure of programs in these adjacent states and decide whether such programs are suitable for their own state. Moreover, if a policy has been successful in a neighboring state, then state officials, in particular state legislators, can, by using the experiences of neighboring states as an example, boost public and legislative support for similar legislation in their own state. This allows them then to selectively utilize the experiences of neighboring states for their own political gain. The decrease in information costs and political uncertainty can thus increase the motivation of state policy makers to initiate or adopt a policy innovation.[24] While early research on policy innovation diffusion focused on regional influences primarily with respect to policy innovations adopted by states,[8,11,24] later work has also studied factors affecting diffusion at the country level[78] and the local level.[11] On a related note, there has also been research on

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examining the mechanisms of policy diffusion,[79–82] which has moved away from looking simply at the effects of neighboring states to analyzing the more complex and multiple factors affecting diffusion, such as the presence of competition, learning, imitation, and coercion; this research has found that policy mimicry is often tempered by the size of the state and by the presence of other third-party institutions that may aid diffusion.[11] Thus, while regional effects or “external determinants” may be important, it is also necessary to understand the mechanisms of diffusion, which may be varied and complex. There are thus, a number of factors that can influence the adoption and implementation of EPIs; need or the type of problem alone may be a necessary but not a sufficient determinant of innovation adoption. Institutional commitment and capacity, the role played by interest groups, and the effects of innovation adoption by neighboring institutions and entities are also important influences affecting the espousal of EPIs.

Conclusion Complex environmental problems require innovative solutions for effective management. The focus of this entry has been on such initiatives or EPIs. In doing so, this entry began by defining and explaining the nature of such initiatives, including a discussion on past research on state policy innovations. I contend that a number of environmental initiatives have pioneered and are symptomatic of a shift in environmental management to more collaborative ways of solving environmental problems. Need or the severity of the problem has been one of the reasons affecting the adoption of EPIs, but institutional commitment and capacity are also critical to adoption and implementation of policy innovations in general and for EPIs. Given that interest groups are critical in the American policy process, the support or opposition of key groups is also vitally important to innovation adoption. Since a number of EPIs provide new ways to work in partnership rather than in conflict with industries to improve compliance with regulations and do not burden them with any additional regulatory costs,[24] industry opposition to such EPIs may be lessened. Interest group opposition may also be mitigated by public support for certain EPIs, which provide broader benefits than they do costs, to the public at large. Finally, innovation adoption may also be affected by the presence or adoption of innovations by neighboring entities, the external determinants. However, the mechanisms of such diffusion need further study and research. Future research could also address the policy implications and long-term outcomes of innovation adoption. For instance, do the adoptions of EPIs lead to better environmental outcomes in terms of solving environmental problems and improve environmental quality? Do EPIs foster better outcomes in terms of increasing stakeholder participation and provide for more democratic and innovative means of achieving environmental goals? Preliminary evidence does suggest that EPIs have positive outcomes, but more research is needed on this issue. Given that the environmental problems facing us in the coming century are only likely to grow and be exacerbated by the slow-moving yet devastating effects of more large-scale problems such as climate change, the need for innovative environmental solutions is likely to be critically important.

Acknowledgments This research was funded in part by the National Science Foundation, NSF Grant No. SBER-9510308. The author is responsible for any omissions or errors.

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3. Gray, V. Innovation in the States: A diffusion study. Am. Polit. Sci. Rev. 1973, 67, 1174–1185. 4. Berry, W.D.; Berry, F.S. State lottery adoptions as policy innovations: An event history analysis. Am. Polit. Sci. Rev. 1990, 84, 395–415. 5. Berry, W.D.; Berry, F.S. Tax innovation in the States: Capitalizing on political opportunity. Am. J. Polit. Sci. 1992, 36, 715–742. 6. Berry, F.S. Innovations in public management: The adoption of strategic planning. Public Adm. Rev. 1994, 54, 322–330. 7. Mintrom, M. Policy entrepreneurs and the diffusion of innovation. Am. J. Polit. Sci. 1997, 41 (3), 738–770. 8. Mintrom, M. Policy Entrepreneurship in Theory and Practice: A Comparative State Analysis of the Rise of School Choice as a Policy Idea. Unpublished Ph.D. dissertation. State University of New York at Stony Brook, 1994. 9. Mooney, C.Z.; Mei-Hsien, L. Legislating morality in the American states: The case of the pre-Roe abortion regulation reform. Am. J. Polit. Sci. 1995, 39 (3), 599–628. 10. Berry F.S.; Berry, W.D. Innovation and diffusion models in policy research. In Theories of the Policy Process; Sabatier, P., Smith, H.-J., Eds.; Westview Press: CO, 2007; 169–180. 11. Shipan, C.R.; Volden, C. The mechanisms of policy diffusion. Am. J. Polit. Sci. 2008, 52 (4), 840–857. 12. Downs, G.W., Jr.; Mohr, L.B. Conceptual issues in the study of innovation. Adm. Sci. Q. 1976, 21 (December), 700–714. 13. Siegelman, L.; Smith, R. Consumer regulation in the American states. Soc. Sci. Q. 1980, 61, 58–76. 14. Menzel, D.; Feller, I. Leadership and interaction patterns in the diffusion of innovation among the American states. West. Polit. Sci. Q. 1977, 30 (4), 528–536. 15. Regens, J.L. State policy responses to the energy issue: An analysis of innovation. Soc. Sci. Q. 1980, 61 (June), 44–57. 16. Canon, B.C.; Baum, L. Patterns of adoption of tort law innovations: An application of diffusion theory to judicial doctrines. Am. Polit. Sci. Rev. 1981, 75, 975–990. 17. Glick, H.R. Innovation in state judicial administration: Effects on court management and organization. Am. Polit. Q. 1981, 9, 49–69. 18. Sigelman, L.; Roeder, P.W.; Sigelman, C.K. Social service innovation in the American states: Deinstitutionalization of the mentally retarded. Soc. Sci. Q. 1982, 62 (3), 503. 19. Glick, H.R.; Hays, S.P. Innovation and reinvention in state policy-making: Theory and the evolution of living will laws. J. Polit. 1991, 53 (3), 835–850. 20. Rabe, B. Power to the states: The promise and pitfalls of decentralization. In Environmental Policy in the 1990s; Vig, N., Kraft, M.E., Eds.; Congressional Quarterly Press: Washington, DC, 1995; 31–52. 21. Rabe, B.G. Federalism and entrepreneurship: Explaining American and Canadian innovation in pollution prevention and regulatory integration. Policy Stud. J. 1999, 27 (2), 288–306. 22. Rabe, B. Statehouse and Greenhouse: The Stealth Politics of America Climate Change Policy; Brooking Institution Press: Washington, DC, 2004. 23. Rabe, B. Racing to the top, the bottom, or the middle of the pack? The evolving state government role in environmental protection. In Environmental Policy: New Directions for the Twenty-First Century; Vig, N., Kraft, M.E., Eds.; Congressional Quarterly Press: Washington, DC, 2010; 27–51. 24. Sapat, A. Devolution and innovation: The adoption of state environmental policy innovations by administrative agencies. Public Adm. Rev. 2004, 64 (2), 141–151. 25. Davis, C.E. The Politics of Hazardous Waste; Prentice- Hall: Englewood Cliffs, NJ, 1993. 26. Lester, J.P.; Bowman, A. Subnational hazardous waste policy implementation: A test of the Sabatier-Mazmanian model. Polity 1989, 21 (4), 731–753. 27. Hird, J. Superfund: The Political Economy of Environmental Risk; John Hopkins University Press: Baltimore, MD, 1994.

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28. Environmental Defense Fund. Using Economics to Solve Eco-Challenges, available at http:// www.edf.org/approach/markets (accessed November 2011). 29. Ingram, H.; Colnic, D.; Mann, D.E. Interest groups and environmental policy. In Environmental Politics and Policy: Theories and Evidence, 2nd Ed.; Lester, J.P., Ed.; Duke University Press: Durham, NC, 1995. 30. Barke, R.P.; Jenkins-Smith, H.C.; Silva, C. Consistency and Controversy in the Translation of Scientific Knowledge into Policy Recommendations. Paper Presented at the Annual Meeting of the Midwest Political Science Association, Chicago, IL, April 6–8, 1995. 31. United States Environmental Protection Agency. Environmental Innovation Portfolio. U.S. EPA, National Center for Environmental Innovation, Washington, DC, available at http://www.epa. gov/osem/pdf/portfolio.pdf (accessed October 2011). 32. Morandi, L. Groundwater Protection Legislation: Survey of State Action 1988–1992; National Conference of State Legislatures: Washington, DC, 1994. 33. Multi-State Working Group on Environmental Performance and Regulatory Policy Program, Mossavar-Rahmani Center for Business and Government. Environmental Innovation: A Dialogue on the Role of Government, Law, and Regulatory Approaches. University of Massachusetts, Lowell and the John F. Kennedy School of Government, Harvard University, January 2006. 34. National Academy of Public Administration. Resolving the Paradox: EPA and the States Focus on Results; NAPA: Washington, DC, 1997. 35. O’Leary, R.; Durant, R.F.; Fiorino, D.; Weiland, P.S. Managing for the Environment: Understanding the Legal, Organizational, and Policy Challenges; Jossey-Bass: San Francisco, 1999. 36. Durant, R.F.; Fiorino, D.; O’Leary, R., Eds. Environmental Governance Reconsidered: Challenges, Choices, and Opportunities; The MIT Press: Cambridge, MA, 2004. 37. Koontz, T.M.; Thomas, C.W. What do we know and need to know about the environmental outcomes of collaborative management? Public Adm. Rev. 2006, 66, 111–121. 38. Weber, E.P. Pluralism by the Rules: Conflict and Cooperation in Environmental Regulation; Georgetown University Press: Washington, DC, 1998. 39. Rogers, E.; Weber, E.P. Thinking harder about outcomes for collaborative governance arrangements. Am. Rev. Public Adm. 2010, 40 (5), 546–567. 40. John, D. Civic Environmentalism: Alternatives to Regulation in States and Communities; Congressional Quarterly Press: Washington, DC, 1994. 41. Massachusetts Department of Environmental Protection. MassDEP Environmental Results Program, available at http://www.mass.gov/dep/service/envrespr.htm (accessed November 2011). 42. United States Environmental Protection Agency. About EPA Region 4, available at http://www. epa.gov/Region4 (accessed November 2011). 43. United States Environmental Protection Agency. Environmental Innovation Portfolio, available at http://www.epa.gov/osem/portfolio/protection.htm (accessed November 2011). 44. Savage, R.H. When a policy’s time has come: Cases of rapid policy diffusion, 1983–1984. Publius 1985, 15 (3), 111–125. 45. Nice, D.C. Policy Innovation in State Government; Iowa State University Press: Ames, Iowa, 1994. 46. Karch, A. National intervention and the diffusion of policy innovations. Am. Polit. Q. 2006, 34 (4), 403–426. 47. Mills, C.W. The Power Elite. Oxford University Press: New York, 1956. 48. Lindblom, C. Politics and Markets: The World’s Political Economic Systems; Basic Books: New York, 1977. 49. Schattschneider, E.E. The Semisovereign People; A Realist’s View of Democracy in America; Holt, Rinehart and Winston: New York, 1960. 50. Noll, R.; Owen, B.M. The Political Economy of Deregulation: Interest Groups in the Regulatory Process; American Enterprise Institute for Public Policy Research: Washington, DC, 1983.

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51. Moe, T. The politics of bureaucratic structure. In Can the Government Govern? Chubb, J.E., Petersen, P.E., Eds.; Brookings Institution: Washington, DC, 1989. 52. North, D.C. Institutions, Institutional Change and Economic Performance; Cambridge University Press: New York, 1990. 53. Chubb, J.E.; Petersen, P.E. Can the Government Govern? Brookings Institution: Washington, DC, 1989. 54. Van Horn, C.E. The State of the States, 3rd Ed.; Congressional Quarterly Press: Washington, DC, 1996. 55. Herbert A. Decision Making and Problem Solving. Research Briefings 1986: Report of the Research Briefing Panel on Decision Making and Problem Solving. National Academy Press: Washington, DC, 1986. 56. Shipan, C.R.; Volden, C. Bottom-up federalism: The diffusion of antismoking policies from U.S. cities to states. Am. J. Polit. Sci. 2006, 50 (4), 825–843. 57. Getz, M.; Walter, B. Environmental policy and competitive structure: Implications for the hazardous waste management program. Policy Stud. J. 1980, 9 (Winter). 58. Feiock, R.C.; Rowland, C.K. Environmental regulation and economic development: The movement of chemical production among states. West. Polit. Q. 1991, 561–576. 59. Lowry, W. The Dimensions of Federalism; Duke University Press: Durham, NC, 1992. 60. Feiock, R.C.; Davis, C. Can state hazardous waste regulation be reconciled with economic development: A test of the Rowland–Goetze model. Am. Polit. Q. 1992, 19. 61. Bernstein, M. Regulating Business by Independent Commission; Princeton University Press: Princeton, NJ, 1955. 62. Stigler, G. The theory of economic regulation. Bell J. Econ. Manage. Sci. 1971, 2, 3–21. 63. Peltzman, S. Toward a more general theory of regulation. J. Law Econ. 1974, 19, 211–240. 64. Meier, K.J. The Political Economy of Regulation: The Case of Insurance; SUNY Press: Albany, NY, 1988. 65. Derthick, M.; Quirk, P.J. The Politics of Deregulation; Brookings Institution: Washington, DC, 1985. 66. Truman, D. The Governmental Process; Knopf: New York, 1951. 67. Dahl, R. A Preface to Democratic Theory; University of Chicago Press: Chicago, IL, 1956. 68. Verba, S.; Nie, N.; Kim, J. The Modes of Democratic Participation: A Cross-National Comparison; Sage Publications: Beverly Hills, CA, 1971. 69. Bachrach, P.; Baratz, M.S. Two faces of power. Am. Polit. Sci. Rev. 1962, 56 (4), 947–952. 70. Becker, G. A theory of competition among pressure groups for political influence. Q. J. Econ. 1983, 98 (3), 371–400. 71. Wilson, J.Q. The Politics of Regulation; Basic Books: New York, 1980. 72. Bernstein, M. Regulating Business by Independent Commission; Princeton University Press: Princeton, NJ, 1955. 73. Posner, R. Theories of economic regulation. Bell J. Econ. Manage. Sci. 1974, 5 (3), 337–352. 74. Aggarwal, V.K.; Keohane, R.O.; Yoffie, D.B. The dynamics of negotiated protectionism. Am. Polit. Sci. Rev. 1987, 81 (2). 75. Wilson, J.Q. Bureaucracy: What Government Agencies Do and Why They Do It; Basic Books: New York, 1989. 76. Ringquist, E. Environmental Protection at the State Level: Politics and Progress in Controlling Pollution; M.E. Sharpe: Armonk, NY, 1993. 77. Rogers, E.M. Diffusion of Innovations, 5th Ed.; The Free Press: New York, 2003. 78. Simmons, B.A.; Dobbin, F.; Garrett, G. Introduction: The international diffusion of liberalism. Int. Organ. 2006, 60 (4), 781–810. 79. Berry, W.D.; Baybeck, B. Using geographic information systems to study interstate competition. Am. Polit. Sci. Rev. 2005, 99 (4), 505–519.

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80. Boehmke, F.J.; Witmer, R. Disentangling diffusion: The effects of social learning and economic competition on state policy innovation and expansion. Polit. Res. Q. 2004, 57 (1), 39–51. 81. Weyland, K. Theories of policy diffusion: Lessons from Latin American pension reform. World Polit. 2005, 57, 262–295. 82. Weyland, K. Bounded Rationality and Policy Diffusion; Princeton University Press: Princeton, NJ, 2007.

31 Food Quality Protection Act

Christina D. DiFonzo

Tolerances ...................................................................................................... 371 The Risk Cup ................................................................................................. 371 Endocrine Disruption .................................................................................. 372 Consumer Right-to-Know .......................................................................... 372 Potential Impacts of FQPA .......................................................................... 372 Bibliography.................................................................................................. 372

Tolerances FQPA fundamentally changed the way EPA sets tolerances for pesticide residues in food. EPA must review all (nearly 10,000) pesticide tolerances under new FQPA guidelines. The tolerance assessment schedule developed by EPA called for examining 33% within 3 years after August 1996, 66% within 6 years, and 100% within 10 years. EPA initially took a “worst-first” approach, to review the pesticides it considered to be of greatest risk, particularly to children, by August 1999. Three major pesticide groups, organophosphates (OPs), carbamates, and probable human carcinogens (B2s), were targeted under the worst-first approach. OPs and carbamates, the majority of which are insecticides, are neurotoxins structurally related to nerve gas. They affect the enzyme acetylcholinesterase in animals, including humans. B2 carcinogens are pesticides classified by EPA as having sufficient evidence for causing cancer in lab animals (usually at very high dose levels), but human evidence is lacking. Several important fungicides, plus a few herbicides and insecticides, are classified in this category. Before FQPA, a single tolerance was established for each pesticide/crop combination, based only on dietary exposure to pesticide residue. Under FQPA, EPA must consider the combined (aggregate) exposure to a pesticide through dietary, drinking water, and nondietary sources (for example, structural, turf, garden, and pet uses) as well as the cumulative exposure to related pesticides with a common mechanism of toxicity. Furthermore, FQPA directs EPA to consider sensitive subpopulations, especially children, when setting tolerances. To insure that sensitive groups are adequately protected, EPA can require a safety factor of up to 10-fold on existing tolerances.

The Risk Cup An analogy of a “risk cup” is used by EPA to explain changes in the establishment of tolerances under FQPA. Before FQPA, there was a separate risk cup for each pesticide/crop combination, containing only dietary exposure to residue. FQPA creates a separate risk cup for each group of related pesticides with common toxicity. Multiple pesticides, as well as multiple residues from all sources—food, water, and nonfood—of each pesticide, go into the same cup. Under this scenario, the cup gets crowded, and

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individual tolerances for each pesticide/crop combination in the group must get smaller. Furthermore, safety factors for children may reduce the overall size of each cup, potentially by a factor of 10.

Endocrine Disruption Under FQPA, all pesticides and pesticide additives must be tested for effects on the endocrine system. This may require in vitro and in vivo screening for three different types of endocrine effects: estrogenic (mimics or blocks estrogen), androgenic (mimics or blocks androgens), and thyroid. Of the potential targets of a screening program, these three hormone groups are important in human development, are fairly well studied, and some laboratory methodology is already available to detect changes in level and function. Estimates are that up to 70,000 pesticides and other chemicals will be screened under FQPA and a second law, the Safe Drinking Water Act.

Consumer Right-to-Know Another issue addressed in FQPA is consumer right-to-know about pesticide residues in food. FQPA mandated that EPA create a brochure to inform consumers about pesticide risks and benefits, and ways to remove residues from food they purchase. The brochure was completed and distributed to supermarkets in early 1999. However, FQPA did not mandate that stores actually display the publication.

Potential Impacts of FQPA Pesticides that do not meet FQPA standards must either be mitigated (use patterns changed) or eliminated (some or all uses dropped). Thus, as FQPA is implemented, it potentially will have a tremendous impact on American agriculture. • Changes in labeling or use patterns (number, frequency, and timing of applications) of pesticides to mitigate residue. • Loss of critical pesticide uses, particularly for so-called minor (specialty) crops. These commodities represent smaller markets for pesticide manufacturers and thus are often “expendable.” • Increases in production costs. Traditional broad-spectrum products might be replaced by more expensive, reduced-risk alternatives that control a narrower range of pests. • Increased complexity of production and pest management systems. Broad-spectrum pesticides will be replaced by narrower spectrum tactics that require better knowledge and more intense management of the production system on the part of the producer. • Potential for pesticide resistance. Loss of certain classes of pesticides could lead to resistance to remaining products, which are being relied on too heavily.

Bibliography 1. Colborn, T.; Dumanoski, D.; Myers, J. P. Our Stolen Future; Penguin Books U.S.A.: New York, 1996; 316. 2. Public Law 104–170 to Amend the Federal Insecticide, Fungicide, and Rodenticide Act and the Federal Food, Drug, and Cosmetic Act, and for Other Purposes, H.R. 1627; Federal Register: Washington, DC, Aug. 3, 1996. 3. Proceedings National Pesticide Impact Assessment Program Workshop, USDA CSREES NAPIAP Program, Sacramento, CA, May 5–7, 1998; Melnicoe, R., Ed.: Washington, DC, 1998; 76. 4. National Research Council. Pesticides in the Diets of Infants and Children; National Academy Press: Washington, DC, 1993; 386. 5. U.S. Environmental Protection Agency, Office of Pesticide Programs. The Food Quality Protection Act (FQPA) of 1996. http://www.epa.gov/oppfead1/fqpa/

32 Food: Cosmetic Standards Introduction .................................................................................................. 373

David Pimentel and Kelsey Hart

History of Cosmetic Standards • Environmental and Health Effects of Pesticide Exposure • Health Effects of Eating Insects/Insect Parts in Food • Conclusions and Future Directions

References ..................................................................................................... 376

Introduction The American marketplace features nearly perfect fruits and vegetables. Gone are apples with an occasional blemish and fresh spinach with a leaf miner. This increase in the “cosmetic standards” of fruits and vegetables has resulted from the efforts of the Food and Drug Administration (FDA) and the U.S. Department of Agriculture (USDA) to limit the levels of insects and mites in produce, and new standards established by food wholesalers, processors, and retailers. Meeting more stringent standards has led to significant increases in the amounts and toxicity of pesticides used in crops. Increased pesticide use has negative environmental and public health consequences. In comparison, the health risks from consuming herbivorous insects/insect parts in food do not exist and certainly do not justify the increase in pesticide use and the associated problems. Recent research indicates that pesticide use can be reduced by 35% to 50% without any substantial increase in food prices or loss of crop yields.[1] Surveys suggest that the public would support relaxation of cosmetic standards if it decreases pesticide residues in its food.

History of Cosmetic Standards The FDA sets defect action levels (DALs) for insects and mites allowed in fruits and vegetables and in products made from them. These DALs were established to reduce insect and mite infestation in foods to a reasonable and safe level, because their presence in food products was thought to indicate that crops had insufficient insect and mite control, were improperly washed, were unsatisfactorily inspected, and contained insects and mites harmful to health. Besides visual prejudice against insects in food, there is the well- placed concern that insects such as nonherbivorous houseflies and cockroaches may transmit disease. During the past 40 years, the FDA has steadily lowered DALs.[2] For example, a fivefold decrease in the number of leaf miners permitted in spinach occurred from 1930 to 1974. As tolerance levels for insects in food have fallen, wholesalers, processors, and retailers have increased their “cosmetic standards” for produce and other food products so that most marketed U.S. produce is visually perfect. Produce distributors encourage high cosmetic standards because their contracts enable them to visually inspect produce before buying and reject it when the supply is excessive. Growers are motivated to produce cosmetically perfect produce to ensure its sale. However, to meet these increasingly stringent regulations, farmers have had to use greater amounts of increasingly toxic pesticides and implement other pest control strategies. Synthetic pesticide use in 373

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FIGURE 1 The amount of synthetic pesticides produced in the United States. About 90% is sold in the United States. The decline in total amount produced since 1975 is in large part due to the 10- to 100-fold increased toxicity and effectiveness of the newer pesticides.[1,3]

the United States has increased about 33-fold since 1945, and the toxicity of pesticides used has increased 10- to 100-fold in the past 25 years (Figure 1). More pesticides need to be used to produce the blemish-free produce distributors and consumers expect. There is little evidence that eating herbivorous insects or insect parts is hazardous to human health.[4] However, solid data suggest that the adverse health and environmental impacts of pesticide exposure are substantial.[5] Given the direct correlation between increases in cosmetic standards and increases in pesticide use, why are cosmetic standards and DALs growing increasingly severe and perpetuating further increases in pesticide use? The increase in cosmetic standards and more stringent DALs are based on the premise that consumers demand unblemished, insect-free food. Clearly, cosmetic appearance of produce is a primary factor consumers use in assessing the quality of produce. Unfortunately, this assessment is often made without more substantive quality information, such as nutritional values or pesticide residue levels. Recent evidence suggests that when consumers are aware of the trade-offs between blemish-free produce and pesticide use, they will purchase produce that is not cosmetically perfect because it has less pesticide residue.[6]

Environmental and Health Effects of Pesticide Exposure An estimated 617,000 tons of more than 600 different kinds of pesticides are used annually in the United States, at a cost of approximately $9 billion.[7] Still, pests such as insects, plant pathogens, and weeds destroy 37% of all potential food and fiber crops.[8] Typically, each dollar invested in pesticides returns about $4 in crops saved. However, this economic evaluation does not take into account the impacts of pesticide use on public health and the environment. Approximately 0.1% of applied pesticides reach target pests, leaving 99.9% of the pesticides to impact the environment.[8] Environmental effects of pesticides can be significant: Domestic animals and wildlife can be poisoned or adversely affected by pesticide exposure; beneficial natural enemies of harmful pests can be destroyed by pesticide use; heavy pesticide

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use can result in pesticide resistance and subsequently even heavier or more toxic pesticide use; and already limited natural resources such as soil, groundwater, and surface water can be contaminated by pesticide residues or drift.[5] The human health effects of pesticide exposure through food are also diverse and significant. About 35% of foods purchased by American consumers have detectable levels of pesticides, and about 1%–3% of these foods have residue levels over the legal tolerance level 8. These estimates are conservative because detection methods currently detect only about one-third of the pesticides now in use in the United States. The contamination rate is undoubtedly higher for fruits and vegetables because they receive the highest levels of pesticides. One USDA study indicates that some pesticide residue remains in produce even after it is washed, peeled, and cored.[9] Both the acute and chronic health effects of pesticide exposure are significant. Worldwide, about 26.5 million acute pesticide poisonings occur each year, resulting in about 3 million hospitalizations, approximately 220,000 fatalities, and 750,000 cases of chronic pesticide-related illness.[10] Chronic effects can adversely affect most systems of the human body. U.S. data indicate that 18% of all insecticides and about 90% of all fungicides are carcinogenic.[11] Many pesticides are also estrogenic, linked to increased breast cancer among some women in the United States. Pesticide exposure can also damage the respiratory and reproductive systems, leading to conditions like asthma and infertility.[12–14] In the United States, EPA[15] reports that 300,000 non-fatal pesticide poisonings occur each year. The negative health effects that pesticides can have are more significant in children. Children have higher metabolic rates than adults, and their ability to detoxify and excrete toxic compounds is different. Also, because of their smaller size, children are typically exposed to higher levels of pesticides than adults. Finally, certain types of pesticides, such as carbamates and organophosphates, are more dangerous for children than adults.[10] Given the significant environmental and public health impacts that pesticides can have, it appears desirable to limit pesticide exposure to minimize these adverse effects. However, the increasingly stringent DALs and cosmetic standards have resulted in considerable increases in pesticide use. Do the health effects of eating herbivorous insects, insect parts, or blemished produce warrant the risks and the substantial health consequences of increasing pesticide exposure to meet these standards?

Health Effects of Eating Insects/Insect Parts in Food Even under the current stringent DAL regulatory guidelines, a few insects and mites do remain in or on produce. For instance, the DAL for apple butter is an “average of 5 whole insects or equivalents per 100 grams not counting mites, aphids, thrips, or scale insects.” DALs for many other food products are similar. Many insects commonly found in foods and food products are so minute in size that they are practically impossible to eliminate. Although the numbers of insects are strictly limited by FDA regulations, some do remain and are eaten. This, however, is not a cause for concern. In contrast with the well-documented acute and chronic negative health effects resulting from pesticide exposure, there is not one known case of human illness from ingesting insects and mites in or on foods. In addition, though some insects do carry disease or present health risks—houseflies, for example—all herbivorous insects/mites found on harvested produce are harmless to humans. While ingesting insects or insect parts in our food may seem distasteful to many Americans, many cultures eat insects by choice. Insects are a substantial source of protein, with digestible protein content ranging from 40% to 65%,[16] and insects, shrimp, lobster, and crawfish are all arthropods; the latter three are often considered food delicacies. Given that herbivorous insects found on produce are not a health hazard, consumers must decide whether they are willing to tolerate the presence of a few insects rather than insisting on visually “perfect” produce that requires high levels of pesticides.

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Conclusions and Future Directions If the health effects of herbivorous insects found on or in food products are not cause for concern, then the need for strict DALs might be relaxed. Relaxing cosmetic standards for some fruits and vegetables might be feasible. Approximately 10% to 20% of pesticides applied to fruits and vegetables are used only to comply with the current strict cosmetic standards established by the FDA, USDA, wholesalers, and retailers that result in blemish-free pro-duce. Rigorous cosmetic standards are probably unnecessary, since surface blemishes on fruits and vegetables generally do not affect nutritional content, storage life, or flavor. However, will the American public purchase produce that appears less than perfect? Research on public preferences shows that 97% of Americans prefer food without pesticide residues. In addition, 50%–66% are willing to pay more for food with less pesticide residue.[17] It is estimated that in the United States, pesticide use can be reduced by about 50% without reducing crop yields. The estimated increase in the consumer’s food costs would be only 0.6%.[10] This marketplace cost increase does not take into account the positive environmental and health benefits that would be realized if pesticide use were reduced. Sweden, for example, has reduced pesticide use 68% and reduced pesticide poisonings 77%.[18] The small increase in consumer cost would be more than offset by these benefits. Therefore, given the environmental and health tradeoffs related to high cosmetic standards for produce, it appears that human health and the environment would be best protected by less stringent DALs and relaxed cosmetic standards for produce, to minimize unnecessary pesticide use and related adverse effects.

References 1. Pimentel, D.; Kirby, C.; Shroff, A. The relationship between “cosmetic standards” for foods and pesticide use. In The Pesticide Question: Environment, Economics, and Ethics; Pimentel, D., Lehman, H. Eds.; Chapman and Hall: New York, 1993; 85–105. 2. FDA. The Food Defect Action Levels: Levels of Natural or Unavoidable Defects in Foods That Present No Health Hazards for Humans; ed.; Department of Health and Human Services, Public Health Service, Food and Drug Administration, Center for Food Safety and Applied Nutrition: Washington, DC, 1995. 3. KRS Network. U.S. Pesticide Industry Report Executive Summary. Covington, GA, 2005; 15 pp. available at http://www.knowtify.net/2005USPestIndReptExecSum.pdf. 4. Defoliart, G.R. The human use of insects as food and as animal feed. Bull. Entomol. Soc. Am. 1986, 35, 22–35. 5. Pimentel, D. Environmental and economic costs of the application of pesticides primarily in the United States. Environ. Dev. Sustainability 2005, 7, 229–252. 6. Pimentel, D.; Terhune, E.; Dritschilo, W.; Gallahan, D.; Kinner, N.; Nafus, D.; Peterson, R.; Zareh, N.; Misiti, J.; Haber-Schaim, O. Pesticides, insects in food, and cosmetic standards. BioScience 1977, 27, 178–185. 7. Environmental Protection Agency. Pesticide Market Estimates: Usage, 2000–2001; 3; available at http://www.epa.gov/opp bead1/pestsales/usage2001.htm (accessed February 11, 2010). 8. Pimentel, D.; Greiner, A. Environmental and socioeconomic costs of pesticide use. In Techniques for Reducing Pesticide Use; Pimentel, D., Ed.; John Wiley and Sons: Chichester, U.K., 1997; 51–78. 9. Wiles, R.; Campbell, C. Washed, Peeled—Contaminated: Pesticide Residues in Ready-To-Eat Fruits and Vegetables; Environmental Working Group: Washington, DC, 1994. 10. Pimentel, D.; Hart, K.A. Pesticide use: ethical, environmental, and public health implications. In New Dimensions in Bioethics: Science, Ethics and the Formation of Public Policy; Galston, A.; Shurr, E., Eds.; Kluwer Academic Publishers: Boston, 2001; 79–108. 11. Eastmond, D.A. Mechanisms of Carcinogenesis of the Fungicide and Rat Bladder Carcinogen o-Phenylphenol. EPA Grant No.: R826408 Project Period: May 1, 1998 through April 30, 2001; 5; available at http://cfpub.epa.gov/ncer_abstracts/index.cfm/fuseaction/display.abstractDetail/abstract/876.

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12. Weisenburger, D.D. Human health effects of agrichemical use. Hum. Pathol. 1993, 24, 571–576. 13. Schneider, E.P.; Dickert, K.J. Health costs and benefits of fungicide use in agriculture. J. Agromed. 1993, 1, 19–37. 14. Yang, L.; Kemadjou, J.R.; Zinsmeister, C.; Bauer, M.; Legradi, J.; Muller, F.; Pankratz, M.; Jakel, J.; Strahle, U. Transcriptional profiling reveals barcode-like toxicoge-nomic responses in the zebrafish embryo. Genome Biol. 2007, 8, R227. 15. Environmental Protection Agency. Hired Farm Workers Health and Well-Being at Risk. United States General Accounting Office Report to Congressional Requesters. February, 1992. 16. Gorham, J.R. Foodborne filth and human disease. J. Food Prot. 1989, 52, 674–677. 17. Anon. Appearance of produce versus pesticide use. Chem-ecology 1991, 20 (4), 11. 18. Pesticide News. Persistence pays—lower risks from pesticides in Sweden. Pesticide News 34, 10–11; available at http://www.pan-uk.org/pestnews/Issue/pn54/pn54p10.htm (accessed February 11, 2010).

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33 Laws and Regulations: Food Tolerance Limits for Pesticide Residues .................................................... 379 Regulatory Inspection and Enforcement

Ike Jeon

Tolerance Limits for Insect Fragments ...................................................... 380 Responsibility of Food Manufacturers ...................................................... 381 Potential Consumer Benefits ...................................................................... 381 Acknowledgments ........................................................................................ 381 References ..................................................................................................... 382

Tolerance Limits for Pesticide Residues The responsibility for ensuring that pesticide residues in foods are not present above the limits is shared by three major government agencies.[1] The Environment Protection Agency (EPA) determines the safety of pesticide products and sets tolerance levels for pesticides. The Food and Drug Administration (FDA) enforces the tolerances in all foods except meat and poultry products. The U.S. Department of Agriculture’s Food Safety and Inspection Service (FSIS) regulates commercially processed egg, meat, and poultry products including combination products (e.g., stew, pizza). In addition, any products containing 2% or more poultry or poultry products, or 3% or more red meat or red meat products are also under jurisdiction of the FSIS. The pesticides of concern usually include insecticides, fungicides, herbicides, and other agricultural chemicals. Table 1 illustrates examples of tolerance levels for pesticide residues in several food categories.[2,3] These tolerance levels are extremely low, usually below parts per million, but do not represent permissible levels of contamination where it is avoidable. In addition, blending of a food (or feed) containing a substance in excess of an action level or tolerance with another food (or feed) is not permitted, and the final product from blending is unlawful, regardless of the level of the contaminant.

Regulatory Inspection and Enforcement The FDA monitors the levels of pesticide residues in processed foods. For imported products, the FDA checks a sample of the food at entry into the United States and can stop shipments at the entry. If illegal residues are found in domestic samples, FDA can take regulatory actions, such as seizure or injunction. The U.S. Department of Agriculture also monitors pesticide residues in food.[4] The Department was charged in 1991 with implementing a program to collect data on pesticide residues on various food commodities. The program has become a critical component of the Food Quality Protection Act of 1996 and currently is known as the Pesticide Data Program. The data on pesticides in selected commodities are

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used by the EPA to support its dietary risk assessment process and pesticide registration and by the FDA to refine sampling for enforcement of tolerances. If a product is in violation of the tolerance limits, it is adulterated under the food law. The product may be destroyed or recalled from the market by the manufacturer or shipper. The recall may be initiated voluntarily by the manufacturer (or shipper) or at the request of the regulatory agency. The responsible agency also may seize the product on orders obtained from the Federal courts and may prosecute persons or firms responsible for the violation.

Tolerance Limits for Insect Fragments Many food materials may contain natural but unwanted debris that cause no health hazards for humans. These debris may include insects, insect fragments, and rodent hairs and are considered unavoidable defects in foods with the current agricultural practices. In fact, the use of chemical substances to control insects, rodent, and other contaminants has little, if any, impact on natural and unavoidable defects in foods. The FDA contends that the use of pesticides does not effectively reduce the presence of these food defects. This has led the regulatory agencies to establish maximum levels of natural or unavoidable defects allowable in foods for human use. The FDA currently lists over 100 products from fruits to fish,[5] and Table 2 shows only several examples. If no defect action level exists for a product, the FDA evaluates and decides on a case-by-case basis using criteria of reported findings such as length of hairs and size of insect fragments. The FDA sets these action levels under the premise that it is economically impractical to grow, harvest, or process raw products that are totally free of nonhazardous, naturally occurring, unavoidable defects. It is incorrect, however, to assume that because the FDA has an established defect action level for a food, the manufacturer needs only keep defects just below that level. The defect levels do not represent averages of the defects that occur in any of the products. The levels represent limits at which FDA will regard the food product as adulterated and, therefore, subject to enforcement action. Like pesticide residues, blending of food with a defect at or above the current defect action level with TABLE 1

Examples of Tolerance Limits for Pesticide Residues in Human Food Action Level

Substance

Commodity

Aldrin and dieldrin

Asparagus

Chlordane

DDTa

Lindane

Source: FDA and USDA. a Dichlorodiphenyltrichloroethane. [2]

[3]

(Parts per Million)

Remark

0.03

Fish

0.3

Peanuts

0.05

Carrots

0.1

Fish

0.3

Lettuce

0.1

Poultry

0.3

Carrots

3.0

Citrus fruits

0.1

Tomatoes

0.05

Beans

0.5

Edible portion

Edible portion Fat basis

Corn

0.1

Milk

0.3

Fat basis

Beef

7.0

Fat basis

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another lot of the same or another food is not permitted. That practice renders the final food unlawful regardless of the defect level of the finished food.

Responsibility of Food Manufacturers Food manufacturers are required to follow the standard manufacturing procedures under a federal regulation, known as good manufacturing practice (GMP), during food production.[6] The GMP guidelines imply that all food materials used must not exceed the tolerance limits set for pesticide residues or any other poisonous or deleterious substances. The GMP also calls for the same regulatory requirement for natural or unavoidable defects in all food materials. The food materials susceptible to contamination may be tested for compliance or relied on a supplier’s guarantee or certification that they are in compliance. In addition, the GMP regulation stipulates that food manufacturers and distributors must utilize at all times quality control operations that reduce natural or unavoidable defects to the lowest level feasible with the current technology.

Potential Consumer Benefits Through conducting a monitoring program, the federal government agencies work together to improve consumer protection. The EPA will continue to review scientific data on all pesticide products, while the FDA and U.S. Department of Agriculture will closely monitor levels of pesticide residues in all foods including both domestic and imported products. The U.S. Department of Agriculture’s data for 1998 suggest that violation of the pesticide tolerance limits was very low in all raw products including fruit and vege, wheat, and milk samples. In 1993, the FDA reported that no pesticide residues were found in infant formulas, and no residues over EPA tolerances or FDA action levels were found in any of the foods that were prepared as consumers normally would prepare them at home.[7]

Acknowledgments Contribution No. 00–231-B, Kansas Agricultural Experiment Station, Manhattan, Kansas 66506, U.S.A. TABLE 2

Examples of Tolerance Limits for Natural or Unavoidable Defects in Foods

Product

Defect

Action Level

Sweet corn, canned Macaroni

Insect larvae Insect filth Rodent filth

2 or more 3 mm or longer larvae 225 insect fragments or more per 225 g 4.5 rodent hairs or more per 225 g

Peaches, canned and frozen

Mold/insect damage Insects

Wormy or moldy on 3% or more fruits 1 or more larvae and/or larval fragments whose aggregate length

Peanut butter

Insect filth Rodent filth

30 or more insect fragments per 100 g 1 or more rodent hairs per 100 g

Popcorn

Rodent filth

Tomato juice

Drosophila fly Mold

1 or more rodent excreta pellets or rodent hairs in 1 or more subsamples 10 or more fly eggs per 100 g 24% of mold counts in 6 subsamples

Wheat flour

Insect filth Rodent filth

75 or more insect fragments per 50 g 1 or more rodent hairs per 50 g

exceeds 5 mm in 12 one-pound cans

Source: FDA.[5]

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References 1. FDA. FDA’s Food and Cosmetic Regulatory Responsibilities; U.S. Food and Drug Administration: Washington, DC, 1998; 1–5, http://vm.cfsan.fda.gov/~dms/regresp.html (accessed June 2000). 2. FDA. Action Levels for Poisonous or Deleterious Substances in Human Food and Animal Feed; U.S. Food and Drug Administration: Washington, DC, 1998; 1–17, http://vm.cfsan.fda.gov/~lrd/fdaact. html (accessed June 2000). 3. USDA. Domestic Residue Book (Appendix I); U.S. Department of Agriculture, Food Safety and Inspection Service: Washington, DC, 1998; 1–30, http://www.fsis.usda.gov:80/OPHS/redbook1/ appndx1.htm (accessed June 2000). 4. USDA. Pesticide Data Program Annual Summary—Calendar Year of 1998; U.S. Department of Agriculture, Agricultural Marketing Service: Washington, DC, 2000; 1–19. 5. FDA. The Food Defect Action Levels—Levels of Natural or Unavoidable Defects in Foods that Present No Health Hazards for Humans. In FDA/CFSAN Food Defect Action Level Handbook; U.S. Food and Drug Administration: Washington, DC, 1998; 1–36, http://vm.cfsan.fda.gov/~dms/ dalbook.html (accessed June 2000). 6. CFR. Current good manufacturing practice in Manufacturing, Packing, or Holding Human Food. In Code of Federal Regulations, Title 21, Part 110; U.S. Government Printing Office: Washington, DC, 1999; 206–215. 7. FDA. FDA Reports on Pesticides in Foods; U.S. Food and Drug Administration: Washington, DC, 1993; 1–5, http://vm.cfsan.fda.gov/~lrd/pesticid.html (accessed June 2000).

34 Laws and Regulations: Pesticides Introduction .................................................................................................. 383 Why Regulate Pesticides?

History ........................................................................................................... 383 Available International Guidelines .............................................................384 Implementation Problems ...........................................................................384 Steps Undertaken .........................................................................................384 Present Scenario and Probable Remedies ................................................. 385 Future Global Policy .................................................................................... 385 Appendix 1 .....................................................................................................386 Toxicological and Other Data Requirements for Pesticide Registration

Praful Suchak

Bibliography .................................................................................................. 387

Introduction Why Regulate Pesticides? Chemical or biological pesticides have target specific toxicity that controls or eradicates pests falling under different groups. These products, though developed for specific usage, could have adverse effects on living beings and the environment and unchecked use can cause havoc. Regulating pesticides, therefore, would assure reasonable safety in use of these toxic substances and ensure that risks from pesticides to humans and their environment are minimized and are consistent with the benefits achieved by their use in terms of reduced losses. Regulating pesticides at the international and national level should consider social costs in line with social benefits. Pesticides impose costs on society, such as health risks and environmental degradation, which are not borne by the user. The available policy remedies include bans on individual or classes of chemicals that prohibit the introduction of hazardous compounds into the environment, and economic instruments such as taxes, registration fees, and import duties that work to redistribute and adjust the social costs occurring for pesticide use and also provide the government with revenues that can be used to cover health costs and environmental clean-up activities.

History The United States in 1910 introduced the Federal Pesticide Act that underwent complete metamorphosis to become the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) in 1947, which since 1970 is under the auspices of the Environmental Protection Agency.

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Australia initiated pesticide legislation with one state in 1925 and by 1945 all states had their individual laws. The Industry Association brought law common to all states in 1995. By the end of 1999 about 95% of the countries in the world had adopted full/partial regulatory systems. Early in-depth studies were not carried out on the long-term effects of: 1) repeated exposures, 2) residual toxicity, 3) accumulated toxicity, and 4) the impact on environment. With additional knowledge on the cumulative toxicity of chlorinated hydrocarbons such as DDT having come to light, the regulating authorities have started demanding the generation of additional critical toxicological data to assess short-term, long-term, and environmental toxicity of earlier registered pesticides. The European Union has already undertaken reviews of 90 molecules in the first phase by a Commission regulation dated December 11, 1992, to be completed in 12 years, and a further 148 molecules in the second phase effective March 1, 2000. The remaining substances in the European Union would be included in third phase. Regulatory requirements for pesticides have undergone a change over the past half a century. With the advent of highly sophisticated testing equipment, more knowledge about harmful effects of the toxic chemicals has come to light. Consistent watch by environmentalists and organizations like the Pesticides Action Network (PAN), Greenpeace, Save the Planet groups, and other nongovernmental organizations has resulted in added awareness resulting in hosts of data requirements for registration/ reregistration of pesticides. Although all developed countries and most of the developing countries have their own legislation to regulate pesticides, there have been vast variations in data requirements for registrations between these countries. With globalization it has became imperative to have harmonized data requirements so that the registrant can hope for faster registration in different (pesticide consuming) countries.

Available International Guidelines 1. 2. 3. 4. 5. 6.

Agenda 21 of the United Nations Conference on Environment and Development (UNCED) The Codex Alimentarius The FAO International Code of Conduct and Prior Informed Consent (PIC) WTO and International Trade with respect to pesticides Agreement on Persistent Organic Pollution (POP) Guidelines of Minor Donor Institutions on the purchase of pesticidess

Implementation Problems Although FAO took the lead to harmonize data requirements in participating nations for registrations of pesticides, certain problems and practical difficulties have occurred such as 7. The original registrant, having invested huge amounts in data generation, is unable to protect the data 8. Absence of confidentiality assurance by the registering country, creating difficulties in multiple country registrations 9. Recommended uses differ from country to country, resulting in difficulties 10. Unchecked dumping of unsafe or banned pesticides in less-developed countries 11. New registrations by a company other than the original registrant by providing data generated by such a company could not be checked

Steps Undertaken Although though PIC entry of banned pesticides could be prevented, this instrument has not been fully effective. Once it becomes fully operational legally things should improve.

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With the United States implementing the Food Quality Protection Act and fixing maximum residue limits for 3000 toxic compounds, countries worldwide would need to harmonize their registrations on toxic chemicals so as to meet the residue levels in food. The formation of the European Union with 15 member countries, OECD with 29 members, and the Technical Working Group having EPA, Canada, and Mexico, has accelerated the pace towards harmonization. However, since a vast disparity exists between developed countries on one side and developing countries on the other side, it is rather difficult to have a unified data requirement, particularly in case of risk assessment. Acceptance of electronic data submission and dossier/ monograph submissions and joint reviews by EU would also pave the way toward harmonization and would address questions in the nondietary exposure area. Apart from studies related to bioefficiency of the product, the toxicological studies of the toxicant, its analogues, impurities and breakdown products, residual toxicity, etc., as listed in Appendix 1 would help understanding and regulating pesticides.

Present Scenario and Probable Remedies Substantial evidence exists that pesticides are being applied in a technically and economically inefficient manner. Many developing countries subsidize pesticides and equipment, resulting in excessive use of pesticides. Also in developing countries, the current legal environment and enforcement capabilities have been inadequate and dysfunctional, thus exerting a significant impact on current levels of pesticide use. This is partly due to lack of resources and partly due to manipulation by vested interests. The inadequacies of the existing regulatory framework, institutional rigidities, and a bias in favor of pesticide-dependent paths also contribute to improper use of pesticides. A major problem confronting many countries is the absence of well-established procedural mechanisms for public involvement in the decision making process including crop protection policy. Competing interests with a stake in the process, including farmers, the pesticide industry, and policy makers responsible for food security, argue for a more liberal regulatory stance. On the other hand, environmentalists, public health workers, and consumers demand strict regulation and reduced pesticide volumes. To be more effective, pesticide regulation and implementation should be handled by a neutral agency like the Ministry of Environment or similar organization and not the Ministry of Agriculture or other interested ministry. Pesticide policy needs to be integrated into the broader public policy debate concerning the nations’ agricultural, environmental, and health strategies. Nevertheless, two general principles should apply. First, dispassionate analysis of the costs and benefits of pesticide use would provide a useful tool for the formulation of normal policies; and second, the broader and more inclusive the debate, the more likely it is that the outcome will serve the public rather than specific private interests.

Future Global Policy A uniform global regulatory system needs to ensure 1. 2. 3. 4. 5. 6.

Agricultural chemical use increases agricultural output Food supplies are safe from harmful toxicants/ residues Reduced-risk chemical pesticides, biopesticides, and nonchemical alternatives are encouraged Uniform MRLs to eliminate trade barriers Uniform health-based safety standards for pesticide residues Special provisions for certain groups of the population including infants and children

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Appendix 1 Toxicological and Other Data Requirements for Pesticide Registration 7. Identity of active substance Chemical name Empirical and structural formula Molecular mass Method of manufacture (synthesis pathways) Purity Identity and content of isomers Impurity and additives 8. Physical and chemical properties Melting point Boiling point and relative density Vapor pressure Volatility Appearance Absorption spectra-molecular extinction at relevant wavelength Solubility in water/organic solvents Partitioning coefficient N-octanol/water Stability and hydrolysis rate in water Photochemical degradation on surface, in water, and in air Thermal stability and stability in air 9. Analytical method Analytical method for the determination of the pure active substance in the technical grade. For breakdown products and additives in plant products, soil, water, animal body fluids, and tissues. 10. Toxicological and metabolism studies Studies on acute toxicity—oral, percutaneous, inhalation, intraperitoneal, skin and, where appropriate, eye irritation, and skin sensitization. Short-term toxicity—oral, cumulative toxicity, and other routes inhalation or dermal. Chronic toxicity—oral, long-term toxicity, and carcinogenicity. Mutagenicity—reproductive toxicity-teratogenicity and multigeneration studies in mammals. Metabolism studies in mammals—absorption, distribution, and excretion studies, elucidation of metabolic pathways. Supplementary studies—neurotoxicity studies— toxic effects of metabolites from treated plants and toxic effects on livestock and pests. Medical data—medical surveillance on manufacturing plant personnel, clinical cases, poisoning incidents from industry and agriculture sensitization/ allergenicity observations, observations on exposure of the general population, and epidemiological studies if appropriate. Diagnosis and specific signs of poisoning, clinical tests, and prognosis of expected effects of poisoning. Proposed treatment: first aid measures, antidotes, and medical treatment. Summary of toxicological studies and conclusions, critical scientific evaluation with regard to all toxicological data, and other information concerning the active substance. 11. Residues in or on treated products, food and feed metabolism in plants and livestock In treated plants (distribution, metabolism, binding constituents, etc.). In livestock (uptake, distribution, metabolism, binding constituents, etc.).

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12. Fate and behavior in the environment Studies on aerobic and anaerobic degradation under laboratory conditions in different soil types. Adsorption and desorption in different soil types including metabolites. Mobility of the active ingredients in different soil types. Behavior in water and air, rate and route of degradation. 13. Ecotoxicological studies Effects on birds, fish, aquatic organisms such as Daphnia magna, algae, honeybees, earthworms, other nontarget macroorganisms and microorganisms. 14. Information concerning the labeling including indication of danger and safety measures.

Bibliography 1. Pesticides Policies in Developing Countries—Do They Encourage Excessive Use? In World Bank Discussions Paper No. 238; 1994. 2. Asian Development Bank. In Handbook on the Use of Pesticides in Asia Pacific Region; ADB: Manila, Philippines, 1987. 3. Pesticide Policy Project Hannover; Publication Serial No. 1, January 1995; No. 2, November 1995; No. 3, December 1995; No. 4, December 1996; No. 5, December 1996; No. 6, 1998; No. 7, April 1999; No. 8, April 1999. 4. EC Directives 91/414/EEC and Subsequent Directives Including 1999/80/EC. 5. Proceedings of Asia Pacific Crop Protection Conference 1997 and 1999, PMFAI: Mumbai, India. 6. Global Pesticides Directory, 2nd Ed.; Suchak’s Consultancy Services: Mumbai, India, 1997. [email protected] 7. Pesticides News; No. 20–47, Pesticides Action Network (PAN): London, 1993 to 1999. 8. Guidelines on the Operation of Prior Informed Consent (PIC) Rome FAO 1990, Guidance to Government in PIC Rome 1991, and Other FAO Publications. 9. U.S. EPA Pesticides Information Network. http://www.cdpr.ca.govt/docs/epa/epachim.htm.

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35 Laws and Regulations: Rotterdam Convention Introduction .................................................................................................. 389 History of PIC ............................................................................................... 390 From Voluntary to Legally Binding • Banning Exports of Banned Pesticides

How the Convention Is Operated .............................................................. 391 Designated National Authorities • Notifying Regulatory Actions • Chemical Review Committee • Two Routes to Be “PIC-ed” • Import Decisions, Information, and Website • The Convention—More than PIC

Barbara Dinham

Building Capacity/Improving Regulations ............................................... 393 References ...................................................................................................... 393

Introduction When chemical pesticides were introduced 50 years ago, little attention was paid to the environmental and health impacts. With the rapid expansion of use in the 1950s, understanding gradually increased of the consequences of exposure to certain chemicals. Wide-ranging impacts began to be identified, including: environmental persistence and effects on birds and wildlife; residues in soil, water, and air; residues in food; human poisonings from acutely toxic pesticides or long-term health impacts such as cancer; and pest resistance, often leading to dramatic crop losses. With almost 1000 different pesticides and thousands of formulations on the market to control insects, diseases, weeds, and other pests, action was clearly needed to protect human health and the environment. International standards recommended that governments establish a registration system to authorize each formulation of a pesticide for each specific crop or other use. Concern with some pesticides led governments to ban or restrict them to a limited number of uses. Few developing countries can fully implement a registration scheme, and they are often unaware of bans imposed elsewhere. Recognizing these problems, in the early 1980s, governments, international organizations, and public interest groups began to demand action to provide a warning system to help developing countries regulate or ban the use of hazardous pesticides. The Rotterdam Convention on Prior Informed Consent Procedure for Certain Hazardous Chemicals and Pesticides in International Trade[1] is the outcome of 15 years of activity on trade in hazardous chemicals. Adopted on 10 September 1998 in Rotterdam, the Netherlands, the Convention was signed by 73 countries[2] and by June 2001 had been ratified by 14 parties. It will become legally binding after 50 countries have ratified. The Convention takes an important step toward protecting humans and the environment from highly toxic chemicals. For the first time, it will help monitor and control trade in dangerous substances, circulate better information about health and environmental problems of chemicals, and prevent unwanted imports of certain hazardous chemicals. 389

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Central to the Rotterdam Convention is the system of Prior Informed Consent (PIC), a means of obtaining and disseminating decisions of importing countries about their willingness to receive shipments of certain chemicals, and ensuring compliance to these decisions by the exporter. To be included in PIC, a pesticide must be banned or severely restricted for health or environmental reasons by two countries in two different regions of the world—indicating that its adverse effects are a “global concern.” But focusing on banned or severely restricted pesticides may only touch the tip of the iceberg. Industrialized countries rely on trained and informed users able to apply good practice as safeguards: in developing countries where pesticides are often used under conditions of poverty, these measures cannot be applied. Furthermore, older—and often more hazardous—pesticides are often cheaper, making them attractive to poorer farmers. The Convention recognizes that “severely hazardous pesticide formulations” should be included in PIC if they cause health or environmental problems in developing countries or in Eastern Europe—termed “countries with economies in transition”— in the Convention.

History of PIC A PIC system was first proposed in the early 1980s as part of the International Code of Conduct on the Distribution and Use of Pesticides, negotiated by governments in the Food and Agriculture Organization (FAO) of the UN. Some governments resisted the concept, and the Code was adopted in 1985 without any reference to PIC. But intense pressure from nongovernmental organizations (NGOs) and others won support, and the principle was accepted in 1987. It took until 1989 to establish the wording and issue a revised version of the Code.[3] That same year, the UN Environment Programme (UNEP) included an identical provision in the London Guidelines on the Exchange of Information on Chemicals in International Trade, and a voluntary system was put in place with the FAO acting as the Secretariat for pesticides and UNEP for industrial chemicals. The first pesticides were added in 1991, and by 1995, 22 pesticides and five industrial chemicals were included.

From Voluntary to Legally Binding The issue of transforming the voluntary scheme into a legally binding international Convention was first mooted in 1992 at the United Nations Conference on Environment and Development (UNCED).[4] In November 1994, the FAO Council meeting agreed to proceed, and this was followed in May 1995 by a decision of the UNEP Governing Council. The two organizations convened an Intergovernmental Negotiating Committee (INC) to draft and agree international legally binding instrument.

Banning Exports of Banned Pesticides An alternative to PIC strongly advocated at the time was to stop all exports of banned pesticides. However, unless action to limit the market for a banned pesticide could be taken, banning exports could encourage companies to relocate production, possibly in a country with less stringent controls. Preventing the export of banned pesticides would have no effect on severely restricted chemicals. Without a PIC system, a developing country could unwittingly allow the import of banned or severely restricted pesticides, ignorant of action taken by some governments. Many developing countries maintained that an export ban could limit their development, as alternatives were more expensive, and that import decisions should rest with them. PIC does not prevent individual countries from deciding that their banned pesticides should not be exported, but does ensure that regulatory actions are widely shared.

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How the Convention Is Operated In negotiating the text of the Rotterdam Convention, governments built on the experience gained in the voluntary PIC. As a mark of its importance, the Convention began immediately on a voluntary basis, with FAO and UNEP continuing as an interim Joint Secretariat.

Designated National Authorities To participate in PIC, governments must appoint a Designated National Authority (DNA). By December 2000, 170 governments had appointed a DNA or a focal point. When ratifying the Convention, DNAs must be authorized to carry out administrative functions such as receiving, transmitting, and circulating information.

Notifying Regulatory Actions When a government bans or severely restricts a pesticide, it must notify the Joint Secretariat within 90 days. Governments need to demonstrate that their action is final and that it was based on a risk evaluation, including a review of scientific data, and the Secretariat will validate the notification. Once two valid notifications from different PIC regions have been received for the same pesticide, it becomes a candidate for PIC.

Chemical Review Committee The Convention set up a Chemical Review Committee to consider notifications, and advise the Conference of the Parties (CoP—this will replace the INC after ratification). A parallel structure operates in the voluntary phase, with an Interim Chemical Review Committee (ICRC). The Committee will review PIC notifications, and—when they meet the agreed criteria—draft a Decision Guidance Document (DGD).

Two Routes to Be “PIC-ed” Pesticides in the voluntary PIC were carried forward, and new pesticides continue to be added. By June 2001, the process included 26 pesticides and five industrial chemicals (Table 1). There are two routes for adding pesticides to the Convention. Under Article 5, a ban or severe restriction in any two regions triggers PIC if the action is taken for health or environmental reasons. Governments have decided that the PIC regions would be: Africa (48 countries), Latin America, and the Caribbean (33 countries), Asia (23 countries), Near East (22 countries), Europe (49 countries), North America (2 countries: Canada and US), Southwest Pacific (16 countries). The second route is covered in Article 6, and addresses “severely hazardous pesticide formulations.” This category applies only to pesticide formulations found to be causing health or environmental problems under conditions of use in developing countries, or countries with economies in transition. These pesticides may not have been banned, but—generally because of high toxicity— cause poisonings and deaths when used without extreme caution. Governments must submit evidence based on a “clear description of incidents related to the problem, including the adverse effects and the way in which the formulation was used.” Nevertheless, this kind of evidence is rare, and collecting information is difficult: incidents take place far from medical facilities; many farmers are unaware of the active ingredients of pesticides they use; and it is common to use mixtures of several pesticides. The ICRC is investigating how to deal with these problems.

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TABLE 1 Pesticides Covered by the Interim PIC Procedure, November 2000 Banned or severely restricted pesticidesa 2,4,5-T (dioxin contamination) Aldrin Binapacryl (INC6)a Captafol Chlordane Chlordimeform Chlorobenzilate DDT Dieldrin Dinoseb and dinoseb salts 1,2-Dibromoethane (EDB, or ethylene dibromide) Ethylene dichloride (INC7)a Ethylene oxide (INC7)a Fluoroacetamide HCH, mixed isomers Heptachlor Hexachlorobenzene Lindane Mercury compounds mercuric oxide mercurous chloride, Calomel other inorganic mercury compounds alkyl mercury compounds alkoxyalkyl/aryl mercury compounds Pentachlorophenol Toxaphene (INC6)a Severely hazardous pesticide formulationsb Monocrotophos Methamidophos Phosphamidon Methyl parathion Parathion Indicates that these four pesticides were added to the PIC list at the 6th and 7th International Negotiating Committee meetings. b Only certain formulations of these severely hazardous pesticides are included. Source: http://www.pic.int/[5] a

Import Decisions, Information, and Website Once a pesticide is included in PIC, the DGD is circulated to all governments who must decide whether to consent to or prohibit its import. Import decisions are posted on the PIC website, and circulated biannually. Governments in exporting countries must ensure that their exporters comply. Of course, many countries are both importers and exporters and under the rules of international trade, a country cannot ban the import of a pesticide that is manufactured and used nationally. An important tool is the PIC Circular, updated every six months by the Secretariat. Circulated in hard copy and on the website,[5] it includes new bans and severe restrictions, importing country responses,

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and general progress reports. For the first time, it is easy to access sound information on government regulatory actions, even if these do not meet all the full PIC criteria.

The Convention—More than PIC Information exchange is an important principle promoted under Article 14 of the Convention. Developing countries lack resources to undertake extensive evaluations of pesticides and governments are encouraged to share scientific, technical, economic, and legal information on chemicals within the scope of the Convention, as well as other information on their regulatory actions.

Building Capacity/Improving Regulations The process of identifying problem pesticides through PIC will be slow, and there are limitations. In some cases, for example, governments will have no easy substitute, although this may increase the incentive to seek safer and more appropriate alternatives, including Integrated Pest Management strategies. Financial resources are needed, not only to allow the Secretariat to meet its obligations, but also to ensure that regulators in developing countries can participate in workshops and training sessions. In poorer countries, with competing demands on scarce resources, chemical regulation is not always a priority. The status of an international Convention gives PIC the attention it requires to be effective, and should help attract the necessary funds. PIC is just one tool, although an important one, in the regulation of pesticides. With good training and additional resources, PIC can play a central role as part of capacity building initiatives to help governments improve their ability to regulate pesticides, and to look for products and strategies that reduce the dependence on hazardous chemicals.

References 1. Rotterdam Convention on the Prior Informed Consent Procedure for Certain Hazardous Chemicals and Pesticides in International Trade, UNEP and FAO, Text and Annexes, January 1999. 2. The signatory countries can be found on the PIC website: http://www.pic.int/. The Convention closed for signatures in September 1999: countries which have not signed accede to, rather than ratify, the Convention, to the same effect. 3. International Code of Conduct on the Distribution and Use of Pesticides (Amended Version); FAO, 1989. The Code is currently being revised and updated. 4. United Nations Conference on Environment and Development, Agenda 21, Chapter 19, Environmentally Sound Management of the Toxic Chemicals, Including Prevention of International Illegal Traffic in Toxic and Dangerous Products, UNEP, Nairobi, 1992. 5. Convention text and PIC website (http://www.pic.int/).

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36 Laws and Regulations: Soil Introduction .................................................................................................. 395 Why Law for Soil? ........................................................................................ 395 Soil • Soil Degradation • Sustainable Use of Soil

International Law and Soil .......................................................................... 396 Declarations • International Conventions, Covenants, Treaties, and Agreements

Ian Hannam and Ben Boer

National Soil Law ......................................................................................... 397 Effectiveness of Soil Law .............................................................................. 397 IUCN Commission on Environmental Law ............................................. 398 Conclusions ................................................................................................... 398 References ...................................................................................................... 398

Introduction At a national level, soil law means a body of law to promote soil conservation enacted by a legislature, e.g., an act, decree, regulation, or other formal legal instrument that is legally enforceable. Soil law, or “soil legislation” as it may also be referred, includes those laws that have primary responsibility for soil conservation, soil and water conservation, and land rehabilitation. They are generally characterized by provisions to mitigate and manage soil erosion and soil degradation and methods to conserve soil resources. Internationally, the legal framework for the conservation of soil can include conventions, protocols, agreements, and covenants, which are expressed to be legally binding. Worldwide, soil law is managed by a variety of legal and institutional systems, which are the individual organizational and operational regimes that have the administrative authority over soil.

Why Law for Soil? Soil bodies are effectively large ecosystems and comprise fundamental components of the earth’s biodiversity. Soil is thus seen as the basis for the conservation of terrestrial biological diversity and the sustenance of all terrestrial organisms, including people. The ongoing and widespread soil degradation as a result of human use of soil provides the imperative for enactment of soil law. The ever-increasing demand for food by rapidly growing populations in many countries in the past few decades has exerted increasing environmental stress on the soil leading to widespread soil degradation.[1] The following definitions provide the context for soil law.

Soil Soil forms an integral part of the earth’s ecosystems and is situated between the earth’s surface and bedrock. It is subdivided into successive horizontal layers with specific physical, chemical, and biological 395

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characteristics. From the standpoint of history of soil use, and from an ecological and environmental point of view, the concept of soil also embraces porous sedimentary rocks and other permeable materials together with the water that these contain and the reserves of underground water.[2]

Soil Degradation Soil degradation is a loss or reduction of soil functions or soil uses. It includes aspects of physical, chemical, and biological deterioration, including loss of organic matter, decline in soil fertility, decline in structural condition, erosion, adverse changes in salinity, acidity, or alkalinity, and the effects of toxic chemicals, pollutants, or excessive flooding.[1]

Sustainable Use of Soil The sustainable use of soils preserves the balance between the processes of soil formation and soil degradation while maintaining the ecological functions and needs of soil. In this context, the use of soil means the role of soil in the conservation of biological diversity and the maintenance of human life.[3]

International Law and Soil International environmental law is an essential component for setting and implementing global, regional, and national policy on environment and development. There is an increasing recognition of the role of international environmental law to overcome the global problems of soil degradation, including its ability to provide a juridical basis for action by nations and the international community.[4] A number of international and regional instruments introduced in the past 10 years contain elements that can contribute to achieving sustainable use of soil. None are sufficient on their own. Some of the instruments could assist by promoting the management of some of the activities that can control soil degradation. However, this role is not readily apparent except for those that include provisions specifically directed to soil (e.g., see Article IV “Soil”— 1968 African Convention on the Conservation of Nature and Natural Resources, final revision text adopted by the African Union Assembly on July 11, 2003).

Declarations A number of nonbinding declarations and charters draw attention to the fact that soil degradation and desertification are reaching alarming proportions and seriously endangering human survival. They call on states to cooperate and develop the tools to conserve soils. Key declarations relevant to soil include the 1972 Stockholm Declaration on the Human Environment, the 1981 FAO World Soil Charter, the 1982 World Charter for Nature, the 1982 Nairobi Declaration, the 1992 Rio Declaration on Environment and Development, and the 2002 Johannesburg Declaration on Sustainable Development. Also of relevance is the Programme for the Development and Periodic Review of Environmental Law for the First Decade of the 21st Century, known as the Montevideo Programme; this program includes provisions to improve the conservation, rehabilitation, and sustainable use of soils.[5]

International Conventions, Covenants, Treaties, and Agreements Many multilateral agreements include provisions that could be used to promote sustainable use of soil, but the provisions are generally tangential to the needs of soil as such. Key global instruments relevant to soil include the 1992 Convention on Biological Diversity, the 1992 United Nations Framework Convention on Climate Change and the 1997 Kyoto Protocol, and the 1994 United Nations Convention to Combat Desertification. Relevant regional instruments include the 1968 African Convention on the Conservation of Nature and Natural Resources (Revised July 2003), the 1985 ASEAN Agreement on the Conservation

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of Nature and Natural Resources, the 1986 Convention for the Protection of the Natural Resources and Environment of the South Pacific Region, the 1986 European Community Council Directive, the 1995 Convention Concerning the Protection of the European Alps, and the 1998 Protocol for the Implementation of the Alpine Convention of 1991 in the Area of Soil Protection.[6]

National Soil Law Legislation has been used for some 60 years in many countries to control soil degradation problems and to manage soil. A worldwide examination of national legal and institutional frameworks indicates that most countries approach the management of soil in a fragmented manner. The term “soil law” also covers those situations where comprehensive provisions for soil protection and management have been integrated in legislation that protects other aspects of the environment, such as forests, water, biodiversity, and desertification. In general, soil law thus provides for farm planning, implementation of soil erosion control measures, establishing community groups, planning catchment schemes, and compliance and enforcement. Some jurisdictions, such as the United Kingdom, have multiple soil legislation mechanisms that cover a broad range of functions including soil planning, access to sensitive land types, organic farming practices, nitrate sensitive areas, and soil restoration. On the other hand, federally organized countries often have a system where each state or province has its own soil legislation and supportive legal mechanisms. Hybrid situations also exist, such as in the People’s Republic of China, which has enacted the Water and Soil Conservation Law 1991 and the Desertification Law 2002 at a national level, but causes them to be implemented through a comprehensive provincial system of law and regulations. There is a wide variety of types of legal mechanisms used to protect and manage soil, including acts, decrees, resolutions, ordinances, codes, regulations, circulars, decisions, orders, and bylaws. Whereas these are generally appropriate, many need to be applied in more inventive ways to effectively manage the soil in an ecosystem context.[3]

Effectiveness of Soil Law The effectiveness of international and national soil law is generally dependent on two matters: first, the capacity of a legal and institutional framework to manage soil—which is measured by the ability of a legislative and institutional system to achieve sustainable use of soil—and second, by the number and type of essential legal and institutional elements present in a soil statute in a format that enables soil degradation issues to be identified. These need to be backed by the legal, administrative, and technical capability in the particular instrument as a basis of some form of effective action. Capacity is also represented in the form of legal rights, the type of legal mechanisms, and importantly, the number and comprehensiveness of the essential elements and their functional capabilities. Legal and institutional “elements” for soil are the basic, essential components of a legal and institutional system. An individual law can include a number of legal mechanisms in a well thought-out structure that gives an organization the power it needs, through its executive and administrative structure, to address soil degradation. It is also possible that the necessary elements may be distributed among a number of individual laws within a comprehensive national legal and institutional system.[7] Most key soil management issues are multifactorial (i.e., many include a sociological, a legal, and a scientific component), so it is obvious that generally more than one piece of environmental legislation (along with detailed regulations) and many types of legal and institutional elements will be needed to effectively manage soil degradation issues.[7] Legal and institutional elements can be used to assist in the evaluation of an existing law or legal instrument to determine its capacity to meet certain prescribed standards of performance for the sustainable use of soil. They can also be used to guide the reform of an existing soil law or to develop new legislation for the sustainable use of soil. The manner and degree in which an “essential element” is applied will vary according to the particular type of legal mechanism

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concerned and its expected role in a particular jurisdiction. For example, an international legal instrument may include a provision for dispute resolution, but the actual implementation of this provision between states might not rely on, or be influenced by, the existence of similar provisions within a law of either of the disputing states.[7]

IUCN Commission on Environmental Law The Commission on Environmental Law of IUCN (The World Conservation Union) has carried out extensive investigation into the options for a new international instrument focusing on soil. The commission has also identified a variety of ways available for states to approach the task of a detailed legal and institutional analysis and the design of appropriate legal and institutional systems that provides for the effective management of soil. Arising from this work, in which the authors have been centrally involved, two principal strategies can be considered for the development of legal and institutional arrangements for soil. These are: • A nonregulatory strategy which is characterized by elements for education, participatory approaches, soil management, and incentive schemes • A regulatory approach that is characterized by statutory soil use plans that prescribe legal limits and targets of soil and land use, issue of licenses or permits to control soil use, and the use of restraining orders and prosecution for failure to follow prescribed standards of sustainable soil use. These strategies can be approached on a short-term time frame for implementation or a longer-term time frame, which involves substantial reform of existing laws, policies, and institutional and sectoral change.[7]

Conclusions Soil law in the past has been neglected at the international level and, in many of the world’s regions, at the domestic level. However, the growing recognition of soil degradation as a major international environmental issue in the context of the conservation of biological diversity is gradually being addressed, and this is starting to change attitudes toward the benefits of improved international and national legal and institutions for soil.[8] Soil bodies represent complex terrestrial ecosystems. They require careful management of their ecological characteristics through the medium of soil law at a national and international level. This approach is essential for the long-term sustainable use of soil and to meet the food production requirements of the expanding human population of the world, as well as to meet the needs of all flora and fauna that depend on the soil for sustenance.

References 1. Bridges, E.M., Hannam, I.D., Oldeman, L.R., Penning deVries, F., Scherr, S.J., Sombatpanit, S., Eds.; Response to Land Degradation; Science Publishers Inc.: Enfield, NH, 2001. 2. Council of Europe European conservation strategy. Recommendations for the 6th European Ministerial Conference on the Environment; Council of Europe: Strasbourg, 1990. 3. Hannam, I.D.; Boer, B.W. Legal and Institutional Frameworks for Sustainable Soils; The World Conservation Union: Gland, U.K., 2002. 4. Khan, R. International law of land degradation. In International Studies. 30:3; Sage Publications: New Delhi, India, 1993. 5. UNEP. The Montevideo Programme III—The Programme for the Development and Periodic Review of Environmental Law for the First Decade of the 21st Century; 2001 Decision 21/23 of the Governing Council of UNEP, UNEP: Nairobi, February 2001.

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6. Sands, P. Principles of International Environmental Law; Cambridge University Press: Cambridge, 2003. 7. Boer, B.W.; Hannam, I.D. Legal aspects of sustainable soils: International and National. Rev. Eur. Commun. Int. Environ. Law 2003, 12 (2), 149–163. 8. WSSD (World Summit on Environment and Development). A Framework for Action on Agriculture; WEHAB Working Group: New York, 2002.

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37 LEED-EB: Leadership in Energy and Environmental Design for Existing Buildings Introduction .................................................................................................. 401 Introduction to LEED ..................................................................................402 United States Green Building Council • LEED Green Building Rating System • Benefits of Green Building Certification • LEED and Existing Buildings • LEED for Existing Buildings: Operations and Maintenance • Issues Addressed by LEED-EB O&M • Key Distinctions between LEED-EB O&M and Other LEED Products

Overview of LEED-EB O&M .....................................................................405 Minimum Program Requirements.............................................................405 Prerequisites ................................................................................................. 406 Credits and Points • Certification Levels • Registration Process

Rusty T. Hodapp

Implementation Process ............................................................................. 408 Benefits of Green Buildings........................................................................ 408 Energy Efficiency Potential of Green Buildings ...................................... 409 Conclusion .................................................................................................... 410 References ...................................................................................................... 410

introduction Existing buildings comprise a significant proportion of the total building stock in the United States and building operations consume large amounts of resources (energy, water, building materials, land, etc.), while generating great amounts of waste. For example, in the United States, commercial and industrial buildings alone are estimated to be responsible for the following[1]: • • • • • •

38.9% of primary energy use 38% of all CO2 emissions 72% of all electricity use 13.6% of all potable water use 170 million tons of construction and demolition debris Using 40% of raw materials globally

Furthermore, because the average person spends 90% of their time indoors, the quality of a building’s interior environment impacts virtually everyone. This suggests a very personal interest in better buildings 401

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in addition to the national implications of large-scale resource consumption. Issues such as these have driven government, corporate, and personal interest in sustainability and “green” topics. Applied to the building industry, this interest is forcefully seen in the rise of green building certification programs and, in particular, the Leadership in Energy and Environmental Design (LEED) Green Building Rating System™ of the United States Green Building Council (USGBC). Intended to guide the development and verify the performance of green buildings, LEED rating systems have become well accepted as a national standard. Consisting of a number of products, the LEED Green Building Rating System largely focuses on design and construction of new buildings. One rating system, however, known as LEED for Existing Buildings: Operations and Maintenance (LEED-EB O&M) is oriented at the operation, maintenance, and management of existing buildings. LEED-EB O&M is of particular interest to facility managers, energy managers, owners, or others interested in reducing operating costs, improving indoor environmental quality, and minimizing the environmental impact of buildings as a growing body of case study and research evidence suggests that these outcomes are linked to green building practices. This entry presents an overview of the LEED-EB O&M Green Building Rating System, its benefits and distinctions from other LEED rating systems, how it is organized and implemented, and the value of high-performance green buildings.

introduction to LeeD LEED stands for the USGBC’s family of standards for rating “green buildings.” The LEED Green Building Rating System is USGBC’s effort to provide a national standard to define a green building. Used as guideline for design, construction, operation, and maintenance and with third-party certification, LEED provides a consistent, credible means of developing and operating high-performance, environmentally sustainable buildings.

United States Green Building council Formed in 1993, the USGBC has become perhaps the most prominent “green building” organization in America. With more than 18,000 member organizations, a network of 78 local affiliates, and more than 140,000 LEED Professional Credential holders, the non-profit organization works through leaders in all sectors of the building industry to advance buildings that are environmentally responsible, profitable, and healthy places to live and work. Driving its mission to transform the building marketplace to sustainability is the Council’s LEED Green Building Rating System and related training and professional accreditation programs. USGBC also supports education, research, and advocacy programs as well as strategic alliances with key industry, research organizations, and federal, state, and local government agencies to transform the building market.[2]

LeeD Green Building Rating System The LEED Green Building Rating System is a voluntary, consensus-based, market-driven building rating system. LEED assesses sustainability from a whole-building perspective by evaluating five key areas of a building’s performance in terms of economic, environmental, and human health impact: sustainable site development, water savings, energy efficiency, materials selection, and indoor environmental quality. Buildings are awarded different levels of certification (Certified, Silver, Gold, Platinum) based upon the amounts of credits satisfied and points earned. These credits are performance oriented and intended to address specific impacts inherent in the design, construction, operation, and maintenance of buildings. The initial LEED rating system (referred to as LEED Version 1.0) was released in August 1998. LEED 1.0 was extensively modified and released as Version 2.0 in March 2000 with LEED Version 2.1 following

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in 2002 and Version 2.2 in 2005. LEED has continued to evolve, undertaking new initiatives and expanding into a family of products. As of April 2009, the portfolio of LEED rating systems consists of several products targeting specific sectors of the buildings market: • • • • • • • • •

LEED 2009 for New Construction and Major Renovation LEED 2009 for Commercial Interiors LEED 2009 for Core and Shell Development LEED 2009 for Existing Buildings: Operations and Maintenance LEED 2009 for Schools LEE 2009 for Retail LEED for Healthcare LEED for Homes LEED 2009 for Neighborhood Development

With new products, technologies, and design innovations coming into the green building marketplace daily, the Rating Systems and Reference Guides will continue to evolve as necessary to stay current and relevant.[3]

Benefits of Green Building Certification LEED provides a guidebook for the design, construction, operation, and maintenance of green buildings, the general benefits of which will be described in more detail later in this entry. The LEED rating system is flexible in order to provide owners and design teams the ability to accommodate circumstances or goals specific to their project. The rigorous and independent certification process provides firm and compelling proof that the building has achieved the sustainable goals established for it and is performing as intended. The credible assurance that a building is in fact green can be valuable to owners, occupants, investors, and other key stakeholders in the industry as well as the public at large.

LEED and Existing Buildings With the exception of LEED-EB O&M, the LEED family of products is intended to address the design and construction phases of buildings. The primary users of these products are architects, engineers, construction contractors, and building owners. A building designed and constructed to LEED standards has verifiably incorporated green or sustainable features and, therefore, should perform better in the key impact areas (economic returns, environmental impact, and occupant health and comfort) than a typical building built to basic code standards. Addressing the design and construction phase of a building’s life is extremely important because it is in these phases that many irreversible decisions with long-term impacts on the building’s performance are made. Figure 1 depicts graphically what has become well established regarding the life-cycle cost of buildings: 1) the majority (75% or more) of total life-cycle costs occur after construction (i.e., during the O&M phase); and 2) many of the decisions that drive long-term cost occur during programming and design.[4] However, the fact is that new buildings represent a very small percentage of the total commercial building stock in the United States. According to some sources, new construction amounts to 2% of the total stock of buildings in any one year.[5] No doubt it is safe to assume that the majority of these existing buildings were not designed or are being operated and maintained to green standards. In order to improve the building sector’s performance in terms of sustainability, clearly the existing stock must be addressed in addition to new construction. Similarly, since the post-construction phase of a building’s life cycle contributes disproportionately to its total cost, resource consumption, and impact on users, standards for operations and maintenance are necessary to maximize the benefits of green practices in the building sector. This is exactly the focus of the LEED-EB O&M Rating System.

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FIGURE 1

Building life-cycle cost curve.

LEED for Existing Buildings: Operations and Maintenance The LEED-EB O&M Rating System is a set of performance standards for sustainable operations and maintenance of existing buildings of various types and all sizes. It is intended to advance high-performance, healthy, durable, affordable, and environmentally sound practices in existing buildings. LEED-EB O&M provides an entry point into the LEED certification process for the existing building stock. It can be used for buildings new to LEED certification as well as those previously certified under LEED-NC. The USGBC began developing LEED-EB in 2000 and it was tested in a pilot phase involving more than 100 buildings in 2002. The final version (Version 2.0) was released in October 2004. The current version, LEED-EB O&M, was released in April 2009 under the suite of 2009 LEED rating systems and has been further updated in April 2010. The introduction to the LEED 2009 for Existing Buildings: Operations and Maintenance Green Building Rating System states “LEED for Existing Buildings Operations and Maintenance encourages owners and operators of existing buildings to implement sustainable practices and reduce the environmental impacts of their buildings over their functional life cycles.” To achieve this, LEED-EB O&M addresses exterior building site maintenance, water and energy use, environmentally preferred products and practices for cleaning and alternations, waste stream management, and ongoing indoor environmental quality.[6]

Issues Addressed by LEED-EB O&M LEED-EB O&M addresses all the key facets of building operations and maintenance that impact total cost of ownership, the environment, and building occupants. Some examples include the following: • • • • • • •

Energy use Water use Building operations and maintenance Building systems (e.g., mechanical, electrical, plumbing) performance Maintenance of building exterior and site Ventilation and indoor air quality Lighting quality

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Thermal comfort of spaces Green cleaning Recycling programs Green product purchasing programs Management of indoor pollutants and toxic substances Systems upgrades[7]

LEED-EB O&M seeks to address sustainability on an ongoing basis. This largely falls under the scope of those involved in managing and operating buildings, and clearly, their involvement and expertise are necessary to successfully certify a building under LEED-EB O&M. To the extent the benefits of green buildings and popularity of standards like LEED continue growing, market forces will create new opportunities for facility managers, energy managers, etc., demand for their services, and highlight the overall value of their contributions.

Key Distinctions between LeeD-eB o&M and other LeeD Products Although sharing many common features in terms of structure and process with other LEED products, LEED-EB O&M is fundamentally distinct in three key ways. LEED-NC (and the other new buildingoriented products) is essentially a onetime event whereas LEED-EB O&M represents an ongoing process. Second, with LEED-NC, the green building process ends after the design and construction phase. For LEED-EB O&M, the green building phase is a continuous process that deals with ongoing operations, maintenance, and upgrades of a building over its life cycle. Buildings certified under LEED-EB O&M require recertification at least once every 5 years. Finally, given their focus on different phases of a building’s life cycle, LEED-NC is primarily a capital budget event while LEED-EB O&M deals with operating budgets.[8]

overview of LeeD-eB o&M In the same manner as all LEED products, the LEED-EB O&M Rating System is based on evaluations of a building in seven categories: 1. 2. 3. 4. 5. 6. 7.

Sustainable sites Water efficiency Energy and atmosphere Materials and resources Indoor environmental quality Innovation (in this case, in operations) Regional priority

Minimum Program Requirements All projects must meet certain minimum program requirements (MPRs) to be eligible for certification under the LEED 2009 rating systems. MPRs define the minimum characteristics that a project must possess in order to be certified and are intended to: 1) provide clear guidance to users; 2) protect the integrity of LEED program; and 3) reduce challenges during the certification process. The LEED 2009 MPRs for EB O&M are as follows: 1. 2. 3. 4.

Must comply with environmental laws Must be a complete, permanent building or space Must use a reasonable site boundary Must comply with minimum floor area requirements

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5. Must comply with minimum occupancy rates 6. Must commit to sharing whole-building energy and water usage data 7. Must comply with a minimum building area-to-site area ratio The ongoing performance data from buildings required as part of the certification will be compiled and used to establish benchmarks for building performance and provide operators an idea of how their building compares on water and energy use. To further its commitment to improving building performance, the USGBC launched the Building Performance Initiative (BPI) in August 2009 to complement the MPR for ongoing performance data. The BPI will make the data collected available to building owners for analysis and feedback.

Prerequisites Also consistent with other LEED rating systems, LEED-EB O&M requires every project to meet certain prerequisites in order to be considered for certification (see the list of prerequisites by category in Table 1). All prerequisites must be satisfied for a project to be eligible for certification. The prerequisites include such items as minimum levels of water and energy efficiency, building commissioning, no CFC refrigerants, no-smoking policy, and other basic elements of a high-performance, green building operation. A key prerequisite involves a minimum performance period for the building. LEED-EB O&M requires buildings to be in operation for a minimum of 12 continuous months before certifying (3 months for all prerequisites and credits except Energy and Atmosphere Prerequisite 2 and Credit 1, which require a minimum of 12 months).

Credits and Points Buildings achieve certification under all LEED products by accumulating a certain number of credit points. Points can be obtained in any combination within and among the credits and categories (see Table 2 for the credit and point breakdown for LEED-EB O&M). All LEED rating systems have 100 base points, and up to 10 bonus points can be earned through Innovation and Regional Priority credits. TABLE 1

LEED 2009 for EB O&M Prerequisites

Category

Prerequisites

Sustainable sites Water efficiency Energy and atmosphere Materials and resources Indoor environmental quality Total

TABLE 2

0 1 3 2 3 9

LEED-EB O&M Credits and Points

Category Sustainable sites Water efficiency Energy and atmosphere Materials and resources Indoor environmental quality Innovation in operations Regional priority Total

Credits

Points

9 4 9 9 15 3 1 50

26 14 35 10 15 6 4 110

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Each LEED rating system uses the same format for prerequisites and credits. The sections include the following: • Intent—describes the main goal of the prerequisite or credit. • Requirements—specifies the criteria needed to satisfy the prerequisite or credit. • Submittals—specifies the documentation required to demonstrate compliance with the prerequisite or credit. • Potential Technologies and Strategies—identifies means and methods that project teams may consider to achieve the prerequisite or credit.

Certification Levels LEED rating systems allow buildings to achieve various levels of certification based on points achieved (see Table 3 for the certification levels for LEED-EB O&M).

Registration Process With the launch of LEED Version 3 in 2009, USGBC implemented a new certification model. LEED v3 consists of three components: • LEED 2009 rating systems. • An upgrade to LEED Online to make it faster and easier to use. • New building certification model—an expanded infrastructure based on ISO standards, administered by the Green Building Certification Institute (GBCI) for improved capacity, speed, and performance. LEED Online is the primary resource for managing the LEED documentation process. It allows project teams to manage project details, complete documentation requirements for LEED credits and prerequisites, upload supporting files, submit applications for review, receive reviewer feedback, and ultimately earn LEED certification. The GBCI is an independent, third-party organization that has assumed administration of LEED certification for all commercial and institutional projects registered under any of the LEED rating systems. The process of certifying a building under any of the LEED rating systems is essentially the same— the project is first registered with GBCI using LEED Online. Once a project is registered, access to software tools, supplemental resources, sample documentation, credit interpretation rulings, and other essential information is provided. For LEED-EB O&M, the initial application (application for standard review) must be submitted within 60 calendar days of the performance periods used. The application has to include complete documentation for all prerequisites and enough points for certification. GBCI reviews the application and designates each credit and prerequisite as anticipated pending or denied. This preliminary standard review is targeted (but not guaranteed) for completion within 25 business days of receipt of the application. Within 25 days of receiving GBCI’s preliminary standard review, the owner may submit a response including any revised documentation. GBCI will then review and return comments for all credits and prerequisites in response to the preliminary Standard Review and TABLE 3

LEED-EB O&M Certification Levels

Certification Level Certified Silver Gold Platinum

Points Required 40–49 50–59 60–79 80+

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designate each as awarded or denied. This final Standard Review is targeted for completion within 15 business days of receipt of the completed application. The owner then accepts or appeals the final review. Following acceptance of the final certification review, LEED projects: 1. Will receive a formal certificate of recognition 2. Will receive information on how to order plaque and certificates, photo submissions, and marketing 3. May be included (at the owner’s discretion) in the online LEED Project Directory of registered and certified projects 4. May be included (along with photos and other documentation) in the U.S. Department of Energy High Performance Buildings Database[9]

Implementation Process From a practical standpoint, the process of implementing LEED-EB O&M should generally involve the following steps: 1. 2. 3. 4. 5. 6. 7. 8. 9. 10.

Become familiar with the LEED-EB O&M Rating System Gain the support of key decision makers and stakeholders Form a project team Conduct a preliminary building audit and identify corrective actions required to meet prerequisites and/ or opportunities Establish project goals related to target certification level, credits to be pursued, and budget Register the project Create and adopt policies and procedures, implement upgrades, make operational changes, etc., in accordance with the project goals Track performance Assemble and submit required documentation (preliminary and any required corrections or resubmittals) Achieve certification

The minimum performance period for initial certifications under LEED-EB O&M is 12 months. During this period, actual operational performance must be tracked and reported. The performance tracking period can be as long as 2 years depending upon the project goals and/or implementation strategy. The USGBC provides project teams with numerous resources including the LEED Online system for managing and preparing the certification application, credit templates that define supporting documentation needed and compliance calculations, credit interpretation rulings that can help answer questions on credits and implementation strategies, and the LEED Reference Guides.[10,11]

Benefits of Green Buildings The premise inherent in the LEED rating systems is that “green” buildings provide superior value to owners, occupants, and other stakeholders. Typically, the value proposition construct for sustainability, green buildings, etc., is the well-known triple bottom line of economic returns, environmental impact, and social benefit. Green buildings, also known (perhaps more accurately) as high-performance buildings, are premised as providing superior economic returns, reduced environmental impact, and enhanced social benefits. Such buildings in theory: • Were properly built and/or are well operated and maintained • Use resources (e.g., energy, water, building materials, O&M supplies) more efficiently • Provide a safer, more comfortable, and productive working environment for occupants

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In fact, there is a robust and growing body of evidence in research and case study that supports these claims. The following examples are illustrative. A report commissioned by California’s Sustainable Building Task Force found that energy savings alone exceeded the average cost increase associated with 33 different LEED buildings studied. When adding the life-cycle cost benefits of water savings, reduced emissions, operations and maintenance efficiencies, and improved occupant productivity and health, the 20 years net present value of the financial benefits of green buildings exceeded the implementation costs by as much as 10–15 times.[12] Case studies on commissioning alone show that construction and operating costs can be reduced from 1 to 70 times the initial cost of commissioning.[13] Improved thermal comfort, reduced indoor pollutants, enhanced ventilation rates, and other characteristics of green buildings have been found to have positive impacts on occupant productivity, student test scores, absenteeism, and incidences of various sicknesses.[14] Other benefits continue to be demonstrated in case studies and research, including the following: • • • • •

Increased building value Risk mitigation Employee loyalty and recruitment Brand image and public relations Environmental stewardship[15]

Energy Efficiency Potential of Green Buildings In LEED 2009, points are allocated among credits based on the potential environmental impacts and human benefits of each credit. As a result, the allocation of points significantly changed in comparison to previous versions of the LEED rating systems. These changes increased the relative importance of reducing energy consumption and building-related greenhouse gas emissions. Reflecting this, one credit in LEED 2009 EB O&M comprises the largest potential amount of points—Energy and Atmosphere Credit 1 Optimize Energy Efficiency Performance provides the opportunity for 18 possible points. Furthermore, with its emphasis (and associated requirements) on demonstrated performance, LEED-EB O&M presents tremendous potential to reduce energy consumption throughout the commercial building sector. When considering the fact that existing buildings comprise some 95% of the commercial building stock in the United States, the magnitude of potential reduction is immense. To put this potential in context, consider U.S. Energy Information Administration (EIA) projections of the impact of energy efficiency on per capita commercial energy consumption. In their Annual Energy Outlook 2010, EIA estimates per capita commercial energy consumption in 2035 could be decreased by as much as 17.5% depending upon the degree to which technology-based efficiency improvements are deployed throughout the sector.[16] McKinsey and Company estimates that, by 2020, the United States could reduce annual energy consumption by 9.1 quadrillion British thermal units (BTUs) (23%) of end-use energy (18.4 quadrillion BTUs in primary energy) from a business-as-usual baseline by deploying an array of energy efficiency measures with the commercial sector accounting for 25% of this potential. At full potential, the projected efficiency improvements could reduce greenhouse gas emissions by as much as 1.1 gigatons of CO2 by 2020 and serve as a bridge to low-carbon energy sources.[17] Similarly, the Electric Power Research Institute (EPRI) estimates that the combination of energy efficiency and demand response programs has the potential to reduce summer peak electric demand in the United States by 157 GW to 218 GW (14% to 20%) by 2030.[18] Finally, consider the potential national impact of LEED energy savings presented in the Green Building Market and Impact Report 2009. This report projects that given the acceleration of LEED adoption, energy savings in the United States could reach 1.75 quadrillion BTUs by 2020 and 3.9 quadrillion BTUs by 2035 (8.3% and 17.3%, respectively, of national annual commercial building energy use). A best-case scenario could see those savings rise to 22.3% by 2030.[19]

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Conclusion The USGBC’s LEED Green Building Rating System has become a well-recognized standard for guiding the development of green, high-performance buildings and for verifying that established green building goals have been accomplished. The LEED family of rating systems is focused on the design and construction process for new buildings. However, one rating system—LEED-EB O&M— focuses on operations and maintenance of existing buildings. LEED-EB O&M is of particular interest to facility managers, energy managers, and other professionals involved in building operation and management. LEED-EB O&M provides a guidebook for those interested in “greening” their existing building stock. Implementing these green processes and practices can be an effective means of reducing a building’s life-cycle costs, reducing its environmental impact, and improving occupant health and productivity.

References 1. United States Green Building Council. Green Building Facts. Available at http://www.usgbc.org/ (accessed August 2007). 2. United States Green Building Council. About USGBC. Available at http://www.usgbc.org/ (accessed August 2007). 3. United States Green Building Council. LEED Reference Guide for Green Building Design and Construction 2009 Edition; USGBC: Washington, 2009; xii. 4. National Research Council. Investments in Federal Facilities, Asset Management Strategies for the 21st Century; The National Academies Press: Washington, 2004; 27. 5. Architecture2030. Available at http://architecture2030.org/the_ solution/buildings_solution_how (accessed November 2010). 6. United States Green Building Council. LEED 2009for Existing Buildings: Operations and Maintenance Rating System (Updated April 2010); USGBC: Washington, 2010; xvi. 7. Opitz, M. What LEED-EB Is and Why to Use It, available at http://www.fmlink.com/ProfResources/ Sustainability/Articles/article.cgi?USBGC:200604-01.html (accessed August 2007). 8. United States Green Building Council. LEED-EB Presentation. Available at http://www.usgbc.org/ (accessed August 2007). 9. Green Building Certification. GBCI LEED Certification Manual. Available at http://www.gbci. org/main-nav/building-certification/leed-certification.aspx (accessed July 2010). 10. Opitz, M. Starting and Managing Your LEED-EB Project, available at http://www.fmlink.com/ ProfResources/Sustainability/Articles/article.cgi?USBGC:200609-20.html (accessed September 2007). 11. United States Green Building Council. LEED 2009 for Existing Buildings: Operations and Maintenance Rating System (Updated April 2010); USGBC: Washington, 2010; xvi–xxiv. 12. Kats, G.; Alevantis L.; Berman, A.; Mills, E.; Perlman, J. The Costs and Financial Benefits of Green Buildings: A Report to California’s Sustainable Building Task Force, available at http://www.cap-e. com/ewebeditpro/items/O59F3259.pdf (accessed September 2006). 13. ASHRAE. ASHRAE GreenGuide; Elsevier: Burlington, 2006; 14. 14. Callan, D. Green Building Report: Studies Relate IAQ and Productivity. Build. Operating Manage. 2006, 52 (11), 6–8. 15. Yudelson, J. Marketing Green Buildings, Guide for Engineering, Construction and Architecture; The Fairmont Press: Lilburn, 2006; 5–7. 16. U.S. Energy Information Administration, U.S. Department of Energy. Annual Energy Outlook 2010, 2010, 59. 17. Choi Granade, H.; Creyts, J.; Derkach, A.; Farese, P.; Nyquist, S.; Ostrowski, K. Unlocking Energy Efficiency in the U.S. Economy, available at http://www.mckinsey.com/USenergyefficiency (accessed July 2010).

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18. Faruqui, A.; Hledik, S.; Rohmund, I.; Sergici, S.; Siddiqui, O.; Smith, K.; Wikler, G.; Yoshida, S. Assessment of Achievable Potential from Energy Efficiency and Demand Response Programs in the U.S. (2010–2030). EPRI: Palo Alto, CA, 2009; xi. 19. Watson, R. Green Building Market and Impact Report, 2009, available at http://www.greenbiz.com/business/research/report/2009/11/05/green-building-market-and-impact-report-2009 (accessed July 2010).

Taylor & Francis Taylor & Francis Group http://taylorandfrancis.com

38 LEED-NC: Leadership in Energy and Environmental Design for New Construction

Stephen A. Roosa

Introduction .................................................................................................. 413 Concept of Green Buildings ........................................................................ 414 Rating Systems for Buildings ...................................................................... 415 LEED-NC Rating System ............................................................................ 415 LEED Prerequisites Categories and Criteria............................................. 416 Assessing LEED-NC .................................................................................... 418 Conclusion .................................................................................................... 419 References ..................................................................................................... 420

Introduction Land development practices have yielded adverse environmental consequences, urban dislocation, and changes in urban infrastructure. Urban development in particular has long been associated with reduced environmental quality and environmental degradation.[2] The rate at which undeveloped land is being consumed for new structures—and the growing appetite of those structures for energy and environmental resources—has contributed to ecosystem disruption and has fostered impetus to rethink how buildings are sited and constructed. While urban developmental patterns have been associated with environmental disruptions at the local and regional scales, the scientific assessments of global impacts have yielded mixed results. In part as a reaction to U.S. development patterns that have traditionally fostered suburbanization and subsidized automobile-biased transportation infrastructure, design alternatives for structures with environmentally friendly and energy efficient attributes have become available. According to the United Nations Commission on Sustainable Development, “air and water pollution in urban areas are associated with excess morbidity and mortality … Environmental pollution as a result of energy production, transportation, industry or lifestyle choices adversely affects health. This would include such factors as ambient and indoor air pollution, water pollution, inadequate waste management, noise, pesticides and radiation.”[3] It has been demonstrated that a relationship exists between the rates at which certain types of energy policies are adopted at the local level and select indicators of local sustainability.[4] As more urban policies focus on the built environment, buildings continue to be the primary building blocks of urban infrastructure. If buildings can be constructed in a manner that is less environmentally damaging and more energy efficient, then there is greater justification to label them as “green” buildings. 413

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The concept of sustainability has evolved from considerations of land development, population growth, fossil fuel usage, pollution, global warming, availability of water supplies, and the rates of resource use.[5] Thankfully, a vocabulary of technologies and methodologies began to develop in the 1970s and 1980s that responded to such concerns. Driven by ever increasing energy costs, energy engineers began to apply innovative solutions, such as use of alternative energy, more efficient lighting systems and improved electrical motors. Controls engineers developed highly sophisticated digital control systems for heating, ventilating and air conditioning systems. With growing concerns about product safety and liability issues regarding the chemical composition of materials, manufacturers began to mitigate the potential adverse impacts of these materials upon their consumers. Resource availability and waste reduction became issues that began to influence product design. In the span of only 25 years, local governments made curbside recycling programs in larger U.S. cities nearly ubiquitous. Terms and phrases such as “mixed use planning,” “brownfield redevelopment,” “alternative energy,” “micro-climate,” “systems approach,” “urban heat island effect,” “energy assessments,” “measurement and verification,” and “carrying capacity” created the basis for a new vocabulary which identifies potential solutions. All of these concerns evolved prior to the 1992 U.N. Conference on the Environment and Development, which resulted in the Rio Agenda 21 and clarified the concept sustainability. In regard to the built environment, architectural designers renewed their emphasis on fundamental design issues, including site orientation, day lighting, shading, landscaping, and more thermally cohesive building shells. Notions of “sick building syndrome” and illnesses like Legionnaires’ disease, asthma and asbestosis, jolted architects and engineers into re-establishing the importance of the indoor environmental conditions in general and indoor air quality (IAQ) in particular when designing their buildings. The decisions as to what sort of buildings to construct and what construction standards to apply are typically made locally. Those in the position to influence decisions in regard to the physical form of a proposed structure include the builder, developer, contractors, architects, engineers, planners, and local zoning agencies. In addition, all involved must abide by regulations that apply to the site and structure being planned. The rule structure may vary from one locale to another. What is alarming is that past professional practice within the U.S. building industry has only rarely gauged the environmental or energy impact of a structure prior to its construction. Prior to the efforts of organizations like the U.S. Green Building Council (USGBC) (established in 1995), the concept of what constituted a “green building” in the United States lacked a credible set of standards.

Concept of Green Buildings Accepting the notion that sustainable, environmentally appropriate, and energy efficient buildings can be labeled “green,” the degree of “greenness” is subject to multiple interpretations. The process of determining which attributes of a structure can be considered “green” or “not green” is inconclusive and subjective. Complicating the process, there are no clearly labeled “red” edifices with diametrically opposing attributes. While it is implied that a green building may be an improvement over current construction practice, the basis of attribute comparison is often unclear, subjective, and confusing. It is often unclear as to what sort of changes in construction practice, if imposed, would lead the way to greener, more sustainable buildings. If determinable, the marketplace must adjust and provide the technologies and means by which materials, components, and products can be provided to construction sites where greener buildings can arise. Since standards are often formative and evolving, gauging the degree of greenness risks the need to quantify subjective concepts. There are qualities of structures, such as reduced environmental impact and comparatively lower energy usage, which are widely accepted as qualities of green construction practices. For example, use of recycled materials with post-consumer content that originates from a previous use in the consumer market and post-industrial content that would otherwise be diverted to landfills is widely considered an issue addressable by green construction practices. However, evaluation of green building attributes or standards by organizations implies the requirement that decisions be based on stakeholder consensus.

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This process involves input to the decision-making processes by an array of representative stakeholders in often widely diverse geographic locations. For these and other reasons, developing a rating system for green buildings is both difficult and challenging.

Rating Systems for Buildings Rating systems for buildings with sustainable features began to emerge in embryonic form in the 1990s. The most publicized appeared in the United Kingdom, Canada, and the United States. In the United Kingdom, the Building Research Establishment Environmental Assessment Method (BREEAM) was initiated in 1990. BREEAMTM certificates are awarded to developers based on an assessment of performance in regard to climate change, use of resources, impacts on human beings, ecological impact, and management of construction. Credits are assigned based on these and other factors. Overall ratings are assessed according to grades that range from pass to excellent.[6] The International Initiative for a Sustainable Built Environment, based in Ottawa, Canada, has its Green Building Challenge program with more than 15 countries participating. The collaborative venture is geared toward the creation of an information exchange for sustainable building initiatives and the development of “environmental performance assessment systems for buildings.”[7] In the United States, agencies of the central government co-sponsored the development of the Energy StarTM program, which provides “technical information and tools that organizations and consumers need to choose energy-efficient solutions and best management practices.”[8] Expanding on their success, Energy StarTM developed a building energy performance rating system which has been used for over 10,000 buildings. Entering the field at the turn of the new century, the USGBC grew from an organization with just over 200 members in 1999 to 3500 members by 2003.[9] The LEEDTM rating system is a consensus-developed and reviewed standard, allowing voluntary participation by diverse groups of stakeholders with interest in the application and use of the standard. According to Boucher, “the value of a sustainable rating system is to condition the marketplace to balance environmental guiding principles and issues, provide a common basis to communicate performance, and to ask the right questions at the start of a project.”[10] The first dozen pilot projects using the rating system were certified in 2000.

LEED-NC Rating System The USGBC’s Green Building Rating System is a voluntary, consensus-developed set of criteria and standards. This rating system evolved with a goal of applying standards and definition to the idea of high-performance buildings. The use of sustainable technologies is firmly established within the LEED project development process. LEED loosely defines green structures as those that are “healthier, more environmentally responsible and more profitable.”[1] LEED-NC 2.1 is the USGBC’s current standard for new construction and major renovations. It is used primarily for commercial projects such as office buildings, hotels, schools, and institutions. The rating system is based on an assessment of attributes and an evaluation of the use of applied standards. Projects earn points as attributes are achieved and the requirements of the standards are proven. Depending on the total number of points a building achieves upon review, the building is rated as Certified (26–32 points), Silver (33–38 points), Gold (39–51 points) or Platinum (52 or more points).[11] Theoretically, there are a maximum of 69 achievable points. However, in real world applications, gaining certain credits often hinders the potential of successfully meeting the criteria of others. While achieving the rating of Certified is relatively easily accomplished, obtaining a Gold or Platinum rating is rare and requires both creativity and adherence to a broad range of prescriptive and conformance-based criteria. The LEED process involves project registration, provision of documentation, interpretations of credits, application for certification, technical review, rating designation, award, and appeal. Depending on variables such as project square footage and USGBC membership status, registration fees can range up to $7500 for the process.[12]

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LEED Prerequisites Categories and Criteria To apply for the LEED labeling process, there are prerequisite project requirements which earn no points. For example, in the Sustainable Sites category, certain procedures must be followed to reduce erosion and sedimentation. In the category of Energy and Atmosphere, minimal procedures are required for building systems commissioning. Minimal energy performance standards must be achieved (e.g., adherence to ANSI/ASHRAE/IESNA Standard 90.1–1999, Energy Standard for Buildings Except LowRise Residential Buildings, or the local energy code if more stringent), and there must be verification that CFC refrigerants will not be used or will be phased out. In addition, there are prerequisite requirements outlining mandates for storage and collection of recyclable material, minimum IAQ performance (the requirements of ASHRAE Standard 62–1999, Ventilation for Acceptable Indoor Air Quality must be adhered to), and the requirement that non-tobacco smokers not be exposed to smoke. In addition to the prerequisite requirements, the LEED process assigns points upon achieving certain project criteria or complying with certain standards. The total points are summed to achieve the determined rating. Projects can achieve points from initiatives within the following sets of categories: Sustainable Sites (14 points), Water Efficiency (5 points), Energy and Atmosphere (17 points), Materials and Resources (13 points), and Indoor Environmental Quality (15 points). Use of a LEED Accredited Professional (1 point) to assist with the project[13] earns a single point. Additional points are available for Innovation and Design Process (maximum of 4 points). Within each category, the specific standards and criteria are designed to meet identified goals. In the category of Sustainable Sites, 20.2% of the total possible points are available. This category focuses on various aspects of site selection, site management, transportation and site planning. The goals of this category involve reducing the environmental impacts of construction, protecting certain types of undeveloped lands and habitats, reducing pollution from development, conserving natural areas and resources, reducing the heat island impacts, and minimizing light pollution. Site selection criteria are designed to direct development away from prime farmland, flood plains, habitat for endangered species and public parkland. A development density point is awarded for projects that are essentially multistory. If the site has documented environmental contamination or is designated by a governmental body as a brownfield, another point is available. In regard to transportation, four points are available for locating sites near publicly available transportation (e.g., bus lines or light rail), providing bicycle storage and changing rooms, provisions for alternatively fueled vehicles and carefully managing on-site parking. Two points in this category are obtained by limiting site disturbances and by exceeding “the local open space zoning requirement for the site by 25%.”[14] In addition, points are available by following certain storm water management procedures, increasing soil permeability, and attempting to eliminate storm water contamination. Potential urban heat island effects are addressed by crediting design attributes such as shading, underground parking, reduced impervious surfaces, high albedo materials, reflective roofing materials, or vegetated roofing. Finally, a point is available for eliminating light trespass. Water efficiency credits comprise 7.2% of the total possible points. With the goal of maximizing the efficiency of water use and reducing the burden on water municipal systems, points are credited for reducing or eliminating potable water use for site irrigation, capturing and using rainwater for irrigation, and using drought tolerant or indigenous landscaping. This section of the LEED standard also addresses a building’s internal water consumption. Points are available for lowering aggregate water consumption and reducing potable water use. Reducing the wastewater quantities or providing on-site tertiary wastewater treatment also earns points. Energy and Atmosphere is the category that offers the greatest number of points, 24.6% of the total possible. The intents of this category include improving the calibration of equipment, reducing energy costs, supporting alternative energy, reducing the use of substances that cause atmospheric damage, and offering measurement and verification criteria. Optimizing the design energy cost of the regulated energy systems can achieve a maximum of ten points. To assess the result, project designs are modeled against a base case solution which lacks certain energy-saving technologies. Interestingly,

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the unit of measure for evaluating energy performance to achieve credits is not kilocalories or million Btus, but dollars. Points are awarded in whole units as the percentage of calculated dollar savings increases incrementally. In addition to the ten points for energy cost optimization, a maximum of three additional points is available for buildings that use energy from onsite renewable energy generation. Purchased green power is allocated a single point if 50% of the electrical energy (in kWh) comes from a two year green power purchasing arrangement. This category provides points for additional commissioning and elimination of the use of HCFCs and halon gases. Measurement and Verification (M&V) is allowed a point, but only if M&V options B, C, and D, as outlined in the 2001 edition of the International Measurement and Verification Protocol (IPMVP), are used. The Materials and Resources category represents 18.8% of the total possible points. This category provides credit for material management; adaptive reuse of structures; construction waste management; resource reuse; use of material with recycled content; plus the use of regionally manufactured materials, certain renewable materials and certified wood products. A point is earned for providing a space in the building for storage and collection of recyclable materials such as paper, cardboard, glass, plastics and metals. A maximum of three points is available for the adaptive reuse of existing on-site structures and building stock. The tally increases with the extent to which the existing walls, floor, roof structure, and external shell components are incorporated into the reconstruction. LEED-NC 2.1 addresses concerns about construction waste by offering a point if 50% of construction wastes (by weight or volume) are diverted from landfills and another point if the total diversion of wastes is increased to 75%. A project that is composed of 10% recycled or refurbished building products, materials, and furnishings gains an additional two points. Another two points are available in increments (one point for 5%, two points for 10%) if post-consumer or post-industrial recycled content (by dollar value) is used in the new construction. To reduce environmental impacts from transportation systems, a point is available if 20% of the materials are manufactured regionally (defined as being within 500 miles or roughly 800 km of the site), and an added point is scored if 50% of the materials are extracted regionally. A point is available if rapidly renewable materials (e.g., plants with a ten year harvest cycle) are incorporated into the project, and yet another point is earned if 50% of the wood products are certified by the Forest Stewardship Council. The category of Indoor Environmental Quality allows 21.7% of the possible total points available. The goals include improving IAQ, improving occupant comfort, and providing views to the outside. With ASHRAE Standard 62–1999 as a prerequisite, an additional point is available for installing CO2 monitoring devices in accordance with occupancies referenced in ASHRAE Standard 62–2001, Appendix C. A point is also available for implementing technologies that improve upon industry standards for air change effectiveness or that meet certain requirements for natural ventilation. Systems that provide airflow using both underfloor and ceiling plenums are suggested by LEED documentation as a potential ventilation solution. Points are available for developing and implementing IAQ management plans during construction and prior to occupancy. The requirements include using a Minimum Efficiency Reporting Value (MERV) 13 filter media with 100% outside air flush-out prior to occupancy. There are points available for use of materials that reduce the quantity of indoor air pollutants in construction caused by hazardous chemicals and by volatile organic compounds in adhesives, sealants, paints, coatings, composite wood products, and carpeting. A point is offered for provision of perimeter windows and another for individual control of airflow, temperature, and lighting for half of the non-perimeter spaces. Points are available for complying with ASHRAE Standard 55–1992 (Thermal Environmental Conditions for Human Occupancy), Addenda 1995, and installing permanent temperature and humidity control systems. Finally, points are gained for providing 75% of the spaces in the building with some form of daylighting and for providing direct line of-sight vision for 90% of the regularly occupied spaces. In the category of Innovation and Design Process, 7.2% of the total possible points are available. The innovation credits offer the opportunity for projects to score points as a result of unusually creative design innovations, such as substantially exceeding goals of a given criteria or standard.

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Assessing LEED-NC The LEED-NC process has numerous strengths. Perhaps the greatest is its ability to focus the owner and design team on addressing select energy and environmental considerations early in the design process. The LEED design process brings architects, planners, energy engineers, environmental engineers, and IAQ professionals into the program at the early stages of design development. The team adopts a targeted LEED rating as a goal for the project. A strategy evolves based on selected criteria. The team members become focused on fundamental green design practices that have often been overlooked when traditional design development processes were employed. Furthermore, the LEED program identifies the intents of the environmental initiatives. Program requirements are stated and acceptable strategies are suggested. Scoring categories attempt to directly address certain critical environmental concerns. When appropriate, the LEED-NC program defers to engineering and environmental standards developed outside of the USGBC. The components of the program provide accommodation for local regulations. Case study examples, when available and pertinent, are provided and described in the LEED literature. To expedite the process of documenting requirements, letter templates and calculation procedures are available to program users. The educational aspects of the program, which succinctly describe select environmental concerns, cannot be understated. A Web site provides updated information on the program with clarifications of LEED procedures and practice. The training workshops sponsored by the USGBC are instrumental in engaging professionals with a wide range of capabilities. These considerations bring a high degree of credibility to the LEED process. Advocates of the LEED rating system have hopes of it becoming the pre-eminent U.S. standard for rating new construction that aspires to achieve a “green” label. To its credit, it is becoming a highly regarded standard and continues to gain prestige. Nick Stecky, a LEED accredited professional, firmly believes that the system offers a “measurable, quantifiable way of determining how green a building is.”[15] Despite its strengths, the LEED-NC has observable weaknesses. The LEED-NC registration process can appear to be burdensome, and has been perceived as slowing down the design process and creating added construction cost. Isolated cases support these concerns. Kentucky’s first LEED-NC school, seeking a Silver rating, was initially estimated to cost over $200/ft 2 ($2152/m 2) compared to the local standard costs of roughly $120/ft 2 ($1290/m 2) for non-LEED construction. However, there are few comparative studies available to substantially validate claims of statistically significant cost impact. Alternatively, many case studies suggest that there is no cost impact as a result of the LEED certification process. It is also possible that the savings resulting from the use of certain LEED standards (e.g. reduced energy use) can be validated using life-cycle costing procedures. Regardless, LEED-NC fails as a one-size-fits-all rating system. For new construction, Kindergarten to 12th-grade (K-12), school systems in New Jersey, California, and elsewhere have adopted their own sustainable building standards. There are other valid concerns in regard to the use of LEED-NC. In an era when many standards are under constant review, standards referenced by LEED are at times out of date. The ASHRAE Standard 90.1–1999 (without amendments) is referenced throughout the March 2003 revision of LEED-NC. However, ASHRAE 90.1 was revised, republished in 2001, and the newer version is not used as the referenced standard. Since design energy costs are used to score Energy and Atmosphere points, and energy use comparisons are baselined against similar fuels, cost savings from fuel switching is marginalized. In such cases, the environmental impact of the differential energy use remains unmeasured, since energy units are not the baseline criteria. There is no energy modeling software commercially available that has been specifically designed for assessing LEED buildings. LEED allows most any energy modeling software to be used, and each has its own set of strengths and weaknesses when used for LEED energy modeling purposes. It is possible for projects to comply with only one energy usage prerequisite, applying a standard already widely adopted, and still become LEED certified.

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In  fact, it is not required that engineers have specialized training or certification to perform the energy models. Finally, LEED documentation lacks System International (SI) unit conversions, reducing its applicability and exportability. A number of the points offered by the rating system are questionable. While indoor environmental quality is touted as a major LEED concern, indoor mold and fungal mitigation practices, among the most pervasive indoor environmental issues, are not addressed and are not necessarily resolvable using the methodologies prescribed. It would seem that having a LEED-accredited professional on the team would be a prerequisite rather than an optional credit. Projects in locations with abundant rainfall or where site irrigation is unnecessary can earn a point by simply documenting a decision not to install irrigation systems. The ability of the point system to apply equally to projects across varied climate classifications and zones is also questionable and unproven. While an M&V credit is available, there is no requirement that a credentialed measurement and verification professional be part of the M&V plan development or the review process. Without the rigor of M&V, it is not possible to determine whether or not the predictive preconstruction energy modeling was accurate. The lack of mandates to determine whether or not the building actually behaves and performs as intended from an energy cost standpoint is a fundamental weakness. This risks illusionary energy cost savings. Finally, the M&V procedures in the 2001 IPMVP have undergone revision and were not state-of-the-art at the time that LEED-NC was updated in May 2003. For example, there is no longer a need to exclude Option A as an acceptable M&V alternative. The LEED process is not warranted and does not necessarily guarantee that in the end, the owner will have a “sustainable” building. While LEED standards are more regionalized in locations where local zoning and building laws apply, local regulations can also preempt certain types of green construction criteria. Of greater concern is that it is possible for a LEED certified building to devolve into a building that would lack the qualities of a certifiable building. For example, the owners of a building may choose to remove bicycle racks, refrain from the purchase of green energy after a couple of years, disengage control systems, abandon their M&V program, and remove recycling centers—yet retain the claim of owning a LEED certified building.

Conclusion The ideal of developing sustainable buildings is a response to the environmental impacts of buildings and structures. Developing rating systems for structures is problematic due to the often subjective nature of the concepts involved, the ambiguity or lack of certain standards, and the local aspects of construction. While there are a number of assessment systems for sustainable buildings used throughout the developed world, LEED-NC is becoming a widely adopted program for labeling and rating newly constructed “green” buildings in the United States. Using a point-based rating system, whereby projects are credited for their design attributes, use of energy, environmental criteria, and the application of select standards, projects are rated as Certified, Silver, Gold, or Platinum. The LEED-NC program has broad applicability in the United States and has been proven successful in rating roughly 150 buildings to date. Its popularity is gaining momentum. Perhaps its greatest strength is its ability to focus the owner and design team on energy and environmental considerations early in the design process. Today, there are over 1700 projects that have applied for LEED certification. Due to the program’s success in highlighting the importance of energy and environmental concerns in the design of new structures, it is likely that the program will be further refined and updated in the future to more fully adopt regional design solutions, provide means of incorporating updated standards, and offer programs for maintaining certification criteria. It is likely that the LEED program will further expand, perhaps offering a separate rating program for K-12 educational facilities. Future research will hopefully respond to concerns about potential increased construction costs and actual energy and environmental impacts.

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References 1. U.S. Green Building Council. LEED green building rating system. 2004. 2. Spirn, A.W. The Granite Garden: Urban Nature and Human Design; Basic Books: New York, 1984. 3. United Nations. Indicators of Sustainable Development: Guidelines and Methodologies; United Nations: New York, 2001; 38. 4. Roosa, S.A. Energy and Sustainable Development in North American Sunbelt Cities; RPM Publishing: Louisville, KY, 2004. 5. Koeha, T. What is Sustainability and Why Should I Care? Proceedings of the 2004 world energy engineering congress, Austin, TX, Sept 22–24, 2004; AEE: Atlanta 2004. 6. URS. http://www.urseurope.com/services/engineering/engineering-breeam.htm (accessed Feb 2005). 7. iiSBE. International Initiative for a Sustainable Built Environment http://iisbe.org/iisbe/start/ iisbe.htm (accessed Feb 2005). 8. United States Environmental Protection Agency. Join energy star—Improve your energy efficiency. http://www.energystar.gov. (accessed June 2003). 9. Gonchar, J. Green building industry grows by leaps and bounds; Engineering News-Record; 2003. 10. Boucher, M. Resource efficient buildings—Balancing the bottom line. Proceedings of the 2004 world energy engineering congress, Austin, TX, Sept 22–24, 2004; AEE: Atlanta 2004. 11. U.S. Green Building Council. Green Building Rating System for New Construction and Major Renovations Version 2.1 Reference Guide; May 2003; 6. 12. U.S. Green Building Council. LEED—Certification Process. http://www.usgbc.org/LEED/Project/ certprocess.asp. (accessed Oct 2004). 13. U.S. Green Building Council. Green Building Rating System for New Construction and Major Renovations Version 2.1: Nov 2002. 14. U.S. Green Building Council. Green Building Rating System for New Construction and Major Renovations Version 2.1: Nov 2002; 10. 15. Stecky, N. Introduction to the ASHRAE greenguide for LEED, Proceedings of the 2004 world energy engineering congress, Austin, TX, Sept 22–24 2004; AEE: Atlanta 2004.

39 Nanomaterials: Regulation and Risk Assessment Introduction .................................................................................................. 421 Registration, Evaluation, and Authorization of Chemicals (Reach)......422 EU Water Framework Directive .................................................................423 Pharmaceutical Regulation ........................................................................ 424 Nanofood Regulation ...................................................................................425 Risk Assessment of Nanomaterials ........................................................... 426

Steffen Foss Hansen, Khara D. Grieger, and Anders Baun

Hazard Identification of Nanomaterials • Dose–Response Relationship in Regard to Nanomaterials • Exposure Assessment • Risk Characterization • Revisions of the Technical Guidance of Risk Assessment

Conclusion ....................................................................................................430 Refer erences......................................................................................................430

Introduction The topics of regulation and risk assessment of nanomaterials have never been more relevant and controversial in Europe than they are at this point in time. As the first major piece of legislation to be amended in Europe, the cosmetics legislation was adopted in 2009 requiring all nanomaterial-containing cosmetics to be labeled after 2013 and producers to provide a safety assessment of the nanomaterial used.[1,2] Concurrently with the recasting of various pieces of legislation, such as the Novel Foods Regulation, the European Commission has commissioned an expert-/multistakeholder investigation of whether nanospecific amendments are needed to the current technical guidelines on substance identification and chemical safety assessment, which lie at the core foundations of the European Chemical legislation known as REACH—Registration, Evaluation, Authorization, and Restriction of Chemicals.[3,4] It is the major piece of legislation concerning regulating the manufacturing and applications of nanomaterials, although the text in REACH itself has only been subject to minor changes thus far. A number of other pieces of legislation relevant to the manufacturing, use, and disposal of nanomaterial and products have furthermore not been subject to any nanospecific changes, although they might be revised in the future. In the following, some of the major pieces of legislation relevant for the regulation of nanomaterials in Europe will be presented. Examples of both horizontal regulation as well as subject-specific legislation will be given. Some of these have yet to take nanospecific issues into consideration, and the focus will therefore be at explaining the limitations of these in handling nanomaterials. For others, nanospecific aspects have recently been taken into consideration and for these the focus will be at explaining how this has been done and what kinds of challenges still need to be addressed. Chemical risk assessment plays a crucial role in many of these pieces of legislation, and hence a short introduction and discussion of the applicability of chemical risk assessment to nanomaterials will be included. 421

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Registration, Evaluation, and Authorization of Chemicals (Reach) One of the key pieces of European legislation affecting nanomaterials is the European chemical regulation known as REACH, which went into force in mid-2007.[5] REACH prescribes 1. The registration of chemicals commercialized by manufacturers and importers in Europe as well as the collection of data on their use and toxicity. 2. The evaluation and examination by governments of the need for additional testing and regulation of chemicals. 3. That authorization has to be sought and given to manufacturers in order for them to use chemicals of high concern. 4. European Union (EU)-wide restrictions or complete ban of certain chemicals that cannot be used safely. The REACH regulation replaced more than 40 other directives and subsequently shifted the responsibility in the registration and authorization process of REACH onto manufacturers and importers (including downstream users of chemicals) to provide data of uses and hazard information. Industry, furthermore, has to show that chemicals of high concern can be used safely. The evaluation and restriction process is still the responsibility of the national authorities, the newly established European Chemical Agency, and the European Commission. Registration of all the commercialized chemicals in the European market is a tremendous task that is expected to occur gradually. Substances produced or imported in the highest volumes or of the greatest (known) concern are to be registered first. Substances produced or imported in more than 1000 tonnes per year per manufacturer or importer had to be registered by November 30, 2010, by the latest date. This was also the case for substances marketed in 100 tons/yr that have been classified as very toxic to aquatic organisms and for substances produced/imported in more than 1 ton/yr and which have been classified as Category 1 or Category 2 carcinogens, mutagens, or reproductive toxicants. Furthermore, substances entering the European market in yearly quantities above 100 tons, and 10 tons per producers or importers have to be registered by June 1, 2013, and June 1, 2018, respectively.[5] REACH does not specifically mention nanomaterials, but does cover chemicals in all their physical– chemical states, using the following definition of a substance: “a chemical element and its compounds in the natural state or obtained by any manufacturing process, including any additive necessary to preserve its stability and any impurity deriving from the process used, but excluding any solvent which may be separated without affecting the stability of the substance or changing its composition.”[5] Therefore, REACH is formally the relevant legislative frame for industrially used nanomaterials, and the exemption registration of carbon and graphite was redrawn in 2008 to address concerns raised about carbonaceous nanomaterials.[6,7] Companies will now have to register these materials if produced in quantities above 1 ton per producer or manufacturer per year. However, for a number of nanomaterials, it is not evident whether a nanoequivalent of a substance with different physicochemical and (eco)toxicological properties from the bulk substance would be considered as the same or as another substance under REACH.[8] If a nanomaterial is considered to be a different substance under REACH, hazard information specifically related to the nanoform of the substance would have to be generated for the registration, if produced in more than 1 ton/yr. On the other hand, if a nanomaterial is considered to be the same as a registered bulk material, hazard information data generated for the registration might not be directly relevant for the nanoform of the substance and hence open to discussion.[8,9] In response to these concerns, the European Commission has launched a multistakeholder project on nanomaterials to look into substance identification under REACH in order to get recommendations on whether the nanoform of a substance should be considered different from the bulk form of the substance.[3,4]

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If manufacturers and importers produce or import nanomaterials in volumes of more than 10 tons/ yr and if it meets the criteria for classification as dangerous or a PBT (persistent, bioaccumulative, and toxic) or vPvB (very persistent and very bioaccumulative), a chemical safety assessment is required that includes information about uses, (eco)toxicological information, exposure assessment, and risk characterization(s). Thus far, no nanomaterial has been classified as PBT or as vPvB, but if it was to be it is highly unclear how companies should do a chemical safety assessment. Both the Commission of the European Communities[10] as well as the its Scientific Committee on Emerging and Newly Identified Health Risks (SCENIHR)[11] have pointed out that current test guidelines in REACH are based on conventional methodologies for assessing chemical risks and may not be appropriate for assessing risks associated with nanomaterials. It should be noted that a chemical safety assessment can also be required if a nanomaterial is selected for further evaluation by a member state or by the European Chemicals Agency due to specific concerns; or if a substance is a CMR (carcinogenic, mutagenic, or toxic for reproduction), PBT, vPvB, ED (endocrine disrupting), or substance of equivalent concern.

EU Water Framework Directive Whereas REACH deals with the manufacturing and import of chemicals, the EU Water Framework Directive (WFD) deals with improving water quality and reducing dangerous chemicals in European river basins. The key aim of the WFD, which was adopted in 2000, is to promote long-term sustainable water use, preventing further deterioration of surface waters, transitional waters, coastal waters, and groundwater, and to protect and enhance the status of aquatic ecosystems with regard to their water needs, terrestrial ecosystems, and wetlands directly depending on the aquatic ecosystems,[12] Article 1. The WFD establishes water management by a river basin approach with cooperation and joint objective setting across member state borders and even in some cases beyond the EU territory. Geographical and hydrological formation of each river basin determines which member states need to establish and implement a so-called river basin management plan. The river basin management plan, which needs to be updated every 6 years, specifies the measures to be taken to meet the environmental objectives for surface waters, for groundwater, and for protected areas. The WFD prescribes the setting of the environmental quality standards ensuring the general protection of the aquatic ecology, specific protection of unique and valuable habitats, and protection of drinking water resources and bathing water. For instance, for surface waters, member states shall implement necessary measures to prevent deterioration, and promote restoration of artificial and heavily modified water bodies with the aim of achieving “good ecological potential” and “good surface water chemical status” in 2015 by the latest. This has to be done along with a progressive reduction of pollution from a set of “priority substances” and discontinue emissions of priority hazardous substances,[12] Article 4. For all surface waters, the WFD set a number of “general requirements for ecological protection” as well as a “general minimum chemical standard” and defines “good ecological status” and “good chemical status” in terms of the quality of the biological community, the hydrological characteristics, and the chemical characteristics,[12] Article 4. The definition of “good chemical status” is especially relevant in regard to nanomaterials as it is defined in terms of compliance with all the quality standards established for chemical substances at the European level. For “priority substances,” member states are required to set environmental quality standards (EQSs) to monitor the chemical status of a water body (European Parliament and the Council of the European Union (EP & CEU),[12] Article 16). Thus, the EQS is taken as concentration below which the chemical status is referred to as “good” in the WFD terminology (European Parliament and the Council of the European Union (EP & CEU),[12] Article 2). Even for the so-called priority hazardous substances, only a few EQSs have been set, but more substances will follow with a specific focus at substances that

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are toxic toward humans and/or aquatic organisms, compounds with a widespread environmental distribution, and those that are discharged in significant quantities. A key question in regard to WFD is whether nanomaterials are possible candidates as priority substances.[13] In favor of this speaks the widespread and diffuse use of nanoparticles in a range of consumer products along with the hazard characteristics of some nanomaterials such as functionalized carbon nanotubes (CNTs), nanoscale silver, and zinc oxide. Some applications of nanomaterials furthermore involve direct contact with the water cycle, e.g., in relation to their use for water disinfection[14] and wastewater treatment,[15] as well as in regard to the direct use for treating soil and groundwater contamination.[16] If a given type of nanoparticles is included in the list of priority substances in the future based on environmental occurrence or hazard information, an EQS will have to be defined.[12] To derive an EQS for a priority substance, the WFD outlines that test results from both acute and chronic ecotoxicological standard tests should be used for the “base set” organisms, i.e., algae and/or macrophytes, crustacean, and fish. Estimating EQS for nanoparticles is currently hampered by lack of ecotoxicological data even for the most tested nanoparticles such as C60, CNTs, TiO2, ZnO, and Ag. For instance, the degradability of C60 and CNTs and their ability to bioaccumulate in the aquatic environment remains to be studied, making it virtually impossible to set an EQS for these two kinds of nanoparticles.[13] Not only are the number of studies very limited, but the number of tested taxa is also too few to be used in the context of setting an EQS. The reliability and interpretation of the available ecotoxicity data is furthermore impeded as a result of factors such as particle impurities, suspension preparation methods, release of free metal ions, and particle aggregation.[13,17,18] Besides these issues, mainly related to the lack of relevant data, it is also questionable whether the principles for deriving EQSs for chemicals can be directly transferred to nanoparticles. The setting of EQS is based on a chemical safety assessment similar to the one required under REACH and, as noted above, the European Commission’s SCENIHR have pointed out that amendments have to be made to the guidelines for chemical safety assessment.[11,19] Another manner in which nanomaterials could meet the criteria to be included in a WFD list of priority substances is if there is “evidence from monitoring of widespread environmental contamination”.[12] However, when it comes to nanomaterials, monitoring in natural waters represents some profound challenges.[20,21] While applicable methods for in situ monitoring remain to be developed and refined,[22] it is also challenging to set up a reliable monitoring program for nanoparticles since a number of issues still remain to be resolved, e.g., choice of suitable sampling materials, preconcentration/fractionation methods, and analytical methods to characterize and quantify collected particles.[21] Despite significant progress in recent years, reliable methods are not yet available to determine nanoparticle identity, concentrations, and characteristics in complex environmental matrices, such as water, soil, sediment, sewage sludge, and biological specimens.

Pharmaceutical Regulation Liposomes, polymer–protein conjugates, polymeric substances, or suspensions are examples of welldescribed and understood medicinal products containing nanoparticles and have been given marketing authorizations within the EU under the existing regulatory framework.[23–26] As in the case of REACH, nanomedicine and nanomaterials are not specifically mentioned in the EU legislation on medicinal products and devices, tissue engineering, and other advanced therapies. Although the scope of the various EU regulations and directives that constitutes this framework covers nanomedicine, they have been accused of being too general, non-specific, and fraught with difficulties in case of complex drugs.[27,28] Given this, it does seem that it is generally believed that the regulatory framework for medicine covers medical products based on nanotechnology, and that the extensive premarket safety assessment of medicine in general is sufficient to ensure that the benefits outweigh any identified risks or the adverse side effects.[29,30]

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Concerns have, however, been raised that the risk assessment, safety, and quality requirements for medicine may not be designed to address nanomedicine and medical devices based on nanotechnology, as these have to be fulfilled by conformity to established quality systems and published product standards. This might be especially true for novel applications such as nanostructure scaffolds for tissue replacement, nanostructures enabling transport across biological barriers, remote control of nanoprobes, integrated implantable sensory nanoelectronic systems, and chemical structures for drug delivery and targeting of disease.[31] Currently, the mechanism of action is key to decide whether a product should be regulated as a medicinal product or as a medical device. This could be problematic when it comes to many novel applications of nanomedicinal products as they are likely to span regulatory boundaries between medicinal products and medical devices.[29,31] This is due to the notion that they may exhibit a complex mechanism of action combining mechanical, chemical, pharmacological, and immunological properties, and combining diagnostic and therapeutic functions. For new marketing authorization applications of pharmaceuticals, an environmental risk assessment has to be provided, which involves a rough calculation of the predicted environmental concentration (PEC) for surface water. Actions have to be initiated if the PEC is predicted to surpass 0.01 ppb.[32] However, this threshold cannot be interpreted as a safe concentration, and it is not based on a scientific evaluation.[32] It could furthermore be problematic when it comes to nanomedicine, as concentration in terms of mass per volume might not be the relevant metric to characterize the environmental hazard of nanomaterials.[33–35]

Nanofood Regulation Food and food packaging are regulated by a number of directives and regulations in the EU, such as the EU Food Law Regulation and the EU Novel Foods Regulation.[36] As an overarching principle, all food are required to be safe and this overarching principle of safety applies to all foods and food packaging that contain nanomaterials. This has, however, been criticized for being too loose.[37] During the recent discussion related to the update of regulation regarding food additives, the European Parliament’s Committee on Environment, Public Health, and Food Safety stated that it wanted separate limit values for nanotechnologies and that the permitted limits for an additive in nanoparticle form should not be the same as when it is in traditional form.[38] This demand, however, never made into the actual regulation and the final adopted regulation on food additives is limited to requiring that food additives that have been produced via nanotechnology or consist of/or include materials fulfill a number of criteria before it can be include on the list of approved food, food additives, food enzymes, and food flavorings. Nanotechnology and nanomaterials are not defined in the regulation, but these criteria include what use does not pose a safety concern to the health of the consumer at the level of use proposed on the basis of available scientific evidence. Furthermore, there has to be a reasonable technological need that cannot be achieved by other economically and technologically practicable means, and using the food additives should entail consumer advantages and benefits.[39] Another important piece of legislation in regard to food regulation in the EU is the Novel Foods Regulation. This regulation requires mandatory premarket approval of all new ingredients and products. In 2008, the European Commission adopted a proposal to revise the Novel Foods Regulation with the purpose of improving the access of new and innovative foods to the EU market.[40] The definition of novel foods was broadened to include those modified by new production processes, such as nanotechnology and nanoscience, which might have an impact on the food itself. This proposal is currently being discussed in the European Parliament and is going through what is known as a “third reading,” and has to be adopted after a co-decision wherein both the Council of the European Union and the Parliament has to agree on the final text of the regulation. If agreement cannot be reached, it goes to conciliation. There are a number of areas on which the European Parliament and the Council of the European Union disagree in regard to nanomaterials. The requirement of having mandatory labeling is also

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controversial and so is the issue of whether to have premarket safety testing of nanotechnology and nanomaterials in food and packaging.[41] In the first line of revisions suggested to the Novel Foods Regulation, both the Council and the Parliament mention the lack of adequate information and lack of test methods for assessing the risks of nanomaterials.[41] Once the European Commission receives an application for authorization of a novel food, the European Food Safety Authority (EFSA) is responsible for the evaluation of whether a novel food and its use as an ingredient presents a danger to or misleads consumers. By regulation, the EFSA is required to provide assessment on the composition, nutritional value, metabolism, intended use, and the level of microbiological and chemical contaminants. Studies on the toxicology, allergenicity, and details of the manufacturing process may also be considered. No distinction is, however, made in regulation in regard to particle size, and hence nanoparticles will not require new safety assessments if the substance has already been approved in bulk form.

Risk Assessment of Nanomaterials Three different kinds of limitations have been identified in various independent analysis of the applicability of existing regulatory frameworks when it comes to nanomaterials. The first category of limitations are related to the limitations of definitions of what qualifies as a “substance,” “novel food,” etc., when it comes to nanomaterials. For instance, does the definition of a chemical substance cover both the bulk from as well as the nanoform of the substance, and does any given application of nanotechnology to manufacture a given food fall under the definition of a novel food? This issue is currently being discussed in a multistakeholder expert working group; however, this has failed to reach a consensus.[42] In the second category fall requirements triggered by thresholds values not tailored to the nanoscale, but based on bulk material. For instance, for pharmaceuticals, the environmental concentration of medical products has to be estimated before marketing, and if it is below 0.01 ppb and “no other environmental concerns are apparent,” no further actions are to be taken for the medical product in terms of environmental risk assessment.[32] Such a predefined action limit could potentially be problematic since the new properties of nanobased products are expected to also affect their environmental profiles, and this problem has yet to be addressed.[35] The third category of limitations are related to lack of metrological tools, (eco)toxicological data, and environmental exposure limits as required by, e.g., REACH, the pharmaceuticals regulation, and the recast of the Novel Foods Regulation. The availability of (eco)toxicological data and chemical risk assessments is necessary to support existing legislation. In regard to REACH, companies are urged to use already existing guidelines when performing chemical risk assessments, despite the fact that both the European Commission[10] and its SCENIHR,[11] as well as others,[9,10] have pointed out that current test guidelines supporting REACH are based on conventional methodologies for assessing chemical risks and may not be appropriate for the assessment of risks associated with nanomaterials. Chemical risk assessment consists of four elements i.e., hazard identification, dose–response assessment, exposure assessment, and risk characterization. In Europe, legislation for controlling the production, use, and release of chemical substances is based on chemical safety assessment or risk assessment, as described in detail in the “Guidance on Information Requirements and Chemical Safety Assessment”.[43] The guidance totals a staggering number of pages and is issued by the ECHA to help companies carry out chemical safety assessments. It includes extensive technical details for conducting hazard identification, dose (concentration)–response (effect) assessment, exposure assessment, and risk characterization in relation to human health and the environment.[43] Each of these four elements holds a number of limitations that are not easily overcome despite the fact that a lot of effort is being put into investigating the applicability of each of these four elements.

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Hazard Identification of Nanomaterials Toxicity and ecotoxicity have been reported on for multiple nanoparticles (metal and metal oxide nanoparticles, carbonaceous nanomaterials, and quantum dots) in scientific studies; however, many of these need further confirmation. Univocal hazard identification is currently impossible as it is hard to systematically link reported nanoparticle properties to the observed effects. For instance, in regard to multiwalled CNTs (MWCNTs), Poland et al.[44] compared the toxicity of four kinds of MWCNTs of various diameters, lengths, shape, and chemical composition by exposing the mesothelial lining of the body cavity of three mice with 50 mg MWCNT for 24 hr or 7 days. This method was used as a surrogate for the mesothelial lining of the chest cavity. It was found that long MWCNTs “produced length-dependent inflammation, FBGCs, and granulomas that were qualitatively and quantitatively similar to the foreign body inflammatory response caused by long asbestos.” Only the long MWCNTs caused significant increase in polymorphonuclear leukocytes or protein exudation. The short MWCNTs failed to cause any significant inflammation at 1 day or giant cell formation at 7 days. Poland et al.[44] also found that the water-soluble components of MWCNT did not produce significant inflammatory effects 24 hr after injection, which rules out that residue metals were the cause of the observed effects, as others previously had speculated on the basis on in vitro studies.[45,46] The findings by Poland et al.[44] have since then been supported by Ma-Hock et al.[47] and Pauluhn et al.[48] in 90-day inhalation toxicity studies. Less work has been done in regard to exploring the ecotoxicological aspects of nanomaterials, but a number of significant studies have been published. In 2004, Oberdorster[49] published the first ecotoxicological study and reported observed significant increase in lipid peroxidation of the brain of juvenile largemouth bass after exposure to uncoated fullerenes (99.5%) in concentrations of 0.5 and 1 ppm after exposure for 48 hr. C60 was dissolved in tetrahydrofuran (THF), which have since then led to some discussion about whether C60 or the THF was responsible for the effects observed.[50,51] The use of THF is no longer recommended.[18] In regard to CNTs, Templeton et al.[52] compared “as prepared” single-walled CNTs (SWCNTs) with electro- phorectically purified SWCNTs and the fluorescent fraction of nanocarbon by-products. They observed an average cumulative life cycle mortality of 13 ± 4%, while mean life cycle mortalities of 12 ± 3%, 19 ± 2%, 21 ± 3%, and 36 ± 11% were observed for 0.58, 0.97, 1.6, and 10 mg/L. Exposure to 10 mg/L showed: 1) significantly increased mortalities for the naupliar stage and cumulative life cycle; 2) a dramatically reduced development success to 51% for the nauplius to copepodite window, 89% for the copepoddite to adult window, and 34% overall for the nauplius to adult period; and 3) a significantly depressed fertilization rate averaging only 64 ± 13%. A number of studies have furthermore highlighted the need to investigate the potential interactions with existing environmental contaminants or what has become known as the “Trojan horse effect.” For instance, Baun et al.[53] found that the toxicity of phenanthrene was increased toward algae and crustaceans following sorption to C60 aggregates. In contrast, Baun et al.[53] found that the toxic effect of pentachlorophenol decreased when C60 was added. After studying the ecotoxicity of cadmium to algae in the presence of 2 mg/L TiO2 nanoparticles of three different sizes, Hartmann et al.[17] found that the presence of TiO2 in algal tests reduced the toxicity of cadmium. This is thought to be due to decreased bioavailability of cadmium resulting from sorption/complexation of Cd 2+ ions to the TiO2 surface. However, the observed growth inhibition was, however, greater for the 30 nm TiO2 nanoparticles than could be explained by the concentration of dissolved Cd(II) species alone, which indicates a possible carrier effect, or combined toxic effect of TiO2 nanoparticles and cadmium.

Dose–Response Relationship in Regard to Nanomaterials In regard to the second element of chemical risk assessment, it is fundamental that a dose–response relationship can be established so that no-effect concentrations or no effect levels need to be predicted or derived. It is unclear whether a no-effect threshold can be actually be established and what the best

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hazard descriptor(s) of nanoparticles is, and what the most relevant dose metrics and the what the most sensitive endpoints are. Several studies have reported observing a dose–response relationship. This goes for, especially, in vitro studies on, among others, C60, SWCNTs and MWCNTs, and various forms of nanometals. Normally, dose refers to “dose by mass”; however, based on the experiences gained in biological tests of nanoparticles, it has been suggested that biological activity of nanoparticles might not be mass dependent, but dependent on physical and chemical properties not routinely considered in toxicity studies.[54] For instance, Oberdorster and col-leagues[55,56] and Stoeger and colleagues[57,58] found that the surface area of the nanoparticles is a better descriptor of the toxicity of low-soluble, lowtoxicity particles, whereas Wittmaack[59,60] found that the particle number worked best as dose metrics. Warheit et al.[61,62] found that toxicity was related to the number of functional groups in the surface of nanoparticles.

Exposure Assessment Completing a full exposure assessment requires extensive knowledge about, among others, manufacturing conditions, level of production, industrial applications and uses, consumer products and behavior, and environmental fate and distribution. Such detailed information is not available, and thus far no full exposure assessment has been published for any one or more nanomaterials. This may partly be due to difficulties in monitoring nanomaterial exposure in the workplace and the environment, and partly due to the fact that the biological and environmental pathways of nanomaterials are still largely unexplored.[63] Some efforts have been made to assess occupational, consumer, and environmental exposure, however, both to assess the level of exposure and to assess the applicability of current exposure assessment methods and guidelines. These are, however, hampered by the paucity of knowledge, lack of access to information, difficulties in monitoring nanomaterial exposure in the workplace and the environment, and by the fact that the biological and environmental pathways of nanomaterials are still largely unexplored. Hence, they should be seen as “proof of principle” rather than actual assessment of the exposure.[64–67]

Risk Characterization All the information from the first three elements of the risk assessment come together in the fourth and final element of chemical risk assessment, namely risk characterization.[63] In the risk characterization process, exposure levels are compared with quantitative or qualitative hazard information, then suitable predicted no-effect concentrations or derived no-effect levels are determined in order to decide if risks are adequately controlled.[43] Often, risk characterization boils down to the estimation of a risk quotient. For the environment, this is, for instance, defined as the PEC/predicted no effect concentration (PNEC). If the risk quotient is 1, further testing can be initiated to lower the PEC/PNEC ratio. If that is not possible, risk reduction could be implemented. A number of studies reported having completed—or attempted to complete—risk assessments of various nanomaterials such as CeO2, TiO2, C60, and CNTs.[18,67–69] For instance, in regard to the use of CeO2-based diesel fuel additive in the United Kingdom, Park et al.[67] assessed the risk of CeO2 causing pulmonary inflammation. First, they estimated an internal dose of 3.8 × 10−7 cm2/cm2 by converting the retained dose into surface area units and then dividing by the area of the proximal alveolar region of the lung. Then, they compared this value to the highest noobserved-effect level found in a number of in vitro toxicity studies. This value was 26.75 cm2/cm2. Assuming that in vitro exposure data can be accurately projected to the in vivo situation, Park et al.[67] concluded that “it is highly unlikely that exposure to cerium oxide at the environmental levels (from both monitored and modeled experimental data) would elicit pulmonary inflammation.”

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Mueller and Nowack[69] reported having completed the first quantitative risk assessment of nanoparticles in the environment. In a first attempt to derive PEC values, Mueller and Nowack used the threshold concentrations of 20 and 40 mg/L reported in the literature for nano-Ag on Bacillus subtilis and Escherichia coli, and considered it to be equivalent to a no-observed-effect concentration. For nano-TiO2 and CNT, the lowest value found in the literature was 40 ha), crop yield and thus crop water and nutrient use are notoriously variable.[12,14,16] The sources of this variation are related to soil physical and chemical properties, pests, microclimate, genetic and phenological responses of the crop and their interactions.[16] The technology for crop yield mapping is more advanced than current methodologies for determining and understanding the causes of yield variability. Prevailing and traditional management practices treat fields uniformly as one unit. However, reports[15–18] show that to understand underlying soil processes that explain crop yield variability, research must be done at the landscape level and using appropriate statistical tools for large-scale studies.[18]

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Crop Yield and Water Use There is a linear relation between crop yield and water use when the only limiting factor is water[19]; however, root water uptake is synergistically related to nutrient uptake, and the two processes cannot be separated. Precision farming has the potential to improve water and nutrient use efficiency on large fields provided there is quantitative understanding of what factors affect crop water and nutrient use and how they vary across the field.[16,20,21] It is known that crop water and nutrient use are a function of many biotic and abiotic variables, including managed inputs, and harvestable yield is a manifestation of how these variables and inputs interact and are integrated during the growing season. However, it is difficult to determine a hierarchy on the contribution of each input and variable to the measured crop yield using classical statistics.[22–25] Often, variables that affect water and nutrient supply to the plant contribute to crop yield at a high level assuming an adequate plant stand and weed control. The cause-and-effect relation between a single state variable and crop yield is site specific and is difficult to establish without considerable sampling of the soil and/or crop. The establishment of response functions, that is, crop water and nutrient use as a function of variable xi, gives only a partial answer to explain crop water use, nutrient use, and yield based on inputs. The general idea of PA is to optimize input application to the measured crop yield at each sampling location using the law of the optimum formulated by Liebscher, which states that a production factor that is in minimum supply contributes more to production, the closer other production factors are to their optimum.[26] This is a simple premise; however, the decisions for variable-rate application of any agronomic input must consider temporal and spatial variability of the soil’s properties affecting crop growth, water and nutrient use, and yield. Soil factors that affect water storage, such as depth of the root-restricting layer and soil textural differences, must be considered in any precision farming operations that attempt to improve crop water use and yield related to agronomic inputs. Similarly, to improve the use of any micro- and macronutrient by the crop, the overall cycle of the nutrient must be considered, including its availability in the soil and demand by the crop. Examples of managing nitrogen fertilization and irrigation at a site-specific and farm scale for cotton production is given by Bronson et al.,[17] Li et al.,[18] and Booker et al.[27] Precision farming must incorporate the inherent spatial and temporal variability of soil physical, chemical, and biological factors within a field for input management. Accurate representation of spatial and temporal variability in a field requires taking and analyzing many samples. Sampling is normally done on a grid with a scale that can vary from one to several hundred meters.[22,23] Once properties are measured, geostatistical tools (e.g., semivariogram, kriging, cokriging), and other spatial statistical tools (e.g., autocorrelation, cross-correlation, state-space analysis), can be used to establish statistical relations in space and to minimize the number of soil samples to characterize and map fields.[22,23,25] The number of samples required a priori to determine spatial and temporal variability is perhaps the single largest deterrent in the application of precision farming practices to manage and improve crop water and nutrient use. Collection of field data to characterize the spatial variability must remain a priority for any study that attempts to understand how to maximize crop yield across the landscape. The trend of using data generated from pedotransfer functions[28] rather than field measured data is of concern and results thus obtained are preliminary at best and should be verified using measured field data. Further, input data generated from pedotransfer functions and used to assess the spatial variability of soil properties can be used as a guide to establish a sampling scheme to measure the properties of interest. Input data generated from pedotransfer functions are not a substitute to field measured data. There is not much information published on combined crop water and nutrient use across large fields at the landscape level and in the context of precision farming.[17,29–31] An exception is the study by Li et al.[18] where cotton water and total nitrogen uses were measured along a 700-m transect to illustrate the landscape pattern of cotton water and total nitrogen uses and to determine the underlying soil processes governing cotton lint yield variability. In this study, state-space analysis[18,22,25] was used to formulate management decisions that may improve crop water and nitrogen use and lint yield using precision farming practices. An additional study regarding variable-rate nitrogen at different locations within a

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48-ha field is given by Bronson et al.,[17] and in a 14-ha field, by Booker et al.[27] A global-scale assessment showing that global yield crop variability is a function of fertilizer use, irrigation, and climate is given by Mueller et al.[32] and a review of integrated nutrient management to sustain crop productivity while minimizing the environmental impact is given by Wu and Ma.[33]

Landscape Crop Water Use The concept of crop water and nitrogen use in a 60-ha field study is given by Li et al.[18] and Li and Lascano.[34] In 1999, a field experiment was conducted near Lamesa, Texas, on a research farm of Texas A&M University on the southern edge of the High Plains of Texas. The soil was classified as an Amarillo sandy loam. The field was 60 ha with slopes ranging between 0.3% and 6.3%.[18] To evaluate the effect of soil water, nitrate-nitrogen (NO3-N), and topography on cotton lint yield across the landscape, two irrigation levels were used. The irrigation treatments consisted of water applications at a 50% and 75% grass reference evapotranspiration (ETo) with a center-pivot low-energy precision application irrigation system.[35] At each irrigation level, one transect was established following the circular pattern of the center pivot. The two transects were instrumented with 50 neutron access tubes, each 15 m apart, and soil volumetric water content (θv) was measured periodically throughout the growing season. At each point, θv was measured in 0.3 m depth increments to 2.0 m depth using a neutron probe calibrated for this soil. In addition, at each transect point soil texture, soil and plant NO3-N, leaf area index, cotton lint yield, slope, plant density, and other parameters were measured.[18,34] A comprehensive study to evaluate site-specific management of cotton production systems at a landscape scale was done by Booker et al.[36,37] using the Precision Agricultural Landscape Modeling System (PALMS),[38] which was integrated with the cotton simulation model GOSSYM.[39] The combined model, PALMSCot, was applied to simulate a cotton crop irrigated with a 400-m center pivot system and covering a ¼-section of land (~65 ha). The PALMSCot model is grid-based (10-m resolution) and defines up to 26 soil depth layers and for each “cell” the water, energy, nutrient, and carbon balance is calculated on a ¼-h time interval for the length of the growing season. This model provides a detailed calculation of the water and nitrogen use for each cell across the field. For example, for an irrigated field with a 400-m-long center pivot sprinkler system, PALMSCot calculates the cotton lint yield, evapotranspiration, and nitrogen use for each of the 5,026 cells defined by the 10 × 10 m grid system (length and width) and the assigned soil depth.

Statistical Calculations It has been shown that the use of classical statistics, such as regression analysis and analysis of variance, is designed to describe the strength of the covariance structure between variables and fails to completely explain the cause and effect between, for example, crop yield and measured soil variables in precision farming experiments.[18,22,23,25,31] These techniques, in general, account for spatial and temporal variability through blocking and do not describe the spatial and temporal structure. Instead, there are other more appropriate statistical tools for relating the variability of soil and plant parameters measured in space and time. For example, the structure of the spatial (or less often, temporal) variance between measurements may be derived from the sample semivariogram, which is the average variance between neighboring measurements separated by the same distance. Spatial or temporal structure between variables is often determined using autocorrelation and cross-correlation functions. Although these techniques can be used to evaluate the temporal variability structure, they are most often used in PA to analyze spatial variables. Autocorrelation measures the linear correlation of a variable in space along a transect. Cross-correlation is the comparison of two variables measured along a transect and is used to describe the spatial correlation between two landscape variables, that is, where one variable, the tail variable, lags behind the head variable by some distance. The spatial association between several variables can be described using state-space analysis, which is a multivariate autoregressive technique.[18,22]

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Spatial Analysis of Crop Water Use To illustrate the variability of crop water use or crop evapotranspiration (ETc), values measured along the 50% irrigation transect were selected.[18] In Figure 1, the relation between the scaled ETc and elevation, both as a function of distance along the transect, is shown. The ETc data are scaled to the maximum value of 426 mm of water, which was measured 210 m from the south end of the transect. These results show, as expected and, that in general higher ETc was measured at lower elevations and that ETc tended to decrease at the higher elevations. Higher elevations are eroded and the depth of the root zone is shallower holding less water. In contrast, lower elevations tend to have more clay and thus hold more water. Calculated values of soil volumetric water content for the surface 0.05 m for a center pivot irrigated cotton field in Floydada, Texas, are shown in Figure 2.[36,37] The four circles shown in Figure 2 illustrate the corresponding wetted soil area (blue) by the center pivot, with values of 45% soil water content over a 6-day period, starting on the 10 July and ending on the 16 July. This is an example of the spatial resolution, 20 × 20 m, and temporal variability that can be used to analyze crop water use across the landscape that can be obtained with a model such as PALMSCot. Spatial cross-correlation between cotton lint yield and soil water, cotton lint yield and site elevation, and soil water and site elevation are shown in Figure 3. For a 95% confidence interval, the cotton lint yield was positively cross-correlated with soil θv across a lag distance of ±30 m. Cotton lint yield and θv were negatively cross-correlated with elevation at a lag distance of ±30 m. These results show the effect of topography on the θv and of crop water use measured along the transect. Similar results are given in other reports.[17,30,31,40] In this example, the cross-correlation between θv and elevation shows the spatial structure of measured variables and, further, that more water was stored in lower elevations, resulting in higher ETc.

FIGURE 1

Scaled crop evapotranspiration and elevation as a function of distance along a 700-m transect.

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FIGURE 2 Calculated values of soil volumetric water content for the surface 0.05 m obtained with the PALMSCot model using 20 × 20 m grid cells for the period from 10 to 16 July 2011. This is a center pivot irrigated cotton production field located in Floydada, Texas. The soils at the site are classified as a Pullman clay loam with 0%–1% slope and an Olton clay loam with a 1%–3% slope. The blue zones correspond to irrigation water applied by the center pivot sprinkler system. Source: Booker et al.[37]

Linear regression analysis between θv and cotton lint yield and relative site elevation is shown in Figure 4, and the state-space analysis for the relation between cotton lint yield and three measured parameters is shown in Figure 5. Results in Figure 4 show the shortcomings of using an inappropriate statistical tool to understand underlying processes explained with the state-space analysis. This analysis (Figure 5) quantified how cotton lint yields varied as a function of distance and showed that by using θv, soil NO3-N and elevation the variation in cotton lint yield can be explained with a high level of confidence. While studies like those of Li et al.,[18] Bronson et al.,[17] and Booker et al.[27] are empirical in design, the relationships that are evaluated provide important validation and field testing of the more mechanistic mass and energy balance accounting provided by models such as PALMS.[38] These studies also provide foundational information for developing PA management strategies at the crop production scale. Benefits of PA to improve crop water and nutrient use may be obtained by an economic analysis of maximizing crop yield as a function of application of nitrogen fertilizer and irrigation water as given by the state-space equation.[18,34] In the example given, the decision can be made to apply more

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FIGURE 3 Cross-correlation as a function of lag distance. (a) Lint yield and soil water, (b) lint yield and elevation, and (c) soil water and elevation. Shown is the 95% confidence for the cross-correlation distance. Source: Li et al.[18]

FIGURE 4

Soil water content (θv) and cotton lint yield as a function of site relative elevation.

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FIGURE 5 State-space equation relating cotton lint yield (Y) to water content (W), nitrogen level (N), and elevation (E) as a function of distance and location (i) along a 700-m transect. Source: Li et al.[18]

nitrogen fertilizer to the lower areas of the field that also hold more water and increase crop water use and nitrogen and lint yield. With the introduction of variable rate planters, it is possible to discriminate site locations and plant more “drought”-tolerant varieties or change the seeding rate in areas that are prone to have less soil water. This implies the delineation of management zones[14,15] within a field that are defined based on potential crop water use and their interaction with other input variables to maximize economic yield across the field. This type of precision farming is slowly being adopted, and wide use remains within the realm of possibilities that this type of farming has to offer. The introduction and use of computer models of cropping systems will likely expedite and facilitate the adoption of PA management techniques.[38] Y(50% ET )i = −0.201 Yi −1 + 1.107 Wi −1 + 0.332  N i −1 D 49.54  Ei −1 + ε i A final consideration is the cost/benefit of PA practices and its impact on agriculture. Currently, hardware for variable-rate application of agronomic inputs is relatively expensive and in many cases unavailable; however, with increased adoption and use of these practices, the cost will be reduced. For example, tractor guidance systems[41,42] were quickly adopted by producers, and high demand reduced its cost. Further, environmental and material cost concerns for a given area will probably place limits on the amount of certain nutrients, for example, nitrogen fertilizer, used for crop production. This will force producers to apply nitrogen and other nutrients across the field according to site-specific needs and position along the landscape. These practices will be beneficial from both an environmental and an economical point of view.[10,11,20,33]

Future of PA Considerable PA-related research, similar to that presented above, has been conducted over the past decade, studying empirical relationships and attempting to better understand the underlying processes controlling crop yield variability. Much of this research has focused on grid soil and crop sampling,[43] surface characterization (e.g., apparent electrical conductivity),[44] and ground- or aerial-based remote sensing.[3,4,45,46] Such research has described numerous useful process relationships but has been somewhat less successful in providing broad-based tools to support production-scale management. The lack of development of decision-support systems to implement precision decisions is the major impediment to the adoption of PA.[47] Recent advances in the availability of soil data provided by the U.S. Department of Agriculture, elevation data provided by the U.S. Geological Survey, and weather data from weather Mesonets (networks)

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provide the necessary input to model the water, energy, nutrient, and carbon balance of large-scale agricultural fields. An example of such a model is the Precision Agricultural-Landscape Modeling System (PALMS) given by Molling et al.[38] The integration of PALMS with crop growth models[36,37] provides a framework whereby site-specific management of crops is an achievable goal. The concept of using simulation models to manage crops was introduced in the 1970s. Many of the theoretical algorithms related to model crop photosynthesis and transpiration were formulated by C.T. de Wit and coworkers at Wageningen University, Netherlands.[48] An example of such model is the simulation of field water use and crop yield given by Feddes et al.[49] An example of a crop-specific model, that is, cotton, known as GOSSYM/COMAX was developed by McKinion et al.[39] This model was used by crop consultants in the Texas High Plains to provide services on irrigation scheduling and application of nitrogen fertilizer, growth regulators, and chemicals to terminate the crop.[50] The biggest drawback ofthe application of these models was that the required inputs, soil and weather, were both difficult and expensive to obtain. Furthermore, these models provided only an average estimate of crop yield for the entire field regardless of size. The models could be run separately for different parts of a field, but this increased the demand on limited computer and input resources and even then did not represent the interaction between various parts of the field. In retrospect, we now recognize that these models were ahead of their time. Given the current availability of soil and weather data that is required by these models, along with the increased computer speed and reduced cost, a resurgence in the application of simulation models to manage crops and cropping systems is anticipated. In coming years, a likely scenario to emerge to manage cropping systems will be based on the combination of three factors. First, is the realization from site-specific management that shows that crop yield varies temporally and spatially and that increases in crop yield are possible by targeting different amounts of an input, for example, irrigation water and fertilizer, to specific parts of the field. Second, crop management, from planting to harvesting, is complex, and simulation models can be used as a decision aid. Use of crop simulation models, developed in the 1970s–1980s, is facilitated due to the increased availability of required soil, elevation, and weather input data to execute the models and reduced cost of computer hardware. The third factor is an increased awareness of producers on production efficiency and environmental concerns. For example, in many agricultural areas, the amount of nitrogen fertilizer that can be applied is restricted and linked to the residual nitrogen in the soil and its potential effect on contamination of surface water and groundwater.[11,13,14,33,51] Advances in management information systems, development of computer software, and communication via the Internet provide us with the tools to manage a crop in real time.[1,6,36,52] We are currently working toward the development of a PA model that includes all of the above factors.[52,53] The integration of a landscape-scale model such as PALMS[38] with a cotton growth model[39] using site-specific management of water and nitrogen[17,18] can give us the tools to manage, for example, a 50-ha irrigated cotton field. The model provides three key features important to real-time production-scale management: (1) it represents the variability in space and time within the entire field and accounts for hydrologic interactions between areas within the field; (2) it provides water, energy, nutrient, and carbon balance information without reliance on field-installed hardware that must be avoided during management operations; and (3) it can provide predictive information that can support various what-if scenario evaluations (something that physical field measurements do not provide). For example, this field can be divided into 5,000 (100 m2) or 20,000 (25 m2) cells, and the model will calculate a cotton lint yield value for each cell, using weather data collected at or near the field and previously collected soil and elevation data, both of which are stable and once developed can be used for many growing seasons. Further, the estimate of cotton lint yield is based on interactions of soil–plant–weather parameters, and the model itself can be used to explore what combination of inputs (e.g., water and nitrogen) would give the highest economic yield while minimizing leaching of nitrogen below the root zone. This is a current topic of research of our cropping system research unit.[52,53] The input and output terms of the water balance of a cropping system are usually well quantified except for information on the input variable rain, that is, frequency and rate of rainfall events across the

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landscape.[53] Rainfall events are normally measured at a single point in space and seldom is the rate of individual rain events measured and/or recorded. Determining, the amount of water from a rain event that is stored in the soil, that is, effective rain, and runs off is key to correctly model the water balance across the landscape. For example, in the semiarid Texas High Plains, about 86% of annual rain events are 10 micron PM10 PM2.5 Polycyclic aromatic hydrocarbons (PAH) CO, heavy metals Inorganic emissions, others Radionuclides (e.g., radon-222) Pollutants emitted to water BOD, COD, DOC, TOC, inorganic compounds (NH4, PO4, NO3, Cl, Na) Individual hydrocarbons, PAH Heavy metals Pesticides NO3, PO4 Pollutants emitted to soil Oil, hydrocarbon total Heavy metals Pesticides

1.40 1.20 2.00 1.50 2.00 3.00

1.50 3.00 1.50 3.00 5.00

1.50 1.50

Source: Data from Frischknecht et al.[8]. c, combustion emissions; p, process emissions; a, agricultural emissions.

1.80 1.50 1.50

1.50 1.20

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Case Study I—System Analysis of BiofuelBased Electricity Generation Process This example demonstrates the application of a systems approach in assessing the merits of biofuels as a green process. The LCA generates a profile of all the steps involved in the process chain—both onsite and upstream—to provide quantitative information on potential impacts of an industrial activity (i.e., a system). Usually, this is done in terms of the released emissions (i.e., burdens), using a linear model. It takes into account the emission factors of all the known pollutants, inventoried during controlled analysis following standard monitoring protocols. These are then scaled by the volume of the industrial activity to provide their actual burdens. Also, in the case that LCA data are available for multiple similar installations, their respective operational performances can be benchmarked and links between operational efficiency and environmental impacts can be established.[9] A number of LCA studies available in the literature compare the environmental impacts of energy production (heat and/or electricity) from co-firing different biofuels in an existing coal-fired power plant, showing their overall greenhouse gas benefits. These cover the use of agricultural residues such as straw and residual wood, short rotation coppiced (SRC) wood,[10] and perennial rhizomatous grasses.[11,12] A more recent study quantifies the airborne emissions from different biomass-based electricity production systems using different technologies, feedstocks, and scales in order to establish the extent to which offsite emissions may contribute to overall environmental impact.[13] In LCA, the emission factors are adjusted according to the pollution abatement technologies used in the industry to compensate for the release of acidic gases, although they do not reflect the fate of the emissions once they are out of the stack. Thus, this approach remains capable of providing realistic emission scenarios as long as the fate of the released emissions is not altered significantly by the dispersion and chemical transformation in the surrounding media (air, water, or soil). Whereas it allows successful prediction of greenhouse gas emissions in a fairly straightforward manner (assuming minimal phase alterations), the modeling of the gas–particle interactions leading to quantification of total particulate matter (PM) loading is far from complete. For example, the PM emissions calculated in the LCA from a biofuel combustion plant represent mainly the dust emissions from fly ash.[14] Any additional aerosols generated from gas-phase interactions of the resulting emissions, either during biofuel cultivation or combustion, would not be adequately quantified within this approach. This case study utilizes the power of LCA as a diagnostic tool to track the pollutants and carbon emissions over the whole life cycle and mainly focuses on feasible management options for mitigating secondary aerosol generation potential from photochemical neutralization of the acidic emissions with ammonia using the following precursor chemistry[15]: NH3(g) + HCl (g) ↔ NH+4(aq ) + Cl−(aq ) A cradle-to-gate system is applied to all the energy systems modeled in this study, accounting for all the relevant flows involved in the extraction of resources (renewable/non-renewable) and cultivation of biofuels leading up to production of the required amount of energy outputs. As shown in Figure 2, the system comprises a series of foreground and background activities. The foreground activities are considered as the focal point of the system. Shown in the central part as shaded region in the figure, it involves storage and preprocessing of the fuel and its combustion to produce electricity. The background activities primarily include the cultivation and harvesting of the biofuel, production of required chemicals, and their transport to the power plant. The other end of the process chain accounts for the disposal of wastes generated. The atmospheric emissions from all these stages have been accounted for in the models. Electricity production from different renewable biofuel sources has been modeled using scenario analysis. Representative biofuel options currently feasible have been considered. The analysis in this

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study is based on the functional unit defined as “1 terajoule (TJ) electricity produced from biomass and delivered to the grid.” It is achieved by using available LCI data for a 50-MW electricity steam turbine/ back pressure cogeneration plant firing solid biomass.[16] It is important to note that, in this study, the modeled system has not been credited for the cogenerated heat in the process. The following five different electricity production scenarios, representative of the technology for 2010 (base scenarios), have been considered, each using different types of biofuels: Scenario A: Perennial rhizomatous grass (Miscanthus giganteas). Scenario B: SRC chips. Scenario C: Residual/waste wood. Scenario D: SRC chips–grass blend (by energy); SRC chips (80%) and perennial rhizomatous grass (20%). • Scenario E: Waste wood–grass blend (by energy); waste wood (80%) and perennial rhizomatous grass (20%). • • • •

In scenarios D and E, a fuel mix of perennial grass and wood in 1:4 ratio (by energy) has been considered. This is meant to improve the combustion quality and minimize excessive atmospheric emissions. In all the base scenarios, the biofuels have been assumed to be sourced locally (50 km distance to the combustion plant) and transported from their origin to the energy production site using 40-ton payload trucks. This is mainly aimed to investigate potential local air quality degradation from interactions of emissions from the power plant and the cultivation sites. The atmospheric burdens for almost all the base scenarios (A–E) are dominated by pollutant emissions from the power plant (Figure 3). Overall, there is no clear winner among the scenarios. For example, scenarios C and E (waste wood) are best for CO2, NH3, non-methylated volatile organic carbons (NMVOCs), and CH4, and scenario B (SRC chips) has the lowest HCl emissions. Scenario A (miscanthus) is best for N2O and PM; however, it is worst for HCl and SO2. For both miscanthus and SRC wood (scenarios A, B, and D), only biomass cultivation is considered to be the main source of NH3 release to the local environment. However, it can be noted that a considerable amount of CO2 is emitted during biomass production, transport, storage, and drying processes for these scenarios, which are mainly associated with the use of fossil-based energy sources in these stages. It has been reported that the forest logging operations involved in harvesting SRC wood contribute to significant atmospheric emissions

FIGURE 3 Environmental burdens of atmospheric emissions for base scenarios A–E. Source: Tiwary and Colls.[3]

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of CO2, CO, NMVOCs, and PM due to fuel, chainsaw, and hydraulic oil consumption by heavy-duty diesel engine vehicles (Athanassiadis, 2000). This is reflected for the base scenarios B and D whose emissions of these pollutants from the “biomass production and local transport” stage are much higher compared with the rest of the scenarios. In case of waste wood, no mechanical chipping was assumed to be involved and hence the corresponding emissions from scenarios C and E have been relatively much lower. However, the power plant loadings of NOx, SO2, HCl, and PM in these two scenarios are found to be much higher than in scenarios B and D. This could be due to their incomplete combustion, which has been reported in earlier studies to result in highly variable emissions.[17,18] Emissions of CO and NOx per terajoule electricity output from power plant alone show comparable values for all the base scenarios but on a life cycle basis, that is, including the biofuel sourcing and storage stages, waste wood (scenarios C and E) seems to have up to 25% higher NOx burdens, whereas SRC wood (scenarios B and D) seems to have up to 100% higher CO burdens. The corresponding emissions of SO2 and HCl from miscanthus plant (scenario A) are higher by as much as 120% and 350%, respectively, compared with the rest of the scenarios. Therefore, scenario A poses the maximum likelihood of secondary aerosol generation potential through interactions of the acidic gas emissions from the power plant with the ammonia released from nearby harvest fields. On the other hand, SRC wood combustion plant (scenario B) has much lower HCl emissions (40 wt.%)a Early winter harvest (20–40 wt.%)a Late winter/early spring harvest (