NATURAL and ENHANCED REMEDIATION SYSTEMS
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NATURAL and ENHANCED REMEDIATION SYSTEMS
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NATURAL and ENHANCED REMEDIATION SYSTEMS Suthan S. Suthersan
LEWIS PUBLISHERS Boca Raton London New York Washington, D.C.
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Library of Congress Cataloging-in-Publication Data Suthersan, Suthan S. Natural and enhanced remediation systems / by Suthan S. Suthersan. p. cm. — (Arcadis Geraghty & Miller science and engineering) Includes bibliographical references and index. ISBN 1-56670-282-8 1. Soil remediation. 2. Groundwater–Purification. 3. Hazardous wastes–Natural attenuation. 4. Bioremediation. I. Title. II. Geraghty & Miller environmental science and engineering series. TD878.S873 2001 628.5—dc21 2001029566 CIP
This book contains information obtained from authentic and highly regarded sources. Reprinted material is quoted with permission, and sources are indicated. A wide variety of references are listed. Reasonable efforts have been made to publish reliable data and information, but the author and the publisher cannot assume responsibility for the validity of all materials or for the consequences of their use. Neither this book nor any part may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, microfilming, and recording, or by any information storage or retrieval system, without prior permission in writing from the publisher. The consent of CRC Press LLC does not extend to copying for general distribution, for promotion, for creating new works, or for resale. Specific permission must be obtained in writing from CRC Press LLC for such copying. Direct all inquiries to CRC Press LLC, 2000 N.W. Corporate Blvd., Boca Raton, Florida 33431. Trademark Notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation, without intent to infringe.
Visit the CRC Press Web site at www.crcpublications.com © 2002 CRC Press LLC Lewis Publishers is an imprint of CRC Press LLC No claim to original U.S. Government works International Standard Book Number 1-56670-282-8 Library of Congress Card Number 2001029566 Printed in the United States of America 1 2 3 4 5 6 7 8 9 0 Printed on acid-free paper
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Sincere Thanks To: Sumathy, Shauna, and Nealon for their enthusiastic support and unending patience. STP, MTP, MLM, and SBB for their insight, support, inspiration, and trust.
Dedicated with utmost humility to the heroes and heroines of Eelam who have put their lives in the line of fire to express their intellectual freedom.
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Foreword I have worked with Dr. Suthersan for the past 13 years and have seen firsthand the impact he has had on the evolution of our business. Over this period, environmental remediation has moved from a world of standard operation and application of proven technology to one where more innovative concepts can be applied, tested, and developed for the benefit of the environment, the regulatory community, and industry. Dr. Suthersan has worked assiduously to develop new remediation technologies, move them to pilot testing in cooperation with industry, and make them demonstrated techniques. As our industry has matured, the pressures on all parties have increased: pressure to assure protection of human health and the environment, to remediate faster, to rapidly return sites to beneficial use, to reduce costs, etc. Finding a solution to these competing objectives has become more and more intricate and must include the impacts of social, economic, business, and environmental factors. Dr. Suthersan is one of the most talented purveyors of remediation technology as a tool to solve these complex problems in a world where competing priorities are the rule not the exception. The author has focused on finding these total business solutions for our industry, using the innovative technical solutions he or others have created. Finding total business solutions to multifaceted environmental problems is one of the hallmarks of Dr. Suthersan’s career. In this book, Dr. Suthersan explains some of the pioneering remediation technologies developed over the past few years. The focus is on those techniques that modify or enhance the natural environment to aid in the remediation of contaminants. When applied correctly, these engineered, natural systems have proven to be more efficient and cost effective than their more intrusive predecessors. Assuring that these techniques are applied correctly and tailored to each particular setting is a key component of any system’s success. The impact of biological, chemical, and hydrogeologic settings on these technologies is thoroughly discussed. Dr. Suthersan describes each technique in detail: its processes, the science behind it, its application, and the constraints. This book will be an invaluable resource to the practicing remediation engineer, the regulatory community charged with evaluating these techniques, and the industry applying them. It has been a privilege to have worked with Dr. Suthersan for these past years and to have seen the influence of his knowledge and skill in our industry. I believe that those who read this book will gain from his wisdom. Steve Blake Executive Board, ARCADIS, N.V. Denver, Colorado
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Preface Remediation of hazardous wastes present in the subsurface has evolved with time and has been influenced by various factors over the years. During the early years, direction and efforts were mostly influenced by the regulations in place and the need for compliance and protection of human health and environment. The contaminants primarily focused upon during this time were the petroleum-related contaminants stemming from leaking underground storage tanks (USTs). In later years, remediation efforts were driven by a combination of economic and regulatory factors. During this time contaminants that caught most of the attention were the chlorinated solvents, heavy metals, and chlorinated and nonchlorinated polynuclear aromatic hydrocarbons (PAHs). The current focus seems to be taking a different direction: instead of focusing on the type of contaminants, emphasis is on evaluating the damage to the environment (and thus the risk) and repairing that damage in a cost-effective manner. Evolution of remediation technologies was influenced not only by changing regulatory and economic factors, but also by the type and chemical characteristics of contaminants under focus. An example is the shift in emphasis from engineered aerobic bioremediation systems of the 1980s to engineered anaerobic bioremediation systems of the 1990s. Significant reliance and dependence on natural remediation systems have increased as a result of recent acceptance that landfills behave as bioreactors and the very recent focus on dealing with ecological risks and natural resources damage (NRD) assessments. Ever increasing understanding of the behavior of most contaminants in the natural environment has also led to the effort of maximizing the remediation potential of natural systems. The thematic focus of this book is to highlight the current phase in the evolution of remediation technologies. All the technologies discussed in the book utilize or enhance the natural biogeochemical environment for remediation of hazardous contaminants. The discussion throughout the book is focused towards helping practitioners of remediation to engineer remediation systems utilizing the natural environment. These natural systems or reactors still have to be properly designed and engineered to optimize the performance and maximize contaminant removal efficiencies. The basic understanding of environmental and contaminant characteristics required to design these systems is provided in Chapter 2. I had just coined the phrase “in situ reactive zones (IRZ)” when I wrote my previous book in 1996 and was able to provide only an introduction of the technology. I have made a significant effort in Chapter 4 to describe the IRZ technology and its various modified applications. The manner in which the application of this technology is exploding may justify a book of its own. I am proud to see the advances and expansion of this technology pioneered by my colleagues and me at ARCADIS G & M, Inc. Due to the shortage of space I could not present data from all the successful sites using this technology. Technical advances and theoretical insights on the application of in situ chemical oxidation are also presented in Chapter 4 (special thanks to Dr. Fred Payne).
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I also had the privilege of being involved in some of the earliest phytoremediation and phyto-cover applications. Some contributions to the science of designing phytocovers are presented in Chapter 7 (special thanks to Dr. Scott Potter). I have provided only a summary on the current state of the science of phytoremediation in Chapter 5. Basic concepts of treatment wetlands are provided in Chapter 6. I truly believe that this technology will have more applications in the field of hazardous waste remediation. I wrote this book to reach a wide audience: remediation design engineers, scientists, regulatory specialists, graduate students in environmental engineering, and people from the industry who have general responsibility for site cleanups. I have tried to provide a general, basic description of the technologies in all chapters in addition to detailed information on basic principles and fundamentals in most chapters. Readers who are not interested in basic principles can skip these passages and still receive the general knowledge they need. Suthan S. Suthersan Yardley, Pennsylvania
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Acknowledgments First and foremost, I would like to thank members and colleagues from the Innovative Strategies Group (ISG) of ARCADIS — Frank Lenzo, Mike Hansen, and Jeff Burdick — for their enthusiasm and hard work in trying to experiment with innovative and cutting edge technologies in the field. Insights and advice provided by Drs. Scott Potter and Fred Payne in formulating the theoretical and mathematical foundations behind the technical concepts are immense. In addition, the patience and excitement exhibited by Chris Lutes and David Liles during the laboratory “proof of concept” experiments always boosted my confidence to proceed to the next level in implementing many of the technologies. Taking these technologies from the conceptual level to field scale applications would not have been possible without these individuals. I have to thank Eileen Schumacher and Ben Tufford for patiently drafting all the figures and Amy Weinert and Gail Champlin for typing the manuscript. The management of my employer ARCADIS G & M, Inc. deserves special mention for all the support given to me over the years. The opportunities and encouragement provided to me in order to “think out of the box” are a reflection of the company’s culture. I owe a special debt to all the engineers and project managers who helped me to implement many innovative and challenging remediation projects. This list is a long one, but special mention is due to the following: Mike Maierle, Don Kidd, Gary Keyes, Steve Brussee, Jack Kratzmeyer, Mark Wagner, Jim Drought, Tina Stack, Eric Carman, Al Hannum, John Horst, Kurt Beil, Dave Vance, Nanjun Shetty, and Pat Hicks. The encouragement, support, and feedback on the state of the science approaches in phytoremediation by Drs. Steve Rock and Steve McCutcheon, of the USEPA, are very much appreciated.
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The Author Suthan S. Suthersan, Ph.D., P.E., is senior vice president and director of Innovative Remediation Strategies at ARCADIS G & M, Inc., an international environmental and infrastructure services company. In his 12 years with the company, Dr. Suthersan has helped make AG&M one of the most respected environmental engineering companies in the U.S., specifically in the field of in situ remediation of hazardous wastes. Many of the technologies he pioneered have since become industry standards. His biggest contribution to the industry, beyond the technology development itself, has been to convince the regulatory community that these innovative technologies are better than traditional ones, not only from a cost viewpoint, but also for technical effectiveness. His experience is derived from working on at least 500 remediation projects in design, implementation, and technical oversight capacities during the past 15 years. Dr. Suthersan’s technology development efforts have been rewarded with seven patents awarded and more pending. His most important recent contributions are reflected by the following patents: Engineered In Situ Anaerobic Reactive Zones, US Patent 6,143,177; In Well Air Stripping, Oxidation, and Adsorption, US Patent 6,102,623; In Situ Anaerobic Reactive Zone for In Situ Metals Precipitation and to Achieve Microbial De-Nitrification, US Patent 5,554,290; In Situ Reactive Gate for Groundwater Remediation, US Patent 6,116,816. Dr. Suthersan has a Ph.D. in environmental engineering from the University of Toronto, a M.S. degree in environmental engineering from the Asian Institute of Technology, and a B.S. degree in civil engineering from the University of Sri Lanka. In addition to his consulting experience Dr. Suthersan has taught courses at several universities. He is the founding editor in chief of the Journal of Strategic Environmental Management and is a member of the editorial board of the International Journal of Phytoremediation.
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Contents Chapter 1 Hazardous Wastes Pollution and Evolution of Remediation....................................1 1.1 Introduction ......................................................................................................1 1.2 The Concept of Risk ........................................................................................2 1.2.1 The Decision Making Framework .......................................................3 1.3 Evolution of Understanding of Fate and Transport in Natural Systems ...............................................................................................4 1.4 Evolution of Remediation Technologies .........................................................7 References................................................................................................................11
Chapter 2 Contaminant and Environmental Characteristics ....................................................13 2.1 Introduction ....................................................................................................14 2.2 Contaminant Characteristics ..........................................................................18 2.2.1 Physical/Chemical Properties ............................................................18 2.2.1.1 Boiling Point.......................................................................18 2.2.1.2 Vapor Pressure ....................................................................18 2.2.1.3 Henry’s Law Constant ........................................................19 2.2.1.4 Octanol/Water Partition Coefficients..................................20 2.2.1.5 Solubility in Water..............................................................20 2.2.1.6 Hydrolysis ...........................................................................22 2.2.1.7 Photolytic Reactions in Surface Water...............................24 2.2.2 Biological Characteristics ..................................................................26 2.2.2.1 Cometabolism .....................................................................27 2.2.2.2 Kinetics of Biodegradation.................................................32 2.3 Environmental Characteristics .......................................................................38 2.3.1 Sorption Coefficient ...........................................................................38 2.3.1.1 Soil Sorption Coefficients...................................................43 2.3.1.2 Factors Affecting Sorption Coefficients .............................48 2.3.2 Oxidation-Reduction Capacities of Aquifer Solids ...........................51 2.3.2.1 pe and pH............................................................................51 2.3.2.2 REDOX Poise .....................................................................52 2.3.2.3 REDOX Reactions ..............................................................53 References................................................................................................................58
Chapter 3 Monitored Natural Attenuation ...............................................................................63 3.1 Introduction ....................................................................................................64 3.1.1 Definitions of Natural Attenuation ....................................................64 3.2 Approaches for Evaluating Natural Attenuation ...........................................65 3.3 Patterns vs. Protocols .....................................................................................70
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3.3.1 3.3.2
Protocols for Natural Attenuation......................................................70 Patterns of Natural Attenuation .........................................................71 3.3.2.1 Various Patterns of Natural Attenuation.............................72 3.4 Processes Affecting Natural Attenuation of Compounds..............................79 3.4.1 Movement of Contaminants in the Subsurface .................................79 3.4.1.1 Dilution (Recharge) ............................................................79 3.4.1.2 Advection ............................................................................81 3.4.1.3 Dispersion ...........................................................................83 3.4.2 Phase Transfers ..................................................................................85 3.4.2.1 Sorption...............................................................................85 3.4.2.2 Stabilization ........................................................................88 3.4.2.3 Volatilization .......................................................................89 3.4.3 Transformation Mechanisms..............................................................89 3.4.3.1 Biodegradation ....................................................................90 3.5 Monitoring and Sampling for Natural Attenuation .....................................109 3.5.1 Dissolved Oxygen (DO) ..................................................................113 3.5.2 Oxidation–Reduction (REDOX) Potential (ORP)...........................117 3.5.3 pH .....................................................................................................119 3.5.4 Filtered vs. Unfiltered Samples for Metals .....................................120 3.5.4.1 Field Filtration and the Nature of Groundwater Particulates..................................................121 3.5.4.2 Reasons for Field Filtration..............................................122 3.5.5 Low-Flow Sampling as a Paradigm for Filtration ..........................124 3.5.6 A Comparison Study........................................................................125 References..............................................................................................................126
Chapter 4 In Situ Reactive Zones...........................................................................................131 4.1 Introduction ..................................................................................................132 4.2 Engineered Anaerobic Systems ...................................................................135 4.2.1 Enhanced Reductive Dechlorination (ERD) Systems .....................135 4.2.1.1 Early Evidence..................................................................135 4.2.1.1.1 Biostimulation vs. Bioaugmentation ................136 4.2.1.2 Mechanisms of Reductive Dechlorination .......................138 4.2.1.3 Microbiology of Reductive Dechlorination .....................142 4.2.1.3.1 Cometabolic Dechlorination .............................142 4.2.1.3.2 Dechlorination by Halorespiring Microorganisms.................................................144 4.2.1.4 Electron Donors ................................................................147 4.2.1.4.1 Production of H2 by Fermentation ...................149 4.2.1.4.2 Competition for H2 ...........................................152 4.2.1.5 Mixture of Compounds on Kinetics.................................155 4.2.1.6 Temperature Effects ..........................................................158 4.2.1.7 Anaerobic Oxidation.........................................................158
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4.2.1.8 4.2.1.9
Electron Acceptors and Nutrients.....................................158 Field Implementation of IRZ for Enhanced Reductive Dechlorination...................................................................160 4.2.1.10 Lessons Learned ...............................................................163 4.2.1.11 Derivation of a Completely Mixed System for Groundwater Solute Transport of Chlorinated Ethenes...170 4.2.1.12 IRZ Performance Data......................................................177 4.2.2 In Situ Metals Precipitation .............................................................183 4.2.2.1 Principles of Heavy Metals Precipitation.........................187 4.2.2.2 Aquifer Parameters and Transport Mechanisms ..............195 4.2.2.3 Contaminant Removal Mechanisms.................................196 4.2.3 In Situ Denitrification.......................................................................197 4.2.4 Perchlorate Reduction ......................................................................199 4.3 Engineered Aerobic Systems .......................................................................200 4.3.1 Direct Aerobic Oxidation.................................................................200 4.3.1.1 Aerobic Cometabolic Oxidation.......................................202 4.3.1.2 MTBE Degradation ..........................................................204 4.4 In Situ Chemical Oxidation Systems...........................................................205 4.4.1 Advantages .......................................................................................206 4.4.2 Concerns...........................................................................................207 4.4.3 Oxidation Chemistry ........................................................................208 4.4.3.1 Hydrogen Peroxide ...........................................................211 4.4.3.2 Potassium Permanganate ..................................................213 4.4.3.3 Ozone ................................................................................216 4.4.4 Application .......................................................................................218 4.4.4.1 Oxidation of 1,4-Dioxane by Ozone................................222 4.4.4.2 Biodegradation Enhanced by Chemical Oxidation Pretreatment.....................................................223 4.5 Nano-Scale Fe (0) Colloid Injection within an IRZ ...................................223 4.5.1 Production of Nano-Scale Iron Particles .........................................228 4.5.2 Injection of Nano-Scale Particles in Permeable Sediments............231 4.5.3 Organic Contaminants Treatable by Fe (0) .....................................231 References..............................................................................................................233
Chapter 5 Phytoremediation ...................................................................................................239 5.1 Introduction ..................................................................................................240 5.2 Chemicals in the Soil–Plant System............................................................241 5.2.1 Metals ...............................................................................................241 5.2.2 Organics............................................................................................242 5.3 Types of Phytoremediation ..........................................................................244 5.3.1 Phytoaccumulation ...........................................................................245 5.3.2 Phytodegradation..............................................................................248 5.3.3 Phytostabilization .............................................................................250 5.3.4 Phytovolatilization............................................................................251
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5.3.5 Rhizodegradation..............................................................................252 5.3.6 Rhizofiltration...................................................................................256 5.3.7 Phytoremediation for Groundwater Containment ...........................259 5.3.8 Phytoremediation of Dredged Sediments ........................................260 5.4 Phytoremediation Design .............................................................................261 5.4.1 Contaminant Levels .........................................................................265 5.4.2 Plant Selection..................................................................................265 5.4.3 Treatability .......................................................................................266 5.4.4 Irrigation, Agronomic Inputs, and Maintenance .............................266 5.4.5 Groundwater Capture Zone and Transpiration Rate .......................267 References..............................................................................................................267
Chapter 6 Constructed Treatment Wetlands...........................................................................269 6.1 Introduction ..................................................................................................270 6.1.1 Beyond Municipal Wastewater ........................................................272 6.1.2 Looking Inside the “Black Box” .....................................................273 6.1.3 Potential “Attractive Nuisances”......................................................274 6.1.4 Regulatory Uncertainty and Barriers ...............................................275 6.2 Types of Constructed Wetlands ...................................................................276 6.2.1 Horizontal Flow Systems.................................................................276 6.2.2 Vertical Flow Systems......................................................................277 6.3 Microbial and Plant Communities of a Wetland.........................................278 6.3.1 Bacteria and Fungi ...........................................................................278 6.3.2 Algae ................................................................................................279 6.3.3 Species of Vegetation for Treatment Wetland Systems...................279 6.3.3.1 Free-Floating Macrophyte-Based Systems.......................282 6.3.3.2 Emergent Aquatic Macrophyte-Based Systems ...............284 6.3.3.3 Emergent Macrophyte-Based Systems with Horizontal Subsurface Flow ...............................................................285 6.3.3.4 Emergent Macrophyte-Based Systems with Vertical Subsurface Flow ...............................................................285 6.3.3.5 Submerged Macrophyte-Based Systems ..........................285 6.3.3.6 Multistage Macrophyte-Based Treatment Systems..........287 6.4 Treatment-Wetland Soils..............................................................................287 6.4.1 Cation Exchange Capacity...............................................................289 6.4.2 Oxidation and Reduction Reactions ................................................290 6.4.3 pH .....................................................................................................292 6.4.4 Biological Influences on Hydric Soils.............................................292 6.4.5 Microbial Soil Processes..................................................................292 6.4.6 Treatment Wetland Soils ..................................................................293 6.5 Contaminant Removal Mechanisms ............................................................294 6.5.1 Volatilization ....................................................................................294 6.5.2 Partitioning and Storage...................................................................295 6.5.3 Hydraulic Retention Time................................................................297
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6.6
Treatment Wetlands for Groundwater Remediation....................................299 6.6.1 Metals-Laden Water Treatment........................................................300 6.6.1.1 A Case Study for Metals Removal ..................................302 6.6.2 Removal of Toxic Organics .............................................................306 6.6.2.1 Biodegradation ..................................................................306 6.6.3 Removal of Inorganics .....................................................................309 6.6.4 Wetland Morphology, Hydrology, and Landscape Position............309 References..............................................................................................................310
Chapter 7 Engineered Vegetative Landfill Covers .................................................................313 7.1 Historical Perspective on Landfill Practices................................................314 7.2 The Role of Caps in the Containment of Wastes........................................315 7.3 Conventional Landfill Covers ......................................................................316 7.4 Landfill Dynamics........................................................................................317 7.5 Alternative Landfill Cover Technology .......................................................321 7.6 Phyto-Cover Technology..............................................................................321 7.6.1 Benefits of Phyto-Covers over Traditional RCRA Caps.................326 7.6.2 Enhancing In Situ Biodegradation...................................................326 7.6.3 Gas Permeability ..............................................................................327 7.6.4 Ecological and Aesthetic Advantages..............................................327 7.6.5 Maintenance, Economic, and Public Safety Advantages ................329 7.7 Phyto-Cover Design .....................................................................................329 7.7.1 Vegetative Cover Soils .....................................................................330 7.7.2 Nonsoil Amendment ........................................................................331 7.7.3 Plants and Trees ...............................................................................331 7.8 Cover System Performance..........................................................................332 7.8.1 Hydrologic Water Balance ...............................................................332 7.8.2 Precipitation .....................................................................................335 7.8.3 Runoff...............................................................................................335 7.8.4 Potential Evapotranspiration — Measured Data .............................337 7.8.5 Potential Evapotranspiration — Empirical Data .............................339 7.8.6 Effective Evapotranspiration ............................................................340 7.8.7 Water Balance Model.......................................................................343 7.9 Example Application....................................................................................344 7.10 Summary of Phyto-Cover Water Balance....................................................347 7.11 General Phyto-Cover Maintenance Activities .............................................348 7.11.1 Site Inspections ................................................................................348 7.11.2 Soil Moisture Monitoring ................................................................349 7.11.2.1 Drainage Measurement.....................................................350 7.11.3 General Irrigation Guidelines ..........................................................352 7.11.4 Tree Evaluation ................................................................................356 7.11.4.1 Stem ..................................................................................356 7.11.4.2 Leaves ...............................................................................356 7.11.5 Agronomic Chemistry Sampling .....................................................357
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7.11.6 Safety and Preventative Maintenance..............................................359 7.11.7 Repairs and Maintenance.................................................................359 7.12 Operation and Maintenance (O&M) Schedule............................................359 7.12.1 Year 1 — Establishment ..................................................................360 7.12.2 Years 2 and 3 — Active Maintenance.............................................360 7.12.3 Year 4 — Passive Maintenance .......................................................361 7.13 Specific Operational Issues..........................................................................362 7.13.1 Irrigation System Requirements ......................................................362 7.13.2 Tree Replacement.............................................................................362 References..............................................................................................................362
Appendix A Physical Properties of Some Common Environmental Contaminants .................365 Appendix B Useful Information for Biogeochemical Sampling...............................................383 Appendix C Common and Scientific Names of Various Plants ................................................405 Index......................................................................................................................409
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CHAPTER
1
Hazardous Wastes Pollution and Evolution of Remediation CONTENTS 1.1 1.2
Introduction ......................................................................................................1 The Concept of Risk ........................................................................................2 1.2.1 The Decision Making Framework .......................................................3 1.3 Evolution of Understanding of Fate and Transport in Natural Systems ........4 1.4 Evolution of Remediation Technologies .........................................................7 References................................................................................................................11
The earth was made so various that the mind of desultory man, studious of change and pleased with novelty, might be indulged.
1.1
INTRODUCTION
Among the many environmental problems that have received attention in recent decades is subsurface contamination caused by hazardous wastes. This has been due to the growing concern over short and long term health and environmental effects of toxic substances released into the environment. The public policy maker is faced with particular difficulties in regulating hazardous pollutants, most notably because of the high levels of uncertainty surrounding the issue. Such uncertainty exists in determining the precise impacts in relation to both human health effects and long term effects on the environment, especially with recalcitrant pollutants, or pollutants with extremely slow degradation rates. Nevertheless, policy makers have been required to formulate environmental regulations using some dependable basis. While theoretical methods of decision making such as dose-response and risk-benefit analysis may be employed to assist regulators,
1
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2
NATURAL AND ENHANCED REMEDIATION SYSTEMS
they are still faced with conflicting pressures — for example, between political and economic priorities or between public demands and technical expertise. Estimating the damage function of a pollutant is an exercise that underpins regulatory formulation for hazardous waste management. Figure 1.1 outlines the basic steps involved in this estimation. Determining the transfer function of a hazardous pollutant raises several problems. Persistent, nondegradable substances will tend to accumulate in the environment often becoming concentrated in the food chain. In the past, the potential life span of persistent substances in the subsurface was considered to be decades or even centuries. Research performed and scientific advancements, specifically in the last decade, indicate that compounds deemed to be persistent or nondegradable in the past are considered to be less persistent or at least partially degradable under natural conditions.
Release of Pollutants
Transfer Function
Rate and Mass at a Particular Place and Time
Figure 1.1
Ambient Conditions
Exposure Pathways
Damage Effects
Concentrations of Pollutants in Different Media
Dose-Response Function
Physical, Ecological, Health, Property, Natural Resources
Monetary
Evaluation of pollution damage.
1.2
THE CONCEPT OF RISK
Many of the problems associated with hazardous waste management, such as uncertainty, irreversibility, and persistence make the concept of risk relevant to this discussion. From an engineering or scientific standpoint, “risk” may be defined in quantitative terms by applying probabilistic measures. If “hazard” is defined as the potential for adverse consequences of some event, then “risk” may be defined as the chance of a particular hazard occurring. It combines two aspects — the probabilistic measure of the occurrence of the event with a measure of the consequences of the event (in this case the level of toxicity of the pollutant). Further aspects of risk are highlighted by social scientists who examine risk perception in recognition that a particular risk or hazard may mean different things to different people in different contexts. The concept of risk is not without problems, particularly in relation to the issue of hazardous pollutants. For example, an initial problem is determining the probability of such risks; there have been only a few decades of experience in dealing with many pollutants. Their effects on human beings had been largely unknown, and thus the probabilistic calculations of risk on exposure and associated health and ecological impacts were mostly conservative. In relation to the assumed, perceived, or calculated risks associated with hazardous pollutants until recently, it is important to highlight two significant features: 1) the subjective probability of the hazard (caused by toxicity of the released pollutant) occurring may be very low, but, 2) the consequence of the hazard was assumed or perceived to be very high, often as irreversible because of assumptions
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HAZARDOUS WASTES POLLUTION AND EVOLUTION OF REMEDIATION
3
of persistence and negligible degradation of many pollutants in the natural environment. Thus the low probability tends to cancel out the assumed or perceived impacts associated with the risk. Recent research shows that pollutants and other organic chemicals present in the subsurface become less available or create lesser levels of hazard (in other words become less toxic) due to interactions between the compound and the subsurface environment. This drop in availability and toxicity lowers the risk of these chemicals to human and ecological receptors. Furthermore, the availability of an organic chemical in the subsurface is not a function of its measured concentration; rather, it depends upon the geologic and biogeochemical characteristics of the subsurface, the physicochemical properties of the chemical itself, and the time of contact between the chemical and the subsurface media, i.e., aging, as well as the type and extent of treatment, natural or anthropogenic, to which it has been subjected. 1.2.1
The Decision Making Framework
In the face of the many uncertainties surrounding hazardous waste management with respect to the assumed, perceived, or calculated risks, the regulatory authorities are faced with an initial decision about the appropriate framework for decisionmaking: should it be a balancing approach such as cost-benefit or risk-benefit analysis, or should it be an approach which emphasizes the protection of human health and natural resources regardless of costs? Three types of approaches have been utilized to implement hazardous waste management policies in the U.S. during the past three decades: 1. Health-based approaches — zero risk, significant risk, or acceptable risk 2. Balancing approaches — cost-benefit, risk-benefit, or decision analysis 3. Technology-based approaches — best available technology, risks as low as reasonably practicable
Environmental threats, rather than the scientific evidence and theory from which they may be deduced, have been ill-defined during the past three decades. The evidence from which a threat is deduced has been challenged by conflicting evidence or placed into a context of associations which heightens its significance. A scenario for an exposure pathway typically used in the past, where a kid climbing an eightfoot fence and eating a few grams of soil every day for a decade is an example of such an association. For many years, there was a widespread but unfounded assumption that some toxic pollutants stemming from industrial releases and/or accidents and landfills would not be degraded in the natural environment. The measurement of damage, and thus the risk, requires an understanding of the physical processes of transportation and of the distribution and deposition of pollutants, including their chemical and biological transformations on the way. The creation of new knowledge usually involves institutions very different from those concerned with its acceptance, application, and dissemination. A genuine science-based environmental policy should be a dynamic one and evolve via continuous monitoring of pollutants in many media, as well as of their impacts on the ecosystem and human health (or any other selected target organisms). The technical
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
means for such monitoring must, of course, be available, as must the baselines for the establishment of a time series so that change can be observed over time in the natural environment. There must be agreement on which pollutants to monitor and how to synthesize and use the masses of data that will accumulate. Even complete understanding of how the subsurface works as a bio-geo-physico-chemical system cannot give ready answers as to the proper regulatory response, i.e., how to use the earth in the common interest of humanity and without degrading it for future generations. This is probably why the government, more out of frustration than intent, has come to rely less on science, engineering, and economics and more on caution and law. Figure 1.2 describes the shortcomings of the health-based, conservative approach of the past and the more credible, balanced approach still evolving. Understanding the contribution by Mother Nature towards a natural remediation process has had a significant influence on this evolution.
10 -4
Remediation expenditure which justifies reasonable risk reduction
10 -5
Associated Risk
Cost of Remediation ($)
10 -3
10 -6
10 -7
Figure 1.2
A hypothetical analysis of cost to risk reduction benefit ratios during remediation activities.
1.3
EVOLUTION OF UNDERSTANDING OF FATE AND TRANSPORT IN NATURAL SYSTEMS
Predicting the hazard of an organic contaminant to humans, animals, and plants requires information not only on its toxicity to living organisms but also on the degree of exposure of the organisms to the compound. The mere release or discharge of a pollutant does not, in itself, constitute a hazard; the individual human, animal, or plant must also be exposed to it. In evaluating exposure, the transport of the
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chemical and its fate must be considered. A molecule that is not subject to environmental transport is not a health or environmental problem except to species at the specific point of release. Information on dissemination of the chemical from the point of its release to the point where it could have an effect is of great relevancy for risk calculations. However, the chemical may be modified structurally or totally destroyed during its transport, and the fate of the compound during transport, that is, its modification or destruction, is crucial to defining the exposure. A compound modified to yield products that are less or more toxic, or totally degraded to harmless end products, or bio-magnified — factors associated with the fate of the molecule — represents greater or lesser hazard to the species potentially exposed to injury. At the specific site of discharge or during its transport, the pollutant molecule or ion may be acted on by abiotic mechanisms. Photochemical transformations occur in the atmosphere and at or very near the surfaces of water, soil, and vegetation, and these processes may totally destroy or appreciably modify a number of different types of organic chemicals. Nonenzymatic, nonphotochemical reactions are also prominent in soil, sediment, and surface and groundwater, and these may bring about significant changes; however, such processes rarely, if ever, totally convert organic compounds to harmless end products or mineralized compounds in nature. Many of these nonenzymatic reactions only bring about a slight modification of the molecule so that the product is frequently similar in structure, and often in toxicity, to the precursor compound. However, biological processes may modify organic molecules at the site of their release or during their transport. Such biological transformations, which involve enzymes as catalysts, frequently bring about extensive modification in the structure and toxicological properties of pollutants or potential pollutants. These biotic processes may result in the complete conversion of the organic molecule to inorganic products, cause major changes resulting in new organic products, or occasionally lead to only minor modifications. The available body of information suggests that the major agents causing the biological transformations in soil, sediment, surface and groundwater, and many other sites are the microorganisms that inhabit these environments. The earth is thought to be about 4.6 × 109 years (4.6 eons) old.2 The original atmosphere surrounding the earth was reducing and probably included the gases CH4, CO2, CO, NH3 and H2O. Although abiotic organic synthesis probably occurred since the earth’s beginnings, life probably did not appear until about 0.5–1 billion years later, according to present thinking. The first form of life that was established on the “infant” earth was anaerobic.1 As anaerobic life became more firmly established, the organic nutrients must have begun to be depleted at a faster rate than they could be replenished by abiotic synthesis. Hence, an alternative mechanism for producing organic matter was required to sustain life. The subsequent evolutionary developments led to the emergence of photosynthesis and eventually resulted in the emergence of aerobic heterotrophic organisms. These aerobic organisms ended up much more efficient than their anaerobic counterparts in sustaining life.
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Billions of years of evolution by Mother Nature have shown that the natural communities of microorganisms in the various habitats have an amazing physiological versatility. These communities are adaptable, flexible, versatile, and robust. They are able to metabolize and often mineralize an enormous number of organic molecules. Probably every natural product, regardless of its complexity, is degraded by one or another species in some particular environment; if not, this long after the appearance of life on earth, such compounds would have accumulated in enormous amounts. The compounds that caught the most attention in the remediation industry, initially during the 1970s and 1980s, were the BTEX compounds released via petroleum spills — natural products formed as the result of decomposition of plants and other organic materials over millions of years. It has been proven during the last decade that the BTEX compounds will naturally attenuate in the groundwater through microbial degradation. Although certain bacteria and fungi act on a broad range of organic compounds, no organism known to date is sufficiently omnivorous to destroy a very large percentage of the natural chemicals.2 Bioremediation is now a widely accepted technique for contaminant cleanup. But a few short decades ago, its use for anything as effective as the in situ cleanup of groundwater contamination was considered laughable. “At that time, there was a myth, widely held by the geological and hydrological community, that the subsurface was sterile, that there were no bacteria, and therefore no biological processes of consequence”.3 This thinking was mainly due to the information available from the textbooks at that time. Microbial ecology is the study of interrelationships between different microorganisms; among microorganisms, plants, and animals; and between microorganisms and their environment. Microbial biogeochemistry is the study of microbially catalyzed reactions and their kinetics with emphasis on environmental mass transfer and energy flow. In subsequent chapters, this book summarizes and systematizes current understanding of abiotic and biotic transformations of organic and inorganic pollutants in the natural environment. Knowledge of abiotic transformations can provide a conceptual framework for understanding biologically mediated transformations. Most abiotic transformations are slow, but they can still be significant within the time scales commonly associated with groundwater movement. In contrast, biotic transformations typically proceed much faster, provided the biogeochemical environment is conducive to mediate such transformations. The ability to predict the behavior of a chemical substance in a biological or environmental system largely depends on knowledge of the physical-chemical properties and reactivity of that compound or closely related compounds. Chemical properties frequently used in environmental fate assessment include melting/boiling temperature, vapor pressure, various partition coefficients, water solubility, Henry’s Law Constant, sorption coefficient, and diffusion properties. Reactivities by processes such as biodegradation, hydrolysis, photoysis, and oxidation/reduction are also critical determinants of environmental fate. Unfortunately, measured values often are not available and, even if they are, the reported values may be inconsistent or of doubtful validity. In this situation it may be appropriate or even essential to
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use estimation methods. The evolution of understanding of half-lives of chlorinated aliphatic compounds and the refinement of those values by precisely measured values have been remarkable during the last decade. Half-lives which were estimated to be in the two to three year range have been measured in the field to be in the range of three to six months. Later chapters describe the marked difference in the accepted half-lives during the last decade.
1.4
EVOLUTION OF REMEDIATION TECHNOLOGIES
Remediation technologies have undergone many changes over the last two decades during which they have been applied to clean up subsurface and hazardous waste contamination problems. These changes have occurred at a relatively rapid pace; during this period some of the most profound changes have occurred in how we apply remedial technologies as a result of pressure from the industry to continuously improve technical efficiency and cost effectiveness of the preferred technologies. Initially contaminated groundwater was a driving concern because it was mobile and, as a result, transported the liability off site. Also, the need to contain the contamination on site led to universal application of pump and treat systems for source control and mass removal. A decade of experience has taught that pump and treat is not the solution and, in fact, is an inefficient technology for fast and cheap site cleanup. The realization that mass removal efficiencies can be significantly enhanced using air as an extractive media instead of water led to the development and application of in situ extractive technologies such as soil vapor extraction and in situ air sparging. While it can be argued that the initial motive for applying these technologies has been one of saving money, the end result is much quicker cleanup times to more acceptable cleanup levels (Figure 1.3). This win-win situation for the entire remediation industry fostered continuous innovation, which led to 1) faster, cheaper solutions, 2) less invasive in situ technologies, and 3) technologies complementary to the natural environment which took advantage of nature’s capacity to degrade the pollutants. Thus holistically, environmentally, and economically sound and sustainable solutions were provided. Figure 1.4 illustrates the evolutionary reduction in remediation costs from the late 1970s to the present time. Ex situ extractive techniques such as pump and treat systems were replaced by in situ extractive techniques, namely, soil vapor extraction (SVE) and in situ air sparging. Subsequently these in situ extractive techniques gave way to in situ nonextractive techniques such as funnel and gate systems and, eventually, to in situ mass destruction techniques such as in situ reactive zones (IRZ) as the preferred remediation technologies. This evolutionary pattern has focused towards more natural solutions and/or enhancing existing subsurface biogeochemical conditions that contribute to remediation. The most recent shift occurred approximately 5 years ago, with the recognition and demonstrated value of natural mechanisms that contributed towards the containment, control, and mass reduction of contaminants in soil and groundwater. Under a host of names — including natural attenuation, bioattenuation, natural remediation,
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
MNA with Source Reduction
MNA - Monitored Natural Attenuation
Conventional Pump and Treat
Concentration
In Situ Reactive Zones (IRZ) In Situ Air Sparging
Only When Contaminants Are Aerobically Biodegradable Clean-Up Standards
Time Figure 1.3
Evolution of in situ remediation technologies and improvements in efficiencies.
monitored natural attenuation (MNA) — this remediation approach has taken root as a viable remediation approach at the appropriate site and under the right biogeochemical conditions. Used in conjunction with already ongoing remediation systems or as a stand-alone remedy, MNA can increase significantly the probability of a successful, cost-effective, and well-documented restoration of a contaminated site. The development of in situ reactive zones (IRZ), which are engineered in situ anaerobic or aerobic systems, is essentially an outgrowth of the efforts to enhance the natural processes which contribute towards degradation of many contaminants. For example, the use of an engineered IRZ to reductively dechlorinate chlorinated aliphatic hydrocarbons, such as PCE and TCE, in essence, enhances the rate of natural degradation by providing the optimum biogeochemical conditions (Figure 1.5). At many contaminated sites, the bulk of the contaminant mass may still be present in what remediation professionals call “source areas.” Even though the plume length has reached a stable equilibrium and the contaminant concentrations have reached steady or declining concentrations at the compliance points, it may be desirable to enhance the rates of natural degradation if the plume has crossed the property boundary (Figure 1.6a). Surgical reduction of the mass at the source areas and enhancement of natural degradation along the property boundary will enable such properties to be restored within a reasonable time frame (Figure 1.6b). The duration of the containment IRZ at the property boundary will be significantly longer if mass removal is not accomplished at the source area.
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Capital Costs O&M Costs
Cost ($)
MNA Monitored Natural Attenuation
Ex Situ Extractive Techniques Late 1970s - Early 1980s
Figure 1.4
In Situ Extractive Techniques Early 1980s - Late 1980s
In Situ Extractive Techniques Early 1990s - Present
In Situ Mass Destruction Techniques Mid 1990s - Present
MNA Current
Evolution reduction in remediation costs.
IRZ In Situ Reactive Zones MNA Monitored Natural Attenuation
Cost ($)
Creation of IRZ
Natural Rate of Decline
Enhanced Rate of Decline
Time Figure 1.5
Implementation of an IRZ for enhanced biodegradation has an impact on the time towards closure in comparison to reliance on MNA.
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
IRZ In Situ Reactive Zones
Engineered IRZ
Source Area
Stable Plume Compliance Points
Property Boundary Figure 1.6a
Implementation of a containment IRZ and a source reduction IRZ to reduce the cleanup time.
Concentraation at Compliance Point(s)
IRZ In Situ Reactive Zones
Creation of IRZ
Time Figure 1.6b
Reduction in cleanup time as a result of enhanced rate of mass removal.
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The evolutionary trend towards more natural remediation solutions should not be surprising due to the recognition for decades, even centuries, of the value of natural wetlands to buffer the effects of human activity in waterways. Today engineered wetlands, phyto-covers and phytoremediation are our attempts to mimic natural systems by “engineering” remediation solutions using nature as the material of construction — trees and microorganisms instead of pumps and air compressors.
REFERENCES 1. Ehrlich, H.L., Geomicrobiology, Marcel Dekker Inc., New York, 1981. 2. Alexander, M., Biodegradation and Bioremediation, 2nd ed., Academic Press, New York, 1999. 3. Harvey, M.A., Quotation by John Wilson in Germ Warfare, Environ. Prot., 23-26, October 1999.
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CHAPTER
2
Contaminant and Environmental Characteristics CONTENTS 2.1 2.2
Introduction ....................................................................................................14 Contaminant Characteristics ..........................................................................18 2.2.1 Physical/Chemical Properties ............................................................18 2.2.1.1 Boiling Point.......................................................................18 2.2.1.2 Vapor Pressure ....................................................................18 2.2.1.3 Henry’s Law Constant ........................................................19 2.2.1.4 Octanol/Water Partition Coefficients..................................20 2.2.1.5 Solubility in Water..............................................................20 2.2.1.6 Hydrolysis ...........................................................................22 2.2.1.7 Photolytic Reactions in Surface Water...............................24 2.2.2 Biological Characteristics ..................................................................26 2.2.2.1 Cometabolism .....................................................................27 2.2.2.2 Kinetics of Biodegradation.................................................32 2.3 Environmental Characteristics .......................................................................38 2.3.1 Sorption Coefficient ...........................................................................38 2.3.1.1 Soil Sorption Coefficients...................................................43 2.3.1.2 Factors Affecting Sorption Coefficients .............................48 2.3.2 Oxidation-Reduction Capacities of Aquifer Solids ...........................51 2.3.2.1 pe and pH............................................................................51 2.3.2.2 REDOX Poise .....................................................................52 2.3.2.3 REDOX Reaction ...............................................................53 References................................................................................................................58
13
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Water is scientifically very different in comparison to other liquids. With its rare and distinctive property of being denser as a liquid than as a solid, it is different. Water is different in that it is the only chemical compound found naturally in solid, liquid, or gaseous states at ambient conditions. Water is sometimes called the universal solvent. This is a fitting name, especially when you consider that water is a powerful reagent, which is capable in time of dissolving everything on earth.
2.1
INTRODUCTION
The primary management goal during remediation of a contaminated site is to obtain closure, that is, to achieve a set of conditions that is considered environmentally acceptable and which will ensure that no future action will be required at the site. A substantial ongoing national debate associated with site closure centers on the definition of “how clean is clean” for contaminated subsurface media. The key issue in this debate is, “What concentration of residual contaminant in the subsurface, particularly adsorbed to the soil, is environmentally acceptable?” In this context, the term contaminant availability becomes an important concept; it refers to the rate and extent to which the chemical will be released from the subsurface into the environment and/or is bioavailable to ecological and human receptors. The dissemination of a contaminant after its release into the environment is determined by its partition among the water, soil and sediment, and atmospheric phases, and its degradability via biotic and/or abiotic means. These processes determine both the impact and the extent of its dissemination. Within the context of overall site management, measurements of contaminant availability are not intended to replace other approaches, required regulatorily, to achieve site closure; rather, they are meant to broaden the range of options or tools available to environmental professionals. This chapter will discuss the basis and parameters for the development of procedures and determination of partitioning, transport, and fate of various types of contaminants in the subsurface. These parameters will also provide the basis for the development of the tools to determine contaminant availability and incorporate those estimations into a decision framework to define environmentally acceptable endpoints for the different media. In addition, how these parameters and characteristics influence contaminant fate and transport and how they impact remediation system design are woven together in the discussions in subsequent chapters. The reactions that contaminants undergo in the natural environment, such as sorption, desorption, precipitation, complexation, biodegradation, biotransformation, hydrolysis, oxidation-reduction, and dissolution, are critical in determining their fate and mobility in the subsurface environment. Reaction time scales can vary from microseconds for many ion association reactions microseconds and milliseconds for some ion exchange and sorption reactions, to days, weeks, or months for some microbially catalyzed reactions, or years for many mineral solution and crystallization reactions. Both transport and chemical reaction processes can affect the reaction rates in the subsurface environment. Transport processes include: (1) transport in the solution phase, across a liquid film at the particle/liquid interface (film diffusion), and in
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liquid-filled macropores, all of which are nonactivated diffusion processes and occur in mobile regions; (2) particle diffusion processes, which include diffusion of sorbate occluded in micropores (pore diffusion) and along pore-wall surfaces (surface diffusion) and diffusion processes in the bulk of the solid, all of which are activated diffusion processes (Figure 2.1).1 Pore and surface diffusion within the immediate region can be referred to as intra-aggregate (intraparticle) diffusion and diffusion in the solid can be called interparticle diffusion. The actual chemical reaction at the surface, e.g., adsorption, is usually instantaneous. The slowest of the chemical reaction and transport process is the ratelimiting reaction.
1
2 3 4 6
Film
4
5 6
Solid (Soil Grain)
Liquid (Groundwater) 1 2 3 4 5 6
Figure 2.1
Transport in the Soil Solution (Macro Pores) Transport Across a Liquid Film at the Solid-Liquid Interface Transport in a Liquid-Filled Macropore Diffusion of a Sorbate at the Surface of the Solid Diffusion of a Sorbate Occluded in a Micropore Diffusion in the Bulk of the Solid
Transport processes in solid-liquid soil reactions (adapted from Sparks, 1998).
As an introduction to the various organic compounds which end up as contaminants once discharged into the environment, Table 2.1 gives the basic structure of the different compounds.
C
C
R RNH
O
NH 2
R
C
R C
C
OH
O
O OH
R C O
H
C
-
R
R C O
+
R 4N X
H
C O
NR 2 X
O
OH
CH 3CH2OH CH 3CH 2OCH2CH 3
Bromide Alcohol Ether
Carboxylic acid
Ketone
CH3 C
O
CH3CH2C
O
OH
CH 3
O
CH3CH2CH
Aldehyde H
CH3(CH2)9N(CH3)3Cl
Quaternary ammonium salt
CH3CH2CH2NH2
CH
+
CH3Br
Chloride
2
CH 2CH2Cl
Alkyne
Amine
HC
Alkene CH 2
CH 3CH 3 H 2C
Alkane
Formula
-
Ethanoic acid
(Continued)
Acetic acid
Methyl ethyl ketone
Propionaldehyde
Propanal
2-Butanone
DecyltrimethylAmmonium chloride
DecyltrimethylAmmonium chloride
Propylamine
Diethyl ether
Ethoxyethane 1-Aminopropane 3
Ethyl alcohol
Methyl bromide
Ethyl chloride
Ethanol
Bromomethane
Chloroethane
Acetylene
Ethyne
Ethane
Common Name
Ethylene
1
Ethene
Ethane
IUPAC Name
Example
16
+
R
OH
Br
R
Br
Cl
R
Cl
-
CnH 2n
C CnH 2n-2
CnH 2n+2
C
None
General Name
General Formula
Functional Group
Table 2.1 Some Common Functional Groups.
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3
2
R
R
OH
Sulfone
Sulfoxide
O
CH3 S
O
CH3 S
O
O
CH 3 S
O
CH 3
CH 3
CH3
OH
S
CH3
The italicized portion indicates the group. A primary (1°) amine; there are also secondary (2°), R NH, and tertiary (3°), R N,3 amines. 2 Another name is propanamine.
O
S
O
O R
S
S
O
O R
S
S
O
OH O
S
O
S Sulfonic acid
CH 3 S
Disulfide
S
O
CH 3
CH3 S
R
Thioether (sulfide)
S
O
R
R
S
O
R
S
CH3 SH
Thiol
N
SH
3
CH3 NO2
Nitrile
Cl
NH 2
OC2H5
CH3 C O C
O
CH3 C
O
CH 3 C
O
C
Nitro
R
Acid anhydride
Acid chloride
Amide
CH
O
Formula
NO 2
R
R
SH
S
R
N
NO 2
C
R
N
O
C
O R C O C
O
Cl
C O C
O
R C
C
Cl
O
O
NH 2
OR'
Ester
General Name 1
Dimethyl disulfide Methanesulfonic acid
Dimethyl disulfide Methanesulfonic acid
Dimethyl sulfone
Dimethyl sulfone
Dimethyl sulfoxide
Dimethyl sulfide
Dimethyl thioether
Dimethyl sulfoxide
Methyl mercaptan
Nitromethane
Acetonitrile
Acetic anhydride
Acetyl chloride
Acetamide
Acetic acid
Common Name
Methanethiol
Nitromethane
Ethanenitrile
Ethanoic anhydride
Ethanoyl chloride
Ethanamide
Ethyl ethanoate
IUPAC Name
Example
CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
1
O R C
C
NH 2
O
R C
C
OR'
O
General Formula
O
Functional Group
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
2.2 2.2.1
CONTAMINANT CHARACTERISTICS
Physical/Chemical Properties
2.2.1.1 Boiling Point The boiling point is defined as the temperature at which a liquid’s vapor pressure equals the pressure of the atmosphere on the liquid.2 If the pressure is exactly 1 atmosphere (101,325 Pa), the temperature is referred to as “the normal boiling point.” Pure chemicals have a unique boiling point, and this fact can be used in some laboratory investigations to check on the identity and/or purity of a material. Mixtures of two or more compounds have a boiling point range. For organic compounds, boiling points range from –162 to over 700°C, but for most chemicals of interest the boiling points are in the range of 300 to 600°C.2 Having a value for a chemical’s boiling point, whether measured or estimated, is significant because it defines the uppermost temperatures at which the chemical can exist as a liquid. Also, the boiling point itself serves as a rough indicator of volatility, with higher boiling points indicating lower volatility at ambient temperatures. The boiling point is associated with a number of molecular properties and features. Most important is molecular weight; boiling points generally increase with this parameter. Next is the strength of the intermolecular bonding because boiling points increase with increasing bonding strength. This bonding, in turn, is associated with processes and properties such as hydrogen bonding, dipole moments, and acid/base behavior. 2.2.1.2 Vapor Pressure The vapor pressure of a chemical is the pressure its vapor exerts in equilibrium with its liquid or solid phase.2 Vapor pressure’s importance in environmental work results from its effects on the transport and partitioning of chemicals among the environmental media (air, water, and soil). The vapor pressure expresses and controls the chemical’s volatility. The volatilization of a chemical from the water surface is determined by its Henry’s Law Constant, which can be estimated from the ratio of a chemical’s vapor pressure to its water solubility. The volatilization of a chemical from the soil surface is determined largely by its vapor pressure, although this is tempered by its sorption to the soil matrix and its Henry’s Law Constant between the soil water content and air. A substance’s vapor pressure determines whether it will occur as a free molecule in the vapor phase or will be associated with the solid phase. For volatile substances that boil at or below 100°C, the vapor pressure is likely to be known, but, for many high-boiling substances with low vapor pressure, the value may be unknown or poorly known. An estimation procedure may be needed to help convert the known vapor pressure at the normal boiling point (i.e., 1 atmosphere) to the vapor pressure at the lower temperatures of environmental importance. For some of these high boiling compounds, the actual boiling point may also be unknown, since the substance may decompose before it boils.
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2.2.1.3 Henry’s Law Constant Along with the octanol-water and octanol-air partition coefficients, the Henry’s Law Constant determines how a chemical substance will partition among the three primary media of accumulation in the environment, namely air, water, and organic matter present in soils, solids, and biota. Volatile organic compounds (VOCs) with large values of Henry’s Law Constant evaporate appreciably from soils and water, and their fate and effects are controlled primarily by the rate of evaporation and the rate of subsequent atmospheric processes. For such chemicals, an accurate value of this parameter KAW is essential. Even a very low value of KAW for example, 0.001, can be significant and must be known accurately, because the volume of the accessible atmosphere is much larger than that of water and soils by at least a factor of 1000; thus even a low atmospheric concentration can represent a significant quantity of chemical. Further, the rate of evaporation from soils and water is profoundly influenced by KAW because that process involves diffusion in water and air phases in series, or in parallel, and the relative concentrations which can be established in these phases control these diffusion rates.2,3 Accurate values of KAW are thus essential for any assessment of the behavior of existing chemicals or prediction of the likely behavior of new chemicals. Air-water partitioning can be viewed as the determination of the solubility of a gas in water as a function of pressure, as first studied by William Henry in 1803. A plot of concentration or solubility of a chemical in water expressed as mole fraction x, vs. partial pressure of the chemical in the gaseous phase P, is usually linear at low partial pressures, at least for chemicals which are not subject to significant dissociation or association in either phase. This linearity is expressed as Henry’s Law. The Henry’s Law Constant (H) which in modern SI units has dimensions of Pa/(mol fraction). For environmental purposes, it is more convenient to use concentration units in water CW of mol /m3 yielding H with dimensions of Pa m3/mol. P (Pa) = H (Pa m3/mol) CW (mol/m3)
(2.1)
The partial pressure can be converted into a concentration in the air phase CA by invoking the ideal gas law: CA = n/V = P/RT
(2.2)
Where n is mols, V is volume (m3), R is the gas constant (8.314 Pa m3/mol K) and T is absolute temperature (K). CA = P/RT = (H/RT) CW = KAWCW
(2.3)
The dimensionless air-water partition coefficient KAW (which can be the ratio in units of mol/m3 or g/m3 or indeed any quantity/volume combination) is thus (H/RT). A plot of CA vs. CW is thus usually linear with a slope of KAW as Figure 2.2 illustrates. For organic chemicals which are sparingly soluble in water, these concentrations are limited on one axis by the water solubility and on the other by the
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Concentration in Air CA
(Vapor Pressure/RT)
Slope = Kaw = H/RT [Solubility of Compound]
Concentration in Water Cw Figure 2.2
Description of Henry’s Law Constant.
maximum achievable concentration in the air phase which corresponds to the vapor pressure, as Figure 2.2 shows. To the right of or above the saturation limit, a separate organic phase is present. Strictly speaking, this saturation vapor pressure is that of the organic phase saturated with water, not the pure organic phase.2,3 2.2.1.4 Octanol/Water Partition Coefficients The usefulness of the ratio of the concentration of a solute between water and octanol as a model for its transport between phases in a physical or biological system has long been recognized.2,4,5 It is expressed as POCT = CO /CW = KOW . This ratio is essentially independent of concentration, and is usually given in logarithmic terms (log POCT or log KOW). The importance of bioconcentration in environmental hazard assessment and the utility of this hydrophobic parameter in its prediction led to an intense interest in the measurement of POCT and also its prediction from molecular structure. (So many calculation methods have been published in the last five years that it is not possible to examine them all in detail.) 2.2.1.5 Solubility in Water Solubility in water is one of the most important physical chemical properties of a substance, having numerous applications to the prediction of its fate and its effects in the environment. It is a direct measurement of hydrophobicity, i.e., the tendency of water to exclude the substance from solution. It can be viewed as the maximum concentration which an aqueous solution will tolerate before the onset of phase separation.
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Substances which are readily soluble in water, such as lower molecular weight alcohols, will dissolve freely in water if accidentally spilled and will tend to remain in aqueous solution until degraded. On the contrary, sparingly soluble substances dissolve more slowly and, when in solution, have a stronger tendency to partition out of aqueous solution into other phases. They tend to have larger air–water partition coefficients or Henry’s Law Constants, and they tend to partition more into solid and biotic phases such as soils, sediments, and fish. As a result, it is common to correlate partition coefficients from water to those media with solubility in water. Solubility normally is measured by bringing an excess amount of a pure chemical phase into contact with water at a specified temperature, so that equilibrium is achieved and the aqueous phase concentration reaches a maximum value. It is rare to encounter a single compound as the contaminant present in the groundwater at a contaminant site. C *i = C 0i x i γ i
(2.4)
where, Ci* Ci0 xi γi
= = = =
equilibrium solute concentration for component i in the mixture equilibrium solute concentration for component i as a pure compound mole fraction of compound i in the mixture activity coefficient of compound i in the mixture.
Possible equilibrium situations may exist, depending on the nature of the chemical phase, each of which requires separate theoretical treatment and leads to different equations for expressing solubility. These equations form the basis of the correlations discussed later. Single compound is an immiscible liquid (e.g., Benzene) C* = Co x γ
(2.5)
In this case, C* is also C°. Thus the product xγ is 1.0 and x is 1/γ. Sparingly soluble substances act in such a way because the value of γ is large.2 For example, at 25°C benzene has a solubility in water of 1780 g/m3 or 22.8 mol/m.3 Since 1 m3 of solution contains approximately 106/18 mol water (1m3 is 106 g and 18 g /mol is the molecular mass of water), the mole fraction x is 22.8/(106/18) or 0.00041. The activity coefficient γ is thus 2440; i.e., a benzene molecule in aqueous solution behaves as if its concentration were 2440 times higher. Substances such as polychlorinated biphenyls (PCBs) can have activity coefficients exceeding 1 million. Hydrophobicity thus is essentially an indication of the magnitude of γ. Some predictive methods focus on estimating γ, from which solubility can be deduced.
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Single compound is a miscible substance (e.g., Ethanol) If the activity coefficient is relatively small, i.e., < 20, it is likely that the liquid is miscible with water and no solubility can be measured. The relevant descriptor of hydrophobicity in such cases is the activity coefficient. Correlations of other environmental partitioning properties with solubility are then impossible.2 Solubility is a function of temperature because both vapor pressure and γ are temperature dependent. Usually γ falls with increasing temperature, thus solubility increases. This implies that the process of dissolution is endothermic. Exceptions are frequent and in some cases, such as benzene, there may be a solubility minimum as a function of the temperature at which the enthalpy of dissolution is zero.2 Under natural conditions, dissolved organic matter such as humic and fulvic acids frequently increases the apparent solubility. This is the result of sorption of the chemical to organic matter which is sufficiently low in molecular mass to be retained permanently in solution. The true solubility or concentration in the pure aqueous phase probably is not increased. The apparent solubility is the sum of the true or dissolved concentration and the quantity which is sorbed. The solubility of substances such as carboxylic acids, which dissociate or form ions in solution, is also a function of pH, a common example being pentachlorophenol. Data must thus be at a specific pH. Alternatively, the solubility of the parent (nonionic) form may be given, and pKa or pKb given, to permit the ratio of ionic to nonionic forms to be calculated as Ionic/non-ionic = 10(pH–pKa)
(2.6)
The total solubility is then that of the parent and ionic forms. 2.2.1.6 Hydrolysis Hydrolysis is a bond-making, bond-breaking process in which a molecule, RA, reacts with water, forming a new R–O bond with the oxygen atom from water and breaking the R–A bond in the original molecule. One possible pathway is the direct displacement of –A with –OH, as Equation 2.7 shows. RA + H2O → ROH + HA
(2.7)
Hydrolytic processes provide the baseline loss rate for any chemical in an aqueous environment. Although various hydrolytic pathways account for significant degradation of certain classes of organic chemicals, other organic structures are completely inert. Strictly speaking, hydrolysis should involve only the reactant species water provides — that is, H+, OH– and H2O — but the complete picture includes analogous reactions and thus the equivalent effects of other chemical species present in the local environment, such as HS– in anaerobic bogs, Cl– in seawater, and various ions in laboratory buffer solutions. Hydrolysis results in reaction products that may be more susceptible to biodegradation, as well as more soluble. The likelihood that a halogenated solvent will
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undergo hydrolysis depends in part on the number of halogen substituents. More halogen substituents on a compound will decrease the chance for hydrolysis reactions to occur and will therefore decrease the rate of the reaction. Hydrolysis rates can generally be described using first-order kinetics, particularly in groundwater where water is the dominant nucleophile. Bromine substituents are more susceptible to hydrolysis than chlorine substituents. As the number of chlorine atoms in the molecule increases, dehydrohalogenation may become more important.12,47 Dehydrohalogenation is an elimination reaction involving halogenated alkanes in which a halogen is removed from one carbon atom, followed by subsequent removal of a hydrogen atom from an adjacent carbon atom. In this two-step reaction an alkene is produced. Although the oxidation state of the compound decreases due to the removal of a halogen, the loss of a hydrogen atom increases it. This results in no external electron transfer, and there is no net change in the oxidation state of the reacting molecule.47 Contrary to the patterns observed for hydrolysis, the likelihood of dehydrohalogenation increases with the number of halogen constituents. Under normal environmental conditions, monohalogenated aliphatics apparently do not undergo dehydrohalogenation. The compounds 1,1,1-TCA and 1,1,2-TCA are known to undergo dehydrohalogenation and are transformed to 1,1-DCE, which is then reductively dechlorinated to VC and ethene. Tetrachloroethanes and pentachloroethanes are transformed to TCE and PCE via dehydrohalogenation pathways.47 Methods to predict the hydrolysis rates of organic compounds for use in the environmental assessment of pollutants have not advanced significantly since the first edition of the Lyman Handbook.8 Two approaches have been used extensively to obtain estimates of hydrolytic rate constants for use in environmental systems.2 The first and potentially more precise method is to apply quantitative structure/activity relationships (QSARs).2,9 To develop such predictive methods, one needs a set of rate constants for a series of compounds that have systematic variations in structure and a database of molecular descriptors related to the substituents on the reactant molecule. The second and more widely used method is to compare the target compound with an analogous compound or compounds containing similar functional groups and structure, to obtain a less quantitative estimate of the rate constant. Predictive methods can be applied for assessing hydrolysis for simple one-step reactions where the product distribution is known. Generally, however, pathways are known only for simple molecules. Often, for environmental studies, the investigator is interested in not only the parent compound but also the intermediates and products. Therefore, estimation methods may be required for several reaction pathways. Some preliminary examples of hydrolysis reactions illustrate the very wide range of reactivity of organic compounds. For example, triesters of phosphoric acid hydrolyze in near-neutral solution at ambient temperatures with half-lives ranging from several days to several years,10 whereas the halogenated alkanes such as tetrachloroethane, carbon tetrachloride, and hexachloroethane have half-lives of about 2 hours, 50 years, and 1000 millennia (at pH = 7, and 25ºC), respectively.11,12 On the other hand, pure hydrocarbons from methane through the PAHs are not hydrolyzed under any circumstances that are environmentally relevant. Hydrolysis can explain the attenuation of contaminant plumes in aquifers where the ratio of rate constant to flow rate is sufficiently high. Thus 1,1,1-trichloroethane
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(TCA) has been observed to disappear from a mixed chlorinated hydrocarbon plume over time, while trichloroethene and its biodegradation product cis-1,2-dichloroethene persist. The hydrolytic loss of organophosphate pesticides in sea water, as determined from both laboratory and field studies, suggests that these compounds will not be long-term contaminants despite runoff into streams and, eventually, the sea. 2.2.1.7 Photolytic Reactions in Surface Water Photolysis (or photolytic reaction) can be defined as any chemical reaction that occurs only in the presence of light. Environmental photoreactions necessarily take place in the presence of sunlight, which has significant photon fluxes only above 295 nm in the near ultraviolet (UV) range, extending into the infrared region of the electromagnetic spectrum.2,13 Environmental photoreactions occur in surface waters, on solid ground, and in the atmosphere, sometimes rapidly enough to make them the dominant environmental transformation processes for many organic compounds. In the atmosphere, for example, photooxidation, mediated by hydroxyl radical (OH•), is the dominant removal process for more than 90% of the organic compounds found there. Photolytic reactions are often complex reactions that not only control the fate of many chemicals in air and surface water, but also often produce products with chemical, physical, and biological properties quite different from those of their parent compounds: more water soluble, less volatile, and less likely to be taken up by biota. Photooxidation removes many potentially harmful chemicals from the environment, although occasionally more toxic products form in oil slicks and from pesticides.14 Biogeochemical cycling of organic sulfur compounds in marine systems involves photooxidation on a grand scale in surface waters, as well as in the troposphere.2 Environmental photoreactions can be divided into two broad categories of reactions: direct and indirect. A direct photoreaction occurs when a photon is absorbed by a compound leading to formation of excited or radical species, which can react in a variety of different ways to form stable products. In dilute solution, rate constants for these reactions are the products of the rate constants for light absorption and the reaction efficiencies. An indirect photoreaction occurs when a sunlight photon is absorbed by one compound or group of compounds to form oxidants of excited states, which then react with or transfer energy to other compounds present in the same environmental compartment to form new products. For example, NO2 and O3 in air form hydroxyl radicals (OH•), and humic acids in water form singlet oxygen and oxyradicals, when they absorb sunlight photons. These oxidants react with other chemicals in thermal (dark) reactions, and the rates for these processes follow simple bimolecular kinetics. Direct Photoreactions: Only a small proportion of synthetic organic compounds absorb UV light in the sunlight region of the spectrum (above 295 nm) and then photolyze at significant rates.13 Most aliphatic and oxygenated compounds, such as alcohols, acids, esters, and ethers, absorb only in the far UV region (below 220 nm),
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and simple benzene derivatives with alkyl groups or one heteroatom substituent absorb strongly only in the far and middle UV region. Nitro or polyhalogenated benzenes, naphthalene derivatives, polycyclic aromatics and aromatic amines, nitroalkanes, azaolkanes, ketones, and aldehydes absorb sunlight between 300 and 450 nm; polycyclic and azoaromatics (dyes), as well as quinones, also absorb visible light, in some cases to beyond 700 nm.2,13 The rate of a direct photoprocess depends only on the product of the rate of light (photon) absorption by compound C,(IA) and the efficiency with which the absorbed light is used to effect reaction (quantum yield, Ø):13 Rate =
dc = Efficiency × Photons Absorbed / time = ∅ I A dt
(2.8)
Under most environmental conditions, chemicals are present in surface water or air at low concentrations, so their light absorbing properties lead to simple kinetic expressions for direct photolysis in water.13 Indirect Photoreactions: Indirect photolysis is most important for compounds that absorb little or no sunlight. Light absorption by chromophores (sensitizers) other than the compound of interest begin the process, forming intermediate (and transient) oxidants or excited states that affect chemical changes in the compound of interest.2,15,16 Examples of sensitizers that serve this purpose are dissolved organic matter (DOM or humic acid) and nitrate ion in water, and ozone and NO2 in the atmosphere. Transient species formed by indirect photoreactions in water include singlet oxygen and peroxy radicals, both of which are relatively selective and electrophilic. As a result, only electron-rich compounds, such as phenols, furans, aromatic amines, polycyclic aromatic hydrocarbons (PAHs), and alkyl sulfides can undergo relatively rapid indirect photoprocesses with these oxidants. Nitroaromatics, though not oxidized, appear to be sensitized by triplet DOM or scavenged by solvated electrons. Many of these compounds (e.g., PAHs, nitroaromatics, and aromatic amines) also undergo rapid direct photoreactions.2,16 By contrast, OH• radical, which dominates tropospheric photochemistry, oxidizes all classes of organic compounds (except perhalogenated compounds such as PCE), including alkanes, olefins, alcohols, and simple aromatics.160,166 Aqueous OH•. radical, derived mainly from the photolysis of nitrate ion, plays an important role in converting marine DOM to simpler carbonyl compounds, even though the average concentration is extremely low (<2 × 10–8).17 OH• also appears important in degrading synthetic chemicals in a variety of nitrate-bearing freshwaters, where the OH• concentrations appear to be one to two orders of magnitude higher.2,13 In many cases, detailed pathways for forming these oxidants and reductants remain unclear, but identities of several of the transients are fairly well established.2,13 Transient species are transient because they react rapidly with themselves or with a variety of natural organic and metal species in natural waters,2 balancing formation rates to give low average concentrations.
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2.2.2
Biological Characteristics
It is generally conceded that biological reactions are of the greatest significance in determining the fate and persistence of organic compounds in most natural aquatic ecosystems. It is essential at the start to make a clear distinction between biodegradation and biotransformation. Biodegradation is a process in which the destruction of a chemical is accomplished by the action of a living microorganism. During biotransformation, on the other hand, only a restricted number of metabolic reactions is accomplished, and the basic framework of the molecule remains essentially intact.18 Even though biodegradation and biotransformation, considered as alternatives, are not mutually exclusive. Biodegradation can be categorized into three types that have importance in an ecosystem setting: Primary Biodegradation: biodegradation to the minimum extent necessary to change the identity of the compound. Ultimate Biodegradation: biodegradation to water, carbon dioxide, and inorganic compounds (if elements other than C, H, and O are present). This is also called mineralization. Under anaerobic conditions, methane may be formed in addition to carbon dioxide during fermentation reactions. Acceptable Biodegradation: biodegradation to the minimum extent necessary to remove some undesirable property of the compound, such as toxicity. Conversion of vinyl chloride to ethene is an example and in many instances this can also be considered biotransformation.
Although biological degradation conceivably might be accomplished by any living organism, available information indicates that, by far, the most significant biological systems involved in ultimate biodegradation of contaminants are bacteria and fungi. Critical and necessary conditions necessary for biodegradation of contaminants to take place are summarized below: • A microbial population must exist that has the necessary enzymes to bring about the biodegradation. • This population must be present in the environment where the contaminant is present. • The contaminant must be accessible to the microorganisms having the requisite enzymes, and most of the time this requires the contaminant to be available in the dissolved phase. • If the initial enzyme bringing about the degradation is extracellular, the bonds acted upon by that enzyme must be exposed for the catalytic enzyme to function. • Should the enzyme catalyzing the initial degradation be intracellular, the molecule must penetrate the surface of the cell to the internal sites where the enzyme acts. Alternatively, for the transformation to proceed further, the products of an extracellular reaction must penetrate the cell. • Because the population or biomass of bacteria or fungi acting on many synthetic compounds is initially small, conditions in the environment must be optimum to allow for proliferation of the potentially active microorganisms.
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The initial concentration of the microbial population and the contaminant somewhat affects the growth and proliferation; a lag period often occurs between the addition of a chemical and the onset of biodegradation. This lag period, usually attributed to the need for acclimation,19,20 could result from enzyme induction, gene transfer or mutation, predation by protozoa, or growth in the population of responsible organisms. The initial species present, their relative concentrations, the induction of their enzymes, and their ability to acclimate once exposed to a chemical are likely to vary considerably, depending upon environmental parameters such as temperature, salinity, pH, oxygen concentration (aerobic or anaerobic), redox potential, concentration and nature of various substrates and nutrients, concentration of heavy metals (toxicity), and effects (synergistic and antagonistic) of associated microflora.21 Many of these parameters affect the biodegradation of contaminants in the environment. One important parameter is the chemical substrate concentration. A number of chemicals have biodegradation rates proportional to substrate concentration, but there are also examples of thresholds and inhibitions.22 Recently, the bioavailability of the chemical to the catalytic enzyme has been identified as a major factor in determining biodegradability in nature. Several studies have demonstrated that, although a chemical freshly added to soil is biodegraded at a moderate rate, the biodegradation rate for some chemicals present in the soil sample for a long time is very low.19 Thus, depending on the chemical, the longer a chemical remains in the soil, the greater the potential for it to become sequestered and less bioavailable. Some microorganisms are capable of biodegrading contaminants without population growth. In this process, known as “cometabolism,”19 the microorganism degrades the contaminant from which it derives no carbon or energy; instead, it is sustained on other organic substrates and nutrients. 2.2.2.1 Cometabolism The transformation of an organic compound by a microorganism that is unable to use the substrate as a source of energy or as one of its growth substrate is termed cometabolism. The active populations thus derive no nutritional benefit from the substrates they cometabolize. Energy sufficient to fully sustain growth is not acquired even if the conversion is an oxidation and releases energy. In addition, the C, N, S, or P that may be in the molecule is not used as a source of these elements for growth and energy deriving purposes. Because of the prefix co, which often is appended to a word to indicate that something is done jointly or together (as in copilot or cooperate), there has been some debate regarding the use of the term cometabolism. Specifically, some classical microbiologists argue that the term should be applied only to circumstances in which a substrate that is not used for growth is metabolized in the presence of a second substrate that is used to support multiplication.19 According to this view, the transformation of a substance that is not used as a nutrient or energy source but which occurs in the absence of a chemical supporting growth should be designated by another term, for example, fortuitous metabolism. However, the prefix co also has another meaning, namely, “the same or similar.” The latter usage implies that the cometabolic transformation is similar to some other metabolic
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reaction, which is consistent with one explanation for the phenomenon. Fortuitous metabolism is, indeed, a more attractive term because it suggests an explanation for cometabolism, but the term will be used here as in the original definition, if for no other reason than it has gained wide acceptance. The term cooxidation is sometimes used in studies of pure cultures of bacteria, referring specifically to oxidations of substrates that do not support growth in the presence of a second compound that does support multiplication. Cooxidation has historical precedence in the debate but since it is restricted to oxidation, the word does not have sufficient breadth to include many reactions that are not oxidations.19 In summary, two types of reactions called cometabolism take place in the environment. In one, the cometabolized compound is transformed only in the presence of a second substrate, which indeed may be the compound that supports growth. For heterotrophs, the energy-providing substrate is organic; for autotrophs, it is inorganic. In the other type, the compound is metabolized even in the absence of a second substrate. Important reasons for using the more general definition, and even for maintaining cometabolism as a term apart from bioconversion or biotransformation, are the environmental consequences of cometabolism. Cometabolic reactions have impacts in nature that are different from growth-linked biodegradations, and when the transformations take place, it is usually totally unclear whether the microorganisms do or do not have a second substrate available on which they are growing. A large number of chemicals are subject to cometabolism in nature. Among cometabolic conversions that appear to involve a single enzyme, the reactions may be hydroxylations, oxidations, denitrations, deaminations, hydrolyses, acylations, or cleavages of ether linkages; however, many of the conversions are complex and involve several enzymes. Some of the unique cometabolic reactions brought about by bacteria and fungi in nature come as no surprise in view of the vast array of growth linked biological transformations that heterotrophic bacteria and fungi are capable of in nature. An example which has no significant importance in contaminant removal is the methane monooxygenase of methanotrophic bacteria which is able to oxidize alkanes, alkenes, secondary alcohols, methylene chloride, chloroform, dialkyl ethers, cycloalkanes, and various aromatic compounds.19 Caution needs to be exercised in concluding that cometabolism is occurring merely because an organism cannot be isolated from an environment in which a chemical is undergoing a biological reaction.19 The isolation of bacteria acting on specific substrates is usually performed by enriching the organism in a medium when the only C source is the test chemical, and the agar medium used to plate the enrichments contains that single organic supplement. Yet, many bacteria that are able to grow at the expense of that substrate will not develop in such simple media because they require amino acids, B vitamins, or other growth factors. These essential growth factors are not routinely included in such liquid media, and hence bacteria and fungi needing them fail to proliferate. If the only organisms in the environment able to metabolize a contaminant need these growth factors, no isolate will be obtained, and an erroneous conclusion will be reached that the compound is cometabolized. If a chemical supports the growth of many species, some will undoubtedly require no growth factors (these organisms are called prototrophs), and they will be
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enriched and ultimately can be isolated. If the compound is acted on by only one species, in contrast, it is likely that the responsible organism will need amino acids, B vitamins, or other growth factors; these species are termed auxotrophs. Hence, the failure to isolate a bacterium or fungus capable of using the contaminant as the sole C source for growth is not sufficient evidence for cometabolism. Mechanisms of Cometabolic Reactions: Several reasons have been advanced to explain cometabolism, that is, why an organic chemical that is a substrate does not support growth but is converted to products that accumulate. Some of these reasons have experimental support: (1) the initial enzyme converts the substrate to a product that is not further transformed by other enzymes in the microorganism to yield the metabolic intermediates ultimately used for biosynthesis and energy production; (2) the initial substrate is transformed to products that inhibit the activity of later enzymes in mineralization or that suppress growth of the organisms; and (3) the organism needs a second substrate to bring about some particular reaction. It could be speculated that the first explanation is the most common. The basis for this explanation is the fact that many enzymes act on several structurally related substrates; thus, an enzyme naturally present in the cell will catalyze reactions and alter synthetic chemicals that are not typical cellular intermediates. These enzymes are not absolutely specific for their substrates. Consider a normal metabolic sequence involving the conversion of A to B by enzyme a, B to C by enzyme b, and C to D by enzyme c in a sequence that ultimately yields CO2 energy for biosynthesis reactions and intermediates that are converted to cell constituents.19 a
b
c
A → B → C → D →→→ CO 2 + energy + cell – C
(2.9)
The first enzyme a may have a low substrate specificity and act on a molecule structurally similar to A, namely, A.1 The product (B1) would differ from B in the same way that A differs from A1. However, if enzyme b is unable to act on B1 (because the structural features controlling which substrate it modifies differ from those controlling the substrate specificity of enzyme a), B1 will accumulate:19 a
→ → 1 A B1
(2.10)
In addition, CO2 and energy will not be generated and, because cellular carbon is not formed, the organisms do not multiply. The formation of B1 is thus entirely fortuitous.19 In instances where the contaminant concentration is high, cometabolism may result from the conversion of the parent compound to toxic products. In the sequence just depicted, if the rate of reaction catalyzed by enzyme a is faster than the process catalyzed by enzyme b, B will accumulate because it is not destroyed as readily as it is generated. For example, a strain of pseudomonas that grows on benzoate but not 2-fluorobenzoate converts the latter to fluorinated products that are toxic.23 The inhibitor that accumulates may affect a single enzyme important for the further metabolism of the toxin.
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In some instances, an organism may not be able to metabolize an organic compound because it needs a second substrate to bring about a particular reaction. The second substrate may provide something that is present in insufficient supply in the cells for the reaction to proceed — for example, an electron donor for the transformation.19 The above explanation is linked to the existence of enzymes acting on more than a single substrate. Many enzymes are not absolutely specific for a single substrate. As a rule, they act on a series of closely related molecules, but some carry out a single type of reaction on a variety of somewhat dissimilar molecules. The following are examples of single enzymes acting on a range of substrates: • Methane monoxygenase of methanotrophic bacteria: When grown on methane, methanol, or formate, these aerobic bacteria are able to cometabolize a large array of organic molecules, including several major pollutants. In each instance, methane monooxygenase is the responsible catalyst. Other chlorinated aliphatic hydrocarbons transformed by one such methanotroph, Methylosinus trichosporium, are cisand trans-1,2-dichloroethylene, 1,1-dichloroethylene, 1,2-dichloropropane, and 1,3-dichloropropylene.24 Apparently the same enzyme in other bacteria, after growth on methane, will catalyze the oxidation of n-alkanes with two to eight C atoms, n-alkenes with two to six C atoms, and mono- and dichloroalkanes with five or six C atoms, as well as dialkyl ethers and cycloakanes.19 • Toluene dioxygenase of a number of aerobic bacteria: This enzyme incorporates both atoms of oxygen from O2 (hence, it is a dioxygenase) into toluene as it catalyzes the first step in the degradation of toluene by bacteria grown on that aromatic hydrocarbon (Figure 2.3a). However, that same enzyme has very low specificity and also is able to bring about the degradation of TCE,19,25,26 to convert 2- and 3-nitrotoluene to the corresponding alcohols, and to hydroxylate the ring of 4-nitrotoluene.19,27 • Toluene monooxygenase of several aerobic bacteria: Differing from the dioxygenase, this enzyme incorporates only one atom of oxygen from O2 into toluene to give o-cresol (Figure 2.3b). However, because of this enzyme, bacteria can cometabolize TCE, convert 3- and 4-nitrotoluenes to the corresponding benzyl alcohols and benzaldehydes, and add hydroxyl groups to other aromatic compounds.19,28,29 • Oxygenase of propane-utilizing bacteria: Aerobes using propane as C and energy source for growth also have an oxygenase of broad specificity. This enzyme cometabolizes TCE, vinyl chloride, and 1,1-di- and trans- and cis-1,2-dichloroethylene and has been recently known to degrade MtBE.19,30 • Ammonia monooxygenase of Nitrosomonas europaea: This bacterium, which is a chemoautotroph whose energy source in nature is NH4+ and whose C source is CO2, cometabolizes TCE, 1,1-dichloroethylene, various mono- and polyhalogenated ethanes, and a variety of monocyclic aromaticcompounds and thioethers, as well as methyl fluoride and dimethyl ether.19,31 • An alkane hydroxylase hydroxylates a number of alkylbenzenes and linear, branched, and cyclic alkanes.19,23 • An alkane monooxygenase degrades TCE, vinyl chloride, and dichloroethylenes and propylenes.19 • Naphthalene dioxygenase acts on xylene, isomers of nitrotoluene, and ethylbenzene.19,33 • Biphenyl dioxygenase transforms several PCB congeners.19,34
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CH3 OH 2H H OH
O2 H
TOLUENE CH2OH
CH3 NO2
NO2
2-NITROTOLUENE CH2OH
CH3
NO2
NO2
3-NITROTOLUENE CH3
CH3
CH3 OH
OH + OH
NO2
NO2
NO2
4-NITROTOLUENE O CCI2 CICH TCE Figure 2.3a
HC
COOH + HCOOH + CI-
Reactions catalyzed by toluene dioxygenase (adapted from Alexander, 1999).
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CH3
CH3 OH
TOLUENE Figure 2.3b
o-CRESOL
Reactions catalyzed by toluene monooxygenase (adapted from Alexander, 1999).
The organism containing these enzymes may be able to use one of several of the enzyme’s substrates for growth. However, many of the substrates are transformed but do not support growth. The product of the reactions then accumulates. Because cometabolism generally leads to a slow degradation of the substrate, attention has been given to enhancing its rate.19 The addition of a number of organic compounds to the contaminated zone promotes the rate of cometabolism of a number of chlorinated aliphatic and aromatic compounds and chlorinated phenols, but the responses to such additions are not predictable. No relation is known to exist between the metabolic pathways involved in the degradation of the added mineralized substrate and the compound that is cometabolized in these circumstances. The added substrates were randomly chosen in these trials, and sometimes they do and sometimes they do not stimulate cometabolism. In instances in which stimulation occurs, the benefit probably results from an unpredicted increase in the biomass of organisms, some of which fortuitously cometabolize the compound of interest. An alternative approach is to add mineralizable compounds that are structurally analogous to the compound whose cometabolism one wishes to promote. Presumably, the microorganism that grows on the mineralizable compound contains enzymes transforming the analogous molecule that is cometabolized. This larger biomass thus has more of the degradative enzyme than is present in the unsupplemented water or soil. This method of analogue enrichment has been used to enhance the cometabolism of PCBs by additions of biphenyl. The unchlorinated biphenyl was selected for addition to soil since it is mineralizable, nontoxic, and serves as a C source for microorganisms that are able to cometabolize PCBs.19 Analogue enrichment is a procedure that is similar to the usual means of isolating bacteria that can cometabolize a compound. The enrichment culture contains a C source that supports growth, and the pure cultures thus obtained also cometabolize structurally related compounds that would not support growth. For example, bacteria isolated on diphenylmethane and containing enzymes to degrade it also cometabolize chlorinated diphenylmethanes. Many of the latter do not sustain growth.19 2.2.2.2 Kinetics of Biodegradation Impacts of the Environment: Soil, water, sediment, and wastewater environments have different microbial populations and different available nutrients which may affect considerably the rate of biodegradation. For example, wastewater
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33
treatment plants may have high levels of nutrients and high microbial populations that may have been pre-exposed to a contaminant (acclimated microbial population), but the contact (retention) time is relatively short. In contrast, marine waters are usually fairly low in nitrogen and phosphorous which may limit the biodegradation of chemicals (e.g., oil spills). Sediment often has high levels of organic nutrients, but often is anaerobic, while surface waters tend, by comparison, to have low levels of organic nutrients.18 Digestor sludge from wastewater treatment plants has high organic nutrients and is anaerobic. Surface soils have high concentrations of organic nutrients (depending on the type of soil), but this usually decreases with depth. In the past, it was believed that groundwater aquifers were devoid of microbial life. However, a number of studies have demonstrated that microorganisms are quite plentiful in certain aquifers, and, in some instances, the bacterial concentration and activity in aquifers may be higher than those in surface waters.19,35 In addition, availability of the chemical to the microbial population can be affected considerably by the conditions of the microenvironment (e.g., organic concentration or clay content may bind the chemical tightly). A contaminant may become less available or essentially unavailable for biodegradation if it enters or is deposited in a micropore that is inaccessible even to the microorganisms. These micropores may be filled entirely with water, as in sediments or groundwater aquifers, and the contaminant would have to move out of a micropore by diffusion to be accessible to bacteria for its destruction. The tortuous path the contaminant molecule must traverse before it gets destroyed dramatically affects bioavailability if the contaminant not only is physically remote from potentially active microorganisms, but also is strongly sorbed to solid surfaces associated with that remote micropore.19 Some organic compounds that persist in the subsurface often undergo a time dependent decline in bioavailability. Since this process is slow and time dependent it is appropriately called aging. This modification in bioavailability to microorganisms as a result of aging is also called sequestration.19 In the initial period, the compound gradually disappears as a result of biodegradation, and possibly by other mass removal mechanisms, but little or none of the compound is destroyed after it has resided in the soil or sediment for some time. Witness the finding that although 80% of hexachlorobenzene deposited in the early 1970s in a lake bottom sediment was dechlorinated in the succeeding 20 years, all sediment cores still contained at least 40 ppb of hexachlorebenzene. This time-dependent change in the rate of degradation, which has been observed with a number of insecticides, was the first line of evidence for sequestration.19 Because these aged molecules are solvent extractable, albeit by vigorous treatment, they are presumed to be present in an uncomplexed form and thus considered to be contaminants and subject to the regulations and cleanup standards. In addition to the above mentioned contaminants, PAHs with three or more rings, such as phenanthrene, anthracene, fluorene, pyrene, chrysene, and others will also undergo sequestration. It has been known for some time that it is increasingly difficult to remove strongly hydrophobic compounds from soil with mild extractants as the residence time of
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
those compounds in the soil increases. This phenomenon is not restricted to soils; a similar decline in extractability is witnessed also from sediment samples.36 The amount of a contaminant that is sequestered increases with time. Expressed in another way, the percentage of the chemical that is bioavailable diminishes with increasing persistence. This presumably occurs because more of the contaminant is diffusing into inaccessible sites. However, after a period of time that varies with the soil and the compound, sequestration of additional quantities slows and possibly stops. The reason for this rate of decline is not presently known.36 Based on the preceding discussion, it can be seen that the rates of biodegradation are likely to vary considerably, depending on the environment to which a contaminant is released, the type of contaminant(s), and the age of contamination. Also, the rates under different conditions may vary depending upon the type of chemical structure. For example, nitro aromatic compounds are usually fairly resistant to biodegradation under aerobic conditions but are reduced rapidly to amines under anaerobic conditions. In contrast, degradation of benzene takes place significantly faster under aerobic conditions than under anaerobic conditions. Structural Effects on Biodegradation: In addition to contaminant concentration, chemical structure and physical/chemical properties have considerable impact on the rate and pathways of biodegradation. The chemical structure determines the possible pathways that a substrate may undergo, generally classified as oxidative, reductive, hydrolytic, or conjugative. Figure 2.4 provides some examples of common microbial degradation pathways.37 Recently, a computer program was developed that will predict the most probable metabolites, and another computer program was also developed that simulates the biodegradation of synthetic chemicals through the sequential application of plausible biochemical reactions.38,39 Over the years, structure/biodegradability “rules of thumb” have been developed.40,41 Figures 2.4a and b summarize these. Some of these structure/biodegradibility relationships have some biochemical mechanistic underpinings. For example, highly branched compounds frequently are resistant to biodegradation because increased substitution hinders β-oxidation, the process by which alkyl chains and fatty acids usually are biodegraded. This structural relationship was discovered in the 1950s when detergent scientists found that alkylbenzene sulfonate (ABS) detergents passed through wastewater treatment plants causing foaming problems in rivers and streams. This problem was solved by switching from the highly branched ABS detergents to linear alkylbenzene sulfonate (LAS) detergents, thus illustrating the importance of understanding the relationship between structure and biodegradability. Few other rules of thumb have such mechanistic bases, but there are some general trends. Functional groups commonly seen by microorganisms in natural products usually are degraded easily, probably because the microbes have had eons to develop the required enzyme systems in order to gain carbon and energy from the metabolism. Conversely, functional groups less common in nature or newly synthesized by man usually make a chemical more resistant to biodegradation. Aromatic substituents that are electron withdrawing (e.g., nitro groups and halogens) increase the persistence of a chemical, possibly by making it more difficult for enzymes to attack the aromatic ring, whereas electron donating functionalities (e.g., carboxylic acids, phenols, amines) generally increase biodegradation rates.
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CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
Type of Reactions (not all steps are given)
Example of Chemicals Subject to Reaction
β-oxidation
Fatty acids and straight chain hydrocarbons (after oxidation of chain to carboxylic acid see methyl oxidation)
CH3[CH2]x
CO2H
CH3[CH2]x
CO2H
Methyl oxidation R
CH3
R
CH2OH
R
CHO
R
CH2H
Epoxide formation
35
Aromatic and aliphatic methyl groups Olefins
O R
R
R
R
Hydroxylation and ketone formation
Aromatic to form phenols and hydrocarbons to alcohols and then ketones
OH
OH R
R
O R
R
R
R
Nitrogen oxidation R
NH2
R
Aromatic amines to nitroaromatic
NHOH
R
N=O
R
NO2
NHOH
R
NH2
Nitro reduction R
NO2
R
N=O
R
Nitrile/amide metabolism
Nitroaromatics aromatic amines (e.g., parathion) especially fast under anaerobic conditions Bromoxynil, Dichlobenil
O R
CN
R
CO2H
Sulfides such as aldicarb
O
S R
R
NH2
Sulfur oxidation
O
S R
R
R
R
S
R
O
Thiophosphate ester oxidation
Thiophosphate pesticides
O S
S,OP
S,OP
S
S,O S
S
S,O S,OP
O
S,O
Dehalogenation CI
Aromatic and aliphatic halogens
OH
R
CH2CI
R
CH2OH
R
CCI3
R
CO2H
Hydrolysis S,O
R
S,OP S,O O
R S,O
Figures 2.4a
S,OP S,O O
R R
R
O
Phosphate and carboxylic esters
S,O
R H2O
OH R
H 2O
H R
O
Common microbial degradation pathways (after Boethling and MacKay, 2000).
Physical/chemical properties affect the rate of biodegradation mostly by affecting bioavailability. Compounds which are sparingly soluble in water tend to be more resistant to biodegradation, possibly due to an inability to reach the microbial enzyme site, a reduced rate of availability due to solubilization, or sequestration due to adsorption or trapping in inert material.19,40 Biodegradation Rates: The study of the kinetics of biodegradation in natural environments is often empirical, reflecting the rudimentary level of knowledge about microbial populations and activity in these environments. An example of an empirical approach is the power rate model.19 –dC/dt = kCn
(2.11)
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
More Biodegradable (Less Perisistent)
Less Biodegradable (More Persistent) Branching
H
R
R
R
R
R
R
R
R
H
H
R
Aliphatic functional groups R
CH2OH
R
CO2H
R
NH2
R
SO3R
O
N R
R
R
R R
N R
CI R
R
N R
R
Aromatic functional groups (benzene, naphthalene, pyridine rings) OH
CI
CH3
OMe
SO3H
NO2
CO2H
NH2
CF3
Br
Aliphatic amines R
N
H
R
H
R
H
N
N
R
N
R
O
R
Halophenols CI OH
OH CI
CI
OH
CI OH
OH
OH
CI
CI
CI
OH
CI Br
CI CI OH CI
OH
CI
CI
CI CI
Br
CI
CI
OH
CI
CI
Polycyclic aromatics ≤3 rings
≥3 rings
Triazines R
N
S
CI
N
N
CH2H3 N
N
CH3
N CH3
Figure 2.4b
CH2H3 N
N
CH3
N CH3
Relationship between chemical structure and biodegradability (after Boethling and MacKay, 2000).
where C is substrate concentration, t is time, k is the rate constant for chemical disappearance, and n is a fitting parameter. This model can be fit to substratedisappearance curves by varying n and k until a good fit is achieved. It is evident from this equation that the rate is proportional to a power of the substrate concentration. The power-rate law provides a basis for comparison of different curves, but it gives no insight into the reasons for the shapes. Therefore, often it may have no predictive ability. Moreover, investigators interested in kinetics do not always state whether the model they are using has a theoretical basis or is simply empirical, and whether constants in an equation have physical meaning or are only fitting parameters.19 An appropriate introduction to the kinetics of biodegradation is to consider
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CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
37
a pure culture of a single bacterial population growing on and degrading a single, soluble organic chemical, and to assume that no barriers exist between the substrate and the cells. Biodegradation kinetics, under the conditions described previously, have been reviewed in detail and a number of kinetic models proposed, including the use of screening tests for generating biodegradation kinetics. 19,42,43 Biodegradation rates typically are interpreted through the Monod equation (Equation 2.12), which is analogous to the Michaelis-Menten equation used in enzyme kinetics: µ = µm
[S] K s + [S]
(2.12)
The parameters µ and µm refer to the growth rate and maximum growth rate, respectively, in the presence of substrate concentration S. Ks is the half-velocity coefficient, i.e., the value of S at which µ = 0.5µm (Figure 2.5). Equation 2.13 can express the degradation rate of a substrate:
Growth Rate (µ)
µmax
0.5 µmax
Ks
Figure 2.5
Substrate Concentration (s)
Relationship between the growth rate of bacteria vs. the substrate concentration as described by the Monod kinetics model.
−
µ m [S][B] d[S] = − dt Y(K s + [S])
(2.13)
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
where B represents biomass and Y is the growth yield factor. The Monod equation assumes that the compound of interest sustains growth and is the only source of carbon and, thus, its applicability to a naturally contaminated environment may be limited. For example, for cometabolic processes with µm and Y defined as zero, the equation would not apply. The equation also ignores toxicity and makes no provision for acclimation. Experimentally, rates are measured either at low substrate concentrations where Ks > [S] and Equation 2.13 simplifies to Equation 2.14, or at high substrate concentrations where [S] > Ks and Equation 2.15 follows from Equation 2.13: −
d[S] µ m = [S][B] dt YK s
(2.14)
d[S] µ m = [B] dt Y
(2.15)
−
For the former case (Equation 2.14), which is environmentally more relevant for low contaminant concentrations typical of many sites, the rate obeys first-order kinetics with respect to substrate and biomass (second-order overall), whereas in the latter case (Equation 2.15), the kinetics have a first-order relationship to biomass but are independent of substrate concentration (Figure 2.6). Methods for measuring biomass, B, have varied widely, and, for studies involving mixed populations, in which only a fraction of the organisms can degrade the substrate, a means for quantifying the responsible fraction is not available. The kinetics of cometabolism have received scant attention. If the microbial populations are neither growing nor declining and the concentration of substrate for cometabolism is below the Km of the active organisms, it is likely that the conversion would be first-order. In a biofilm bioreactor inoculated with methane-oxidizing bacteria, the cometabolism of TCE, 1,1,1-trichloroethane, and cis- and trans-1,2dichloroethylene is first-order at concentrations up to 1 mg/liter.19 However, in environments in which the transformations are slow, the C source for growth probably is being depleted, so the kinetic patterns may change with time. Other models have been developed for cometabolism by nongrowing or growing populations.19 2.3 2.3.1
ENVIRONMENTAL CHARACTERISTICS
Sorption Coefficient
Sorption processes play a major role in determining the environmental fate and impact of contaminants, affecting a variety of specific fate processes, including solubilization, volatilization, bioavailability, biodegradability, and hydrolysis. Sorption coefficients quantitatively describe the extent to which an organic contaminant is distributed at equilibrium between an environmental solid (i.e., soil, sediment,
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Log [Cs]
CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
39
Second Order
Zero Order
First Order
Time Figure 2.6
Concentration of substrate vs. time for zero-, first-, and second order biodegradation reactions.
suspended sediment, wastewater solids) and the aqueous phase with which it is in contact. Sorption coefficients depend on (1) the variety of interactions occurring between the solute and the solid and aqueous phases and (2) the effects of environmental variables such as organic matter quantity and type, clay mineral content and type, clay to organic matter ratio, particle size distribution and surface area of the sorbent, pH, ionic strength, suspended particles or colloidal material, temperature, dissolved organic matter (DOM) concentration, solute and solid concentrations, and phase separation process. Adsorption, absorption, and sorption are terms used to describe the uptake of a solute by another phase. Adsorption describes the concentration of a solute at the interface of two phases, while absorption describes the process when a solute is transferred from the bulk state of one phase into the bulk state of the other phase. The term sorption is used frequently in environmental situations to denote the uptake of a solute by a solid (soil or sediment or component of soil) without reference to a specific mechanism, or when the mechanism is uncertain.44 Sorption occurs when the free energy of the interaction between an environmental solid sorbent and contaminant sorbate is negative. The sorption process can be either enthalpy or entropy driven, depending on the properties of the solid sorbent and chemical solute. Enthalpy-related forces include van der Waals interactions, electrostatic interactions, hydrogen bonding, charge transfer, ligand exchange, direct and induced dipole-dipole interactions, and chemisorption, while hydrophobic bonding or partitioning is considered the primary entropy driven force.2,44 Figure 2.7 shows the polarity of the H2O molecule.
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
Hydrogen
"+" "-"
Oxygen Positive side of 105° H 2O molecule
Negative side of H2O molecule
"+" Hydrogen Figure 2.7
The polarity of the H2O molecule. Because of the non-linear position of H+s, water is polar. The H2O molecule has one portion that is more negative than positive and an opposite side that has two hydrogens which are more positive than negative.
The complex and heterogeneous nature of environmental solids makes it difficult, if not impossible, to identify specific sorption mechanisms for most solid-chemical combinations; in most situations, several mechanisms operate simultaneously. In most soils, and under most conditions, organic chemicals are sorbed on both organic and inorganic constituents. The relative importance of organic vs. inorganic constituent depends on the amount, distribution, and properties of those constituents and the properties of the organic chemical. As the polarity, number of functional groups, and ionic nature of the organic chemical increases, so too does the number of potential sorption mechanisms (Figure 2.8). Fortunately, for many solid-organic chemical interactions, one or two mechanisms dominate the sorption process and generalizations regarding sorption behavior can be made.44 For instance, the sorption of most neutral, hydrophobic organic chemicals by environmental solids correlates highly with the organic matter content of the solid. The extent to which clay minerals contribute to sorption depends on both the ratio of clay mineral to organic carbon fractions of the soil or sediment and on the nature of the organic sorbate. A ratio of 40 has been suggested as the cutoff for organic carbon dominated sorption.45 Among the various inorganic soil constituents, smectites have the greatest potential for sorption of organic chemicals, due to their large surface area and abundance in agricultural soils.44,46 Soil is a dynamic and life-sustaining system composed of solids, liquid, and gas, with solids typically accounting for about one-half to two-thirds by volume. Living
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CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
41
NONPOLAR
Cl
POLAR
Cl
Cl C
-
C
Cl
Chloride ion
Cl
H
Tetrachloroethylene (PCE)
δ+
O δ−
H
H
H
d+
Water H
H H
H+
H
Hydrogen ion
Benzene
H
H
H
H
C
C
C
H
H
H
H
H
Propane H
H
H
C
C
C
H
H
H
OH
Propanol
H
H
H
H
H
H H H Naphthalene
Figure 2.8
H
H
H
C
C
H
O
Acetate Ion
Some examples of polar and nonpolar chemical species. Note that unbalanced electrical charge, asymmetry, and the presence of oxygen all tend to make chemicals more polar.
organisms are also very important parts of soil and contribute greatly to its general properties and behavior. The solid phase of soil comprises fragmented mineral matter, derived from the weathering of hard rock at the earth’s surface, and from soil organic matter (SOM) consisting of a mixture of plant and animal residues in various stages of decomposition and substances synthesized microbiologically.1 In its broadest sense, the term SOM encompasses all organic materials contained in soil and is made up of live organisms, their decomposed and partly decomposed remains, and microbially and/or chemically resynthesized products resistant to further biological attack. More specifically, the term SOM refers to the nonliving organic components, which are largely composed of products resulting from microbial and chemical transformations of organic debris. Some scientists have defined SOM as the total of organic components in soil, excluding undecayed plant and
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
animal tissues, their “partial decomposition” products (the organic residues), the soil biomass (living microbial tissues), and macrofauna and macroflora. The terms SOM and humus are thus generally interchangeable.1 To simplify this very complex system, SOM is generally divided into two groups designated as nonhumic and humic substances.1 The nonhumic substance group comprises organic compounds that belong to chemically recognizable classes and are not unique to the soil. These include polysaccharides and simple carbohydrates, amino sugars, proteins and amino acids, fats and waxes, lignin, resins, pigments, nucleic acids, hormones, a variety of organic acids, and so on. Most of these substances are relatively easily degradable and can be utilized as substrates by soil microorganisms, and as such have a transient existence in the soil. In contrast, humic substances comprise a heterogeneous mixture of chemically unidentifiable macromolecules that are distinctive to and synthesized in the soil, and are relatively resistant to chemical degradation and microbial attack. Recent estimates of the average composition of SOM are the following: carbohydrates, 10%; N components, 10%; lipids (including alcanes, fatty acids, waxes, and resins), 15%; humic substances, 65%. However, different soils may contain widely different amounts of nonhumic and humic substances. The amount of carbohydrates can range from 5 to 25%; proteins may vary from 15 to 45%; lipids from 2% in forest SOM to 20% in acid peat soils, and humic substances from 33 to 75% of the total SOM.1 Humic substances are the most widespread and ubiquitous natural nonliving organic materials in all terrestrial and aquatic environments and represent a significant proportion of total organic C in the global C cycle. They constitute the major fraction of SOM (up to 80%) and the largest fraction of natural organic matter (NOM) in aquatic systems (up to 60% of dissolved organic C).1 Soil humic substances comprise a physically and chemically heterogeneous mixture of naturally occurring, biogenic, relatively high-molecular weight, yellowto-black colored, amorphous, colloidal, polydispersed organic polyelectrolytes. These polyelectrolytes are of mixed aliphatic and aromatic natures, formed by secondary synthesis reactions (humification) during the decay process and transformation of organic matter originating from dead organisms and microbial activity. These materials are distinctive of the soil system and exclusive of undecayed plant and animal tissues, their partial decomposition products, and the soil biomass. Soil water acts both as a solvent for the organic chemical and as a solute with which the organic chemical has to compete for sorption sites on the solid surface. Typically, soil water is a solution comprising mainly Ca+2, Mg+2, Na+, K+, SO4–2 , CO3–2 , and HCO3– . Ionic strengths are typically 0.5 mol/L or higher; pH values of 5–8.5 are common.44 The characteristics of the solution phase determine the reaction chemistry and the dissolution/precipitation reactions, and they influence ion activity, ion pairing, and speciation. All these potentially can influence a chemical’s sorptive behavior (Figures 2.9a, b, and 2.10). In large lakes and estuaries, the natural organic material in sediments and suspended sediments is derived from a mixture of the remains of terrestrial and planktonic organisms. Generally, soils and sediments differ in the amount and type of organic matter they contain. Soils typically contain higher percentages of cellulose
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CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
WATER
43
Air Where Water has Drained WATER
SOIL PARTICLE
SOIL PARTICLE
Matric Potential Increases toward Particle Surface
WATER SOIL PARTICLE WATER Air Where Water has Drained
SOIL PARTICLE
Water is Held Strongly near the Soil Particle Surface
Figures 2.9a
A cross section of a soil pore and the solid particles that make up its walls. Water is held strongly as the distance from the soil particle decreases; at some distances from the surface, water is held so weakly that the pull of gravity causes some of it to drain.
and hemicellulose, while sediments contain higher percentages of lipid-like material. For neutral organic compounds, sorption is generally greater in sediments than in soils, even when normalized to organic carbon content. 2.3.1.1 Soil Sorption Coefficients Sorption coefficients quantitatively describe the extent to which an organic chemical distributes itself between an environmental solid (i.e., soil, sediment, suspended sediment, wastewater solids, etc.) and the aqueous phase that it is in contact with at equilibrium. Sorption coefficients generally are determined from an isotherm, a diagram that depicts the distribution of the test chemical between a solid sorbent and the solution in equilibrium with it over a range of concentrations at constant temperature (Figure 2.11). These isotherms can be linear or nonlinear, depending on the properties of the chemical and solid and on the aqueous phase concentration, but tend to become nonlinear (sorption tends to decrease) as the concentration of chemical in the aqueous phase increases, especially for polar or ionizable chemicals or soils that are low in organic carbon and high in clay. Linear sorption isotherms often are observed if the equilibrium aqueous phase organic compound concentrations are below 10–5 M or
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
Solids
Cg = H'C w Trapped gas bubble w
Advection - dispersion
Pore Space C w = Aqueous phase concentration C g = Gas phase concentration H = Dimensionless Henry's Law Constant
Figure 2.9b
Trapped gas in saturated soil.
one-half the aqueous phase solubility (whichever is lower) and the organic content of the solid is greater than 0.1%: Kd = CS /CW
(2.16)
where, CS and CW are the concentrations of the organic chemical sorbed by the solid phase (mg/Kg) and dissolved in aqueous phase (mg/L), respectively. Units of Kd typically are given as L/kg, mL/g, or cm3/g. For nonlinear isotherms, the Freundlich model most often is used to describe the relationship between the sorbed (CS) and the solution phase concentrations (CW): CS = Kf CWN
(2.17)
where, Kf is the Freundlich sorption coefficient and N (values of N are less than one and typically range between 0.75 and 0.95) generally is a constant.47 However, in some cases, N has been observed to exceed one. When N is equal to one, a linear equation results, and Kf and Kd are equivalent. The Langmuir and Brumnauer, Emmett, and Teller (BET) models also have been used to describe nonlinear sorption behavior for environmental solids, particularly
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CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
45
A
Air Meniscus Air C
BA B Soil Grain
C
B C
-
+
0 Porewater Pressure
B
C
Meniscus
Height
A
A only A and B
-
A, B, C 0
+
Water Pressure
Figure 2.10
Soil water under three different values of water content. At high water content (condition C) porewater pressure is rendered negative by the force of surface tension acting over a meniscus of relatively large area. The meniscus may be thought of as a flexible diaphragm that is under tension, thus pulling on the water on its convex side. As water content is decreased and the meniscus retreats into smaller pore spaces (B, then A), surface tension forces act over a smaller area of water, and the resulting water pressure is more negative. The same effect occurs in capillary tubes, where the most suction (more negative pressure, and thus more capillary rise) is developed in the tube of smallest diameter.
for mineral dominated sorption.44 The Langmuir model assumes that maximum adsorption corresponds to a saturated monolayer of solute molecule on the adsorbent surface, that there is no migration of the solute on the surface, and that the energy of adsorption is constant. The BET model is an extension of the Langmuir model that postulates multiplayer sorption. It assumes that the first layer is attracted most strongly to the surface, while the second and subsequent layers are more weakly held.47 The most commonly used method for expressing the distribution of an organic compound between the aquifer matrix and the aqueous phase is the distribution coefficient, Kd, which is described by Equation 2.16. The transport and partitioning of a contaminant are strongly dependent on the chemical’s soil-water distribution coefficient and water solubility. The distribution coefficient is a measure of the sorption/desorption potential and characterizes the tendency of an organic compound to be sorbed to the aquifer matrix. The higher the distribution coefficient, the greater the potential for sorption to the aquifer matrix. The distribution coefficient is the slope of the sorption isotherm at the contaminant concentration of interest. The greater the amount of sorption, the greater the value of Kd. For systems described by a linear isotherm, Kd is a constant.47 In general
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
Linear
Adsorbed Concentration Cs(µg/g)
Freundlich
Langmuir
Dissolved Concentration Cw(µg/mL) Figure 2.11
Characteristic adsorption isotherm shapes.
terms the distribution coefficient is controlled by the hydrophobicity of the contaminant and the total surface areas of the aquifer matrix available for sorption. Thus the distribution coefficient for a single compound will vary with the composition of the aquifer matrix. Because of their extremely high specific surface areas (ratio of surface area to volume), the organic carbon and clay mineral fractions of the aquifer matrix generally represent the majority of sorption sites in an aquifer. Based on literature reports, it appears that the primary adsorptive surface for organic chemicals is the organic fraction of the aquifer matrix.47 However, there is a critical level of organic matter below which sorption onto mineral surfaces is the dominant sorption mechanism.47,48 This critical level of organic matter, below which sorption appears to be dominated by mineral-solute interactions, and above which sorption is dominated by organic carbon-solute interactions, is given by47,48 focc =
As 1 0.84 200 K ow
where focc = critical level of organic matter (mass fraction) As = surface area of mineralogical component of aquifer matrix Kow = octanol-water partitioning coefficient
(2.18)
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CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS
47
From this relationship it is apparent that the total organic carbon content of the aquifer matrix is less important for solutes with low octanol-water partitioning coefficients (Kow).47 Also apparent is the fact that the critical level of organic matter increases as the surface area of the mineralogic fraction of the aquifer matrix increases. The surface area of the mineralogic component of the aquifer matrix is most strongly influenced by the amount of clay. For compounds with low Kow values present in materials with a high clay content, sorption to mineral surfaces could be an important factor causing retardation of the chemical. Several researchers have found that if the distribution coefficient is normalized relative to the aquifer matrix total organic carbon (TOC) content, much of the variation in observed Kd values between different soils is eliminated.49 Distribution coefficients normalized to total organic carbon content are expressed as Koc. The following equation gives the expression relating Kd to Koc: K oc =
Kd foc
(2.19)
where Koc Kd foc
= soil sorption coefficient normalized for total organic carbon content = distribution coefficient = fraction of total organic carbon (mg organic carbon/mg soil)
In areas with high clay concentrations and low TOC concentrations, the clay minerals become the dominant sorption sites. Under these conditions, the use of Koc to compute Kd might result in underestimating the importance of sorption in retardation calculations, a source of error that will make retardation calculations based on the total organic carbon content of the aquifer matrix more conservative. In fact, aquifers that have a high enough hydraulic conductivity to spread organic chemical contamination generally have a low clay content. In these cases the contribution of sorption to mineral surfaces is generally trivial. Sorption coefficients also have been expressed on an organic matter basis (Kom) by assuming that the organic matter content of a soil or sediment equals some factor, usually between 1.7 to 1.9, times its organic carbon content on a mass basis.47,50 Often 1.724 is used as this factor, implying that the carbon content of organic matter is 1/1.724 or 60%. However, Koc is considered a more definite and less ambiguous measure than Kom.47 Assumptions inherent in the use of a Koc (or Kom) are that: sorption is exclusively to the organic component of the soil, all soil organic carbon has the same sorption capacity per unit mass, equilibrium is observed in the sorption–desorption process, and the sorption and desorption isotherms are identical.45 Both Koc and Kd have units of L/kg or cm3/g. Numerous studies have been performed using the results of batch and column tests to determine if relationships exist that are capable of predicting the sorption characteristics of a chemical based on easily measured parameters. The results of
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these studies indicate that the amount of sorption is strongly dependent on the amount of organic carbon present in the aquifer matrix and the degree of hydrophobicity exhibited by the contaminant.47 These researchers observed that the distribution coefficient, Kd, was proportional to the organic carbon fraction of the aquifer times a proportionality constant. This proportionality constant, Koc, is defined as given by Equation 2.19. Because it is normalized to organic carbon, values of Koc are dependent only on the properties of the compound (not on the type of soil). Values of Koc have been determined for a wide range of chemicals. By knowing the value of Koc for a contaminant and the fraction of organic carbon present in the aquifer, the distribution coefficient can be estimated using the relationship Kd = Koc foc
(2.20)
The fraction of soil organic carbon must be determined from site-specific data. Representative values of the fraction of organic carbon (foc ) in common sediments is available in the literature. When predicting sorption of organic compounds, total organic carbon concentrations obtained from the most transmissive aquifer zone unaffected by contamination should be averaged and used for predictions. This is because the majority of dissolved contaminant transport occurs in the most transmissive portions of the aquifer. In addition, because the most transmissive aquifer zones generally have the lowest total organic carbon concentrations, the use of this value will give a conservative prediction of contaminant sorption and retardation. Determination of the coefficient of retardation using sorption coefficients is described in Chapter 3. 2.3.1.2 Factors Affecting Sorption Coefficients Many factors potentially can affect the distribution of a contaminant between an aqueous and solid phase. These include environmental variables, such as temperature, ionic strength, dissolved organic matter concentration, and the presence of colloidal material, surfactants, and cosolvents. In addition, factors related specifically to the experimental determination of sorption coefficients, such as sorbent and solid concentrations, equilibration time, and phase separation technique, can also be important. A brief discussion of several of the more important factors affecting sorption coefficients follows. Temperature: The effect of temperature on sorption equilibrium is a direct indication of the strength of the sorption process. The weaker the interaction between sorbent and sorbate, the less the effect of temperature.47,50 While temperature can influence sorption, the strength and direction of the effect depends on the properties of the sorbent and sorbate and on the sorption mechanism. Adsorption processes are generally exothermic, so the higher the temperature, the less the adsorption. Hydrophobic sorption, however, has been shown to be relatively independent of temperature. Other reviews also indicate that the influence of temperature on equilibrium sorption and have found that, in most cases, equilibrium sorption decreases with increasing temperature.47
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pH: For neutral chemicals, sorption coefficients usually are unaffected by pH. However, for ionizable organic chemicals, sorption coefficients can be affected greatly, since pH affects not only the speciation but also the surface characteristics of natural sorbents. Typically, for weak acids the free acid form (HA) is more strongly sorbed than the anionic form (A–). For example, pentacholorophenol (PCP) sorption decreased with increasing pH over the entire pH range tested (2 to 12). For weak bases the cationic form dominates at low pH and is more highly sorbed than the free base.44 Ionic Strength: Salts can affect sorption of organic compounds by displacing cations from the soil ion exchange matrix, by changing the activity of the sorbate in solution, and by changing the charge density associated with the soil sorption surface. Salt effects are most important for basic sorbates in the cation state, where an increase in salinity can significantly lower the sorption coefficient. Salt effects are least important for neutral compounds, which may show either increases or decreases in sorption as salinity increases.44 Dissolved or Colloidal Organic Matter: The presence of dissolved or colloidal organic matter has been shown to influence sorption depending on the nature of the chemical and the organic matter. Some compounds were found to be associated extensively with the dissolved organic matter; sorption by soil decreased significantly in the presence of dissolved organic matter. Some have characterized several size fractions of water soluble organic carbon and found that the effect of dissolved organic matter on the sorption of pyrene may be limited, but the presence of colloidal organic matter suspended in the soil solution may have significant impact on the sorption of pyrene.44,51 Cosolvents: The effect of nonpolar cosolutes (trichloroethylene, toluene), polar cosolutes (1-octanol, chlorobenzene, nitrobenzene, o-cresol) and polar cosolvents (methanol and dimethyl sulfoxide) on sorption of several polycyclic aromatic hydrocarbons (PAHs) has been investigated.44,52 The nonpolar cosolutes did not significantly influence PAH sorption, while the polar cosolutes (nitrobenzene, o-cresol), having sufficiently high aqueous solubilities, caused a significant decrease in PAH sorption. Miscible organic solvents, such as methanol and ethanol, have been shown to increase solubility of hydrophobic organics and to decrease sorption. This is presumably the result of reducing the activity coefficient of the sorbate chemical in the aqueous phase, and competition for sorbing sites. Competitive Sorption: At concentrations normally encountered in environmental situations, sorption often has been observed to be relatively noncompetitive. For example, it was found that there is no competition in the sorption of binary solutes m-dichlorobenzene and 1,2,4-trichlorobenzene and between parthion and lindane.53 The sorption of methyl and dimethyl naphthalene, individually and as components of JP-8 and synthetic jet fuel mixture, on two sediments and montomorillonite clay in water was measured.54 The sorption coefficients of the naphthalenes generally varied by less than a factor of two. However, there are reports of competitive sorption taking place that is thought to be the result of site-specific sorption occurring in soil organic matter.
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Organic Matter Type and Origin: While the constancy of Koc values suggests a uniformity of organic matter with regard to sorption behavior, it is becoming increasingly apparent that organic matter type can be an important sorption variable for some sorbent/sorbate combinations. For example, it was found that the sorption of naproamide, a nonionic herbicide, was greater in the sediment than in soils, even on an organic carbon basis.44 The increased sorption in sediment was attributed to the fact that soils contained a higher percentage of cellulose and hemicellulose material, whereas the sediments contain a higher lipid-like fraction. Kinetic Considerations: Sorption generally is regarded as a rapid process and, in many laboratory sorption experiments, equilibrium often is observed within several minutes or hours. An equilibration time of 24 hours often is used for convenience. True sorption equilibrium under natural conditions, however, may require weeks to months to achieve depending on the chemical and environmental solid of interest. In many instances, an early period of rapid and extensive sorption, followed by a long slow period, is observed. Experimental determination of sorption coefficients requires preliminary kinetic experiments to determine the time to reach equilibrium. Two processes govern rate-limited or nonequilibrium sorption: transport of the substance to the sorption sites and the sorption process itself.44,50 Transportrelated nonequilibrium typically results from the existence of a heterogeneous flow domain. Sorption-related nonequilibrium, caused by rate-limited interactions between the sorbate and sorbent, may be the result of chemical nonequilibrium (i.e., chemisorption) or diffusive mass transfer limitations (i.e., diffusion of solute within pores of microporous particles or molecular diffusion into macromolecular organic matter). Sorption kinetics are likely to be environmentally important in short contact situations such as sediment resuspension, soil erosion, and infiltrating ground water.44 In general, adsorption processes tend to be rapid and nearly instantaneous, whereas nonsurface sorption tends to be slower. For neutral organic chemicals, the more hydrophobic the compound, the larger the sorption coefficient, and the longer it takes to reach equilibrium between the solid and aqueous phases. This is because the sorbent must remove a chemical from a larger volume of water. Generally, sorption estimates are based on equilibrium conditions only; however, incorporation of kinetic considerations into sorption estimation techniques is likely to be an important area of future work. For example, the assumption of equilibrium sorption in dynamic field systems may result in calculating too much pesticide in the sorbed state. Ionizability: For neutral organic compounds, in soils having a low clay/organic carbon ratio, sorption coefficients tend to increase as the hydrophobicity of the compound increases. Aqueous solubility or octanol/water partition coefficients often are used as indicators of a compound’s hydrophobicity. An increase on polarity, number of functional groups, and ionic nature of the chemical will increase the number of potential sorption mechanisms for a given chemical. For ionizable compounds, pKa is of particular importance because it determines the dominant form of a chemical at the specific environmental pH.
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The entropy change is largely due to the destruction of the highly structured water shell surrounding the solvated organic. The term “partitioning” was used to denote an uptake in which the sorbed organic chemical permeates the network of an organic medium by forces common to the solution, analogous to the extraction of an organic compound from water with an organic liquid. By either description, hydrophobic sorption or partitioning should increase as compounds become less water soluble or more hydrophobic. Additional characteristics typically associated with hydrophobic sorption or partitioning include sorption isotherms that are linear over a relatively wide range of concentrations, and sorption coefficients that are not strongly temperature dependent, and lack a competition between sorbates.44,53 2.3.2
Oxidation-Reduction Capacities of Aquifer Solids
There has been considerable research activity focused on the characterization of REDOX-potential or intensity (Eh) conditions in groundwater systems defined as the REDOX activity of dissolved chemical species. Early observations of significant Eh trends along groundwater flow paths led to hypotheses of successive REDOX zones characterized by the activity of specific thermodynamically favored electron acceptors. These REDOX zones may be classified as oxic (i.e., detectable dissolved O2), suboxic or postoxic (i.e., no detectable O2 or sulfide, detectable Fe2+), and reducing (i.e., detectable Fe2+ and sulfide, no detectable O2).1 Further investigations correctly postulated that oxidation-reduction processes were mediated by natural microbial populations that catalyze electron-transfer reactions. More recent work noted considerable temporal and spatial variability in measured subsurface REDOX conditions and that the succession of electron acceptors under oxic, suboxic, or reducing conditions was not strictly predictable by either chemical equilibrium calculations or platinum electrode measurements. 2.3.2.1 pe and pH A pH is the negative log (p for power) of proton (H+) activity and pe, its energy or work analog, is the negative log of the electron potential. An electron is not a full-fledged analog of a proton. Together, two equal but opposite charges make up a hydrogen atom, but that is about the extent of the equality between an electron and a proton. Without its proton, an electron is no longer an analog of H+, and it no longer has any claim to being part of a hydrogen atom. An electron does not bounce about by itself in the manner of an H+, and therefore it is probably not correct to try to characterize its “activity.” It always is either attached to an atom or radical or in the process of being transferred from one to another. A proton is a cation. It can replace or be replaced by other cations and it is as good as any other cation when it comes to balancing a chemical equation. Electrons receive no recognition in balanced chemical equations because the donated and accepted electrons must always cancel one another on opposite sides of an equation.
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Electrons do not have anion status. They cannot trade places with other negatively charged species. Usually we see release of H+ when metals are oxidized and consumption of H+ with their reduction. Oxidation is furthered in a subsurface environment where protons and electrons are deficient; that is, where acidity and levels of easily degraded (labile) electron donors are low. But there must be a ready supply of available electron acceptors. Reduction is favored by surpluses of both protons and electrons. This means that low pH and high availability of organic substances will promote reduction in soil. Reduction of Fe or Mn oxides, or of nitrate, uses up H+, thereby increasing pH of the soil and, theoretically, lowering the pe. Oxidation of Fe, Mn, or nitrate lowers the pH (measurable) and raises the pe (not measurable in most soils). Measuring changes in concentrations of REDOX species is more reliable for predicting these things in the subsurface than is an attempted measurement of pe with the platinum electrode. The farther apart the electrons, the more proportional work required to bring them together and the higher the respective pe. A low pe system has a surplus of electrons and, therefore, a big tendency to lose some of them and become oxidized. A high pe system is hungry for electrons. As deficient electrons are replenished, the tendency for reduction to occur will increase. If we substitute pe and pH for their defined equivalents in a generic REDOX half-reaction in which activities of oxidized and reduced species are equal, we see that the (pe + pH) sum is equivalent to the equilibrium constant of the half-reaction: Oxidized species + e– + H+ = reduced species
(2.21)
log K = log red – log ox – log e– – log H+
(2.22)
log K = pe = pH
(2.23)
If indeed their sum is constant, then, thermodynamically, pe and pH are on opposite ends of a seesaw. If behavior follows thermodynamic theory, when one goes up, the other will come down, like any sound seesaw. This sum is referred to as the REDOX parameter because, if a soil is at internal equilibrium, the (pe + pH) represents the sums of all of the REDOX equilibrium constants in the soil.1 2.3.2.2 REDOX Poise In the natural environment REDOX seesaws are not so simple. This seesaw-like behavior reflects the interaction between source/sink quantities and electron/proton intensities. If we add reducing reagents or reduced substances such as Fe(II), or Mn(II) or Cr(III) to a soil poised so that its easily reduced substances are in balance with its easily oxidized substances, some of the added reduced species will be quickly oxidized. On the other hand, adding Fe(III), Mn(IV) or Cr(VI) will result in immediate reduction of a portion of the added oxidants. There appears to be a tendency for a soil, if disturbed, to maintain a REDOX balance, that is, poise, by donating
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electrons to surplus electron acceptors or by accepting electrons from surplus electron donors.1 A soil kept near field capacity moisture with occasional mixing, double bagged inside a thin polyethylene bag for several months at 15 to 25ºC, will be close to internal equilibrium. If this metastable equilibrium is disturbed by adding an easily oxidized substance to it, e.g., glucose, the (pe +pH) of the overall system will tend to remain fairly constant as the disturbed soil system moves back toward a new metastable equilibrium. In this instance, the pe will tend to go down, and to the extent that it does, the pH will tend to rise.1 By adding increments of Cr3+ and HCrO4– , respectively, to separate subsamples of the same soil and then determining the amount of Cr reduced [loss of Cr(VI)] and the amount oxidized [gain of Cr(VI)], it is possible to find a point of poise or buffered REDOX region, where the electron donating and electron accepting tendencies cross. There the REDOX seesaw is balanced at dead-level.1 2.3.2.3 REDOX Reactions REDOX is one of those catchy phrases invented by someone unhampered by commitment to the use of scientifically correct terminology. The name is reversed (RED-OX, instead of OX-RED) for the sake of easy pronunciation. The RED stands for reduction and it signifies gain of electrons by a chemical species called electron acceptors; the OX connotes oxidation, or electron loss by a chemical species called electron donors. Oxidation-reduction (REDOX) reactions, along with hydrolysis and acid-base reactions, account for the vast majority of chemical reactions that occur in aquatic environmental systems (soils, sediments, aquifers, rivers, lakes, and many remediation operations). This section provides a survey of the environmental and substrate characteristics that govern REDOX transformations in aquatic systems. The distinction between biotic and abiotic processes is a particularly important issue in defining the scope of this section. Living organisms are responsible for creating the conditions that determine the REDOX chemistry of most aquatic environmental systems. So, in this sense, most REDOX reactions in natural systems ultimately are driven by biological activity. Once environmental conditions are established, however, many important REDOX reactions proceed without further mediation by organisms. These reactions are considered to be abiotic when it is no longer practical (or possible) to link them to any particular biological activity. Assigning Oxidation States: REDOX reactions involve oxidation and reduction; they occur by the exchange of electrons between reacting chemical species.2,55 Electrons (or electron density) are lost (or donated) in oxidation and gained (or accepted) in reduction. An oxidizing agent (or oxidant) that accepts electrons (and is thereby reduced) causes oxidation of a species. Similarly, reduction results from reaction with a reducing agent (or reductant) that donates electrons (and is oxidized). To interpret REDOX reactions in terms of electron exchange, one must account for electrons in the various reacting species. Various textbooks provide simple rules, such as the following, for assigning oxidation states for inorganic REDOX couples:2,55
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• • • • •
For free elements, each atom is assigned oxidation number 0. Monoatomic ions have an oxidation number equal to the charge of the ion. Oxygen, in most compounds, has the oxidation number –2. Hydrogen, in most compounds, has the oxidation number +1. Halogens, in most environmentally relevant compounds, have the oxidation number –1.
These rules, however, are not easily applied to organic REDOX reactions, and this difficulty has led to a steady stream of alternative concepts for assigning oxidation states. For present purposes, familiarity with a method for assigning oxidation states to organic molecules is sufficient. This method reflects the qualitative observations from which the historical concepts of oxidation and reduction originated: oxidation is the gain of oxygen (O), chlorine (Cl) or double bonds, and/or the loss of H; reduction is the gain of H, saturation of double bonds, and/or loss of O or Cl. Thus, for example, mineralization of any hydrocarbon to CO2 and H2O involves oxidation, and dechlorination of any chlorinated compound to hydrocarbon products involves reduction. Oxidations: Organic chemicals that are susceptible to oxidation and are of concern from the perspective of contamination and environmental degradation include aliphatic and aromatic hydrocarbons, alcohols, aldehydes, and ketones, phenols, polyphenols, sulfides (thiols), sulfoxides, nitriles, amines, diamines, nitrogen and sulfur hetercyclic compounds, mono- and di-chlorinated aliphatics and many others. Equations below show example half-reactions for oxidation of some of these chemical groups. Alkanes to alcohols
R – H + H2O → R – OH + 2H+ + 2e– (loss of H+ and e–)
(2.24)
Alcohols to aldehydes
R CH2 OH → RCHO + 2H+ + 2e– (loss of H+ and e–)
(2.25)
Aldehydes to acids
RCHO + H2O → RCOOH + 2H+ + 2e– (loss of H+ and e–)
(2.26)
Reductions: Most interest in reductive transformations of environmental chemicals involves dechlorination of chlorinated aliphatic and aromatic compounds and the reduction of nitroaromatic compounds. Other examples of reductive transformations that may occur abiotically in the environment include reduction of azo compounds, quinines, disulfides, and sulfoxides. An example of a half-reaction is described by the equation:
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Reductive dechlorination
R – Cl + H+ + 2e– → R – H + Cl– (gain of H+, e– and loss of Cl–) (2.27) Dechlorination can occur by several reductive pathways. The simplest results in replacement of a C-bonded halogen atom with a hydrogen and is known as hydrogenolysis or reductive dechlorination. The process is illustrated for trichloroethene, TCE, in Figure 2.12, where complete dechlorination by this pathway requires multiple hydrogenolysis steps. The relative rate of each step is a critical concern because the steps tend to become slower with each dechlorination (and DCE and VC are at least as hazardous as TCE if not more so than with VC). Aryl halogens, such as those in the pesticide chlophyrifos, also are subject to hydrogenolysis, but this reaction rarely occurs abiotically. One notable exception is the rapid abiotic dechlorination of polychlorinated biphenyls (PCBs) by zero-valent iron with catalysis by Pd.2,55 H
CI C
+H+ +2e-CI-
C
CI
H
CI
C CI
TCE
Figure 2.12
H
H +H+ +2e-CI-
C CI
CI C
cis-1,2-DCE
+H+ +2e-CI-
C
H
H
H
H C
C
H
VC
H Ethene
Reductive dechlorination or hydrogenolysis of TCE.
The other major dechlorination pathway involves elimination of two chlorines, leaving behind a pair of electrons that usually goes to form a carbon-carbon double bond. Where the pathway involves halogens on adjacent carbons, it is known as vicinal dehalogenation or reductive β-elimination. The major pathway for reductive transformation of lindane involves vicinal dehalogenation, which can proceed by steps all the way to benzene (Figure 2.13).2,55 Recently, data have shown that this pathway not only can convert alkanes to alkenes, but also can produce alkynes from dihaloalkenes (see Equation 2.28). CI
CI
CI
CI
CI
CI
CI
+2e-2CI-
CI
CI
CI
CI
+2e-2CI-
CI
Lindane
Figure 2.13
+2e-2CI-
Vicinal dechlorination or reductive-elimination of lindane.
Benzene
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Vicinal Dehalogenation
Cl – R – R1 – Cl + 2e– → R = R1 + 2Cl– (formation of double bond)
(2.28)
The contaminant REDOX reactions just summarized only occur when coupled with suitable half-reactions involving oxidants or reductants from the environment. In a particular environmental system, these REDOX agents collectively determine the nature, rate, and extent of contaminant transformation. Under favorable circumstances, the dominant REDOX agent(s) can be identified and quantified, thereby providing a rigorous basis for estimating the potential for, and rate of, transformation by abiotic REDOX reactions.2,55 Such specificity is often possible with systems engineered for contaminant remediation. However, natural systems frequently involve complex mixtures of REDOXactive substances that cannot be characterized readily. The characterization of REDOX conditions in complex environmental media is a long-standing challenge to environmental scientists that continues to be an active area of research. The remainder of this section summarizes what is currently known about the identity of oxidants and reductants relevant to environmental systems, in order to provide a basis for estimating rates of contaminant transformations by specific pathways. With respect to natural reductants, however, a great deal remains to be learned, so substantial developments can be expected as new research in this area becomes available. Oxidants: The best opportunities for predicting REDOX transformations come from engineered systems where a known oxidant is added to achieve contaminant remediation. Well-documented examples include the use of ozone and chlorine in drinking water treatment. In natural systems, important oxidants are oxides of iron and manganese, as well as molecular oxygen and various photooxidants. In engineered remediation systems oxidants used include potassium permanganate, ozone and hydrogen peroxide.1 The presence of molecular oxygen, O2 is used widely as the defining characteristic of oxidizing environments because the overwhelming supply of molecular oxygen makes it the ultimate source of oxidizing equivalents. However, O2 in its thermodynamic ground-state (3O2) is a rather poor oxidizing agent and it is not usually the oxidant directly responsible for oxidative transformations of contaminants. Instead, activated oxygen species may be involved where they are formed by the action of light on natural organic matter (NOM), peroxides, or various inorganic catalysts. Activated oxygen species include singlet oxygen (1O2), protonated superoxide (HO2· ) hydrogen peroxide and hydroperoxide anion (H2O2/HO2– ), hydroxyl radical (OH•), and ozone (O3).1,2,55 O2 + e– + H+ = H2O˚ (protonated superoxide)
(2.29)
O2 + 2e– + 2H+ = H2O2 (hydrogen peroxide)
(2.30)
O2 + 3e– + 3H+ = H2O + HO˚ (hydroxyl free radical)
(2.31)
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O2 + 4e– + 4H+ = 2H2O (water)
57
(2.32)
Equations 2.29–2.32 are half-reactions showing reduction of O2 by single electron additions. Thus, superoxide and hydroxyl, produced by one and three odd electron additions, are free radicals; whereas peroxide and water, with two and four electrons added, respectively, are not. Restricting conditions of interaction between the availabilities of soil O2 and electron donors, for example, at the interface between oxygenated water and anaerobic soil in a wetland, tends to favor transfers of electrons in single steps, and thus such interfaces are likely to be sites for free radical formation. Free radical mechanisms appear to explain why kinetically slow and seemingly unlikely REDOX transformations often occur readily at interfaces. Oxygen free radicals are much more reactive than O2 itself, and both superoxide and the hydroxyl free radical are especially reactive with H2O2, each one capable of being quickly transformed into the other. Aside from oxygen and the activated oxygen species, there are several other oxidants that cause abiotic oxidation reactions involving environmental contaminants. In engineered systems, these include chlorine, chlorine dioxide, permanganate and ferrate. At highly contaminated sites, anthropogenic oxidants such as chromate, arsenate, and selenate may react with co-contaminants such as phenols. In natural anoxic environments, the major alternative oxidants are Fe(III) and manganese (IV) oxides and hydroxides. Both are common in natural systems as crystalline or amorphous particles or coatings on other particles. In the absence of photocatalysis, however, iron and manganese oxides are weak oxidants. As a result, they appear to react at significant rates only with phenols and anilines. In the dissolved phase, few alternative abiotic oxidants are available in the neutral environment. Nitrate, sulfate, and other terminal electron acceptors used by anaerobic microorganisms are thermodynamically capable of oxidizing some organic contaminants, but it appears that these reactions almost always require microbial mediation. Reductants: Abiotic environmental reductants are not well characterized as the oxidants because, until recently, there were fewer remediation applications of reductants, and natural reducing environments are characterized by especially complex biogeochemistry. The most familiar natural reductants are sulfide (present primarily as HS– and H2S), Fe (II) and Mn (II), and natural organic matter (NOM). The transformation of contaminants by sulfur species in anaerobic environments can involve both reduction and nucleophilic substitution pathways. These processes have been studied extensively, but the complex speciation of sulfur makes routine predictions regarding these reactions difficult.1,2,55 A similar situation applies for reduced forms of iron. As with oxidations, some of the best opportunities for reliably estimating rates of redox transformations are afforded by engineered systems where a reductant of known composition and quantity is added to achieve contaminant remediation. In addition to zero-valent iron, other methods for chemical reduction of contaminants involve dithionite and electrolysis (where, in effect, electrons are added directly).1,2,55 The role of natural organic reductants in environmental systems is even more difficult to characterize than the roles of sulfur and iron because most natural organic matter is of indeterminate composition. There are two general categories of NOM:
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high molecular weight organic materials such as humic and fulvic acid, and low molecular weight compounds such as acids, alcohols, etc. Specific examples of the latter include glycolate, citrate, pyruvate, oxalate, and ascorbate. These types of compounds have been studied extensively for their role in global cycling of carbon, but very little work has been done on whether they act as specific reductants of organic contaminants.1,2,55 In contrast, the possibility that high molecular weight NOM acts as a reductant in environmental systems is widely acknowledged. Although most evidence for this involves the reduction of metal ions, several studies have shown that the process extends to various organic contaminants. Presumably, the reducing potential of NOM is due to specific moieties such as complex metals or conjugated polyphenols. Often, REDOX reactions involving these moieties are reversible, which means that NOM may serve as a mediator of REDOX reactions rather than being just an electron donor (or acceptor).1,2,55 In the recent past, the addition of labile electron donors such as molasses, lactate, and methanol is gaining ground to facilitate enhanced reductive dechlorination of chlorinated aliphatic and aromatic compounds. This technology is discussed in detail in Chapter 4. Demonstrating that a REDOX transformation of a contaminant involves mediated electron transfer requires meeting several criteria: 1) the overall reaction must be energetically favorable, 2) the mediator must have a reduction potential that lies between the bulk donor and the terminal acceptor so that both steps in the electron transfer chain will be energetically favorable, and 3) both steps in the mediated reaction must be kinetically fast relative to the direct reaction between bulk donor and terminal acceptor. Most evidence for involvement of mediators in reduction of contaminants comes from studies with model systems, because natural reducing media (such as anaerobic sediments) consist of more REDOX couples than can be characterized readily. Although this is an active area of research, a variety of likely mediator half-reactions can be identified.
REFERENCES 1. Sparks, D. L., Soil Physical Chemistry, CRC Press, Boca Raton, FL, 1998. 2. Boethling, R. S. and D. MacKay, Handbook of Property Estimation Methods for Chemicals, Lewis Publishers, Boca Raton, FL, 2000. 3. MacKay, D., W. Y. Shiu, and K. C. Ma, Henry Law Constant, in Handbook of Property Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000. 4. Leo, A. et al., Partition coefficients and their uses, Chem. Rev., 71, 525–616, 1971. 5. Leo, A. J., Hydrophobicity, the underlying property in most biochemical events, Environmental Health Chemistry, McKinney, J., Ed., Ann Arbor Science, Ann Arbor, MI, 1981, 323–336. 6. Kenage, E., Determination of bioconcentration potential, Residue Rev., 44, 73–113, 1996. 7. Neely, W. B. et al., Partition coefficients to measure bioaccumulation potential of organic chemical in fish, Environ. Sci. Technol., 8, 1113–1115, 1974.
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8. Lyman, W. J., W. F. Reehl, and D. H. Rosenblatt, Handbook of Chemical Property Estimation Methods, McGraw-Hill, New York, 1982. 9. Wolfe, N. L., and P. M. Jeffers, Hydrolysis, in Handbook of Property Estimation Methods for Chemicals, Boethling, R.S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000. 10. Wolfe, N. L., Organophosphate and organophosphorothioate esters: application of linear free energy relationships to estimate hydrolysis rate constants for use in environmental fate assessment, Chemosphere, 9, 571–579, 1980. 11. Mabey, W. R. and T. Mill, Critical review of hydrolysis of organic compounds in water under environmental conditions, J. Phys. Chem. Ref. Data, 7, 383–415, 1978. 12. Jeffers, P. M. et al., Homogeneous hydrolysis rate constants for selected methanes, ethanes, ethenes and propanes, Environ. Sci. Technol., 23, 965–969, 1989. 13. Mill, T., Photoreactions in surface waters, in Handbook of Property Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000. 14. Larson, R. A., L. L. Hunt, and D. W. Blankenship, Formation of toxic products from a No. 2 fuel oil by photooxidation, Environ. Sci. Technol., 11, 492–496, 1977. 15. Atkinson, R. J., A structure-activity relationship for the estimation of rate constants for the gas phase reactions of OH radicals with organic compounds, Int. J. Chem. Kinetics, 19, 799–828, 1987. 16. Hoag, W. R. and T. Mill, Survey of sunlight-produced transient reactants in surface waters, Proceedings of a workshop on effects of solar ultraviolet radiatiaon on geochemical dynamics, Woods Hole, MA, 1989. 16a. Atkinson, R., Atmospheric Oxidation, in Handbook of Property Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000. 17. Mopper, K. and X. Zhou, Hydroxyl radical photoproduction in the sea and its potential impact on marine processes, Science, 250, 661–664, 1990. 18. Howard, P. H., Biodegradation, in Handbook of Property Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000. 19. Alexander, M., Biodegradation and Bioremediation, Academic Press, New York, 1999. 20. Spain, J. C. and P. A. Van Weld, Adaptation of natural microbial communities to degradation of xenobiotic compounds: effects of concentration, exposure time, inoculum, and chemical structure, Appl. Environ. Microbiol., 45, 428–435, 1983. 21. Howard, P. H. and S. Banerjee, Interpreting results from biodegradability test of chemicals in water and soil, Environ. Toxicol. Chem., 3, 551–562, 1984. 22. Alexander, M., Biodegradation of organic chemicals, Environ. Sci. Technol., 19, 106–111, 1985. 23. Taylor, B. F. et al., Arch. Microbio., 122, 301–306, 1979. 24. Oldenhuis, R. et al., Appl. Environ. Microbiol., 55, 2816–2819, 1989. 25. Nelson, M. J. K. et al., Appl. Environ. Microbiol., 54, 604–606, 1988. 26. Li, S. and L. P. Wackett, Biochem. Biophy. Res. Commun., 185, 443–451, 1992. 27. Rebertson, J. B. et al., J. Appl. Environ. Microbiol., 58, 2643–2648, 1992. 28. Delgado, A. et al., J. Appl. Environ. Microbiol., 58, 415–417, 1992. 29. Shields, M. S. et al., J. Appl. Environ Microbiol., 57, 1935–1941, 1991. 30. Wackett, L. P. et al., J. Appl. Environ Microbiol., 55, 2960–2964, 1989. 31. Hyman, M. R. et al., J. Appl. Environ Microbiol., 60, 3033–3035, 1994.
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32. Van Beilen, J. B., J. Kingma, and B. Witholt, Eng. Microb. Technol., 16, 904–911, 1994. 33. Lee, K. and D. T. Gibson, J. Appl. Environ. Microbiol., 62, 3101–3106, 1996. 34. Hernandez, B. S., J. J. Arensdorf, and D. D. Focht, Biodegradation, 6, 75–82, 1995. 35. Ladd, T. I. et al., Heterotropic activity and biodegradation of labile and refractory compounds in groundwater and stream microbial population, Appl. Environ. Microbiol., 44, 321–329, 1982. 36. Neilson, A. H., Organic Chemicals, Lewis Publishers, Boca Raton, FL, 1999. 37. Alexander, M., Biodegradataion of chemicals of environmental concern, Science, 211, 132–138, 1981. 38. Klopman, G. et al., Computer-automated predictions of aerobic biodegradation of chemicals, Environ. Toxicol. Chem., 14, 395–403, 1995. 39. Punch, W. F. et al., Bess, a computerized system for predicting the biodegradation potential of new and existing chemicals, 7th Int. Workshop on QSARS in Env. Sci., June 24-28, Elsinore, Denmark, 1996. 40. Alexander, M., Nonbiodegradable and other recalcitrant molecules, Biotechnol. Bioeng., 15, 611–647, 1973. 41. Howard, P. H. et al., Review and Evaluation of Available Techniques for Determining Persistence and Routes of Degradation of Chemical Substances in the Environment, EPA-560/5-75-006, U.S. NTIS PB 243825, 1975. 42. Simkins, S. and M. Alexander, Models for mineralization kinetics with the variables of substrate concentration and population density, Appl. Environ. Microbiol., 47, 1299–1306, 1984. 43. Schmidt, S. K., S. Simkins, and M. Alexander, Models for the kinetics of biodegradation of organic compounds not supporting growth, Appl. Environ. Microbiol., 50, 323–331, 1985. 44. Doucette, W. J., Soil and Sediment Sorption Coefficients, in Handbook of Property Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000. 45. Green, R. E. and S. W. Karickoff, Sorption estimates for modeling, in Pesticides in the Soil Environment, Cheng, H. H., Ed., Soil Science Society of America, Inc., Madison, WI, 79–101, 1990. 46. Laird, D. A. et al., Adsorption of atrazine on smectites, Soil Sci. Soc. Amer. J., 56 (1), 62–67, 1992. 47. Wiedemeier T. H. et al., Natural Attenuation of Fuels and Chlorinated Solvents in the Subsurface, John Wiley & Sons, New York, 1999. 48. McCarty, P. L., M. Reinhard, and B. E. Rittmann, Trace organics in groundwater, Environ. Sci. Techn., 15, 40–51, 1981. 49. Dragun, J., The Soil Chemistry of Hazardous Materials, Hazardous Materials Control Research Institute, Silver Spring, MD, 1988. 50. Hamaker, J. W. and J. M. Thompson, Adsorption in Organic Chemicals in the Soil Environment, Goring, C. A. I. and J. W. Hamaker, Eds., Marcel Dekker, New York, 1972, 49–143. 51. Herbert, B. E. et al., Pyrene sorption by water-soluble organic carbon, Environ. Sci. Technol., 27 (2), 398–403, 1993. 52. Rao, P. S. C., L. S. Lee, and R. Pinal, Consolvency and sorption of hydrophobic organic chemicals, Environ. Sci. Technol., 24 (5), 647–654, 1990. 53. Chiou, C. T. and T. D. Shoup, Soil sorption of organic vapors and effects of humidity on sorption mechanism and capacity, Environ. Sci. Technol., 19, 1196–1200, 1985.
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54. MacIntyre, W. G., T. B. Stauffer, and C. P. Antworth, A comparison of sorption coefficients determined by batch, column, and box methods on a low organic carbon acquifer material, Ground Water, 29 (6), 908–913, 1991. 55. Tratnyek, P. G. and D. L. Macalady, Oxidation-reduction reactions in the aquatic environment, Handbook of Property Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000.
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CHAPTER
3
Monitored Natural Attenuation CONTENTS 3.1
Introduction ....................................................................................................64 3.1.1 Definitions of Natural Attenuation ....................................................64 3.2 Approaches for Evaluating Natural Attenuation ...........................................65 3.3 Patterns vs. Protocols .....................................................................................70 3.3.1 Protocols for Natural Attenuation......................................................70 3.3.2 Patterns of Natural Attenuation .........................................................71 3.3.2.1 Various Patterns of Natural Attenuation.............................72 3.4 Processes Affecting Natural Attenuation of Compounds..............................79 3.4.1 Movement of Contaminants in the Subsurface .................................79 3.4.1.1 Dilution (Recharge) ............................................................79 3.4.1.2 Advection ............................................................................81 3.4.1.3 Dispersion ...........................................................................83 3.4.2 Phase Transfers ..................................................................................85 3.4.2.1 Sorption...............................................................................85 3.4.2.2 Stabilization ........................................................................88 3.4.2.3 Volatilization .......................................................................89 3.4.3 Transformation Mechanisms..............................................................89 3.4.3.1 Biodegradation ....................................................................90 3.5 Monitoring and Sampling of Natural Attenuation ......................................109 3.5.1 Dissolved Oxygen (DO) ..................................................................113 3.5.2 Oxidation–Reduction (REDOX) Potential (ORP)...........................117 3.5.3 pH .....................................................................................................119 3.5.4 Filtered vs. Unfiltered Samples for Metals .....................................120 3.5.4.1 Field Filtration and the Nature of Groundwater Particulates..................................................121 3.5.4.2 Reasons for Field Filtration..............................................122 3.5.5 Low-Flow Sampling as a Paradigm for Filtration ..........................124 3.5.6 A Comparison Study........................................................................125 References..............................................................................................................126 63
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…natural attenuation (NA) is not a “no action (NA)” alternative. Monitored natural Attenuation (MNA) defines the required monitoring parameters to demonstrate that the ongoing natural processes will continue to meet the remediation objectives…
3.1
INTRODUCTION
The term monitored natural attenuation (MNA) refers to an approach to clean up subsurface contamination, specifically in groundwater, by relying on natural processes and monitoring. MNA is also referred to as natural degradation and intrinsic or passive remediation. Natural attenuation processes include a variety of physical, chemical, or biological processes that, under favorable conditions, act without human intervention to reduce the mass, toxicity, mobility, volume, and concentration of contaminants in groundwater. Depending on the geologic conditions, types of contaminants, and contaminant mass and distribution at a given contaminated site, MNA could emerge as the preferred choice of remediation approach. Natural attenuation relies on the assimilative capacity of the ecosystem for the reduction of contaminant concentration and mass. This approach has been utilized by environmental engineers for a long time to control industrial and municipal wastewater discharges into surface waterbodies and maintain acceptable water quality standards. 3.1.1
Definitions of Natural Attenuation
A variety of organizations have espoused the following definitions of natural attenuation due to the emerging popularity and preference of MNA as the remediation method of choice at many contaminated sites across the country.1 Environmental Protection Agency2: This policy directive defines monitored natural attenuation as the reliance on natural attenuation process (within the context of a carefully controlled and monitored site cleanup approach) to achieve site-specific remediation objectives within a time frame that is reasonable compared to that offered by other more active methods. The “natural attenuation processes” that are at work in such a remediation approach include a variety of physical, chemical, or biological processes that, under favorable conditions, act without human intervention to reduce the mass, toxicity, mobility, volume, or concentration of contaminants in soil or groundwater. These in situ processes include biodegradation; dispersion; dilution; sorption; volatilization; radioactive decay; and chemical or biological stabilization, transformation, or destruction of contaminants. American Society for Testing and Materials (ASTM)3: Its document titled Standard Guide for Remediation of Groundwater by Natural Attenuation at Petroleum Release Sites defines natural attenuation as the “reduction in mass or concentration of a compound in groundwater over time or distance from the source of constituents of concern due to naturally occurring physical, chemical, and biological processes, such as biodegradation, dispersion, dilution, adsorption, and volatilization.”
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Air Force4: The first document, published in 1995, defines the process as resulting “from the integration of several subsurface attenuation mechanisms that are classified as either destructive or nondestructive. Biodegradation is the most important destructive attenuation mechanism. Nondestructive attenuation mechanisms include sorption, dispersion, dilution from recharge, and volatilization.” Army5: Its report defines natural attenuation as “the process by which contamination in groundwater, soils, and surface water is reduced over time…[via] natural processes such as advection, dispersion, diffusion, volatilization, abiotic and biotic transformation, sorption/desorption, ion exchange, complexation, and plant and animal uptake.”
In the past, the first question to be asked in consideration of the potential for natural attenuation at a contaminated site was whether biodegradation of the chemical contaminant had been reported. Oftentimes the question was, “Does the biogeochemistry exist for ongoing degradation?” due to the assumption that the responsible microorganisms are ubiquitous in the subsurface. However, in this chapter the term “natural attenuation” will include all the processes that contribute towards the decrease in contaminant concentrations.
3.2
APPROACHES FOR EVALUATING NATURAL ATTENUATION
Documenting that contaminant concentration has become very low or detectable in groundwater samples is an important piece of evidence that natural attenuation is working. However, such documentation is not completely sufficient to show that natural attenuation is protecting human health and the environment, for three primary reasons: • Monitoring of contaminant concentration reductions is not always precise due to the complex nature of groundwater systems. In some cases the total contaminant mass may have decreased, but the contaminant may have transformed to another, more hazardous chemical form. • In a few instances reactions that initially cause contaminants to attenuate may not be sustainable until reasonable cleanup goals are achieved. • Another situation of concern occurs when natural biogeochemical parameters, such as electron acceptors and electron donors that support attenuation, are used up before the treatment of contamination is complete.
For these reasons, environmental regulators and others should not rely on simple rules of thumb (such as maximum contaminant concentration data or trends in these data over a relatively short time) in evaluating the potential success of natural attenuation. The decision to rely on natural attenuation and the confirmation that it will continue to work depend on linking monitoring data to a site conceptual model and “footprints” of the underlying mechanisms. Footprints are mappings of concentration changes in reactants (contaminant(s), electron acceptors, and donors) or products of the biogeochemical processes (such as Cl– ion, dissolved Fe2+) that degrade or
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immobilize the contaminants (Figures 3.1a, b, and c). Footprints can be measured to document that these transformation or immobilization processes are active at the site. An observation of the loss of a contaminant, coupled to observation of a few footprints, helps to establish which processes are responsible for the decrease in contaminant mass and concentrations. The three basic steps to document natural attenuation are as follows: 1. Develop a conceptual model of the site: The model should show where and how fast the groundwater flows, where the contaminants are located and at what concentrations, and which types of natural processes could theoretically affect the contaminants (Figures 3.2a and b). 2. Analyze site measurements: Samples of groundwater should be analyzed chemically to look for footprints of the natural attenuation processes and to determine whether these processes are sufficient to control the contamination. 3. Monitor the site: The site should be monitored until regulatory requirements are achieved to ensure that documented attenuation processes continue to occur.
Figures 3.1a
Initial vinyl chloride plume at a landfill site in Maryland with radial groundwater flow from the center of the landfill.
Although the basic steps are the same for all sites, the level of effort needed to carry out these steps varies substantially with the complexity of the site. When site characteristics or the controlling mechanisms are uncertain, it will be difficult to develop the site conceptual model; thus, a large amount of data will be required to document natural attenuation. In these complex situations, computer modeling may be necessary, and data on footprints and site characteristics will have to be more than adequate to develop the model.
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1200 1000
67
Three-dimensional perspective plot of observed vinyl choride concentrations in groundwater -1996
800 600
500
400
Landfill boundary
200
200 150
0
100 20 5 1 0
Figures 3.1b
Natural attenuation effects on the vinyl chloride plume. Note: The significant reduction in vinyl chloride concentration and mass due to natural attenuation.
Dissolved Oxygen Redox Fe 2+ Manganese Vinyl Chloride Landfill
Saprolite
Sand/Gravel
Bedrock
Dissolved Oxygen, Redox, and Vinyl Chloride Distribution Figures 3.1c
Effects of the primary electron acceptor dissolved oxygen on the attenuation of VC and Mn along a North-South transect through the middle of the landfill.
Figures 3.2a
Abandoned Well?
A general site conceptual exposure model (adapted from ASTM, 1997).
Confining Layers?
Confining Layers?
Future Domestic Water Supply Well
Current Municipal Water Supply Well
Confining Layers?
Shallow Water Table
Current Domestic Water Supply Well
68
Dissolved Plume
Residual NAPL
Utilities
Utilities
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Figures 3.2b
Dissolved Groundwater Plume
Dissolved Groundwater Plume
Mobile NAPL Migration Stormwater/ Surface Water Transport
Non-Aqueous Phase Liquid (NAPL) Affected Surface Soils, Sediments or Surface Water
Leaching and Groundwater Transport
Volatilization and Atmospheric Dispersion
Wind Erosion and Atmospheric Dispersion
Transport Mechanisms
Affected Subsurface Soils (>3 ft depth)
Affected Surface Soils (<3 ft depth)
Secondary Sources
Site conceptual exposure models.
Chemical Storage Piping / Distribution Operations Waste Management Unit Soil or Waste Piles Lagoons or Ponds Other
Primary Sources
Recreational Use/ Relevant Habitat
SURFACE WATER
Potable Water Use
GROUNDWATER
Inhalation of Vapor or Particulates
AIR
Dermal Contact or Ingestion
SOIL
Exposure Routes
Residential Commercial/Industrial Relevant Ecological Receptor
Residential Commercial/Industrial
Residential Commercial/Industrial Construction Worker Relevant Ecological Receptor
Residential Commercial/Industrial Construction Worker Relevant Ecological Receptor
Receptors
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3.3 3.3.1
PATTERNS VS. PROTOCOLS
Protocols for Natural Attenuation
Within the past few years, many organizations have issued documents providing guidance on evaluating natural attenuation.1 Among the 14 documents developed by a range of organizations from federal and state agencies to private companies and industry associations, the available technical protocols address two classes of organic contaminants only: fuel hydrocarbons and chlorinated solvents (with the exception of the Department of Energy (DOE) document). A large body of empirical evidence and scientific and engineering studies in recent years has been developed to support understanding of natural attenuation of these contaminants — mostly fuel hydrocarbons under certain conditions. However, the natural attenuation of polycyclic aromatic hydrocarbons, polychlorinated biphenyls, explosives, and other classes of persistent organic contaminants is not addressed in any protocol.1 Furthermore, although the DOE document proposes a method for assessing natural attenuation processes for inorganic contaminants such as metals, such processes are extremely complex, and this document does not adequately reflect this complexity.6 A recent effort was made to compare the guidelines currently available on natural attenuation against a list of characteristics of a comprehensive protocol.1 The consensus was that a comprehensive protocol should cover three broad areas: • Community concerns: The protocol should describe a plan for involving the affected community in decision making, maintaining institutional controls to restrict use of the site until cleanup goals are achieved, and implementing contingency measures if natural attenuation fails to continue as expected. • Scientific and technical issues: The protocol should describe how to document which natural attenuation processes are responsible for observed decreases in contaminant concentrations, how to assess the site for contaminant source and hydrogeologic characteristics that affect natural attenuation, and how to assess the sustainability of natural attenuation over the long term. • Implementation issues: The protocol should be easy to follow and should describe the monitoring frequency and various monitoring procedures, in addition to the training and expertise required for the personnel carrying out the field implementation.
None of the current documents fulfills all the criteria defined above.1 To some extent, this reflects the various, and sometimes limited, purposes for which these documents were prepared. Some are detailed technical guides; others are intended to help ensure consistency in site evaluation within a particular organization (such as a private corporation or a branch of the military), and others are intended to guide policy. Nonetheless, key gaps in the existing body of protocols have to be addressed. The existing protocols provide little or no discussion of when and how to involve the public in site decisions and when and how to implement institutional controls. In the few instances where these matters are mentioned, the discussion is typically brief, almost in passing. Although most environmental regulatory agencies have separate policies that specify procedures for community involvement and
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institutional controls, these procedures may be inadequate in cases where natural attenuation is selected as the remedy. Discussion of when and how to implement contingency plans in case natural attenuation does not work is also inadequate in many of the protocols. Further, the protocols do not provide sufficient guidance on when and how engineered methods to remove or contain sources of contamination benefit natural attenuation. A major shortcoming of some of the protocols relates to scoring systems used for initial screening to determine whether a site has potential for treatment by natural attenuation. Such scoring systems yield a numeric value for the site in question. If this value is above a certain level, the site is judged an eligible candidate for natural attenuation. Frequently, such scores are used inappropriately as the key factor in deciding whether natural attenuation can be a successful remedy at the site. Moreover, these scores often lead to erroneous conclusions about whether natural attenuation will or will not succeed, due to the complexity of the processes involved and the tendency of scoring systems to oversimplify them. In addition, the scoring systems developed for evaluating natural attenuation at petroleum sites are erroneously used to evaluate sites with chlorinated solvents by many practitioners of remediation. In summary, the existing body of natural attenuation protocols is limited in several important areas.1 Where and how existing protocols can be used to meet regulatory requirements for documenting site cleanup — and whether such protocols are required at all — is also unclear. Guidance on the use of natural attenuation for remediation has to be developed to cover topics not addressed in existing protocols and to provide for the use of protocols in regulatory programs. 3.3.2
Patterns of Natural Attenuation
Instead of relying on protocols and scoring systems, an educated screening tool should be to observe the patterns in reduction of contaminant concentrations. Naturally attenuating contaminant plumes can take a variety of forms: they might be expanding, stable, or shrinking, depending on the trends in the spatial variations of contaminant concentrations with time (Figures 3.3a, b, and c). Common patterns in all attenuating plumes are a decline in the dissolved contaminant mass with time, and a decline in contaminant concentrations downgradient from the source. Once these patterns are observed initially, the following list of questions should be developed to collect additional data to develop a platform demonstrating that MNA is an ongoing and continuing process to meet the site cleanup objectives: • What chemical, physical, and biological processes are in effect to support natural degradation of the site-specific contaminants? • What site biogeochemical conditions are needed for these chemical, physical, and biological processes to work? Which types of site conditions are optimal? Which conditions inhibit natural attenuation? • What level of information is needed to characterize the site fully? • What breakdown products that may be more toxic, persistent, or mobile are created when the contaminants degrade? How does one prove that contaminants are degrading into harmless substances?
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Contaminated Zone
Monitoring Well
Cross Sectional View
MW-2
MW-1
MW-3 t0 t1
t2
Plan View
"Contaminant plume is continuing to grow and move downgradient from the source area"
Figures 3.3a
Expanding plume.
• What kinds of specific monitoring and testing are needed to determine that the site and the contaminants are suitable for natural attenuation? Is extensive monitoring necessary? • How long is it reasonable to monitor to ensure that natural attenuation is working? • How viable are institutional controls? Can they be enforced? • Is stabilization by natural attenuation irreversible for metals or other substances?
3.3.2.1 Various Patterns of Natural Attenuation Removal of Contaminant Sources: At most contaminated sites, the bulk of the contaminant mass is in what remediation professionals call “source zones.” Examples of source zones include landfills, areas of chemical spills, buried tanks that contain residual chemicals, deposits of tars, etc. Some of these sources can be easily located
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Contaminated Zone
Monitoring Well
Cross Sectional View
t2 MW-3
MW-2
t0
t1
MW-1
Plan View
Contaminant plume is almost stationary over time and concentrations at points within the plume are relatively constant over time with a slight declining trend.
Figures 3.3b
Stable groundwater plume.
and complete or partial removal or containment may be possible. However, other common types of sources often are extremely difficult to locate and remove or contain. One example of a source in this category is chemicals that have sorbed to soil particles but have the potential to dissolve later into groundwater that contacts the soil. Another extremely important example is the class of organic contaminants known as “nonaqueous-phase liquids” (NAPLs). There are two types of NAPLs: those that are more dense than water (dense nonaqueous-phase liquids, or DNAPLs), and those that are less dense than water (light nonaqueous-phase liquids, or LNAPLs). When released to the ground, these types of fluids move through the subsurface in a pattern that varies significantly from that of the water flow because NAPLs have different physical properties than water. As shown in Figure 3.4a, b, and c, LNAPLs can accumulate near the water table, DNAPLs can penetrate the water table and form pools along geologic layers, and both types of NAPLs can become entrapped in soil pores. These NAPL accumulations contaminate the
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Contaminated Zone
Monitoring Well
Cross Sectional View
MW-2
MW-1
MW-3 t2 t1 Plan View
t0
Contaminant plume is receding back toward the source area over time and the concentrations at points within the plume are declining over time.
Figures 3.3c
Shrinking groundwater plume.
groundwater that flows by them by slow dissolution. Common LNAPLs include fuels (gasoline, kerosene, and jet fuel) and common DNAPLs include industrial solvents (trichloroethene, tetrachloroethene, and carbon tetrachloride) and coal tar. Once they have migrated into the subsurface, NAPLs are often difficult or impossible to locate in their entirety. Normally, the total mass of a contaminant within source zones is significantly larger compared to the mass dissolved in the plume. Therefore, the source usually persists for a very long time. The rate at which contaminants dissolve from a typical NAPL pool is so slow that many decades to centuries may be needed to dissolve the NAPL completely by dissolution without any intervention. The potential for success of natural attenuation of various dissolved organic and inorganic compounds is presented in Table 3.1. Given the persistent nature of contaminant sources, removing them would seem like a practical way to speed natural attenuation of the contaminant plume (Figure 3.4). In many cases, environmental regulators require source removal or containment
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Table 3.1 The Potential for Success of Natural Attenuation for Various Compounds (adapted from NCR, 2000) Contaminant Type
Dominant Attenuation Processes
Likelihood of Success
Organic Hydrocarbons BTEX Gasoline, fuel oil Nonvolatile aliphatic compounds PAHs Creosote
Biotransformation Biotransformation Biotransformation, immobilization Biotransformation, immobilization Biotransformation, immobilization
High Moderate Low Low Low
Biotransformation
High
Biotransformation
Moderate
Methylene chloride Vinyl chloride (VC)
Biotransformation Biotransformation, abiotic transformation Biotransformation Biotransformation
Dichloroethene (DCE)
Biotransformation
Moderate Moderate to High High Moderate to High Moderate
Biotransformation, immobilization
Low
Biotransformation Biotransformation
Low High
Biotransformation, abiotic transformation, immobilization
Low
Oxygenated Hydrocarbons Low-molecular-weight Alcohols, ketones, esters MtBE Chlorinated Aliphatics PCE, TCE, carbon tetrachloride Trichloroethane (TCA)
Chlorinated Aromatics Highly chlorinated PCBs, pentachlorophenol, multichlorinated benzenes Less chlorinated PCBs, dioxins Monochlorobenzene Nitroaromatics TNT, RDX
Inorganic Metals Ni Cu, Zn Cd Pb Cr
Immobilization Immobilization Immobilization Immobilization Biotransformation, immobilization
Hg
Biotransformation, immobilization
Moderate Moderate Low Moderate Low to Moderate Low
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Table 3.1 The Potential for Success of Natural Attenuation for Various Compounds (adapted from NCR, 2000) (continued) Contaminant Type
Dominant Attenuation Processes
Likelihood of Success
Biotransformation, immobilization Biotransformation, immobilization
Low Low
Biotransformation Biotransformation
Moderate Moderate
Nonmetals As Se Oxyanions Nitrate Perchlorate
Not Enough Mass Spilled to Form an NAPL a)
t4
t1 t2
t3
LNAPL Release
b)
t1
MNAPL t2
MDISS
t3
t4 MN > M D
DNAPL Release
c)
t1
t2
Adsorbed DNAPL
MNAPL
DNAPL Pool
Figures 3.4
t4
t3 MDISS
MN > M D
Various possibilities of source zone contamination.
as part of a natural attenuation remedy. Although requiring source control or removal is good policy for many sites, expert opinions conflict on whether source removal is advisable when using natural attenuation as a remedy, even when such removal is technically feasible.
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Goals of source removal should be the following: • Remove as much contaminant mass as practical to reduce the mass flux of contaminants emanating from the source zone, thus reducing the concentration of the contaminant plume rapidly and also reducing the longevity of the required monitoring period; and • Avoid any changes that would reduce the effectiveness of natural attenuation, such as disturbing the natural dissolution equilibrium from an NAPL source by drilling through it and thus increasing the mass flux.
In theory, if one can delineate the source completely and succeed in removing most of the mass, then a significant benefit may be achieved. There are many case studies available in the literature even for compounds like polycyclic aromatic hydrocarbons (PAHs) plumes in which it appears that, after removal of the source, the plumes attenuated rapidly. However encouraging this example might be, this kind of success may not always be realized. Particularly, DNAPL sources in fractured bedrock environments cannot be delineated completely and/or cannot be removed to any significant degree at a reasonable cost. Hence, source removal options may be rejected because none are anticipated to be able to warrant the expense and risks of the removal effort by removing all of the source mass without leaving a significant level of residual mass. In some cases, source removal efforts may directly and adversely affect natural attenuation. Most of the negative impacts will be caused mainly by the disturbance of the equilibrium between the moving groundwater and the quiescent mass of NAPL, particularly DNAPL. As a precautionary measure, an outside-in approach to investigating the source zone is recommended in contrast to an inside-out approach. Consideration should be given when looking at removal of the source of one type of contaminant which may adversely affect natural attenuation of another type and thus result in minimal or no overall benefit. A good example is the removal of a petroleum hydrocarbon source zone serving as a nutrition source for microbes involved in degrading a chlorinated solvent plume. Such an action could slow down or completely shut off natural attenuation of the chlorinated solvent. Natural Attenuation Capacity (NAC): The manner in which natural attenuation and active remediation measures (such as source removal, pump and treat, chemical oxidation, or enhanced bioremediation) are combined depends on the natural attenuation capacity (NAC) of the system. If the NAC is small, for example, active remediation measures will need to remove or degrade a high proportion of the contaminant source to protect downgradient receptors. Conversely, if the NAC is large, less source removal may be required to protect downgradient receptors. In either case, it is necessary to quantify the NAC of the biogeochemical system to combine contaminant source-removal methods with natural attenuation effectively. Natural attenuation capacity is a concept that refers to the capacity of a biogeochemical system to lower contaminant concentrations along aquifer flow paths. The NAC of groundwater systems depends on hydrogeologic (dispersion and advection) and biological (biodegradation rates) factors for organic contaminants and precipitation potential also for heavy metals.
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The concept of NAC is useful because it illustrates those characteristics and parameters of a groundwater system that affect the efficiency of natural attenuation.7 For example, if the biodegradation rate constant is small (≅ 0.001 d–1) relative to the groundwater velocity (~3 ft/day) and aquifer dispersivity (30 feet), the NAC of the system also will be small. Because of this small NAC, contaminants will be transported relatively long distances downgradient of the source area (Figure 3.5a). Conversely, if the biodegradation rate is high relative to groundwater velocity and aquifer dispersivity, the NAC will be proportionally higher, and the transport of contaminants will be restricted closer to the source area.
Very Low NAC
Concentration
Moderate NAC
High NAC
Distance Along Flow Path Figure 3.5a
The effect of natural attenuation capacity on contaminant transport.7
Quantitative mathematical techniques in addition to empirical methods are available to estimate NAC. In addition to NAC, the distance that contaminants are transported in a groundwater system also depends on the contaminant concentrations at the source area (Figure 3.5b). High concentration, low NAC High concentration, higher NAC
Concentration
Lower concentration, low NAC Lower concentration, higher NAC
Cleanup Standards Distance Along Flow Path Figure 3.5b
The effect of source area concentrations on the distance required to reach cleanup standards.
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3.4.1
79
PROCESSES AFFECTING NATURAL ATTENUATION OF COMPOUNDS
Movement of Contaminants in the Subsurface
Even in the absence of biotic and/or abiotic transformations of a contaminant, the contaminant always is subject to transport processes — meaning that physical processes cause it to move. All important transport processes for subsurface contaminants can be categorized as dilution, advection, dispersion, or “phase transfer” (from one type of physical medium to another, such as from an NAPL to groundwater or from water to the soil matrix). 3.4.1.1 Dilution (Recharge) Recharge is the amount of water entering the saturated zone of the water table at the water table surface, made available mainly by precipitation events. In recharge areas, flow near the water table is generally downward. Recharge defined in this manner may therefore include not only precipitation that infiltrates through the vadose zone, but also water entering the groundwater system via discharge from surface water bodies. Where a surface water body is in contact with or is part of the groundwater system, the definition of recharge is stretched slightly. However, such bodies often are referred to as recharging lakes or streams.8 The recharge of the water table aquifer has two effects on the natural attenuation of a dissolved contaminant plume: 1) additional water entering the system due to infiltration of precipitation or from surface water will contribute to dilution of the plume and 2) the influx of relatively fresh, electron-acceptor-charged water will alter the geochemical processes and in some cases, facilitate additional biodegradation.8,9 Recharge from infiltrating precipitation is the result of a complex series of processes in the unsaturated zone. Description of these processes is beyond the scope of this chapter; however, it is worth noting that the infiltration of precipitation through the vadose zone brings the water into contact with the soil and thus may allow the introduction of electron acceptors (such as NO3– and SO42– ) in addition to the DO in the recharge water and also dissolved organic carbon (electron donor). Infiltration therefore provides fluxes of water, inorganic species, and possibly organic species into the groundwater. In the case of surface water it may be connected as part of the groundwater system, or it may be perched above the water table. In either case, the water entering the groundwater system will not only aid in dilution of a contaminant plume, but it may also add electron acceptors and possible electron donors to the groundwater. An influx of electron acceptors will tend to increase the overall assimilation capacity of the groundwater system. In addition to the introduction of electron acceptors that may be dissolved in the recharge (e.g., dissolved oxygen, nitrate, or sulfate), the infiltrating water may also foster biogeochemical changes in the aquifer. For example, Fe2+ will be oxidized back to Fe3+ and will be precipitated out. This reprecipitation of Fe3+ could be again available for reduction by microorganisms. Such a shift may be beneficial for biodegradation of contaminants utilized as electron
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donors, such as fuel hydrocarbons or vinyl chloride. However, these shifts can also make conditions less favorable for reductive dechlorination. Evaluating the effects of recharge can be difficult. The effects of dilution might be estimated if one has a detailed water budget for the system in question. However, if a plume has a significant vertical extent, it cannot be known with any certainty what proportion of the plume mass is being diluted by the recharge. In addition, separating the effects of dilution from other processes of mass reduction may be difficult. After recharge, the effects of the addition of electron acceptors may be apparent due to elevated electron acceptor concentrations, differing patterns in electron acceptor consumption, or by-product formation in the area of recharge. However, the effects of short-term variations in such a system (which are likely due to the intermittent nature of precipitation events in most climates) may not easily be quantified. Where recharge is from surface water, the influx of mass and electron acceptors is more steady over time. In this scenario, quantifying the effects of dilution may be less uncertain, and the effects of electron acceptor replenishment may be more easily identified (although not necessarily quantified). In some cases the effects of recharge-diluting contaminant plumes can be estimated with a simple relationship based on the specific discharge of groundwater passing through the point of interest and the amount of recharge entering the plume area. It is imiportant to note that at most sites, recharge will not actually mix with groundwater in an aquifer but will form a stratified layer on top due to the very low amount of vertical dispersion characteristic of aquifer systems. Mixing can be assumed in some cases, such as a very thin, unconfined aquifer: the aquifer discharges into a surface water body, and the groundwater associated with the recharge is assumed to be mixed with the original groundwater flowing past a source zone.8-10 The relationship for estimating the amount of dilution caused by recharge is Ê RWL VD ˆ C L = C 0 expÁ ˜ Ë WTh VD ¯
(3.1)
Eliminating the width and rearranging gives: Ê RL ˆ C L = C 0 expÁ Á T (V )2 ˜˜ Ë ¯ D h
(3.2)
where CL C0 R W L
= concentration at distance L from origin assuming complete mixing of recharge with groundwater (mg/L) = concentration at origin or at distance L = 0 (mg/L) = recharge mixing with groundwater (ft/yr) = width of area where recharge is mixing with groundwater (ft) = length of area where recharge is mixing with groundwater (ft)
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VD Th
81
= Darcy velocity of groundwater (ft/yr) = thickness of aquifer where groundwater flow is assumed to mix completely with recharge (ft)
3.4.1.2 Advection Transport of a contaminant molecule occurring with the groundwater movement is called advection or convection or bulk flow. Advection occurs in any moving fluid. Thus, contaminants can advect when they are in air in soil pores or in a moving NAPL, as well as in water. Advection transport is illustrated simply by considering a contaminant that does not react biotically or abiotically (also known as conservative compound or tracer) in the subsurface and that moves at the average velocity of the groundwater. Figures 3.6a and b describe this phenomenon. The contaminant moves at exactly the same velocity as the water and does not change from its initial concentration of C0′ at the injection point.9
time= t0
Concentration (C)
Q
t
time= t1
time= t2
t1
t2
Distance (x)
Figure 3.6a
Dispersion of a pulse of a tracer substance in a sand column experiment.
1.0
Initial Contaminant Slug
Advection Only
Advection Only
Advection, Dispersion, and Sorption C/
C 0.5 O
Advection and Dispersion
0
Distance From Source
Figure 3.6b
Concentration curves showing plug flow with an instantaneous source from advection only and from a combination of advection, dispersion, and sorption.
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The mass flux rate at which a dissolved contaminant moves across a vertical plane in the subsurface is the product of the contaminant concentration and the velocity of groundwater. Groundwater velocity is governed by three key factors specific to each site: • The hydraulic gradient includes gravity and pressure components and is the driving force for water movement. Water always moves in the direction of higher hydraulic head (which can be thought of qualitatively as elevation) to lower head. • Hydraulic conductivity is the ability of porous rocks or soil sediments to transmit fluids and is measured from field tests or samples. Hydraulic conductivity values for common rocks and sediments vary over ten orders of magnitude from almost impermeable crystalline rocks to highly permeable gravels; the hydraulic conductivity values for fractured rocks, sand, and clay are between these extremes. A contaminant plume moving with the groundwater will travel faster through sand layers, which have high hydraulic conductivity, than through clays of low hydraulic conductivity, under the same hydraulic head gradient. • Porosity is a measure of the volume of open spaces in the subsurfaces relative to the total volume. Like hydraulic conductivity, it depends on the type of geologic material present and can be determined from field tests or samples.
The equation for describing the rate of groundwater flow from one location to another is known as Darcy’s equation: VD = − K H
∆h ∆X
(3.3)
where KH ∆h ∆X VD
= hydraulic conductivity (in units of distance per time) = hydraulic gradient = Darcy velocity (in units of distance per time)
To determine the seepage velocity of a contaminant that travels at the same speed as the groundwater, the Darcy velocity must be divided by the effective porosity ε: V=
VD ε
(3.4)
KH and ε can be estimated using various field test methods or laboratory evaluations of cores taken from the subsurface. Uncertainty is inherent in all such measurements, and this uncertainty must be acknowledged by developing a range of possible flow scenarios.
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3.4.1.3 Dispersion Spreading of contaminants from the main direction of groundwater flow takes place as the groundwater moves, altering concentrations from those that would occur if advection were the only transport mechanism. This mixing is called hydrodynamic dispersion. The mechanisms causing dispersion within the plume include molecular diffusion, different water velocities within individual pores, different water velocities between adjacent pores, and tortuosity of the subsurface flow path (Figure 3.7). Mixing caused by local variations in velocity is also known as mechanical dispersion. Groundwater scientists quantify the combined mixing effect using a hydrodynamic dispersion coefficient DH. Except at very low water velocities, DH increases linearly with the average speed of groundwater. A' A B' Average Water Flow Direction C' B C
Figure 3.7
Seemingly random variations in the velocity of different parcels of groundwater are caused by the tortuous and variable route the water must follow.
The curve labeled “dispersion” in Figure 3.6 a and b illustrates the effects of dispersion for a conservative contaminant that travels precisely with the water molecules. The solute is detected at the observation well before it would be if advection were the only process affecting its movement. Dispersion causes the solute to spread, rather than moving as an unchanged “plug.” Molecular Diffusion: Molecular diffusion takes place as a result of the contaminant gradients created within the zones of contamination. It is significant only when the groundwater velocities are low, and the diffusive flux of a dissolved contaminant, at steady state, can be described by Fick’s first law. F = −D
dc dx
where F D C
= mass flux of solute per unit area of time = diffusion coefficient = solute concentration
dc dx
= concentration gradient
(3.5)
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For systems where the dissolved contaminant concentrations are changing with time, Fick’s second law must be applied. The one-dimensional expression of Fick’s second law is d2 dc = D c2 dt dx
(3.6)
dc is the change in concentration with time. dt The process of diffusion is slower in porous media than in open water because the contaminant molecules must follow more tortuous flow paths. To account for this, an effective diffusion coefficient D* is used. Fetter estimates a range of 1 × 10–9 to 2 × 10–9 m2/S for D* has been estimated.9(a) The effective diffusion coefficient is expressed quantitatively as where,
D* = wD
(3.7)
where w is the empirical coefficient determined by laboratory experiments. The value of w ranges greatly from 0.01 to 0.5.9 Mechanical Dispersion: Mechanical dispersion occurs due to variations in flow velocity because of varying pore throat sizes and tortuosity caused by variations in flow path lengths. An additional cause of mechanical dispersion is variable friction within an individual pore, thus allowing the groundwater flowing in the center of the pore to move faster than groundwater flowing next to the soil particle itself. The component of hydrodynamic dispersion contributed by mechanical dispersion can be described as: mechanical dispersion = ∝x V
(3.8)
where ∝x V
= dispersivitiy = seepage velocity
Advection dispersion equation: The advection-dispersion equation, which includes hydrodynamic dispersion, can be described as:8,9 ∂c ∂2c ∂c = DH −V ∂t Ox 2 ∂x where c t DH x V
= = = = =
contaminant concentration time hydrodynamic dispersion distance along flow path seepage velocity
(3.9)
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85
Phase Transfers
Contaminants will be added or removed from the groundwater when they transfer between phases. The relevant phases in the subsurface are groundwater (dissolved), soil grains (adsorbed), NAPLs (liquid), and soil gas (air) in the vadose zone. Phase transfers can increase or decrease the contaminant concentration within the groundwater plume, depending on the transfer mechanism, the contaminant, and the geochemistry. Although the basic concepts of phase transfer are straightforward, quantification of these transfers often is not easy. 3.4.2.1 Sorption Many contaminants, including chlorinated solvents, BTEX and dissolved metals, are removed from solution by sorption onto the aquifer matrix, thus slowing the movement of contaminants. This slowing of contaminant transport is called retardation of the contaminant relative to the average seepage velocity of groundwater and results in a reduction in dissolved organic concentrations in groundwater. Sorption can also influence the relative importance of volatilization and biodegradation. Figure 3.6b illustrates the effects of sorption on an advancing dissolved contaminant front. Sorption is a dynamic and reversible reaction; thus, at a given solute concentration, some portion of the contaminant is partitioning out of solution onto the aquifer matrix, and some portion is desorbing and reentering solution. As solute concentrations change, the relative amounts of contaminant that are sorbing and desorbing will change. For example, as solute concentrations decrease due to other factors such as biodegradation and dilution, the amount of contaminant reentering solution will probably increase. The affinity of a given compound for the aquifer matrix will not be sufficient to isolate it permanently from groundwater, although for some compounds the rates of desorption may be so slow that the adsorbed mass may be considered as permanent residual within the time scale of interest. Sorption, therefore, does not permanently remove solute mass from groundwater; it merely retards migration. The various mechanisms that cause sorption effects to take place within the aquifer matrix are described in detail in Chapter 2. Because of their nonpolar structure, hydrocarbons most commonly exhibit sorption through the process of hydrophobic bonding. When the surfaces comprising the aquifer matrix are less polar than the water molecule, as is generally the case, there is a strong tendency for the nonpolar contaminant molecules to partition from the groundwater and sorb to the aquifer matrix. This phenomenon, referred to as hydrophobic bonding, is an important factor controlling the fate of many organic pollutants in soils. As described in Chapter 2, two components of an aquifer have the greatest effect on sorption: organic matter and clay minerals. In most aquifers, the organic fraction tends to control the sorption of organic contaminants. Sorption Models and Isotherms: Regardless of the sorption mechanism, it is possible to determine the amount of sorption to be expected when a given dissolved contaminant interacts with the materials comprising the aquifer matrix. Bench-scale experiments are performed by mixing water-contaminant solutions of various concentrations with aquifer materials containing various amounts of organic carbon and
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clay minerals. The solutions are then sealed with no headspace and left until equilibrium between the various phases is reached. (True equilibrium may require hundreds of hours of incubation, but 80 to 90% of equilibrium may be achieved in one or two days.) The amount of contaminant left in solution is then measured. The results are commonly expressed as a plot of the concentration of chemical sorbed (µg/g) vs. the concentration remaining in solution (µg/L). The relationship between the concentration of chemical sorbed (Ca ) and the concentration remaining in solution (Cs ) at equilibrium is referred to as the sorption isotherm because the experiments are performed at constant temperature (Figure 2.11). Sorption isotherms generally exhibit one of three characteristic shapes, depending on the sorption mechanism: the Langmuir isotherm, the Freundlich isotherm, and the linear isotherm (a special case of the Freundlich isotherm). Retardation: As mentioned earlier, sorption tends to slow the transport velocity of contaminants dissolved in groundwater. When the average velocity of a dissolved contaminant is less than the average seepage velocity of the groundwater, the contaminant is said to be retarded. The coefficient of retardation, R, is used to estimate the retarded contaminant velocity. The variation between the velocity of the groundwater and that of the contaminant is caused by sorption and is quantified by the coefficient of retardation, defined as: R=
V Vc
(3.10)
where R V Vc
= coefficient of retardation = average seepage velocity of groundwater parallel to groundwater flow = average velocity of contaminant parallel to groundwater flow
The ratio (V/Vc) describes the relative velocity between the groundwater and the dissolved contaminant. When Kd = 0 (no sorption), the transport velocities of the groundwater and the solute are equal (V/Vc). If it can be assumed that sorption is described adequately by the distribution coefficient (valid when the fraction of organic carbon (foc) > 0.001), the coefficient of retardation for a dissolved contaminant is described by the following equation:9 R = 1+ where R ρb Kd n
= = = =
coefficient of retardation bulk density of aquifer distribution coefficient porosity
ρb K d n
(3.11)
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The bulk density, ρb, of a soil is the ratio of the soil mass to its field volume. Bulk density is related to particle density by the following equation: ρb = (1 – n)ρs
(3.12)
where n is the total porosity and ρs is the density of soil grains comprising the aquifer. In sandy soils, ρb can be as low as 1.81g/cm3. In aggregated loams and clayey soils, ρb can be as low as 1.1g/cm3. The sorption relationship shown above expresses the coefficient of retardation in terms of the bulk density and effective porosity of the aquifer matrix and the distribution coefficient for the contaminant. Substitution of this equation into Equation 3.10 gives ρ K V = 1+ b d Vc n
(3.13)
Solving for the contaminant velocity, Vc , gives Vc =
Vx 1 + ρb K d n
(3.14)
Retardation factors can be calculated for several fuel and chlorinated solventrelated chemicals as a function of the fraction of organic carbon content of the soil. The value of R can vary over two orders of magnitude at a site, depending on the chemical in question and the estimated value of porosity and soil bulk density. Earlier investigations reported distribution coefficients normalized to total organic matter content (Kom ). The relationship between fom and foc is nearly constant, and assuming that the organic matter contains approximately 58% carbon:9 Koc = 1.724 Kom
(3.15)
Two methods are used to estimate the distribution coefficient and amount of sorption (and thus retardation) for a given aquifer-contaminant system. The first method involves estimating the distribution coefficient by using Koc for the contaminants and the fraction or organic carbon comprising the aquifer matrix. The second method involves conducting batches of column tests to determine the distribution coefficient. Because numerous authors have conducted experiments to determine Koc values for common contaminants, literature values are reliable, and it generally is not necessary to conduct laboratory tests.9
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3.4.2.2 Stabilization The transfer of an organic compound from an NAPL source to the surrounding water increases the contaminant concentration in groundwater. The rate of transfer varies depending on the type of NAPL. Computation of this transfer rate can be complex because the transfer rate depends on chemical properties of the contaminant and the NAPL, as well as on resistance at the interface between the water and the NAPL.11 Diffusion of the contaminant within the NAPL itself also can affect the transfer rate for viscous NAPLs. DNAPLs: Dense nonaqueous phase liquids (DNAPLs) present in the form of residual (held under capillary forces) or free phase (mobile) product may result in continued long-term contamination of the surrounding groundwater. The marginally soluble organic contaminants can partition into the aqueous phase at rates slow enough to continue to exist as a nonaqueous phase, yet rapid enough to cause significant groundwater contamination. DNAPLs can migrate to depths well below the water table. As they migrate, they can leave behind trails of microglobules in the pore spaces of the soil matrix, which effectively serve as long-term sources of groundwater contamination. Current conceptual DNAPL transport models suggest that, when sinking free phase DNAPL encounters a confining layer (e.g., competent clay or bedrock zone), it can accumulate, or “pool,” and spread laterally until it encounters a fracture or an alternative path of relatively low flow resistance towards deeper zones.11 In addition, globules can enter pores and be held as a residual phase in capillary suspension. This complex mode of subsurface transport results in unpredictable heterogeneous distribution of nonaqueous product that is difficult to delineate. The current lack of appropriate methods for detecting and delineating widely dispersed microglobules of DNAPL has been identified as one of the most significant challenges today. Investigative techniques that have been used to identify DNAPL source zones are listed below. It should be noted that some of those techniques are well proven and extensively field tested, while others are considered relatively new.12 • • • • • • • • • • •
Soil gas surveys Visual evidence of soil, rock and/or groundwater samples Chemical analyses of soil, rock and/or groundwater samples Enhanced visual identification — shake tests Enhanced visual identification — UV fluorescence with portable light, dye addition with Sudan IV or Oil Red O Accumulation within monitoring wells at target locations Partitioning interwell tracer tests Backtracking using dissolved concentrations in wells (the 1% rule) Surface geophysics Subsurface geophysics Cone penetrometer testing (CPT) methods: • Permeable membrane sensor, membrane interface probe (MIP) • Hydrosparse • Laser induced fluorescence (LIF) techniques • GeoVis
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• Raman spectroscopy • Electrochemical sensor probe • Cosolvent injection/extraction technique • Precision injection/extraction (PIX) technique • Flexible liner underground technologies everting (FLUTE) membrane technique
It is important to recognize that each of the methods listed presents specific advantages and disadvantages and applicability will be determined by technical and economic challenges encountered at each site. Several methods can be complementary in an overall site management plan, and a hybrid approach could be developed to exploit the strengths of the different techniques at the most appropriate and logical times in the site management process. For example, one can initially screen a site with a laser induced fluoroscence (LIF) technique or with geophysical techniques, then analyze confirmation soil samples in the field visually, with Sudan IV dye, and in the laboratory for chemical constituents. After determining the location of the DNAPL source zone, discreetly screened or multilevel wells can be installed for monitoring and remediation. CPT and/or geophysical techniques, integrated with minimally intrusive direct push technologies, can provide the framework for development of the conceptual site model. Then the refined conceptual site model integrated with hydrogeologic considerations can be used for guidance on a sampling plan to define the spatial extent of the contamination. 3.4.2.3 Volatilization Volatilization reduces the total mass of the contaminant in the groundwater system. The potential for volatilization is expressed by the contaminant’s Henry’s Law Constant and described in detail in Chapter 2. Henry’s Law Constants are widely available for common volatile contaminants (see Appendix A). Although not a destructive mechanism, volatilization does not remove contaminants from groundwater. In addition to Henry’s Law Constant, other factors affecting the volatilization of contaminants from groundwater include the contaminant concentration, the change in contaminant concentration with depth, diffusion coefficient of the compound, temperature, and sorption. Because the soil gas often advects and dispersion also occurs in the gas phase, contaminants transferred to the soil gas often migrate away from the location at which they volatilize. Volatilization itself does not destroy contaminant mass or permanently immobilize it. Volatilized contaminants can biodegrade in some circumstances but also can redissolve in infiltrating groundwater or be transported to the surface, where humans may be exposed to the vapors. 3.4.3
Transformation Mechanisms
A variety of reactions transform contaminants. The possible reactions are called biogeochemical: all are chemical (prefix chem) and occur in a geological setting (prefix geo), but some are catalyzed by microorganisms (prefix bio). Some biogeochemical reactions can degrade or transform a contaminant into benign and
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harmless end products or immobilize it permanently. A contaminant transformed or immobilized in these ways no longer contributes to groundwater contamination. Although other reactions do not directly lead to such positive results, they can control whether or not the transformation or immobilization reactions take place. Often, a suite of chemical reactions (termed a reaction network) leads to contaminant transformation or immobilization. In other instances, the reaction network prevents the contaminants from being transformed or immobilized and may make natural attenuation an ineffective remediation strategy. 3.4.3.1 Biodegradation Microorganisms can cause major changes in the chemistry of groundwater. Their small size and adaptability, as well as the diversity of nutritional requirements for different microbes, enable them to catalyze a wide range of reactions that often are the basis for natural attenuation. Chemical changes brought about by microorganisms can directly or indirectly decrease the concentrations of certain groundwater contaminants. Microorganisms use enzymes to accelerate the rates of certain biochemical reactions. The most important reactions are reductions and oxidations, together known as REDOX reactions. The reactions involve transfer of electrons from one molecule to another, which allow the microorganisms to generate energy and grow (Figure 3.8). More discussions on REDOX reactions and microbial electron transfers are provided in Chapters 2 and 4.
s and
tron Elec
Organic Contaminant
New Cells
C
+
Energy Elec
trons
Figure 3.8
n arbo
Electron Acceptor (e.g., O2)
Conceptual description of microorganisms gaining energy and utilizing the substrate for growth.
Microorganisms reproduce by organizing chemical reactions that create daughter cells composed of cellular components (e.g., membranes, proteins, deoxyribonucleic acid [DNA], cell walls) derived from building blocks that they synthesize or scavenge from the environment.1 The chemical reactions are made possible by enzymes — protein molecules that bring together the chemicals in a way that allows them to react quickly (Figure 3.9). The reactions are driven to completion by the expenditure of cellular energy in the form of a chemical known as adensoine triphosphate (ATP),
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Oxidized Donor Product
Unicellular Microorganism
NADH2
Electron Donor
NADH2
Synthesis and Maintenance ATP NAD ADP+Pi
NAD
Electron Acceptor Respiration
Reduced Acceptor Product
Figures 3.9
Conceptual diagram of microbial activity to derive energy for growth and multiplication (adapted from NRC, 2000).
which can be thought of as a cellular fuel. Like all living organisms, microorganisms generate ATP by catalyzing redox reactions: they transfer electrons from electronrich chemicals to electron-poor chemicals. The technical term for the electron-rich chemical is electron donor substrate. As an analogy, human metabolism involves transfer of electrons from chemicals derived from ingested food (the donor substrate) to oxygen (the acceptor substrate) inhaled from the air.1 When cells remove electrons from the donor substrate, they do not transfer the electrons directly to the acceptor substrate. Instead, they transfer the electrons to internal electron carriers as shown in Figure 3.9. Although electrons held by the carriers can be used for many purposes, the major purpose is to generate ATP through a process called respiration. In respiration, the electrons are passed from carrier to carrier until they reach the electron-acceptor substrate. Since this is the last molecule to receive the electrons, it is called the terminal electron acceptor. The need for ATP production forces all microorganisms to have one or more electron-donor and electron-acceptor pairs, and these materials largely define the metabolism of individual microorganisms. The amount of energy yielded varies depending on the electron donor and electron acceptor used.
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Ideally, all biologically mediated reactions produce energy for microbial growth and reproduction. Biologically mediated electron transfer results in oxidation of the electron donor, reduction of the electron acceptor, and the population of usable energy (quantified by the Gibbs free energy of the reaction ∆Gvo ). Table 3.2 presents a few select electron acceptor and electron donor reactions and calculated ∆Gyo values.9 Negative values indicate an energy-producing reaction, otherwise called an exothermic reaction, and will proceed from left to right. The value of ∆Gyo can be used to estimate how much free energy is consumed or produced during the reaction. Positive values indicate an endothermic reaction; for the reaction to proceed from left to right energy must be put into the system. Microorganisms will not invest more energy into the system than can be released and must couple an endothermic with an exothermic reaction to derive energy and grow. Collectively, microorganisms can use a wide range of electron donors, including both organic and inorganic chemicals. Electron acceptors are more limited. Common electron acceptors include O2, NO–3 , NO2– , SO42– , CO2, Fe(III), and Mn(IV). Oxygen has a special status because of its importance in many environments and reactions. Microbial use of oxygen as an electron acceptor is called aerobic metabolism; microbial use of electron acceptors other than oxygen is called anaerobic metabolism. When biotransformation of a particular contaminant leads directly to energy generation and the growth of more microorganisms, the contaminant is known as a primary substrate (see Figure 3.8). However, the reactions that lead to microbial metabolism of contaminants may not be part of cell-building or energy-generating reactions. An important category of such biotransformations is cometabolism. Cometabolism is the fortuitous degradation of a contaminant when other materials are available to serve as microorganisms’ primary substrates. Cometabolic reactions often occur because the enzymes designed for metabolizing primary substrates fortuitously transform the cometabolic substrate. It is important to note the historic debate on the use of the word cometabolism for the microbially catalyzed process described above.13,14 One school of thought, propagated by classical microbiologists, insists that usage of either the term cometabolism or the term cooxidation to describe conversions of nongrowth substrates by nonproliferating microbial populations in the absence of a metabolizable cosubstrate would be inappropriate. The enzymatic conversion of a substrate by a nonproliferating microbial population because an enzyme of broad specificity and conversion capability is in proximity to the substrate might at best be described as bioconversion. There is no co- (with or together) activity concerned with such an event. First-Order Decay Model: One of the most commonly used expressions for representing the biodegradation of an organic compound involves the use of an exponential decay relationship: C = C0 e–kt where C C0 k
= biodegraded concentration of the chemical at time t = initial concentration = rate of decrease of the chemical (units of 1/time) [T–1]
(3.16)
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Table 3.2 Half-Cell Reactions for Some of the Common Electron Acceptors and Donors (adapted from Wiedemeier et al., 1999) Half-Cell Reaction
∆Gro (kcal/mol e–)
4e– + 4H+ O2 ⇒ 2H2O Aerobic respiration
–18.5
5e– + 6H+ + NO3– ⇒ 0.5N2 + 3H2O Denitrification
–16.9
2e– + 4H+ + MnO2 ⇒ Mn2+ + 2H2O Manganese reduction
–8.6
e– + Fe3+ ⇒ Fe2+ Fe(III) reduction
–17.8
8e– + 9.5H+ + SO42– ⇒ 0.5HS– + 0.5H2S + 4H2O Sulfate reduction
5.3
8e– + 8H+ + CO2 ⇒ CH4 + 2H2O Methanogenesis
5.9
C2Cl4 + H+ + 2e– ⇒ C2HCl3 + Cl– PCE reductive dechlorination
–9.9
C2HCl3 + H+ + 2e– ⇒ C2H2Cl2 + Cl– TCE reductive dechlorination
–9.6
C2H2Cl2 + H+ + 2e– ⇒ C2H3Cl + Cl– cis-DCE reductive dechlorination
–7.2
C2H3Cl + H+ + 2e– ⇒ C2H4 + Cl– VC reductive dechlorination
–8.8
C2H3Cl3 + H+ + 2e– ⇒ C2H4Cl2 + Cl– TCA reductive dechlorination /2 H 2 ⇒ H+ + e – Hydrogen oxidation
1
–10.3 –9.9
/4 CH2O + 1/4 H2O ⇒ 1/4 CO2 + H+ + e– Carbohydrate oxidation
–10.0
12H2O + C6H6 ⇒ 6CO2 + 3O H+ + 3Oe– Benzene oxidation
–7.0
14 H2O + C6H5CH3 ⇒ 7CO2 + 36H+ + 36e– Toluene oxidation
–6.9
20H2O + C10H8 ⇒ 10CO2 + 48H+ + 48e– Naphthalene oxidation
–6.9
1
4H2O + C2H3Cl ⇒ 2CO2 + 11H+ + 10e– + Cl– Vinyl chloride oxidation
–11.4
12H2O + C6H5Cl ⇒ 6CO2 + 29H+ + 28e– + Cl– Chlorobenzene oxidation
–8.0
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First-order rate constants are often expressed in terms of half-life for the chemical: t1 2 =
0.693 k
(3.17)
The first-order decay model shown in Equation 3.16 assumes that the solute degradation rate is proportional to the solute concentration. The higher the concentration, the higher the degradation rate. This method is usually used to simulate biodegradation of contaminants dissolved in groundwater. Modelers using the firstorder decay model typically use the first-order decay coefficient as a calibration parameter and adjust the decay coefficient until the model results match the field data. With this approach, uncertainties in a number of parameters (e.g., dispersion, sorption, biodegradation) are lumped together in a single calibration parameter. Regression methods are commonly used to obtain approximations of site-specific degradation rates (first-order) from log-linear plots of concentration vs. time. This involves fitting an exponential regression to approximate the trend in the data. This type of approximation can be used to evaluate trends at an individual well or for several wells along a flow path. When individual wells are being evaluated, the analytical data should be used from multiple sampling events, and the time element in the plot represents the temporal arrangement of the data. When multiple wells along a flow path are being evaluated, the analytical data from a single sampling event can be used; the time element in the plot represents groundwater travel time between the wells.15 Electron-Acceptor-Limited or Instantaneous Reaction Model: The electronacceptor-limited model (traditionally called the instantaneous reaction model) was first proposed in 1986 for simulating the aerobic biodegradation of petroleum hydrocarbons.9,16 It was observed that microbial biodegradation kinetics are fast in comparison with the transport of oxygen and that the growth of microorganisms and utilization of oxygen and organics in the subsurface can be stimulated as an electron-acceptor-limited or instantaneous reaction between the organic contaminant and oxygen. From a practical standpoint, the instantaneous reaction model assumes that the rate of utilization of the contaminant and oxygen by the microorganisms is very high, and that the time required to biodegrade the contaminant is very short, almost instantaneous, relative to the seepage velocity of the groundwater. Using oxygen as an electron acceptor, for example, biodegradation is calculated using the expression: ∆C R =
O F
(3.18)
where ∆CR = change in contaminant concentration due to biodegradation O = concentration oxygen F = utilization factor, the ratio of oxygen to contaminant consumed
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The variable F is obtained from the oxidation-reduction reaction involving the organic and the given electron acceptor. Biodegradation of Organic Contaminants: Organic contaminants vary widely in their susceptibility to transformation by microorganisms. Some contaminants are highly biodegradable, while others resist degradation. In general, the more degradable contaminants have simple molecular structures (often similar to the structures of naturally occurring organic chemicals), are water soluble and nontoxic, and can be transformed by aerobic metabolism (Figure 3.10). In contrast, organic contaminants that resist biodegradation may have complex molecular structures (especially structures not commonly found in nature), low water solubility or an inability to support microbial growth, or they may be toxic to the organisms.
Not Accessible
Accessible
Gaseous
Sorbed
Dissolved
Nonaqueous Figure 3.10
Schematic diagram describing the mechanisms by which a contaminant becomes available for biodegradation.
Microorganisms can completely convert some organic contaminants to carbon dioxide and water, while they are capable of only partial conversions of others. Complete conversion to carbon dioxide is called “mineralization.” In some cases, the products of partial conversion are more toxic than the original contaminant. Vinyl chloride is an example of a highly toxic chemical that results from incomplete biodegradation of chlorinated solvents. The following discussion explains how microbial transformations occur for various organic contaminant classes. It describes all of the elements of some metabolic pathways because these illustrate the core concepts of biodegradation. Biodegradation pathways for most contaminants are extremely complex, so these pathways are not described in detail. Petroleum hydrocarbons are a highly varied class of naturally occurring chemicals used as fuels in a variety of commercial and industrial processes. Biodegradation potential varies depending on the type of hydrocarbon. Benzene, Toluene, Ethylbenzene, and Xylene (BTEX): Benzene, Toluene, Ethylbenzene, and Xylene are components of gasoline. Because of their widespread use and because BTEX storage tanks commonly leaked in the past, BTEX are common groundwater contaminants. A large body of scientific research exists on the biodegradation and natural attenuation of BTEX. However, the effectiveness of
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MNA via intrinsic bioremediation, as with any other contaminant, depends on the relationship between the contaminant biodecay rate and the groundwater velocity. BTEX are easily biodegraded to carbon dioxide by aerobic microorganisms and can also biodegrade under anaerobic conditions. When the volume of BTEX is small enough and/or the supply of oxygen is large enough, microbes can degrade all of the BTEX components within the aerobic zones of a contaminated site. When oxygen is depleted in an advancing contaminant plume, anaerobic conditions can develop and lead to the formation of as many as five different downgradient zones, each with a different terminal electron acceptor (Figure 3.11). In these zones, BTEX degradation processes are slower and less reliable than when oxygen is present.
Source Area
Methanogenic 5
SO4 2-
Fe3+/Mn4+ NO3 Reduction Reduction
4 Reduction 3
2
Aerobic Zone O2 1
Groundwater Flow Direction 1
Figure 3.11
Encroachment of the Aerobic Fringe
- Aerobic Zone
2 and 3
- Transient Anaerobic Zones
4 and 5
- Core Anaerobic Zones
Conceptualization of the dominant terminal electron acceptor process (TEAP) in advancing BTEX plume.
Of the possible electron acceptors, oxygen yields the most energy. Once oxygen is depleted, nitrate is next as the most energy-yielding terminal electron acceptor. If nitrate is abundant in groundwater, zones in which microbes use nitrates as the electron acceptor will develop. A Mn(IV)-reducing zone may develop next if Mn(IV) is present in the subsurface mineral matrix (although the coupling of Mn reduction to BTEX degradation has not been well studied). Upon depletion of the Mn(IV), Fe(III) reduction will prevail if iron oxide minerals are present. In the next zones, sulfate and CO2 will serve as electron acceptors. Based on electron acceptor abundance, Fe3+, Mn4+, and SO2– 4 reduction by bacteria may play a dominant role in intrinsic bioremediation under certain geologic conditions. Both Fe3+ and SO42– reduction processes involve mineral phases and may not be properly understood by evaluating only groundwater concentrations. Fe and S mineral analyses, from soil samples, should be incorporated in natural attenuation studies, when the geologic conditions are appropriate. Fe and S mineral analyses may not be widely utilized in natural attenuation studies because of the inherent difficulty in solid sample collection, preservation, and analysis of bioavailable minerals.17
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Many field studies of BTEX biodegradation in the subsurface have been carried out. For example, several lines of evidence indicated that all BTEX components were biodegrading mainly in the Fe(III)-reducing zone of an aquifer in Bemidji, Minnesota, that was contaminated with crude oil.18-20 At a petroleum spill site in South Carolina, toluene, but not benzene, was metabolized as it moved through a sulfate-reducing zone.1,21 In a recent study of an anaerobic gasoline-contaminated aquifer in California, researchers injected BTEX components (along with bromide as a tracer) and either sulfate or nitrate into a sandy aquifer. Periodic withdrawal of samples from the injected zones showed that under nitrate-reducing conditions, toluene, ethylbenzene, and m-xylene, (but not benzene) were transformed in less than ten days. Under sulfate-reducing conditions, toluene, m-xylene, and o-xylene were completely transformed in 72 days, while benzene loss was uncertain.1,22 During a recent study24 in short term (< 2 weeks) incubations, addition of sulfate slightly stimulated benzene degradation and caused a small decrease in the ratio of methane to CO2 production from benzene. However, in long term (>100 days) incubations, sulfate significantly stimulated benzene degradation with a complete shift to CO2 as the end product of benzene degradation. The addition of Fe(III) and humic substances had short- and long-term effects that were similar to the effects of sulfate amendments. A novel in situ respiration technique was reported recently to measure and predict natural attenuation of petroleum compounds in the subsurface. Monitoring CO2 and CH4 produced in situ, and their radiocarbon (14C), stable carbon (13C), and deutrium (D) signatures provides a novel method to assess anaerobic microbial processes. The in situ anaerobic respiration test was conducted by injecting a large volume of industrial grade Argon, an inert gas, into the subsurface to replace CO2 and CH4, followed by monitoring the production of CO2 and CH4.23 Figures 3.12a and b show the formula of BTEX compounds. H C HC
CH
HC
CH
OR
OR
C H
Figures 3.12a
Benzene formula and simplified representations.
CH 3
CH 3
CH 3 CH 3
Benzene Figures 3.12b
Toluene
m-Xylene
Structures of single-ring aromatic hydrocarbons.
Ethylbenzene
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Polycyclic Aromatic Hydrocarbons: In contrast to BTEX, Polycyclic Aromatic Hydrocarbons (PAHs) biodegrade very slowly. PAH contamination comes mostly from fossil fuel use and the manufactured-gas industry.1 Groundwater contamination at manufactured gas plants has persisted for decades because of the slow, continuous dissolution of PAHs from subsurface coal tar. PAHs are compounds that have multiple rings in their molecular structure. These compounds have complex molecular structures and low water solubility, and they tend to sorb strongly to solids in the subsurface. However, because PAHs dissolve slowly, natural attenuation could control the contamination even if biodegradation is slow, as long as it occurs at the same rate as or faster than dissolution. The fate of PAHs in subsurface systems is governed largely by their hydrophobic nature (the reason for their low solubility and tendency to attach to surfaces). PAH molecules held within NAPLs or adsorbed to surfaces cannot be biodegraded. Consequently, understanding dissolution and the sorption processes for PAHs often is the key to understanding biodegradation and natural attenuation potential. Biodegradation of PAHs depends on the complexity of the chemical structure and the extent of enzymatic adaptation. In general, PAHs that contain two or three rings such as napthtalene, anthracene, and phenanthrene are degraded at reasonable rates when O2 is present. Studies have shown that some microorganisms can metabolize dissolved PAHs composed of up to five benzene rings. Microorganisms generally use oxygenase enzymes to initiate the biodegradation; these reactions require the presence of oxygen. However, microbial degradation of PAHs with lower molecular weights (fewer benzene rings) can occur under nitrate-reducing and sulfatereducing conditions.28,29 Oxygenated Hydrocarbons: Microbiologists and remediation engineers have long known that low molecular weight alcohols, ketones, esters and ethers biodegrade readily particularly under aerobic conditions. The polar oxygen atom in MtBE (CH3–O–C(CH3)3) causes the molecule to be much more hydrophilic than other gasoline constituents. However, one prominent oxygenated hydrocarbon methyl tertbutyl ether (MtBE) was thought to be resistant to biodegradation because of its stable molecular structure and its reactivity with microbial membranes. MtBE has been used as an octane enhancer in gasoline since the late 1970s and recently has been used up to 15% by volume. MtBE has relatively high water solubility (43,000–54,000 ppm in comparison to 1780 ppm for benzene), a very low Henry’s Law Constant (0.022 in comparison to 0.22 for benzene), and very weak sorbtion to soil (log Koc = 1 to 1.1 in comparison to about 1.5–2.2 for benzene). Until very recently, MtBE was considered nonbiodegradable in the subsurface; a prestigious state of the science report by the National Research Council, in the year 2000, stated that “present knowledge on MtBE biodegradation is limited ….” The report further pointed out that the process was not well understood and therefore the likelihood of success for natural attenuation as a remediation solution for MtBE contaminated sites was low.1 However, a number of recent studies have demonstrated natural MtBE biodegradation in the field.1,30-34 It is still unclear how prevalent this biodegradation is and whether the rates are rapid enough to restrict and eventually shrink groundwater
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plumes. Also, it is unknown if the potential for natural MtBE biodegradation could be reasonably predicted using some indicator parameters. A recent study, which included many field sites, suggested that natural biodegradation of MtBE and TBA under anaerobic subsurface conditions at some sites may control migration of MtBE and TBA plumes. There appeared to be a good correlation between strongly anaerobic plume biogeochemistry and natural biodegradation of MtBE.35 To date no one has shown MtBE biodegradation in the laboratory under sulfate reducing conditions. It is important to note from this study that MtBE and TBA naturally biodegrade only under strongly anaerobic conditions (preferably under methanogenic conditions) and the rates of biodegradation (at the sites where this happens) are comparable to those of benzene.35 Chlorinated Aliphatic Compounds: The chlorine atoms added to aliphatic organic molecules to produce these chemicals significantly change many properties, including solubility, volatility, density, hydrophobicity, stability, and toxicity. These changes are valuable for commercial products, but also can make the compounds less biodegradable. Several good reviews have been published on the biodegradation of the small (one- and two-carbon) chloroaliphatic compounds.25 The biodegradation potentials of many chlorinated aliphatics are discussed extensively in Chapter 4. Researchers first demonstrated the potential for anaerobic biotransformation of chlorinated aliphatic hydrocarbons during the early 1980s.36 Subsequent studies have shown that these compounds can biotransform under a variety of environmental conditions in the absence of oxygen. In general, the biotransformation rates, particularly for chlorinated compounds with more than two chlorine atoms in the molecule, are higher under anaerobic conditions. Exceptions to the general rule that chlorinated aliphatic hydrocarbons require special environmental conditions for biodegradation to occur are methylene chloride, known also as dichloromethane, and vinyl chloride. Methylene chloride and vinyl chloride can support the growth of a wide range of microorganisms (both aerobic and anaerobic) under a range of environmental conditions. Methylene chloride and vinyl chloride therefore are likely to be treated successfully by natural attenuation at a much broader range of sites than other chlorinated aliphatics compounds. In addition to methylene chloride and VC, there are a few other chloroaliphatic compounds which will degrade under aerobic conditions as growth substrates or as cometabolic substrates. Natural biotransformation of chloroaliphatics is most likely where excess organic material is available to serve as an electron donor and biogeochemical conditions support a reducing environment. Successful intrinsic reductive dechlorination has been found to occur in the presence of other electron-donating organic pollutants, such as those from leaking sewage systems, BTEX, and phenol. Reductive dechlorination to VC and ethene appeared to be driven by fuel hydrocarbon co-contaminants in the center of many mixed contaminant plumes. Down gradient, where carbon sources became depleted, VC was oxidized further by iron and aerobic oxidation. When soil organic matter serves as electron donor, reductive dechlorination may also be observed down gradient of the plume and dechlorination products may accumulate.
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Chlorinated Aromatic Compounds: Bacteria able to degrade all but the most complex chloroaromatic compounds have been discovered during the past 20 years.1,25 Chlorobenzenes including hexachlorobenzene can be sequentially dechlorinated to chlorobenzene under methanogenic conditions in soil slurries37 (Figure 3.13). Reductive dechlorination of chlorobenzene has not been reported, but chlorotoluenes are dechlorinated to toluene in the preceding methanogenic systems and it seems likely that chlorobenzene could serve as a substrate for reductive dechlorination. Cl Cl
Cl -
Cl -
Cl -
Cl -
Cl
Cl
Cl
Cl Cl
Figure 3.13
Cl -
Methanogenic Conditions
Reductive dechlorination of hexachlorobenzene under anaerobic conditions.
Chlorobenzenes up to and including tetrachlorobenzene are readily biodegraded under aerobic conditions. Bacteria able to grow on chlorobenzene,25,38 1,4-dichlorobenzene,25,38 1,3-dichlorobenzene,25 1,2-dichlorobenzene,25 1,2,4-trichlorobenzene,25 and 1,2,4,5-tetrachlorobenzene25 have been isolated and their metabolic pathways identified. The pathways for aerobic degradation are remarkably similar and lead to the release of the chlorine as HCl. Chlorobenzenes are very good candidates for natural attenuation under either aerobic or anaerobic conditions. Aerobic bacteria able to grow on chlorobenzene have been detected at a variety of chlorobenzene-contaminated sites but not at adjacent uncontaminated sites,25,39 providing strong evidence that they are selected for their ability to derive carbon and energy from chlorobenzene degradation in situ. Removal of multiple chlorines as HCl consumes a large amount of alkalinity and produces a considerable drop in the pH of unbuffered systems which could lead to a loss of microbial activity at some sites. Although the benzene ring that is the nucleus of chlorinated aromatic compounds is relatively easy for microorganisms to biodegrade, the addition of chlorine atoms completely alters the biodegradability of benzene. The number and position of chlorine atoms on the benzene ring determine how biodegradable the compound will be. Compounds with many chlorine atoms may not be biodegradable at all under aerobic conditions; however, under special environmental conditions, these compounds can be reductively dechlorinated by the same type of microbial dechlorination process that can occur for chlorinated aliphatic compounds.25,40-42 As the reductive dechlorination process removes chlorine atoms from the benzene ring, the molecules become more susceptible to biodegradation by aerobic microbes. When environmental conditions are right, natural attenuation may be able to control halogenated aromatic compounds, but these conditions generally are uncommon. Chlorophenols and chlorobenzoates are dechlorinated under anaerobic conditions in sediments and subsurface material.25,43,44 In some instances the dechlorination clearly yields energy for the growth of the specific bacteria. In other examples the dechlorination is specific and enriched in the community, but has not been rigorously linked
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to energy production. Addition of small fatty acids or alcohols as electron donors or sources of carbon can enhance the process of reductive dechlorination. Aerobic pathways for the degradation of chlorophenols and chlorobenzoates are initiated by oxygenase-catalyzed attack on the aromatic ring and subsequent removal of the chlorine after ring fission or hydrolytic replacement of the chlorine with a hydroxyl group. Bacteria able to grow on chlorophenols and chlorobenzoates are widely distributed and readily enriched from a variety of sources, indicating a high potential for natural attenuation. Chlorophenols are unusual among the synthetic compounds discussed here in that they can be very toxic to microorganisms. They are often used as biocides; therefore, high concentrations can dramatically inhibit biodegradation. Inoculation with specific bacteria has been helpful in overcoming toxicity and stimulating degradation of chlorophenols.25,43 Pentachlorophenol deserves special consideration because it has been widely used as a wood preservative and has been released into the environment throughout the world. Reductive dechlorination under methanogic conditions can lead to mineralization.25,43 Aerobic bacteria catalyze the replacement of the chlorine in the 4 position by a hydroxyl group to form tetrachlorohydroquinone. Subsequent reductive dechlorinations lead to the formation of ring fission substrates. Bacteria able to degrade pentachlorophenol are widely distributed, and both experimental and fullscale bioremediation projects have been successful in field applications43 (Figure 3.14). Adding selected strains has been helpful in some instances; in others, indigenous strains have been used. Wood treatment facilities typically are contaminated with complex mixtures of organic compounds; therefore, investigations of toxicity must be conducted for each site under consideration. Natural attenuation of pentachlorophenol has been reported, because specific bacteria able to use it as a growth substrate are enriched at contaminated sites. However, rates seem to be low at many sites due to toxicity and bioavailability of the pentachlorophenol. Although polychlorinated biphenyl (PCB) use has been banned, these chemicals are still present in the environment, especially in sediment and aquatic systems, and their persistence is due in part to their resistance to biodegradation.1,45 PCBs consist of up to ten chlorine and hydrogen atoms attached to a structure consisting of two benzene rings attached by a bond between carbon atoms. Chemical synthesis can create various possible combinations — called “congeners” — of chlorine and hydrogen atoms in the ten positions (Figure 3.15). PCBs were marketed as mixtures of congeners called Aroclors (the Monsanto Corporation trade name), characterized according to average chlorine content. PCBs have been studied extensively because of their stability, toxicity, and bioaccumulation potential.1,46 Anaerobic transformation of PCBs is catalyzed by bacteria in aquatic sediment from a wide range of contaminated and uncontaminated sites. Higher activities in contaminated sites suggest that the dechlorination reactions provide a selective advantage to the microbial population, indicating the potential for significant natural attenuation. A number of studies have clearly demonstrated that natural attenuation of PCB is taking place in anaerobic sediments at significant rates. Methanogenic conditions in freshwater sediments seem to provide the highest rates of reductive dechlorination.
Figure 3.14
Cl Cl
Cl
Cl
Cl
OH
Cl
Cl
Pathways of pentachlorophenol (PCP) degradation.
Cl
Cl
OCH 3
Cl
Cl
OH
Cl
Cl Cl
Cl
Cl
OCH 3
Cl
Cl
Cl
Cl
Cl
OH
Cl
OH
Cl
OH
Cl
Cl
Cl
Cl
Cl
Cl
Cl
OH
Cl
OH
Cl
OH
Cl
Cl
Cl
HOOC
Cl
OH
Cl
Cl
Cl
Cl
Cl
Cl
Cl
Cl
Cl
Cl
Cl
Ring Cleavage COOH
Cl
OH
OH
102
OCH 3
Cl
Cl
OH
Cl
Cl
OCH 3
Ring Cleavage
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CIm m=1 Figure 3.15
103
CIn 5
n=1
5
Structure of PCB.
Dechlorination converts the more highly chlorinated congeners to less chlorinated products containing one to four chlorines. Complete dechlorination does not occur, but the depletion of the more highly chlorinated congeners dramatically reduces not only the toxic and carcinogenic potential, but also the bioaccumulation potential. A variety of dechlorination patterns have been identified as a function of the microbial community involved. The patterns are constant within a given microbial community or enrichment, supporting the premise that dechlorination provides a selective advantage to the organisms involved. The electron donors for the dechlorination in sediment are unknown. Addition of exogenous carbon sources does not stimulate the reaction. In contrast, “priming” the mixtures with low levels of bromobiphenyl or specific isomers of tetrachlorobiphenyl1,46,47 seems to selectively enrich a population of PCB-dechlorinating bacteria and dramatically stimulate the dechlorination of other congeners. The lower chlorinated PCB congeners, whether part of the original Arochlor mixture or derived from reductive dehalogenation, are biodegraded by aerobic bacteria.25,48 The initial attack is catalyzed by a 2,3- or 3,4-dioxygenase followed by a sequence of reactions that leads to ring cleavage and accumulation of chlorobenzoates readily degraded by a variety of bacteria. The enzymes that oxidize PCBs are produced by bacteria growth or biphenyl, and addition of biphenyl to slurryphase reactors stimulates the growth and activity of PCB degraders. Such stimulation has been shown to be effective in the field. There is also good evidence that aerobic PCB degradation is taking place in contaminated river sediments.48 It seems clear that reductive dechlorination is ongoing at a wide range of PCBcontaminated sites. The strategy of anaerobic dechlorination followed by aerobic degradation seems to be particularly effective with PCB whether in an engineered system or in natural systems occurring during resuspension of anaerobic sediments. To date, the complete biodegradation of PCB is slow and difficult to predict or control in the field. Several new strategies, including construction of novel strains, may increase the potential for effective PCB biodegradation. Nitroaromatic Compounds: The literature on biodegradation of nitroaromatic compounds has been reviewed recently.25,49,50 These compounds are subject to reduction of the nitro groups in the environment under either aerobic or anaerobic conditions. Reduction does not lead to complete degradation in most instances and could
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be considered nonproductive for purposes of natural attenuation. In contrast, aerobic bacteria able to grow in nitrobenzene, nitrotoluenes, dinitrotoluenes, dinitrobenzene, nitrobenzoates, picric acid, and other nitrophenols have been isolated from a variety of contaminated sites, suggesting that natural attenuation is taking place. Mineralization of dinitrotoluenes in aquifer material from a dinitrotoluene-contaminated site was measured recently.51 It was concluded that the indigenous microorganisms provide a significant degradative capacity for the contaminant. The simple nitroaromatic compounds can be considered excellent candidates for natural attenuation as long as the degradation process yields a selective advantage. Some of the compounds, including 3-nitrophenol, nitrobenzene, 4-nitrotoluene, and 4-nitrobenzoate, are degraded via catabolic pathways that minimize the use of molecular oxygen and are particularly well suited for operation in the subsurface where oxygen is limiting. The pathways all involve a partial reduction of the molecule prior to oxygenative ring fission. For example, the first three steps in the pathway for degradation of nitrobenzene can take place in the absence of oxygen,49 which is required only for ring fission and subsequent metabolism. Mixtures of the isomeric nitro compounds can be problematic for microbial degradation. For example, the industrial synthesis of polyurethane produces large amounts of 2,4- and 2,6-dinitrotoluene in a ratio of 4:1. Bacteria able to grow on 2,4-dinitrotoluene have been studied extensively. Unfortunately, 2,6-dinitrotoluene inhibits the degradation of 2,4-dinitrotoluene and may prevent natural attenuation. Bacteria able to grow on 2,6-dinitrotoluene have been isolated recently, and insight about the metabolic pathway might allow better prediction of degradation of the mixture.25 Nitroaromatic organic contaminants are associated uniquely with military activities and include the explosives trinitrotoluene (TNT), royal Dutch explosive (RDX or hexahydro-1,3,5-trinitro-1,3,5-triazine), and octahydro-1,3,5,7-tetranitro-1,3,5,7tetrazocene (HMX).25 Manufacturing, loading, storage, and decommissioning operations have generated large quantities of explosive wastes, some of which were deposited in soils and unlined lagoons and subsequently leached to groundwater. Despite the number of sites contaminated with explosives, only a few rigorous field studies have been conducted to determine the transport, fate, and influence of microbial activity on explosives. Furthermore, the field studies carried out to date are inconclusive in establishing the role of biodegradation in the fate of nitroaromatics.25,51 Laboratory studies clearly show the potential for microorganisms to metabolize nitroaromatic compounds.2,49,51,52 However, microbes apparently cannot readily use TNT, RDX, or HMX as primary substrates for sources of the carbon and energy needed for their growth. Instead, cometabolic reactions generally prevail.1,49 Under aerobic and anaerobic conditions, microorganisms routinely reduce the nitro groups on nitroaromatics to amino nitro groups. These changes can increase toxicity of the molecules and cause them to form polymers, and/or strongly sorb onto soils.1,52 Recent reports have shown that aerobically and anaerobically grown bacteria can use TNT and RDX as nutritional nitrogen sources,2,53,54 but metabolic byproduct accumulation is common. The possibility of natural attenuation of nitroaromatics cannot be precluded, but the kinds of conditions needed are not clearly understood.
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Nitrate Esters: A variety of nitrate esters including glycerol trinitrate, pentaerythritol tetranitrate, and nitrocellulose have been used extensively as explosives. Recent studies have indicated that the nitrate esters can be degraded by bacteria from a variety of sources.25,55,56 Bacterial metabolism releases nitrite which can serve as a nitrogen source and yield a selective advantage for the organisms. The biodegradation of nitrate esters has only recently been studied extensively and little is known about degradation in the environment. The recent laboratory results show considerable promise that natural attenuation is possible, but more information is needed on the bioavailability, toxicity, and kinetics of the process. Pesticides: Most pesticides used in the past 20 years in the U.S. have been formulated to degrade in the environment, and a considerable amount of information is available on degradation kinetics in soil and water. The U.S. Environmental Protection Agency Risk Reduction Engineering Laboratory in Cincinnati, OH has developed an extensive Pesticide Treatment Database that contains information on a variety of compounds.25 Many pesticides hydrolyze and yield compounds that serve as growth substrates or sources of nitrogen or phosphorus for bacteria. Enhanced degradation of pesticides has been studied extensively25,57 and is closely related to natural attenuation. For example, carbamates,25,57 chlrophenoxyacetates,25 dinitrocresol, atrazines,25 and some organophosphates serve as growth substrates for bacteria and would be good candidates for natural attenuation. A variety of other pesticides are hydrolyzed by extracellular enzymes derived from soil bacteria but provide no advantage to the organisms that produce the enzymes. Similarly, some of the organohalogen insecticides can be reductively dehalogenated but provide no advantage to specific organisms. Their biodegradation rates are proportional to the biomass and activity in the soil. Other organohalogens, such as lindane, can serve as growth substrates for specific bacteria,25,58,59 but such bacteria seem not to be widely distributed (Figure 3.16). Microbial Transformation of Inorganic Contaminants: Many research reports have documented that microorganisms can transform inorganic contaminants.1 However, unlike organic compounds, which microbes can destroy completely to CO2, H2O, and other innocuous products, most inorganic contaminants can be changed only to forms with different solubilities and mobilities. Microbial reactions can lead to precipitation, volatilization, sorption, or solubilization of inorganic compounds. These outcomes can be the direct result of enzymes produced by the microbes, or they can be the indirect result of microbiological production of materials that alter the biogeochemical environment. One nearly universal means by which microorganisms lower concentrations of inorganic contaminants in water is adsorption to the microorganisms themselves. Adsorption can be caused by electrostatic attraction between the metals and the microbes or by highly specific scavenging systems that accumulate metals to high concentration within the cells.1,60 Although sorption to microbial biomass probably cannot be harvested from the subsurface, which would be required to prevent later release of contaminants, it is not likely to be a major factor in natural attenuation. Metals: Microbial effects on metals vary substantially depending on the metal involved and the geochemistry of the particular site. The behavior of many toxic metals depends on the microbially mediated cycling of naturally occurring elements,
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Cl
Cl
Cl
Cl
Cl
Cl
Cl
Cl Cl
Cl
γ - 3,4,5,6,- tetrachloro cyclohexane
Lindane
Cl Cl
Cl
Cl
Metabolites Cl Figure 3.16
Pathways of lindane degradation.
especially iron and manganese. The possible fates of chromium and mercury illustrate the variable effects of microbially mediated reactions on metals. Chromium: As with many metals, the effects of microbial transformation on chromium vary with its chemical form (technically, its oxidation state). In groundwater, the predominant form of chromium is the oxidized form, Cr(VI), present as chromate (CrO42– ) and dichromate (Cr2O72– ) ions. Cr(VI) (known as hexavalent chromium) is toxic and mobile. Reduced chromium, Cr(III), is less toxic and less mobile because it precipitates as Cr(OH)3 at groundwater pH values of 4.5 to 10.5. A variety of aerobic and anaerobic microorganisms enzymatically reduce Cr(VI) to Cr(III), but the physiological reason for this ability has not been adequately investigated. Among the hypotheses explaining these reduction reactions are detoxification (to move Cr away from the cells), cometabolism (fortuitous enzymatic
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reactions), and the use of Cr(VI) as a respirator electron acceptor. Microbes also may cause indirect reduction of Cr(VI) by producing sulfide, Fe(II), and reduced organic compounds because Cr(VI) reduction occurs spontaneously in the presence of these substances. Regardless of the mechanism involved, natural attenuation that relies on chromium reduction requires environmental conditions that strongly favor the reduced form of chromium. Mercury: Mercury is sometimes present in soils and sediments at contaminated sites in the form of mercuric ion, Hg(II), elemental mercury, Hg(0), and the biomagnification-prone organic mercury compounds monomethyl- and dimethylmercury (both of which can accumulate at hazardous levels in the food chain). All microbial transformations of mercury are detoxification reactions that microbes use to mobilize mercury away from themselves.1 Most reactions are enzymatic, carried out by aerobes and anaerobes, and involve uptake of Hg(II) followed by reduction of Hg(II) to volatile forms (elemental Hg(0) and methyl- and dimethylmercury) or the formation of highly insoluble precipitates with sulfide. In general, natural attenuation based on microbial mercury reduction and volatilization seems implausible because the volatile forms remain mobile, although immobilization as Hg(II) sulfides may be possible if the electron donors needed to sustain the microbial production of enzymes and the sulfate needed for precipitation are present together. Nonmetals: Arsenic is a relatively common toxic groundwater contaminant, due both to its use in industry and agriculture and to its natural weathering from rocks. Arsenic can exist in five different valence states: As(-III), As(0), As(II), As(III), and As(V), where the roman numerals indicate the charge on the arsenic atom. Depending on its valence state and the environment in which it exists, arsenic can be present as sulfide minerals (e.g., As2S3), elemental As, arsenite (AsO2– ), arsenate (AsO43– ), or various organic forms that include methylated arsenates and trimethyl arsine. The two most common forms of arsenic in natural systems are arsentate As(V) and arsenite As(III). As(V) is less soluble and less toxic than the more soluble As(III) form. The more oxidized arsenate would be expected as the dominant form in aerobic surface waters, and arsenite may be the dominant form in reduced groundwater systems. As(V) (arsenate), like phosphate, exists mainly in its deprotonated forms at natural pH levels, and so is readily adsorbed onto the positively charged surfaces of minerals such as Fe(III) oxides. The more toxic aresenite exists primarily as a neutral dissolved species at pHs typical of natural systems, and its transport is therefore not as much retarded by sorption onto oxide surfaces. Half times for oxidation of an arsenite in the presence of Mn(IV) oxides in laboratory experiments have been measured as 10–20 min, compared to 17 h in natural systems and 8760 h for solution of arsenite and dissolved oxygen without Mn oxides. Arsenic speciation in natural systems is not consistent with thermodynamic equilibrium and the kinetics of redox conversions of arsenic are relevant to its fate and transport. Microorganisms can transform arsenic for one of several physiological reasons. Under anaerobic conditions, microbes can use As(V) as a terminal electron acceptor. Under aerobic conditions, oxidation of reduced As (e.g., arsenite) generates energy for microbes. Under anaerobic and aerobic conditions, microbes transform arsenic by methylation, oxidation, or reduction mechanisms that mobilize it away from
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microbial cells. However, microbial transformation of arsenic is not promising because this element can exist in many mobile forms. Selenium, another nonmetal, is used in a number of commercial and industrial processes (including photocopying, steel manufacturing, glass making, and semiconductor manufacturing) and is sometimes present at contaminated sites. Selenium contamination has also resulted from irrigation practices that led to the accumulation of selenium dissolved from soils. Although selenium is an important micronutrient for plants, animals, humans, and some microorganisms (largely because of its role in some key amino acids) when present at very low concentrations, it is toxic at higher concentrations. In natural environments, selenium has four inorganic species: Se(VI) (selenate, SeO42– ), Se(IV) (selenite, SeO32– ), Se(0) (elemental selenium), and Se(-II) (selenide) and exists primarily as the two soluble species, Se(VI) and Se(IV).1,61 Like arsenic, selenium also has many volatile organic forms. Reduced inorganic selenium compounds can be oxidized under aerobic conditions, although the oxidation does not support microbial growth. Oxidized selenium (selenate) can serve as a final electron acceptor for anaerobic microorganisms, resulting in production of selenide and/or elemental Se. Methylation of the various selenium compounds is a detoxification mechanism that mobilizes Se away from microbial cells, but methylselenium is mobile and highly toxic to mammals. Anaerobic microbial reduction of selenate and selenite to insoluble elemental selenium can immobilize and remove Se from aqueous solution. Nonetheless, given the complex chemical and biological processes that influence the fate of selenium and its many mobile forms, microbial reactions are not a promising means for controlling Se contamination. The speciation of Se in natural systems is dependent on the redox potential, pH, microbial interactions, solubility, complexing ability of soluble and solid ligands, and reaction kinetics. Se(VI) (selenate), the predominant water soluble Se species, mainly occurs in well aerated alkaline soils of higher redox potential, while Se(IV) selenite occurs mostly in natural systems of moderate or reduced redox potential. Although both ions are highly water soluble, the higher adsorption properties of Se(IV) make it less mobile in the subsurface than (Se(VI). Overall the redox status appears to be the most predominant controlling factor over Se speciation.61 Oxyanions: Oxyanions are water-soluble, negatively charged chemicals in which a central atom is surrounded by oxygen. Nitrate (NO3– ) is one such oxyanion. It can come from natural sources or human sources including nitrogen fertilizers. Although NO3– can occur naturally, it is a serious health concern at high concentrations because it can cause the respiratory stress disease methemoglobinemia in infants and because it can produce cancer-forming nitrosamines. The major microbial process that destroys nitrate is reduction to nitrogen gas (N2) via a process called “denitrification.” Microbes can use nitrate as a terminal electron acceptor when oxygen is not available. The denitrification process has been ongoing for millions of years and is widespread among microorganisms; it occurs reliably in every anaerobic habitat with abundant carbon and electron sources. Natural attenuation by denitrification is possible, as long as the supply rate of an electron donor is sufficient to sustain the reaction. Many organic compounds, as well as H2 and H2S, can serve as the electron donor.1
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The oxyanions chlorate (C1O3– ) and perchlorate (C1O4– ) or their precursors (chlorine dioxide, hypochlorite, and chlorite) are produced by a variety of paper manufacturing, fertilizer, water disinfection, aerospace, and defense industries. Although not naturally occurring, these highly oxidized forms of chlorine are energetically favorable electron acceptors for microorganisms. Knowledge of chlorate and perchlorate biodegradation reactions is quite limited compared to understanding of denitrification.61 However, laboratory studies using bacterial cultures and environmental samples (soil, freshwater sediments, and sewage) have shown that microorganisms can reduce perchlorate and chlorate when supplied with common electron donors (such as carbohydrates, carboxylic acids, amino acids, H2, or H2S). Reducing perchlorate and chlorate generates nontoxic chloride iron.1,61 Microbial transformation of perchlorate or chlorate is plausible if the supply rate of electron donors is adequate.62
3.5
MONITORING AND SAMPLING FOR NATURAL ATTENUATION
At long last, natural attenuation has come into its own. Over the past five years, great strides have been made in conceptualizing natural attenuation and developing protocols, field methodologies, guidance documents, and strategies for implementation. However, the most important aspect of a monitored natural attenuation (MNA) evaluation at a site is the need to collect biogeochemical and groundwater quality data of the highest quality to predict the natural attenuation capacity of the system. As typically practiced, natural attenuation studies place heavy emphasis on quantifying aqueous-phase electron acceptors, contaminants, and byproducts by sampling groundwater in monitoring wells. In response to this need, a number of companies offer multiparameter, in situ and down-hole groundwater quality field monitoring devices that can facilitate the collection of the biogeochemical information. It can be easily concluded that groundwater samples from zones in which contaminants are being naturally biodegraded are often in dramatic nonequilibrium with ambient conditions. Furthermore, contact of these samples with the atmosphere can cause significant shifts in aqueous biogeochemistry. The key to minimizing or avoiding shifts in the biogeochemistry of reduced samples, in particular, is minimizing contact with atmospheric air. Associated sampling considerations to avoid include the following: • Purging wells at a high rate may lower the water level in the monitoring well. During recharge, there is significant contact between the groundwater and the atmospheric air as the groundwater trickles into the well. • Use of a bailer for sample collection results in exposure of the sample to the atmospheric air as the sample is poured into the sample bottle. • Sample holding times, typical with many commercial laboratories, offer the opportunity for changes in the biogeochemistry of the sample. • Other than samples for volatile organic compounds (VOCs) analysis, groundwater samples are often collected in such a way that there is headspace in the sample bottle. Agitation of the sample bottle during handling and shipping may result in mixing and thus altering of biogeochemistry.
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Since many of the commonly employed groundwater sample collection techniques in the past presented varying degrees of contact between the sample and the atmosphere, an inherent concern always exists regarding the reliability and representativeness of results obtained through these sampling methods for the biogeochemical parameters of interest. During the past few years, a low-flow, minimal aeration method has evolved which produces the most representative samples for parameters particularly sensitive to artificial aeration and resulting changes in biogeochemistry. This method involves slow purge rates, a down-hole pump, a flow cell for probe measurements, and a sample-bottle filling procedure that minimizes sample aeration (Figures 3.17 and 3.18). Various studies have reported that the lowflow, minimal aeration method using the Grundfos pump produces the most representative results for most parameters. Minor biases in the results of methane, dissolved iron (Fe2+), and sulfide are possible. Many devices have been developed to collect data in situ (down-hole); consequently, data errors related to sampling artifacts associated with above-ground data collection and sequential parameter measurements can be avoided. Furthermore, because many of these units can be coupled to automatic data recorders that provide for immediate data collection and storage in the field and subsequent data transfer to personal computers in the laboratory, errors related to data transcriptions can also be eliminated. Meter 3-Way Valve Probe Flow Cell Probe Measurement Device Fill to Overflowing With Discharge End of Tube Fully Submerged
Valve for Additional Regulation of Pump Discharge Rate Flexible Tube
Beaker (for Probe Measurement) or Sample Bottle
Water Table Monitoring Well (2"Ø or Greater)
Slow Purge Rate to Minimize Water Table Drawdown
Submersible Pump
Figure 3.17
Schematic of minimal aeration, low-flow groundwater sampling technique.
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Atmosphere O2 = 21% CO2 = 0.03% CH4 = 0%
O2
(Eh ) Anaerobic, Reducing Groundwater 0.5 DO Fe2+ = 10 - 50 mg/L CH4 = 2 - 20 mg/L Alkalinity = ± 500 mg/L
Figures 3.18
CO2
CH4
Fe2+
Fe(OH)3 Precipitate
Geochemical consequences due to atmospheric interferences during sampling.
Field portable meters capable of measuring (DO) concentrations are available from a variety of manufacturers. These instruments can record DO levels in fresh water or saltwater and most are equipped to make temperature and salinity corrections. Oxidation/reduction potential (ORP, Eh, or REDOX) can be difficult to measure even with the best available instrumentation. The sensing device (most often a platinum electrode in a circuit with a standard reference electrode) may be unstable in fresh waters with low ionic strength. The time required to obtain a stable reading may be quite long in some cases. Although it is possible to measure REDOX in the field, considerable operator skill and experience are necessary to obtain accurate results. Two types of field measurements for DO and REDOX are possible with the current generation of water quality instrumentation: on-site and in situ. On-site refers to measurements in which a water sample is removed from the aquifer or body of water and a sensor immediately placed in it for measurement. Great care is taken to isolate the sample from the atmosphere. In situ or “down-hole” sensors refer to measurements made by lowering the probe directly into the well or surface water at the desired depth. After a suitable equilibration time, continuous monitoring of water quality can be performed. Two types of on-site measurements are available: discrete sampling and flowthrough sampling. Discrete samples are collected in the appropriate sample container (e.g., 300-mL biological-oxygen-demand (BOD) bottles or other suitable glassstoppered bottles capable of preventing entrainment of atmospheric oxygen). The DO or REDOX sensor is then placed in the sample for measurement. Flow-through cells incorporate the sensor in a cell in line with a pump. DO and/or REDOX and other primary water quality parameters are continuously monitored as the water flows through the cell. The flow-through technique provides immediate results and minimizes problems resulting from the collection and transport of samples to an onsite laboratory or measurement station.
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Two types of in situ measurements are available: short-term continuous monitoring and long-term continuous monitoring. Once calibrated, positioned at the desired depth, and equilibrated to the sample conditions, most probes can send continuous readings to surface instrumentation. Meters may display or log these results for a short period of time. A probe designed for long-term monitoring incorporates features to allow it to be anchored in place and operated unattended for long periods of time. Long-term monitoring can be useful in evaluating groundwater quality before and during corrective action. The membrane electrode has been used to monitor DO levels for a long time during site characterizations. Dual DO/REDOX measurements can be complementary. If REDOX measurements indicate a negative or reducing environment, the corresponding DO reading should be low (e.g., <1 mg/L). In situ DO/REDOX measurements can also be used to evaluate stratification in an aquifer. Despite their data-collection virtues, however, it is important to realize that such units, especially when they are leased, can collect inaccurate data if certain precautions and procedures are not followed prior to and during their use in the field. Therefore, the following sections provide considerations and recommendations regarding the use of such multiparameter, groundwater quality field equipment designed to maximize the accuracy of biogeochemical data collected. It is very important that the field technician has considerable experience with general field monitoring equipment and procedures to avoid human bias and that he or she has been afforded the opportunity to become familiar with the multiparameter monitoring system to be used. Equipment consideration: Many firms that lease multiparameter monitoring units will calibrate the sensing electrodes in the laboratory prior to shipment to the end user. As a consequence, the remediation engineer is typically told that field calibrations of the sensors in the unit are unnecessary. However, there are two considerations with respect to the as-received accuracy and utility of the monitoring system. First, the previous user may have subjected the equipment to severe groundwater environments, improper handling, and inadequate cleanup. If those conditions are not completely corrected by the vendor prior to shipping, erroneous readings can be collected. Second, the monitoring system may have been subjected to harsh treatment during shipping such that the calibrated function of one or more of the parameters has been seriously altered or perhaps completely eliminated. As a consequence, it is a good practice to follow the instructions specified to ensure proper system function and to maximize the probability that the key electrode sensors are indeed providing accurate field measurements. Initial system checkout: Upon receipt of the monitoring equipment, it should be carefully unpacked and inspected for signs of damage or fouling. Inspect both the meter housing for cracks or blemishes and the electrode sensing unit to verify it has been properly cleaned. Expose the electrodes in the sonde according to the manufacturer’s directions and verify that each electrode is intact and has been properly packaged for shipment. Also inspect the cable between the meter and the sonde for damage or contamination. If any evidence of equipment damage is identified, immediately contact the vendor and evaluate the need for a replacement unit.
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Otherwise, attach the cable to the meter, place the sonde in a beaker of distilled water or tap water, and turn on the system. Within moments, the unit typically performs an internal system checkout and the readouts for the different parameters should become activated. If necessary, scroll through the different parameter readouts to verify proper function and reasonable readings. If any spurious readings are observed, one or more of the electrode sensing units may be experiencing problems. Check to make sure the suspect electrode is firmly attached to the sonde. If spurious readings are still observed, contact the vendor for technical advice and evaluate the need to replace the unit. Response time is the most important quality control check to be made in the field for DO/REDOX systems. It is used to determine the condition of the electrodes, especially those in water with high contaminant concentrations, dissolved inorganic salts, sediments, and other insoluble material. Fouling of the sensing electrodes is the most likely cause of errors in measurement of DO/REDOX. 3.5.1
Dissolved Oxygen (DO)
Electrode inspection and membrane replacement: The user should inspect the membrane covering the DO electrode in the sonde for any bubbles that may have formed beneath the membrane during shipment. Although an unlikely development with the current generation of DO electrodes, if any bubbles are present beneath the membrane, spurious DO readings are likely to result. Consequently, if any bubbles are observed, the membrane must be replaced using the materials and instructions provided in the field kit supplied with the multiparameter unit. Once a membrane has been properly replaced, the electrode should be immersed in distilled water for a minimum of five minutes prior to field use to allow the new membrane to equilibrate. After the membrane has been checked, verify the proper function of the DO sensing system by turning on the meter and placing the sonde in a beaker of aerated, room temperature (∼20°C) tap water and observing the DO readout on the meter. After about two minutes, the DO reading at sea level should be about 9 mg/L. Refer to the table provided in Appendix B to determine the appropriate DO reading for the altitude of interest and the current barometric pressure. Probes designed to detect DO consist of reference and sensing electrodes immersed in supporting electrolyte and separated from the sample solution by a selective membrane (Figure 3.19). The oxygen sensor consists of the membrane and a closely fitted electrode. The sensing electrode is considered the cathode where molecular oxygen is reduced, and the reference electrode is considered the anode. Only species that can permeate the membrane and are reduced at the sensing electrode will produce a signal, resulting in the highly selective and sensitive nature of the DO sensor. The cell current is linearly proportional to the DO concentration and can be converted to concentration by simple calibration procedures. Several types of DO sensor designs are available; the most commonly used electrode is the polarographic (commonly called the standard) electrode. This probe utilizes an applied potential to reduce molecular oxygen and requires circulation of the water being analyzed. If the water is not moving at about one foot per second, an error in DO concentration will result. The error is caused by the buildup of a
Figure 3.19
Permeable Membrane
Permeable Membrane
Principle of measurement of dissolved oxygen.
Cathode (Gold)
Cathode (Silver)
Polarographic Electrode
KCI Electrolyte
Alkaline Electrolyte
114
DO Cell of Galvanic Type
Anode (Lead)
e - (Applied Voltage)
Anode (Lead)
e - (Spontaneous Reaction)
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concentration gradient in the vicinity of the DO membrane as oxygen is consumed by the sensor. Circulation of the water replenishes the sample near the sensor. The galvanic cell (Figure 3.19) is less commonly used for detecting DO. The voltage detected by this type of probe is produced by the spontaneous reduction of molecular oxygen at the cathode (analogous to a fuel cell). Because less oxygen is consumed from the sample, this cell is less sensitive to low water flow. For both types of sensors, molecular oxygen is reduced by a noble metal cathode fitted closely to the permeable membrane. Chemical changes that occur in the electrolyte, and on the surface of the electrode as a result of the chemical reactions in the cell, will eventually necessitate cleaning the electrode surfaces and changing the electrolyte. There is a third type of DO sensor which reportedly consumes no oxygen from the sample. Field Calibrations: On the day of the site visit, obtain the current site specific barometric pressure and altitude. Upon arrival at the site, the monitoring unit should be activated and allowed to warm up for at least five minutes before any field calibrations are attempted. For the highest level of accuracy in DO measurements, the multiprobe system should be calibrated at the ambient groundwater temperature. This can be accomplished as follows: Lower the sonde down a monitoring well or a measuring cell that is uncontaminated and upgradient from the other monitoring wells of interest. Avoid aerating the groundwater surface during sonde passage into the groundwater. Monitor the temperature function on the meter and determine when the sonde has attained thermal equilibrium with the groundwater. This may take over 90 seconds. Remove the sonde and immediately calibrate the DO electrode using the standard procedure provided in the operation manual. Because the sonde will tend to maintain the ambient groundwater temperature during calibration, DO calibration will have been essentially accomplished at the ambient groundwater temperature. Remember to adjust the calibration procedure to account for the current barometric pressure and the site’s altitude. For DO calibrations at groundwater sites where low DO conditions are anticipated, the meter and sensing unit should be calibrated at two end points. The first end point, the higher-end calibrated value, is obtained by following the procedures provided in the operation manual. The method provides the high-end value because it calibrates the instrument under oxygen-saturated conditions. To obtain the low-end calibration value, three different methods are recommended: • For the first method, add one tablespoon of sodium sulfite (Na2SO3) crystals to a beaker containing 300 ml of tap water. After three minutes, lower the sonde in the beaker (making sure that the DO electrode is immersed in the solution) and observe the DO reading. It should descend to 0 mg/L within three minutes. If not, adjust the meter to read 0 mg/L. After calibration, rinse the electrode with distilled water. • For the second method, add a cube or packet of baker’s yeast to a beaker containing 300 ml of tap water and allow the yeast to dissolve completely. Lower the DO electrode into the solution and wait three minutes. Again, if the DO reading does
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not descend to zero after three minutes, adjust the meter to read 0 mg/L. After calibration, rinse the electrode as before. • For the third method, bring a small bottle of pressurized nitrogen gas to the field. Attach a small flexible hose to the outlet of the pressure regulator on the nitrogen bottle and slightly open the regulator. This will allow a slow flow of nitrogen gas to exit the tubing. Orient the tubing to within two centimeters of the oxygen electrode to deliver the nitrogen gas directly to the surface of the oxygen electrode membrane. After three minutes, the DO reading on the meter should decline to 0 mg/L. If it does not, adjust the meter to read 0 mg/L.
Regardless of which method is used to obtain the low-end calibration value, the calibration should be conducted immediately after the high-end value has been obtained at the ambient groundwater temperature. If no adjustment to the meter was required to obtain the low-end calibration value, the DO function of the meter is now calibrated. If a meter adjustment was required for the low-end reading, immediately repeat the high-end calibration step under ambient groundwater temperature conditions. The DO function of the meter should now be properly calibrated. Field Measurements: The sonde should be slowly lowered into the groundwater of the well until the DO and other parameter sensing devices are immersed in the groundwater. Again, avoid aerating the groundwater during sonde passage into the groundwater. The sonde’s electrodes should be placed at either: 1) the midpoint of the well screen or, 2) in cases where the water column in the well is shorter than the length of the well screen, the midpoint of the available water column in the well. The appropriate depth should be determined by knowing the depths of the upper and lower ends of the well screen and by obtaining the depth to groundwater in the well on the day of sampling. The DO measurement should be recorded when the reading has stabilized, generally within two minutes. During the collection of field measurements in multiple groundwater monitoring wells, the DO membrane should be checked after each well monitoring event for bubbles or organic fouling due to the presence of nonaqueous phase liquids (NAPLs) or microbial “slimes” in the groundwater in the monitoring well. If bubbles are encountered or the membrane is severely fouled with an organic film or microbial slime, the membrane should be replaced prior to further use. If only light organic fouling is observed, however, the membrane may be cleaned by immersing the sonde in a dilute detergent solution and “swishing” the electrode around in the solution. For electrodes lightly fouled with microbial slimes, the electrodes can be disinfected using a 10% solution of denatured alcohol in distilled water. Once it has been verified that the membrane has been cleaned, it should be rinsed with distilled water. Regardless of the presence or absence of organic films, however, for most groundwater monitoring circumstances, it is good practice to decontaminate the sonde and any cabling exposed to groundwater. Decontaminate using appropriate, standard decontamination procedures prior to the using the equipment in each subsequent wells.
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117
Oxidation–Reduction (REDOX) Potential (ORP)
The oxidation-reduction potential of groundwater is a measure of electron activity and an indicator of the relative tendency of a biogeochemical system to accept or transfer electrons. In a very oxidizing environment the activity of electrons is low and in a very reducing environment the activity of electrons is high. The electron activity is characterized by using the lowercase “p” notation (just as in pH). pe = – log [e–]
(3.19)
The pe of a surface water at pH 7, in equilibrium with atmospheric oxygen, is calculated to be 13.6; it decreases to approximately 4 in an environment where Fe3+ reduction takes place and drops to approximately –4 where sulfate reduction and methanogenic conditions exist. pe can be calculated from the measured concentrations of reactants and products in a REDOX half-reaction (Figure 3.20). A scale equivalent to pe is the Eh scale, which is expressed in volts and is based on the determination of electron activity using electrochemical methods. At temperatures typical of the natural environments: Eh =
pe 16.7
(3.20)
where Eh is in volts. Eh is often confused with the closely related REDOX potential or ORP measurement. ORP or REDOX is measured by placing a REDOX electrode into the water sample; the REDOX electrode is a piece of metallic platinum, which acquires a more negative potential with respect to its reference electrode under reducing conditions where electron activities are higher (Figure 3.21). ORP is the voltage measured between this REDOX electrode and the reference electrode placed in the same environment. If the activities of the oxidizing species are greater than those of the reducing species, a voltage greater than the reference electrode voltage will result. Although REDOX is temperature dependent, temperature corrections are rarely performed because of the lack of theoretical knowledge involving the exact nature of the active species in the sample. ORP provides a useful, approximate characterization of REDOX conditions in the aquatic invironment, although it lacks precise theoretical definition. Although ORP and Eh are both measured in volts (millivolts) and do show some rough correlation, they are defined quite differently and should not be treated as synonymous. Electrode inspection: The surface of the oxidation-reduction potential (ORP) electrode should be inspected and it should be verified that the surface is clean of any organic or inorganic films. If it appears to be fouled, it can be gently cleaned using a slightly abrasive cloth dipped in a mild detergent solution as described previously. After cleaning, it should be rinsed with distilled water.
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5
0
0
-300
-5
O2
NO -3
MnO 2
Fe(OH)3
2SO4
Methanogenesis
300
Sulfate Reduction
10 Iron Reduction
600
Approximate range due in part to variability in reaction/product concentrations
Manganese Reduction
15
Denitrification
900
Aerobic Respiration
Eh(mV)
118
CO2
Oxidant in Use Figure 3.20
The REDOX sequence of different electron acceptor reactions catalyzed by microorganisms.
Laboratory reference checks: ORP sensing systems tend to be quite robust with respect to maintaining calibration and are typically calibrated by the vendor prior to shipment to the user. The ORP electrode tends to respond to a linear fashion over a large range of ORP conditions at groundwater sites. However, where strongly negative ORP readings are possible, it may be advisable to verify the proper operation of the ORP function by conducting high-end and low-end ORP reference checks. In those cases, the proper operation of ORP electrode systems can be evaluated upon receipt of the monitoring unit using the following two-point reference check method: 1. For high-end reference value, obtain the “ZoBell” solution from the vendor or prepare 125 ml of the solution using the method supplied in many chemical texts such as Standard Methods. Regardless of whether you purchase or prepare the solution, make sure you obtain a copy of the ORP temperature table from the vendor. Immerse the sonde’s ORP electrode in the solution and observe the ORP reading on the meter. At 25°C, the ORP reading should be +237mV.
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Figure 3.21
119
Principle of measurement REDOX.
2. At low-end reference value evaluation can be obtained by using the sodium sulfite solution prepared for the low-end DO calibration method described previously. Once the sonde has been immersed in the sodium sulfite solution, the ORP value should be negative.
If the value observed at the high-end check is considerably different from the expected value or if the value obtained during the low-end evaluation is positive, the ORP function may be damaged and the vendor should be contacted for further guidance or perhaps for meter and/or sonde replacement. Field measurement: Once the sonde has been lowered to the appropriate depth in the monitoring well, the ORP measurement can be recorded after it has stabilized, typically after 90 seconds. 3.5.3
pH
The acidity of a given sample is determined by the concentration of hydrogen ions present. This concentration is expressed by the pH. The value of pH is usually expressed by the equation: pH = – log [H+]
(3.21)
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Although this equation is written in terms of concentration, in fact the quantities in the brackets should be activity, which may be thought of as “corrected” concentrations that take into account nonideal effects in aqueous systems. These nonideal effects arise from electrostatic forces between ions dissolved in the water, and they increase as the total concentration of dissolved ions, measured as ionic strength increases. Electrode Inspection: The pH electrode should be inspected and it should be verified that any organic or inorganic films are absent. If the electrode appears to be fouled, the sensing region of the electrode can be gently cleaned using a nonabrasive cloth dipped in a mild detergent solution as described previously. After cleaning, the electrode should be rinsed with distilled water. Laboratory Calibrations: There are several standards available for calibrating a pH electrode. A three-point calibration should be performed using standards anticipated to bracket the range of pH levels expected to be observed in the field. For example, for most groundwater situations, calibrations using standard solutions of pH = 4, 7, and 10 would be appropriate. Calibrate the pH function using the pH = 7 standard first. At sites where alkaline conditions are anticipated, calibrate the pH electrode with the pH = 10 standard last. At sites where acidic conditions are anticipated, calibrate the electrode with the pH = 4 standard last. Once the calibrations have been completed, rinse the pH electrode with distilled water. The proper operation and calibration of the pH function has now been verified. Field Measurements: pH measurements can be recorded once the pH reading has been stabilized, typically within 60 seconds. 3.5.4
Filtered vs. Unfiltered Samples for Metals
One of the most hotly contested protocols in the last few years is whether to fieldfilter samples that are to be analyzed for dissolved metals. While partisans commonly make the case either for always filtering or always not filtering, two essential factors lead inevitably to the conclusion that filtering is technically appropriate in most cases and not in others: • Metal concentration data may be used for many purposes, and the data use dictates whether filtering is appropriate. For example, if direct ingestion from a drinkingwater source is involved, data from an unfiltered sample is appropriate. Alternatively, for assessment of contaminant removal during a remediation project, filtered samples may be appropriate. • Different hydrogeologic conditions may cause groundwater samples to be turbid, no matter what well installation, development, purging, or sampling methods are used. As indicated above, turbid water may give metal concentrations that are not indicative of the concentrations actually moving in the aquifer and would be biased high relative to actual dissolved metals transport.
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3.5.4.1 Field Filtration and the Nature of Groundwater Particulates Monitoring wells are purged and samples are brought to the surface by pumps of many types or by bailer. The types of sampling devices and their advantages and disadvantages are discussed in numerous research papers and regulatory guidance manuals. For a given well sampling event, one device may be used for purging stagnant water from the well, while another may be used to take the sample. Filtration is most often conducted on samples to be analyzed for metals content. The distinction is commonly made between a sample to be analyzed for total metals content (an unfiltered sample) and a sample to be analyzed for dissolved metals content (a filtered sample). As discussed later, the presence of very small particles called colloids adds substantial complexity to the discussion of total vs. dissolved metals content. Many of these colloidal particles typically will not settle out due to gravitational forces. Many types of filters are available with regard to both the type of filter apparatus and the pore size of the filter. Filtration can occur in the open: a water sample is poured into a filter funnel and the sample is pulled by vacuum through a membrane. Alternatively, the filtration can occur in a closed system where the sample is pushed by the sampling pump through its discharge tube and finally through an “in-line” filter unit. In either the open or closed system, the filtrate, once through the filter, is directed into the sample bottle for preservation (usually with nitric acid for metals) and shipment for analysis. A wide range of filter pore sizes is available. Since the late 1970s, the most standard nominal pore diameter used in the environmental business has been 0.45 micrometers (µm). Other common filters that have been used in environmental research and for special applications have pore diameters ranging from 5 µm at the high end to 0.1 and 0.03 µm at the low end. Historically, the groundwater system was viewed as having two parts: the immobile rock/soil phase and the mobile water phase. However, in the recent past, emphasis has been placed on another partially mobile phase — colloids — in evaluating contaminant transport. Colloids are very small organic or inorganic particles that can range from less than 0.1 to 10 µm in diameter. The truly dissolved phase has molecules or polymers that are substantially smaller than 0.1 µm. Colloid composition, charge, and aquifer conditions vary considerably through space and time, but research has shown that particles up to about 2 µm in diameter can move with groundwater.63 Particles at the higher end of the colloid size range may be trapped by small pore spaces in the soil matrix or settle out due to gravity. Colloidal particles may be present naturally in groundwater or may be released from soil or rock surfaces during well installation or sampling or may be deposited at the bottom of the well due to precipitation of the dissolved metals in the well volume itself. Even the action of a bailer in a well can substantially increase the concentration of colloids in a groundwater sample. If the objective of an investigation is to assess the movement of metals in the subsurface, field filtration will
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remove some intermediate-to-large size colloids that are not mobile under natural conditions, but have been mobilized near a well due to well installation or sampling disturbance. Of special importance is that the standard environmental filter has a pore size (0.45 µm) that is in the middle of the size range where colloids can be expected to move in the subsurface. This filter does not provide a clean cutoff between dissolved and colloidal species that are mobile and colloids and larger particles that are not. As a matter of fact, no single filter pore size could provide the correct cutoff between mobile and immobile species because aquifer, well, and sampling conditions and circumstances vary widely. Short of making an intensive research effort for each project site, until recently there has been no way to accurately distinguish mobile from immobile particulates. As discussed later, recent research and experience with low-flow sampling has provided a means, under some circumstances, of retrieving samples with only dissolved, naturally occurring, and mobile colloids.64,65 Before the installation of numerous monitoring wells at contaminated sites during the past 20 years, the thinking about groundwater samples was shaped largely by our experiences with production wells. These wells were installed in “true” aquifers that generally, in the case of unconsolidated formations, were composed of relatively coarse-grained, well-sorted materials with high hydraulic conductivities. These wells could be readily developed to yield water with low turbidity, so filtration was not usually needed to obtain a water sample considered representative of formation water. In addition, filtration to remove any particles was not desirable because water from such a well was expected to be directly consumed. Filtration can be considered undesirable because the extra handling in the field can alter the sample’s chemistry. Aside from the intended removal of particulates larger than the pore size of the filter, dissolved or colloidal chemicals in the sample can adsorb onto the filter membrane or apparatus and the filter pore size may be altered by particle clogging. Alternatively, ions associated with the filtration system can leach into the sample, thereby giving false positive results. Finally, the exposure of a sample to air during filtration can cause metals such as iron to oxidize and precipitate. These iron precipitates may be stopped by the filter, thereby lowering the iron concentration in the sample filtrate. The precipitation may also entrain other metals, resulting again in lowered concentrations in the sample. 3.5.4.2 Reasons for Field Filtration Many monitored sites are located over aquifers with low hydraulic conductivities. By definition, an aquifer is a formation that contains sufficient saturated permeable material to yield sufficient quantities of water to wells and springs. Although the concepts “low hydraulic conductivity” and “aquifer” seem mutually exclusive, as a regulatory matter, the “uppermost aquifer” is frequently the most contaminated and requires intense monitoring. Low hydraulic conductivity formations screened by monitoring wells may be uniformly fine grained, such as in a silty clay unit. Alternatively, a lower permeable unit to be monitored may be characterized by a wide array of grain sizes; a common example of such poorly sorted material is glacial
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till. In either case, fine-grained materials of clay or colloid size (particle diameter of less than 2 µm) will be present. Monitored formations with low hydraulic conductivities present numerous challenges with respect to an excess of colloids being present in a groundwater sample. First, such fine-grained formations intrinsically contain numerous small particles. Second, because of the difficulty of making water flow through the formation, removal of colloidal particles in the zone around the well through well development is difficult. The colloid load present in the sample can still be larger than what is in the formation, even after a substantial well development effort has been made and the well has been sampled numerous times. In contrast, production wells are developed continuously as large volumes of water are pumped for long periods of time. Finally, in low permeability materials, water levels are substantially drawn down to the point that a well may have little or no remaining water in it, even at very low pumping rates. The drawdown causes a high groundwater gradient near the well, with the attendant increase in groundwater flow rate in individual pore spaces. The action of purging and sampling may easily mobilize particles that would not normally be moving with the groundwater. It should be noted that the expected shearing of colloidal particles from large soil particle surfaces is proportional to the square of the groundwater pore velocity. In many geologic units, large amounts of iron oxyhydroxides are present. If the REDOX conditions are oxidizing, these oxyhydroxides are in the low solubility ferric form and are present as coatings on the surfaces of large, immobile aquifer solids like sand and grains. At low or near neutral pH values, these ferric oxyhydroxide coatings are positively charged and can therefore attract and pull colloidal clay particles from the groundwater solution. However, if substantial amounts of natural or contaminant organic matter are present, the REDOX conditions can be reducing, and the ferric oxyhydroxides convert to the more soluble ferrous form, which enters the groundwater phase. This combination of soluble ferrous ions and associated colloidal clay increases groundwater turbidity. Much of the regulatory literature creates the expectation that, regardless of geology, if sufficient care is taken, a well can almost always be designed, installed, and developed so that it produces water with low colloid content, and therefore, low turbidity. There is also the expectation that well purging will be conducted so that the water sample will have low turbidity. The most common criterion for acceptable well construction and development is a well that yields samples with turbidity of less than five nephelometric turbidity units (NTUs). However, some geologic matrices will yield turbid water to a monitoring well, no matter how proficient the well installation and development. As discussed more fully in the section on low-flow, the only recourse in such formations is either to purge and sample a well at a rate that does not exceed the natural recharge to the well or to use a bailer, which has its disadvantages. As a practical matter in tight formations, however, purging and sampling rates cannot be lowered to the level of natural groundwater flow. Substantial drawdown and the attendant disturbance to the formation will occur even at the lowest pumping rate achievable with current technology. Therefore, if low-flow rate sampling is not
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possible, the next best procedure is to sample at a higher flow rate followed by sample filtration. Hydrogeologic and biogeochemical conditions can vary through time at a monitoring location. In addition, sampling method and operator technique can change during the history of a monitoring program. As a result, the turbidity and metal concentrations in a series of samples can vary above and beyond the real changes of concentration in the aquifer. At wells where it is difficult to control the turbidity and colloid concentration in samples, filtration will help to level out the variations. Frequently, the ability to track changes in concentration through time is an important objective of any remediation project. More consistent data will increase the power of statistical analyses such as trend evaluations. 3.5.5
Low-Flow Sampling as a Paradigm for Filtration
The desire to disturb the aquifer as little as possible has led to research and the rapid acceptance of minimum aeration, low-flow sampling of wells. Research in the late 1980s and 1990s has shown that, in many geologic formations, groundwater moves horizontally through the screened portion of the well and interacts little with the water standing in the well above the screened zone.64 If the sample is pumped from the well at a rate that is less than or equal to the natural flow rate through the screen, no stagnant water from above the screen zone will be present in the sample. In contrast to traditional well purging techniques with pumping rates of 4 liters per minute or more, low-flow purging and sampling occurs typically in the range of 0.1 to 1 L/min. Current pump technology allows pumping rates down to about 0.1 L/min; for wells that yield less than this amount, the low-flow methodology is not as applicable. The insertion of a bailer or sampling pump into a well causes substantial disturbance to the water in and around the well. Bailing is particularly troublesome in this regard; comparison studies show that bailed samples have much higher turbidity and order of magnitude or more higher-metals concentrations in samples collected via low-flow pumping rates. As a result, it is essential that low-flow equipment is used during sampling of MNA parameters. Low-flow sampling with dedicated pumping equipment has proven to be an important advance in groundwater sampling. Research and practice indicate that low-flow samples have lower turbidity and are more representative of formation water than samples obtained at higher flow rates with bailers or pumps. The lack of particulates has been demonstrated by showing little change in analyte concentration when low-flow samples are filtered with membranes ranging in pore size from 0.03 to 5 µm in one study65 and 0.1 to 10µm in another.64 In conclusion, filtered samples, even those taken with a bailer, have similar analyte concentrations to unfiltered samples taken by the low-flow methodology, which is state of the art with respect to collecting samples considered representative of aquifer conditions.
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125
A Comparison Study
During the mid-to late 1990s, filtered and unfiltered pairs of samples were collected at two study sites by ARCADIS Geraghty & Miller, Inc. The wells and temporary well points were mostly installed in the shallow, unconsolidated zone, characterized chiefly by fill (cinders and a variety of other natural and artificial materials), till, and meadow mat (high natural organic matter). Lithologic logs for many locations describe the presence of fine-grained materials. REDOX measurements were taken, and because of the presence of dissolved organic matter (both natural and contaminants), conditions were reducing at a large number of the wells in portions of both sites. The presence of fine-grained materials and low REDOX values enhanced the formation of colloids in groundwater. As part of this study, ARCADIS Geraghty & Miller, Inc. reviewed the 1266 pairs (i.e., each location multiplied by the number of analytes) of filtered and unfiltered groundwater sample results. The ratios of filtered (dissolved) to unfiltered (total) concentrations varied widely from the dissolved concentrations being less than 1% of the total concentration to cases where the dissolved concentrations were greater than the total concentration. Of special note is that dissolved concentrations of only certain metals were frequently low (10% or less) compared to total concentrations. Aluminum, iron, copper, chromium, lead, vanadium, and zinc fell into this category. This study showed that filtered, bailer or high-flow samples give results most similar to the unfiltered low-flow samples, which are presumably most representative of undisturbed formation water. A detailed evaluation of the analytical data and geologic conditions revealed that samples with lower values of dissolved concentrations had the following characteristics: • • • •
Higher organic matter concentrations Low REDOX potentials and reducing conditions Fine-grained particles in the lithology Very low yield to the point where they can be pumped or bailed dry
As discussed in detail, filtration has both good and bad aspects. The improved accuracy and consistency derived from the removal of the large, normally immobile particles liberated during well installation and sampling outweighs the loss of some of the smaller particles yielding better overall results. An important argument for filtration is that the sample results from location to location and sampling event to sampling event will be more consistent with filtering. Turbidity and colloid load will vary between samples; by reducing the importance of this variable through filtration, spatial and temporal relationships will emerge more quickly. It is these relationships that help to determine the location of a source area, the effectiveness of a remedial action, or the behavior of dissolved constituents over time. Perhaps the most powerful argument for filtration arises from the results of lowflow sampling. When wells are pumped at no more than the rate of natural recharge
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to the well, analyte concentrations in samples are considered to be the most representative of formation water quality. In cases where low-flow pumping is not practical, the closest results are achieved by filtering samples obtained by bailer or highflow pumps. Unfiltered samples obtained by these methods would be expected to have a substantially high bias. In selected cases where there are data collection objectives different from general site characterization and remediation planning, there may be reason to collect unfiltered samples as well as filtered samples. Furthermore, dedicated, low-flow sampling equipment may be considered at specific locations if frequent sampling is indicated, well yield is adequate, and there is no expectation that data from these wells will need to be compared to data from wells without such equipment.
REFERENCES 1. National Research Council, Natural Attenuation for Groundwater Remediation, National Academy Press, Washington, D.C., 2000. 2. United States Environmental Protection Agency, Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action and Underground Storage Tank Sites, Directive Number 9200.4-17P, Office of Solid Waste and Emergency Response, Washington, D.C., 1999. 3. American Society for Testing and Materials (ASTM), Standard Guide for Remediation of Groundwater by Natural Attenuation at Petroleum Release Sites; Draft, Philadelphia, 1997. 4. Wiedermeier, T. et al., Technical Protocol for Implementing Intrinsic Remediation with Long Term Monitoring for Natural Attenuation of Fuel Contamination Dissolved in Groundwater, Vols. 1 and 2, Air Force Center for Environmental Excellence, Brooks Air Force Base, San Antonio, TX, 1995. 5. U.S. Army Science Board, Infrastructure and Environment Issue Group, Remediation of Contaminated Army Sites: Utility of Natural Attenuation, Draft Report, Department of the Army, Washington, D.C., 1995. 6. Department of Energy (DOE), Final Programmatic Environmental Impact Statement for the Uranium Mill Tailings Remedial Action Ground Water Project, DOE/E150198, NTIS, Springfield, VA, 1998. 7. Chapelle, F. H. and Paul M. Bradley, Selecting remediation goals by assessing the natural attenuation capacity of groundwater systems, Bioremediation J., 2, 227–238, 1998. 8. Potter, S., personal communication, 1999. 9. Wiedemeier, T. H. et al., Natural Attenuation of Fuels and Chlorinated Solvents in the Subsurface, John Wiley & Sons, Inc., New York, 1999. 9a. Fetter, C. W., Contaminant Hydrogeology, Macmillan Publishing, New York, 1993. 10. Hemond, H. F. and E. J. Fechner, Chemical Fate and Transport in the Environment, Academic Press, New York, 1994. 11. Pankow, J. F. and J. A. Cherry, Dense Chlorinated Solvents and Other DNAPLs in Groundwater, Waterloo Press, Portland, OR, 1996. 12. Kram, M. L. et al., Comparison of DNAPL characterization approaches, in Treating Dense Non Aqueous Phase Liquids (DNAPLs), Int. Conf. Remed. Chlorinated and Recalcitrant Compds., Monterey, CA, May 22–25, 2000.
L1282_C03-B_frame Page 127 Monday, June 18, 2001 10:07 AM
MONITORED NATURAL ATTENUATION
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13. Perry, J. J., Microbial cooxidations involving hydrocarbons, Microbiol. Rev., March 1979, 59–72. 14. Alexander, M., Biodegradation and Bioremediation, 2nd ed., Academic Press, New York, 1994. 15. Buscheck, T. E. and G. M. Alcantar, Regression techniques and analytical solutions to demonstrate intrinsic bioremediation, In Situ and On Site Bioremed. Symp. Proc., San Diego, CA, 1995, 109–116. 16. Borden, R. C. and Bedient, P. B., Transport of dissolved hydrocarbons influenced by oxygen limited biodegradation: theoretical development, Water Resour. Res., 22, No. 13, 1973–1982, 1996. 17. Kennedy, L. G. et al., Iron and sulfur mineral analysis methods for natural attenuation assessments, Bioremed. J., Vol. 2, No. 3, 259–276, 1998. 18. Baedecker, M. J. et al., The fate and effects of crude oil in a shallow aquifer: the distribution of chemical species and geochemical facios, US Geo. Surv. Water Resour. Div. Rep. 88–4220, USGS, Reston, Virginia, 1989. 19. Baedecker, M. J. et al., Crude oil in a shallow sand and gravel aquifer: biogeochemical reackons and mass balance modeling in anoxic groundwater, Appl. Geochem., 8, 569–586, 1993. 20. Lovley, D. R. et al., Oxidation of aromatic contaminants coupled to microbial iron reduction, Nature, 339, 297–299, 1989. 21. Chappelle, F. H. et al., Measuring rates of biodegradation in a petroleum hydrocarbon contaminated aquifer, Groundwater, 34, 691–698, 1996. 22. Reinhard, M. et al., In situ BTEX biotransformation under enhanced nitrate and sulfate reducing conditions, Environ. Sci. Technol., 31, 28–36, 1997. 23. Kirtland, B. C., C. Marjorie Aelion, and P. A. Stone, Monitoring anaerobic natural attenuation of petroleum using a novel in situ respiration method in low-permeability sediments, Bioremed. J., 4, 181–187, 2000. 24. Weiner, J. M., T. S. Lauck, and D. R. Lovely, Enhanced anaerobic benzene degradation with the addition of sulfate, Bioremed. J., 4, 159–174, 1998. 25. Spain, J., Organic chemicals with potential for natural attenuation, Bioremed. J., 1, 1–10, 1997. 26. McNally, D. L. et al., Biodegradation of three and four ring polycyclic aromatic hydrocarbons under aerobic and denitrifying conditions, Environ. Sci. Technol., 32, 2633–2639, 1998. 27. Mihelcic, J. R. and R. G. Luthy, Degradation of polycyclic aromatic hydrocarbon compounds under various redox conditions in soil-water systems, J. Appl. Environ. Microbiol., 54, 1182–1187, 1988. 28. Coates, J. D. et al., Anaerobic degradation of polycyclic aromatic hydrocarbons and alkanes in petroleum contaminated marine harbor sediments, J. Appl. Environ. Microbiol., 63, 3589–3593, 1997. 29. Zhang, X. and L. Y. Young, Carboxylation as an initial reaction in the anaerobic metabolism of naphthalene and phenanthrene by sulfidogenic bacteria, J. Appl. Environ. Microbiol., 63, 4759–4764, 1997. 30. Wilson, J. T., J. S. Cho, B. H. Wilson, and J. A. Vardy, National Attenuation of MtBE in the Subsurface Under Methnogenic Conditions, USEPA/600/R-00/006, Washington, D.C., 2000. 31. Salanitro, J. P. et al., Field scale demonstration of enhanced MtBE bioremediation through aquifer bioaugmentation and oxygenation, Environ. Sci. Technol., 34, 4152–4162, 2000.
L1282_C03-B_frame Page 128 Tuesday, June 19, 2001 1:01 PM
128
NATURAL AND ENHANCED REMEDIATION SYSTEMS
32. Bradley, P. M. et al., Aerobic mineralization of MtBE and TBA by stream bed sediment microorganisms, Environ. Sci. Technol., 33, 1877–1879, 1999. 33. Borden, R. et al., Intrinsic biodegradation of MtBE and BTEX in a gasoline contaminated aquifer, Water Resour. Res., 33, 1105–1115, 1997. 34. Schirmer, M. and J. Barker, A study of long term MtBE attenuation in the Borden aquifer, Ontario, Canada, Groundwater Monitor. Remed., 113–122, Spring, 1998. 35. Kolhatkar, R., J. T. Wilson, and L. E. Dunlap, Evaluating Natural Biodegradation of MtBE at Multiple UST Sites, paper presented at the API Petroleum Conf., Anaheim, CA, November, 2000. 36. Bouner, E. J., B. E. Rittmann, and P. L. McCarty, Anaerobic degradation of halogenated 1- and 2-carbon organic compounds, Environ. Sci. Technol., 15, 596–599, 1981. 37. Ramanand, K. et al., Reductive dehalogenation of chlorinated benzenes and toluenes under methanogenic conditions, J. Appl. Environ. Microbiol., 59, 3266–3272, 1993. 38. Reineke, W. and H. J. Knackmuss, Microbial metabolism of haloaromatics: isolation and properties of a chlorobenzene degrading bacterium, J. Appl. Environ. Microbiol., 47, 395–402, 1984. 39. Nishino, S. F., J. C. Spain, and C. A. Pettigrew, Biodegradation of chlorobenzene by indigenous bacteria, Environ. Toxicol. Chem., 13, 871–877, 1994. 40. Cozza, C. L. and S. L. Woods, Reductive dechlorination pathways for substituted benzenes: A correlation with electronic properties, Biodegradation, 2, 265–278, 1992. 41. Halden, R. U. and D. F. Dwyer, Biodegradation of dioxin related compounds, a review, Bioremed. J., 1, 11–25, 1997. 42. Mohn, W. W. and J. W. Tiedje, Microbial reductive dehalogenation, Microbiol. Rev., 56, 482–507, 1992. 43. Haggblom, M. M. and R. J. Valo, Bioremediataion of chlorophenol wastes, in Microbial Transformation and Degradation of Toxic Organic Chemicals, Young, L. Y. and Cerniglia, C. E., Eds., Wiley-Liss, New York, 1995, 435–486. 44. Suflita, J. M. and G. T. Townsend, The microbial ecology and physiology of aryl dehalogenation reactions and implications for bioremediation, in Microbial Transformation and Degradation of Toxic Organic Chemicals, Young, L.Y. and Cerniglia, C. E., Eds., Wiley-Liss, New York, 1995, 243–268. 45. Luthy, R. G. et al., Aqueous solubility of PCB congeners from an Aroclor and an Aroclor – hydraulic oil mixture, Water Res., 31, 561–573, 1997. 46. Bedard, D. L. and J. F. Quensen, Microbial reductive dechlorination polychlorinated biphenyls, in Microbial Transformation and Degradation of Toxic Organics, Young, L. Y. and Cerniglia, C. E., Eds., Wiley-Liss, New York, 1995, 127–216. 47. Bedard, D. L. et al., Stimulation of microbial para-dechlorination of polychlorinated biphenyls that have persisted in Housatonic River sediment for decades, Environ. Sci. Technol., 30, 687–694, 1996. 48. Flanagan, W. P. and R. J. May, Metabolite detection as evidence for naturally occurring aerobic PCB biodegradation in Hudson River sediments, Environ. Sci. Technol., 27, 2207–2212, 1993. 49. Spain, J. C., Biodegradation of nitroaromatic compounds, Ann. Rev. Microbiol., 49, 523–555, 1995. 50. Spain, J. C., Ed., Biodegradation of Nitraromatic Compounds, Plenum Publishing Corp., New York, 1995. 51. Bradley, P. M. et al., Potential for intrinsic bioremediation of a DNT-contaminated aquifer, Ground Water, 35, 12–17, 1997. 52. Funk, S. B. et al., Full scale anaerobic bioremediation of trinitrotoluene (TNT) contaminated soil, Appl. Biochem. Biotechnol., 51–52, 625–633, 1995.
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53. Binks, P. R. et al., Degradation of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) by Stenotrophomonas, Int. Conf. Remed. Chlorinated and Recalcitrant Compds., Monterey, CA, May 22–25, 2000. 54. French, C. E. et al., Aerobic degradation of 2,4,6-trinitrotoluene by Enterobacter cloacae PB2 and by pentaerythritol tetranitrate reductase, J. Appl. Environ. Microbiol., 64, 2864–2868, 1998. 55. White, G. F. and J. R. Snape, Microbial cleavage of nitrate esters: defusing the environment, J. Gen. Microbiol., 139, 1947–1957, 1993. 56. White, G. F. et al., Biodegradation of glycerol trinitrate and pentaerythritol tetranitrate by Agrobacterium radiobacter, J. Appl. Environ. Microbiol., 62, 637–642, 1996. 57. Topp, E. et al., Isolation and characterization of an N-methylcarbamate insecticide degrading methylotrophic bacterium, J. Appl. Environ. Microbiol., 59, 3339–3349, 1993. 58. Siddartha, K. S. et al., Degradation of alpha-, beta-, and gamma- hexachlorocyclohexane by a soil bacterium under anaerobic conditions, J. Appl. Environ. Microbiol., 56, 3620–3622, 1990. 59. Cookson, J. T., Bioremediation Engineering: Design and Application, McGraw-Hill, New York, 1994. 60. Chen, S. and D. B. Wilson, Genetic engineering of bacteria and their potential for Hg2+ bioremediation, Biodegradation, 8, 97–103, 1997. 61. Dungan, R. S. and W. T. Frankenberger, Microbial transformations of selenium and the bioremediation of seleniferous environments, Bioremed. J., 3, 171–188, 1999. 62. Logan, B. E., A review of chlorate and perchlorate respiring microorganisms, Bioremed. J., 2, 69–80, 1998. 63. Puls, R. W. and R. M. Powell, Transport of inorganic colloids through natural aquifer material: Implicaitons for contaminant transport, Environ. Sci. Technol., 26, 614–621, 1992. 64. Puls, R. W. and M. J. Barcelona, Groundwater Sampling for Metals Analyses, EPA/540/4-89/001, VSEPA, 1989. 65. Puls, R. W. and R. M. Powell, Acquisition of representative groundwater quality samples for metals, Groundwater Monitor. Rev., 167–176, Summer, 1992.
CREDIT Materials from pages 100 through 104 contain substantial excerpts from Spain, J., Synthetic chemicals with potential for natural attenuation, Bioremediation Journal, 1(1): 1–9, 1997.
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CHAPTER
4
In Situ Reactive Zones CONTENTS 4.1 4.2
Introduction ..................................................................................................132 Engineered Anaerobic Systems ...................................................................135 4.2.1 Enhanced Reductive Dechlorination (ERD) Systems .....................135 4.2.1.1 Early Evidence..................................................................135 4.2.1.1.1 Biostimulation vs. Bioaugmentation...............136 4.2.1.2 Mechanisms of Reductive Dechlorination .....................138 4.2.1.3 Microbiology of Reductive Dechlorination ...................142 4.2.1.3.1 Cometabolic Dechlorination .........................142 4.2.1.3.2 Dechlorination by Halorespiring Microorganisms ..........................................144 4.2.1.4 Electron Donors ..........................................................147 4.2.1.4.1 Production of H2 by Fermentation ................149 4.2.1.4.2 Competition for H2 ......................................152 4.2.1.5 Mixture of Compounds on Kinetics.................................155 4.2.1.6 Temperature Effects ....................................................158 4.2.1.7 Anaerobic Oxidation ...................................................158 4.2.1.8 Electron Acceptors and Nutrients .................................158 4.2.1.9 Field Implementation of IRZ for Enhanced Reductive Dechlorination ............................................................160 4.2.1.10 Lessons Learned .........................................................163 4.2.1.11 Derivation of a Completely Mixed System for Groundwater Solute Transport of Chlorinated Ethenes ..170 4.2.1.12 IRZ Performance Data......................................................177 4.2.2 In Situ Metals Precipitation .............................................................183 4.2.2.1 Principles of Heavy Metals Precipitation.........................187 4.2.2.2 Aquifer Parameters and Transport Mechanisms ..............195 4.2.2.3 Contaminant Removal Mechanisms.................................196
131
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4.2.3 In Situ Denitrification.......................................................................197 4.2.4 Perchlorate Reduction ......................................................................199 4.3 Engineered Aerobic Systems .......................................................................200 4.3.1 Direct Aerobic Oxidation.................................................................200 4.3.1.1 Aerobic Cometabolic Oxidation.......................................202 4.3.1.2 MTBE Degradation ..........................................................204 4.4 In Situ Chemical Oxidation Systems...........................................................205 4.4.1 Advantages .......................................................................................206 4.4.2 Concerns...........................................................................................207 4.4.3 Oxidation Chemistry ........................................................................208 4.4.3.1 Hydrogen Peroxide ...........................................................211 4.4.3.2 Potassium Permanganate ..................................................213 4.4.3.3 Ozone ................................................................................216 4.4.4 Application .......................................................................................218 4.4.4.1 Oxidation of 1,4-Dioxane by Ozone................................222 4.4.4.2 Biodegradation Enhanced by Chemical Oxidation Pretreatment.....................................................223 4.5 Nano-Scale Fe (0) Colloid Injection within an IRZ ...................................223 4.5.1 Production of Nano-Scale Iron Particles .........................................228 4.5.2 Injection of Nano-Scale Particles in Permeable Sediments............231 4.5.3 Organic Contaminants Treatable by Fe (0) .....................................231 References ...................................................................................................233
Oxidation-reduction process plays a major role in the mobility, transport, and fate of inorganic and organic contaminants in natural waters. Hence, the manipulation of REDOX conditions to create an in situ reactive zone (IRZ) to meet the cleanup objectives was a predictable evolution … .
4.1
INTRODUCTION
The concept of in situ reactive zones is based on the creation of a subsurface zone, where migrating contaminants are intercepted and permanently immobilized or degraded into harmless end products. Figures 4.1a and b pictorially describe the concept of in situ reactive zones (IRZ). The successful design of these reactive zones requires the ability to engineer two sets of reactions: 1) between the injected reagents and the migrating contaminants; and 2) between the injected reagents and the subsurface environment to manipulate the bio-geo-chemistry to optimize the required reactions, in order to effect remediation. These interactions will be different at each contaminated site and, in fact, may vary within a given site. Thus, the major challenge is to design an engineered system for the systematic control of these reactions under the naturally variable or heterogeneous conditions found in the field. The effectiveness of the reactive zone is determined largely by the relationship between the kinetics of the target reactions and the rate at which the mass flux of contaminants passes through it with the moving groundwater. Creation of a spatially
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Source Area IRZ Grid
Contaminant Plume
Individual Reactive Zones Created by Individual Injection Points Providing a Collective In Situ Reactive Zone (IRZ) Curtain
Plan View
Figure 4.1a
Pictorial depiction of an in situ reactive zone (IRZ) formation.
Reagent
Contaminant Zone
Cross Sectional View Figure 4.1b
Cross sectional view of the creation of an IRZ around an individual injection well at a selected location.
fixed reactive zone in an aquifer requires not only the proper selection of the reagents, but also the proper mixing of the injected reagents uniformly within the reactive zone. Furthermore, such reagents must cause few side reactions and be relatively nontoxic in both its original and treated forms. When dealing with dissolved inorganic contaminants such as heavy metals, the process sequence in a pump and treat system required to remove the dissolved heavy metals present in the groundwater becomes very complex, operation- and maintenance-intensive, and costly. In addition, the disposal of the metallic sludge, in most cases as a hazardous waste, is also very cost prohibitive. Therefore, in situ treatment methods capable of achieving the same mass removal reactions for dissolved contaminants in an in situ environment are evolving and gradually gaining prominence in the remediation industry. The advantages of an in situ reactive zone to address the remediation of groundwater contamination are as follows: • An in situ technology enables implementation of most ground treatment processes and eliminates the expensive infrastructure required for a pump and treat system with no disposal of water or wastes
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• Inexpensive installation because primary capital expenditure for this technology is the installation of injection wells at appropriate locations • Inexpensive operation that allows inexpensive reagents to be injected at fairly low concentrations and, hence, should result in insignificant cost; only sampling required is for groundwater quality monitoring and performance monitoring parameters are usually done in the field; remediation of large volumes of contaminated water without any pumping or disposal needs • Can be used to remediate deep sites because cluster injection wells or in-well mixing systems can be installed to address deeper sites • Unobtrusive because once the system is installed, site development and operations can continue with minimal obstructions • In situ degradation of contaminants because organic contaminants and a few inorganics such as NH4+ , NO3– , and CIO4– can be degraded by implementing the appropriate reactions • Immobilization of contaminants because once the dissolved heavy metals are precipitated out, the capacity of the soils and sediments is utilized to adsorb, filter out, and retain inorganic contaminants
Manipulation of the reduction-oxidation (REDOX) potential of an aquifer is a viable approach for in situ remediation of REDOX-sensitive groundwater contaminants. In addition, various microbially induced or chemically induced reactions also can be achieved in an in situ environment. As noted earlier, creation of spatially fixed reactive zones to achieve these reactions is very cost effective in comparison to treating the entire plume as a reaction zone. Since the first IRZ for the precipitation and remediation of hexavalent chromium (Cr6+), was installed in 1993, this technology has advanced by leaps and bounds.1 Currently the application of this technology can be classified into three categories based on the creation of specific bio-geo-chemical and REDOX environments: 1) engineered anaerobic systems, 2) engineered aerobic systems, and 3) in situ chemical oxidation. The engineered anaerobic systems can be further divided into enhanced reductive dechlorination (ERD) systems, in situ denitrification, in situ perchlorate transformation, and in situ heavy metals precipitation. The ERD application has been expanded to many contaminants since the first trichloroethene (TCE) application site. The IRZ technology has been successfully applied to remediate the following chlorinated compounds: • Chlorinated ethenes: tetrochloroethane (PCE), trichloroethane (TCE), dichloroethene (Cis 1,2 DCE, and 1,1 DCE), vinylchloride • Chlorinated ethanes: 1,1,2,2 tetrachloroethane (1,1,2,2 PCA), 1,1,1 trichloroethane, (1,1,1 TCA), 1,1,2 trichloroethane (1,1,2 TCA), 1,1 and 1,2 dichloroethane (DCA), chloroethane (CA) • Chlorinated phenols: pentachlorophenol (PCP), and tetrachlorophenol • Chlorinated pesticides • Perchlorate
In addition, the IRZ technology has been successfully applied to precipitate the following dissolved metals at contaminated sites: Cr6+, Pb2+, Cd2+, Ni2+, Zn2+, Hg2+.
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ENGINEERED ANAEROBIC SYSTEMS
Enhanced Reductive Dechlorination (ERD) Systems
4.2.1.1 Early Evidence The first microbially mediated reductive dechlorination of PCE and TCE was observed in the early 1980s, and this study2,3 reported the degradation of PCE to nonchlorinated end products in an acetate-fed, continuous-flow methanogenic glass bead column. It appeared that the first step in the degradation pathway was dechlorination to TCE. Further anaerobic oxidation of TCE to carbon dioxide and hydrochloric acid was suggested. In 19844, further evidence of dechlorination of PCE beyond TCE came in an experiment where sediments from an aquifer recharge basin were incubated with PCE and methanol as the electron donor. Significant concentrations of TCE, cis1,2 DCE and VC were observed after three weeks, whereas in sterile controls no dechlorination had occurred. Another study in the 1980s demonstrated that dechlorination of PCE to VC in a methanogenic column was achievable.5 Similar studies using 13C-TCE, showed that TCE was dechlorinated exclusively to cis-DCE in soil.6 In 1989, the first evidence of complete dechlorination of PCE to ethene under methanogenic conditions with methanol as electron donor was demonstrated.7 Another study found PCE reduction via ethene to ethane with lactate as electron donor in a flow-through column filled with a mixture of polluted sediment and anaerobic granular sludge.8 Meanwhile, numerous publications showed that microorganisms capable of reductively dechlorinating chlorinated ethenes are abundant in polluted anaerobic environments. (An overview of the biological reductive dechlorination pathway of chlorinated solvents is shown in Figure 4.2.) PCE and TCE are dechlorinated mainly to cis-DCE, although sometimes trans-DCE and 1,1-DCE have also been found as products.9,10 However, the formation of the 1,1-DCE is believed to be a result of abiotic dechlorination in the presence of sulfide.10 Evidence from the earlier studies indicated that the dechlorination of PCE to cis-DCE was found to be a relatively fast process, whereas, subsequent rates of dechlorination of cis-DCE to VC and ethene were significantly slower or even absent.7,11 Dechlorination of 1,1-DCE and trans-DCE was less studied. In some of the earlier reports and studies the dechlorination of chlorinated ethenes was often found to be incomplete, both in the laboratory and in field experiments, resulting in an accumulation of cis-DCE and VC. It was not fully understood at that time why dechlorination beyond these compounds was problematic, other than raising valid questions regarding the required microbial consortia for complete dechlorination. During that time (late 1980s and early 1990s) microorganisms capable of dechlorinating DCE and VC had not been isolated yet, although several enrichment cultures existed. Little was known about the substrate requirements of these bacteria. Later studies reported that PCE dechlorination in a contaminated soil down to ethene was only achieved by adding a complex mixture of organic electron donors. Significant research was focused, during the early to mid 1990s, on the microbial ecology that could perform complete dechlorination of PCE to
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PCE
CT
TCE
cis-1,2 DCE*
1,1- DCE
VC
Ethene
1,1,1-TCE
CF
1,1- DCA
DCM
CA
CM
Acetic Acid
Ethane CO2, H2O, CI-
Biotic Reactions
Figure 4.2
Abiotic Reactions
* Primary Reaction
Biological and abiotic degradation pathways of the common chlorinated compounds encountered at contaminated sites (adapted from McCarty and Semprini, 1994; after Vogel et al., 1987, and Wiedermeier et al., 1999).
ethene and the biogeochemical conditions under which this biotransformation could be achieved. The choice of a suitable electron donor for the stimulation of in situ dechlorination is still a matter of discussion and may be dependent on local conditions; this will be discussed in detail in a later section. When hydrogen is assumed to be the major electron donor for dechlorination, its amendment can only be achieved by using substrates yielding hydrogen after anaerobic degradation.12 Often, short-chain organic acids are produced as intermediate products, which may lead to acidification of the groundwater and soil. Additionally, electron donors that support dechlorination are generally readily degraded by nondechlorinating microorganisms, leading to competition for the substrate and excessive bacterial growth in soil pores near the injection well. As a result, significantly more electron donor mass will be needed than theoretically necessary to reduce all chlorinated ethenes present to ethene. 4.2.1.1.1 Biostimulation vs. Bioaugmentation The first level of the treatment hierarchy for chlorinated ethenes is intrinsic bioremediation, or natural biodegradation, whereby indigenous microflora destroy the contaminant(s) of concern without any stimulation or enhancements. The second choice in this hierarchy, biostimulation or enhanced biodegradation, involves stimulating the indigenous microbial populations and thus enhancing microbial activity
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so that they destroy the target compounds at a rate that meets the cleanup objectives at the site. At almost every contaminated site a natural population of degradative microorganisms exists within the contaminated zone; however, specific nutrients, growth substrates inducers, electron donors, and electron acceptors may be required to create optimal microbial activity.12 Thus, through the introduction of required additional reagents, the native degradative microbial population can be stimulated to grow, multiply, and destroy the target contaminants. Most environments contain microorganisms able to grow on and destroy a variety of chlorinated compounds; at some sites, the persistence of these compounds, is not a consequence of the absence of organisms but rather of the absence of the full set of conditions necessary for the indigenous species to function rapidly.12 In the past there was a significant debate among remediation experts whether the microorganisms responsible for cometabolic degradation and dehalorespiration are ubiquitous. Current belief is that these organisms are nearly ubiquitous. When intrinsic bioremediation or biostimulation is not feasible at a given site due to the absence of an appropriate microbial population, bioaugmentation may be utilized. Bioaugmentation involves injection of selected exogenous microorganisms with the desired metabolic capabilities directly into the contaminated zones along with any required nutrients to effect the rapid biodegradation of target compounds. Two distinct bioaugmentation approaches have been developed for remediating chlorinated ethenes. In the first approach, degradative organisms are added to complement or replace the native microbial population. The added microorganisms can be selected for their ability to survive for extended periods or to occupy a specific niche within the contaminated environment. If needed, stimulants or selective cosubstrates can be added to improve survival or enhance the activity of the added organism. Thus, the goal of this approach is to achieve prolonged survival and growth of the added organisms and degradation of the target contaminants. In the second bioaugmentation approach, large numbers of degradative bacteria are added to a contaminated environment as biocatalysts which will degrade a significant amount of the target contaminant before becoming inactive or perishing.12 Additional microbes can be added as needed to complete the remediation process. Attempts can be made to increase the production of the degradative enzymes or to maximize catalytic efficiency or stability, but long-term survival, growth, and establishment of an active microbial population are not the primary goals of this treatment approach. In the past, bioaugmentation has been implemented frequently and successfully only in bioreactors. The conditions in these bioreactors are controlled and quite different from those in nature, and prior to start-up, no microorganisms are present anyway. Hence, the addition of enriched cultures is essential. Furthermore, bioreactors are engineered and controlled systems where conditions can be readily altered or optimized for a particular process and can be designed to promote the multiplication and activity of the inoculated species — in contrast to contaminated field sites. The record of success of in situ bioaugmentation systems for chlorinated compounds has been rather spotty. On the one hand, the initiation or enhancement of degradation has been reported (far more commonly in samples of the contaminated environments in simulated laboratory experiments) following the addition of
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enriched bacterial cultures that can metabolize and grow on chlorinated ethenes. On the other hand, a number of failures in the field have been reported. Such reports of failure of bioaugmentation came as no surprise to microbial ecologists. Without question, a species with a substrate uniquely available to it has a distinct advantage, yet that advantage may not be sufficient to compensate for many other traits also necessary for survival, no less multiplication, in a natural ecosystem. Possessing the requisite enzymes to metabolize a novel compound is a necessary attribute for the organism, but it is not sufficient for the organism to succeed. Populations of introduced microorganisms are subject to a variety of abiotic and biotic stresses, and these must be overcome for these organisms to be able to express beneficial traits. The reasons for the frequent failures of bioaugmentation are many:12 limiting nutrients and growth factors in the uncontrolled natural environment, suppression by predators and parasites, inability of the introduced bacteria to penetrate significant space, metabolism of other nontarget organic compounds present, concentration of the target chlorinated compound too low to support multiplication, and other inhibitory biogeochemical conditions such as pH, temperature, salinity, and toxins. In summary, the problems usually encountered in scaling up the bioaugmentation successes achieved in laboratory experiments can be summarized as follows:12 • Contaminant rates established in controlled laboratory studies may differ substantially from those in pilot-scale, full-scale, or even other laboratory studies. • Positive biotransformation results from small systems often are not reproduced in different systems. • Instantaneous biotransformation rates vary widely and in an apparently stochastic manner, even in well-operated, steady-state systems.
4.2.1.2 Mechanisms of Reductive Dechlorination Naturally occurring biological processes can degrade organic contaminants in situ or during transport in the subsurface under aerobic and/or anaerobic conditions. Microorganisms catalyze degradation reactions to obtain energy for growth, reproduction, and cell maintenance. Useable energy is recovered through a series of REDOX reactions where the microorganisms act as “electron transport mediators” (Figure 4.3). Biologically mediated electron transfer couples the oxidation of an electron donor (organic compound) with the reduction of an electron acceptor (inorganic or organic) and results in the production of useable energy for microbial consortia.12,13,14 The bulk electron donor acts as a fuel source for the reactions and the reactions proceed as long as there is a source of bioavailable electrons. Fuel sources can be the target chlorinated compounds, native organic carbon, co-contaminants such as fuel hydrocarbons, or organic compounds such as carbohydrates. In aerobic environments, the chlorinated compounds act as electron donors and under anaerobic conditions they act as electron acceptors. There are two primary mechanisms involved in the biodegradation of chlorinated organic contaminants (Table 4.1). First, biodegradation may be growth-linked and provide carbon and energy to support growth when the compound is used as primary
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(mediator)ox
(bulk)ox
+ ne-
+ ne-
(Contaminant)ox
- ne-
+ ne-
(mediator)red
(bulk)red Figure 4.3
139
(Contaminant)red
Description of microorganisms acting as electron transport mediators (after Schwarzenbach et al., 1993).
substrate and directly utilized by the mediating organisms via the processes included in Category 1. Some chlorinated solvents are used as electron donors and some are used as electron acceptors when serving as primary growth substrates. When used as an electron donor (under aerobic and anaerobic conditions) the contaminant is oxidized. Conversely, when used as an electron acceptor, the contaminant is reduced via the reductive dechlorination process called halorespiration.17 Table 4.1 Summary of the Categories of Degradation Pathways for Chlorinated Organic Compounds (Adapted from Wiedemeier et al., 1999)14 Category 1 (used as primary substrate)
Chlorinated Compound Tetrachloroethylene (PCE) Trichloroethylene (TCE) Dichloroethene (DCE) Vinyl Chloride (VC) Trichchloroethane (1,1,1 TCA) Dichloroethane (1,2 DCA) Carbontetrachloride (CT) Methylenechloride (MC)
Category 2 (used as cometabolic substrate) Anaerobic Direct Direct Aerobic Cometabolism HaloAerobic Anaerobic Cometabolism (reductive respiration Oxidation Oxidation (co-oxidation) dechlorination) X
X
X
X
X
X
X
X
X
X
X X
X
X
X X
X X
X
X
X
X
X
X X
X
X
In addition to their use as a primary growth substrate, chlorinated solvents can also be degraded via cometabolic pathways. During cometabolism, microorganisms gain carbon and energy for growth from metabolism of a primary substrate, and chlorinated solvents are degraded fortuitously by enzymes present in the metabolic
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pathways. Cometabolism is a process where the organism receives no direct benefit from the degradation of the organic compound.13,16 There are two types of cometabolic reactions: co-oxidation and reductive dechlorination, described as Category 2 in Table 4.1. Cometabolic reactions tend to be incomplete and can possibly lead to an accumulation of more toxic daughter products. To date, vinyl chloride (VC) and dichloroethene (cis/trans) are the only chlorinated solvents that can be degraded by all aerobic and anaerobic pathways.15 The predominant mechanism for the biodegradation of chlorinated solvents in anaerobic environments is reductive dechlorination, whether the organic compound is a primary electron acceptor (halorespiration) or is cometabolized. Before 1994, reductive dechlorination was thought to be strictly a cometabolic process because the organisms that cause these reactions are ubiquitous at most contaminated sites.14,15 However, research has shown that combetabolic reductive dechlorination is “sufficiently slow and frequently incomplete.”15,18 During reductive dechlorination, the chlorinated solvents act as an electron acceptor and a chlorine atom is replaced with a hydrogen atom (Figure 4.4). Cometabolic reduction of the chlorinated solvents is catalyzed by the reductive dehalogenase and reductase enzymes produced by microorganisms.14,20 Cometabolic degradation occurs under iron reducing, manganese reducing, sulfate reducing, and methanogenic environments.21 The enzymes of these reducing microorganisms are induced to reduce abiotic forms of Fe (III) to Fe (II), Mn (IV) to Mn (II), sulphate to sulfide or hydrogen sulfide, and carbon dioxide to methane. Electrons are transferred to dissolved contaminants coincidentally during the reducing processes. These degradation reactions are often incomplete, resulting in an accumulation of toxic daughter products. Just as aerobic biodegradation systems utilize oxygen as a terminal electron acceptor to stimulate microbial activity, oxidative anaerobic systems require other terminal electron acceptors, such as nitrate or ferric iron (Fe III), to stimulate biodegradation. Anaerobic oxidation occurs when anaerobic bacteria use the chlorinated contaminant as the electron donor and, in most instances, allow the microorganism to derive useful amounts of energy from the reaction. It has been shown that vinyl chloride can be oxidized to carbon dioxide, water, and chloride ion via Fe (III) reduction.22 Significant anaerobic mineralization of DCE, VC, and methylene chloride also have been reported in the literature. While in oxidative anaerobic systems the contaminant is used as an electron donor, in reductive systems highly oxidized contaminants (such as PCE) are used as electron acceptors. The process begins by supplying excess reduced substrate (electron donor) to a microbial consortium, i.e., a cooperative community of microbial species (Figures 4.3 and 4.5). The presence of the substrate expedites the exhaustion of any naturally occurring electron acceptors. As the natural electron acceptors are depleted, microorganisms capable of discharging electrons to other available electron acceptors, such as oxidized contaminants, gain a selective advantage. The intricacies of these microbial communities are complex, but recent research has provided some insight into methods for enhancing populations of contaminantdegrading microorganisms.
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Electron Flow
Perchloroethene CI
e-
CI C
H2
C
CI
CI
Trichloroethene CI
CI C
CI C CI
CI
CI
C
C H
+ H ion
H
(Predominant Biological Reaction) cis-1,2,-Dichloroethene
H
-
C
CI
(Limited Biological Reaction) 1,1-Dichloroethene
+CI
(Limited Biological Reaction) trans-1,2,-Dichloroethene H
C
C H
H
CI
CI
C H
Vinyl Chloride H
CI C
C
H
H
Ethene h
H C
C H
H
Ethane H C
H H
Figure 4.4
H H
C H
Hydrogenolysis reactions of PCE during reductive dechlorination with H2 acting as the electron donor and the chlorinated compounds acting as electron acceptors (adapted from Vogel et al., 1987, and Wiedermeier et al., 1999).
The reductive dechlorination of PCE to ethene proceeds through a series of hydrogenolysis reactions (Figure 4.4). Each reaction becomes progressively more difficult to carry out; subsequently, the DCEs, particularly cis-DCE, and vinyl chloride (VC), tend to accumulate in anaerobic environments under natural conditions due to the absence of sufficiently reducing conditions.
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Reducing Conditions Electron Acceptor Processes
ORP
Hydrogen Generation
Electron Donors
Dissolved Organic Compounds
BTEX
Environmental Conditions
Temperature pH
Fermentable Substrates
Anaerobic Conditions
Reductive Dechlorination
Figure 4.5
Pictorial description of the conditions which control reductive dechlorination.
The oxidation-REDOX potential (ORP) affects the thermodynamics of reductive dechlorination. Microorganisms will facilitate only those oxidation-reduction reactions that have a net yield of energy. For reductive dechlorination to be thermodynamically favorable the REDOX potential must be sufficiently low, thereby excluding the presence of oxygen and nitrate as terminal electron acceptors. Furthermore, the presence of nitrate may have an inhibitory effect on PCE dechlorination.23 The REDOX potential range for reductive dechlorination is shown in Figure 4.6. It is important to note that the values of Eh ranges shown in Figure 4.6 and the values of ORP measured in the field by remediation engineers are not the same. Both parameters have some correlation and do not represent the same conditions. Figure 4.5 summarizes the mechanisms and the required environmental condition for the degradation of chlorinated solvents. 4.2.1.3 Microbiology of Reductive Dechlorination 4.2.1.3.1 Cometabolic Dechlorination A cometabolic process is defined here as a process in which the compound of interest (e.g., PCE) is converted by a biological enzyme system or cofactor in which the compound does not serve as a source of carbon or energy. Pure Microbial Cultures: Reductive dechlorination is the only biodegradative conversion known for PCE. This reaction can occur cometabolically or in a metabolic energy-producing reaction. In both cases, the cofactors of the enzymes involved are metal-containing porphyrins. Examples25 of acetogenic and methanogenic bacteria that dechlorinate PCE cometabolically are listed in Table 4.2. In general, acetogenic bacteria dechlorinate PCE at higher rates than methanogenic bacteria. Metal-containing cofactors have been found to catalyze the in vitro degradation of chlorinated ethenes.24,25 In general, reductive dechlorination rates decrease with a declining amount of chlorine atoms in the molecule. In vivo experiments with methanogenic and acetogenic bacteria indicate that dechlorination rates are low (0.5 to 235 nmol PCE
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1,000mV
REDOX Potential (Eh) in millivolts (mV) at pH=7 and T=25˚C
Aerobic
500mV
NO3- Reduction Mn4+ Reduction Anaerobic
0mV Fe3+ Reduction
Possible Range of Reductive Dechlorination
SO42- Reduction Methanogenic
Optimum Range for Reductive Dechlorination
-500mV
Figure 4.6
Optimal range for reductive dechlorination.
Table 4.2 Examples of Acetogenic and Methanogenic Bacteria that Dechlorinate PCE Cometabolic Dechlorination
Cosubstrate
Methanosarcina sp. Methanosarcina mazei Sporomusa ovata Acetobacterium woodii
Methanol Methanol Methanol Fructose
(mg protein)–1 day–1), compared with those of halorespiring bacteria25,26 (Table 4.3). In vivo usually only one halogen atom is removed. An exception is the reductive dehalogenation of dibromoethene by Methanobacterium and Methanococcus that yields acetylene as a product.27 However, in many studies the possible formation of nonchlorinated products during dechlorination reactions was not included in the carbon balance. Complete dechlorination of PCE to ethene by pure cultures of acetogenic or methanogenic bacteria has not been observed. This is in contrast to
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Table 4.3 Examples of Halorespiring Bacteria Halorespiration Dehalobacter restrictus Dehalospirillum multivorans Desulfitobacterium sp. strain PCE1 Desulfitobacterium sp. strain TCE1 Strain MS-1
Electron Donor H2 Pyruvate Lactate Lactate Yeast extract
findings with mixed anaerobic cultures in which more extensive dechlorination has been observed.7,8,27-29 The latter may be due to the interactions between different microorganisms. Sometimes, however, it is difficult to distinguish between cometabolic and specific dechlorination in these mixed cultures, and often it is not even clear which microorganisms are responsible for the dechlorination. The most often observed degradation pathway of PCE is via reductive dechlorination to cis-DCE.10,25,30-32 Several dechlorination rates for chlorinated ethenes have been reported in the literature, but it is difficult to compare the data because often the numbers of bacteria involved were not known. Nevertheless, it can be stated that dechlorination rates in mixed cultures are generally higher than those found for single acetogenic or methanogenic strains. There are a few reports on the degradation of PCE by granular methanogenic sludge from upflow anaerobic sludge blanket reactors. This sludge is enriched with acetogenic and methanogenic bacteria and contains high concentrations of cofactors. Such bacterial consortia, therefore, are suitable as a source of cometabolic dechlorinating activity. In one of these reactors, fed with a mixture of sucrose, lactic acid, propionic acid, and methanol as primary substrates, granular sludge showed a fast adaptation to high PCE concentrations. Influent concentrations of 360 to 420 µM PCE were completely dechlorinated to ethene.26,33 Average removal rates of 7.6 µmol (g VSS)–1 day–1 were achieved, with a maximum removal rate of 28.3 µmol (g VSS)–1 day–1. A bacterial consortium in a similar reactor operated in batch mode converted PCE to TCE, cis- and trans-DCE and traces of 1,1-DCE with ethanol as the primary substrate. 4.2.1.3.2 Dechlorination by Halorespiring Microorganisms Halorespiration is a type of anaerobic respiration in which a chlorinated compound is used as a terminal electron acceptor. In this reductive dechlorination process, which enables the conservation of energy via electron transport phosphorylation, one or more chlorine atoms are removed and replaced by hydrogen. Examples of halorespiring bacteria species are shown in Table 4.3. Halorespiration, also referred to as dehalorespiration, occurs when the organic compound acts as an electron acceptor (primary growth substrate) during reductive dechlorination. During halorespiration, the chlorinated organic compounds are used directly by microorganisms (termed halorespirators), such as an electron acceptor while dissolved hydrogen serves as an electron donor:15,34 H2 + C – Cl ⇒ C – H + H+ + Cl–
(4.1)
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where C – Cl represents the chlorine bond to the carbon in the chlorinated ethene molecule. Halorespiration occurs as a two-step process which results in the interspecies hydrogen transfer by two distinct strains of bacteria. In the first step, bacteria ferment organic compounds to produce hydrogen. During primary or secondary fermentation, the organic compounds are transformed to compounds such as acetate, water, carbon dioxide, and dissolved hydrogen. Fermentation substrates are either biodegradable, nonchlorinated contaminants (i.e., BTEX — benzene, toluene, ethyl benzene, and xylenes — or sugar) or naturally occurring organic carbon. In the second step, the nonfermenting microbial consortia utilize the hydrogen produced by fermentation for halorespiration.35,36 Denitrifiers, iron reducers, sulfate reducers, methanogens, and halorespirators can all utilize hydrogen as an electron donor.14 Figure 4.7 shows which reducing environment is favored depending on the hydrogen concentration. Although compounds produced during fermentation and hydrogen have been demonstrated to drive halorespiration,32 hydrogen appears to be the most important electron donor for this process.36.37 Halorespiration has been found to be limited if available nutrients are not present. Direct injection of H2 is able to serve as an electron donor for reductive dechlorination of PCE to VC and eventually to ethene in cultures provided with the proper nutritional supplements.34,38 Because reductive dechlorination of chlorinated ethenes is a reductive process, microorganisms may exist that can use chlorinated compounds as a terminal electron acceptor and possibly conserve the concomitant energy gain into ATP. This hypothesis, developed in the early 1990s, proved to be true.25,26 The first evidence that bacteria exist that can couple reductive dechlorination of PCE to growth (halorespiration) under strict anaerobic conditions was presented in the early 1990s.25,39 A highly purified enrichment culture able to grow by the reduction of PCE to cis-DCE using hydrogen as the electron donor was described. The dechlorinating organism, later designated Dehalobacter restrictus, uses only hydrogen as the electron donor and can couple growth to the reduction of PCE or TCE to cis-DCE. A recent study25,40 described a new isolate, strain TEA, which is closely related to Dehalobacter restrictus. Another strict anaerobic bacterium, Dehalospirillum multivorans, capable of coupling dehalogenation of PCE to growth, was identified and described recently.25,41 This bacterium is less restricted concerning both electron donors and acceptors. Several dechlorinating strains belonging to the genus Desulfitobacterium were isolated from different sources. The strictly anaerobic D. dehalogenans, able to grow by the reductive dechlorination of chlorinated phenolic compounds was isolated recently. Another Desulfitobacterium, strain PCE1, was isolated from polluted soil and is reported to couple the reduction of both chlorinated phenols and chlorinated ethenes to growth.43 This strain dechlorinates PCE only to TCE, whereas other known halorespiring microorganisms dechlorinate PCE further. The same authors also described a Desulfitobacterium sp. strain TCE1, which is able to use several electron donors for the reduction of PCE to cis-DCE.43,44 Another researcher has described a Desulfitobacterium sp. (strain PCE-S) that converts PCE to cis-DCE.45 All isolated Desulfitobacterium strains are able to use a number of different electron donors and acceptors for growth. The nature and origin of the dechlorinating enzymes in these organisms are still unknown.
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15
Hydrogen Concentration (nM)
Possible Reactions
10
5
0 Denitrifiers
Figure 4.7
Fe(III) Halorespirators Sulfate Methanogens Reducers Reducers
Range of hydrogen concentrations for the different anaerobic metabolic pathways (after Wiedermeier et al., 1999).
A recent study described two facultative aerobic bacteria, strain MS-1 and the closely related Enterobacter agglomerans biogroup 5, which can reductively dehalogenate PCE to cis-DCE under anaerobic conditions.46 It is not clear yet whether strain MS-1 and E. agglomerans biogroup 5 are actually halorespiring organisms. Recently, an anaerobic bacterium, Desulfuromonas chloroethenica (strain TT4B), has been isolated which can not only reductively dechlorinate PCE to cisDCE with acetate as an electron donor, but also can reduce Fe (III) and polysulfide. These are unique features for PCE-dehalogenating organisms.47 All the above-mentioned organisms are only able to couple growth to the partial reduction of PCE or TCE. An exception is Dehalococcoides ethenogenes strain 195 that couples growth to rapid dehalogenation of PCE to VC, followed by a substantially slower reduction to ethene.37 Growth of this bacterium is restricted to the presence of hydrogen, which is the only electron donor supporting the dechlorination reactions. Dechlorination was only sustained by using hydrogen, acetate, vitamin B12, anaerobic digester sludge supernatant, and cell extracts from mixed cultures in the medium.
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Chloroethene Reductive Dehalogenases: The biochemistry of PCE dehalorespiration has been studied with enzymes that were purified from a reductively dechlorinating pure culture or enrichment, which is able to couple dechlorination to energy conservation (halorespiration). PCE respiration has been studied most extensively in Dehalobacter restrictus42,49 and Dehalospirillum multivorans.26,41,45 In general, a PCE respiration chain should contain an electron-donating enzyme, electron carriers, and a reductive dehalogenase as terminal reductase. Studies with D. multivorans and Desulfitobacterium strain PCE-S indicate that a proton gradient or a membrane potential may also be essential for chloroethene respiration because several ionophores have been found to inhibit dechlorination in whole cell suspensions.45,50 The nature of the electron-donating enzyme depends on the electron donor. In D. restrictus, which uses hydrogen as electron donor for PCE respiration, hydrogenase activity has been localized on the membrane, facing the outside.49 D. multivorans and Desulfitobacterium strain PCE-S are able to use several electron donors for dechlorination, and different electron-donating activities have been found.45,50 The electrons thus generated are transported to the dehalogenase via electron carriers such as quinones and cytochromes. It was demonstrated that menaquinone is involved as electron carrier for PCE respiration in D. restrictus,49 but not in D. multivorans.45,50 Cytochrome b is present in both organisms, but its involvement in PCE respiration has not been established. In contrast to the well studied PCE and TCE dechlorination, little is known about the mechanism of DCE and VC dechlorination. It was found that the enzymes catalyzing VC dechlorination in an enrichment culture are membrane bound and, in contrast to the known PCE reductase, cobalamin independent.51 It remains unclear whether this enrichment is able to use VC as terminal electron acceptor. Recently, an enzyme has been obtained from an enrichment containing D. ethenogenes that catalyzes the dechlorination of TCE to cis-DCE, VC, and ethane. This cobalamincontaining TCE-reductive dehalogenase is membrane bound and dechlorinates its substrates at similar rates, as have been reported for the PCE dehalogenases. More research is needed to know what determines the difference in substrate specificity of the cobalt-containing reductive dehalogenases. 4.2.1.4 Electron Donors The selection of an appropriate electron donor may be the most important design parameter for developing a healthy population of dechlorinating microorganisms during implementation of an IRZ for enhanced reductive dechlorination. Recent studies have indicated a prominent role for molecular hydrogen (H2) in the reductive dechlorination of chloroethenes.34,39,48 Most known dechlorinators can use H2 as an electron donor; some can use only H2. Because more complex electron donors are broken down into metabolites and residual pools of H2 by other members of the microbial community, they may also be used to support reductive dechlorination. From the small but growing pool of knowledge about dechlorinating organisms, it thus appears that H2 may serve an important role in reductive dechlorination of PCE in many environments. The author recently has observed the quick or direct transformation of PCE or TCE to ethene under very reducing conditions leading to
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Complex Organics
PCE
Ethene
4 H+ + 4CI-
H2 Dechlorinators
CO2
Methanogens
Acetic Acid Figure 4.8
Conceptual diagram of microbial activity to derive energy for growth and multiplication.
fs electrons
Carbon Substrate Cell Synthesis Reactions
fs electrons
Target Electron Donor Substrate H electrons
First Intermediate Electron Donor Substrate
fe electrons
Electron Acceptor Substrate Energy Generation Reactions
fe electrons
H electrons
fs electrons
Figure 4.9
Second Intermediate Electron Donor Substrate
fe electrons
Distribution of electrons to generation and cell synthesis during the breakdown of organic electron donors.
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speculations of the probable effect of high H2 concentrations or reductive dechlorination. In natural systems, including contaminated aquifers, most H2 becomes available to hydrogenotrophic microorganisms through the fermentation of more complex substrates by other members of the microbial consortium. The dechlorinators must then compete with other organisms, such as methanogens and sulfate-reducing bacteria, for the evolved H2 (Figure 4.8). Figure 4.9 also describes the distribution of electrons during the microbial breakdown of organic electron donor substrates. During studies in which ethanol or lactate was used to stimulate dechlorination in mixed anaerobic enrichment culture, both active dechlorination and methanogenesis at high H2 levels was observed; however, when H2 levels fell, dechlorination continued, albeit slowly, while methanogenesis ceased entirely. It was speculated that the addition of electron donors fermented only under low H2 partial pressures might give selective advantage to dechlorinators over methanogens. One school of thought in the past was that the rate and quantity of H2 made available to a degrading consortium must be carefully engineered to limit competition for hydrogen from other microbial groups, such as methanogens and sulfatereducers. Competition for H2 by methanogens is a common cause of dechlorination failure in laboratory studies. As the methanogen population increases, the portion of reducing equivalents used for dechlorination quickly drops and methane production increases.17,18,36 Speculation was that the use of slowly degrading nonmethanogenic substrates would help prevent this. Recent thinking on this issue is evolving to be different and is discussed later. Many different compounds may serve as electron donors for the reductive dechlorination of chlorinated solvents (Table 4.4). Several researchers suggest that the microbial reductive dechlorination of chlorinated ethenes depends on the presence of molecular hydrogen as the actual electron donor, either directly available or produced from other substrates by fermentation.32,34,55,56 Although this statement applies to many studies, in several cases it does not hold. Acetate, from which usually no hydrogen is produced during anaerobic metabolism, has been shown to support reductive dechlorination of chlorinated ethenes both in microcosms and environmental samples5,7,26,58 and in pure culture.47 Until recently, most research activities concerning the anaerobic degradation of chlorinated compounds focused on methanogenic systems. Such systems typically involve the introduction of a fermentable organic compound, such as acetate, lactate, hexoses (present in molasses) or even a co-contaminant such as toluene or phenol, which is fermented to produce hydrogen, among other things. It is now clear that these systems probably contained at least two distinct strains of bacteria. One strain fermented the organic carbon to produce hydrogen, and another utilized the hydrogen as an electron donor for dehalorespiration.15 Only in the last two or three years have researchers finally recognized the role of hydrogen as the electron donor in the reductive dechlorination process. 4.2.1.4.1 Production of H2 by Fermentation The production of H2 by different microorganisms is intimately linked with their respective energy metabolisms. The production of H2 is one of the specific
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Table 4.4 Electron Donors That Have Been Used to Enhance Reductive Dechlorination and Relative Costs per lb of PCE51–53 Electron Donor
Bulk Price $/lb
$/lb of PCE
Soluble (Fast Release) Donors Methanol Milk Ethanol Molasses Sugar (Corn Syrup) Sodium Lactate
0.05 0.05 0.20 – 0.25 0.20 – 0.35 0.25 – 0.30 2.20
0.64 0.18 NA 0.16 0.40 NA
Slow Release Donors Whey Edible Oils Flour (Starch) Cellulose Chitin Methyl Cellulose HRC (Regenesis Commercial Material)
0.05 0.20 – 0.50 0.30 0.40 – 0.80 2.25 – 3.00 4.00 – 5.00 12.00
0.04 NA 0.85 NA NA NA NA
NA – Not Analyzed.
mechanisms to dispose excess electrons through the activity of hydrogenase present in H2 producing microorganisms.59 All hydrogen producing microorganisms can be categorized into four groups:60 • Hetertrophic facultative anaerobes that contain cytochromes and lyse formate to produce H2 • Desulfovibrio desulfuricans, which is the only strict anaerobe in this group with a cytochrome system • Photosynthetic bacteria with light-dependent evolution of H2 from reduced NADH • Strict anaerobic heterotrophs that do not contain a cytochrome system (clostridia, micrococci, methanobacteria, etc.)
Production of H2 by obligate anaerobic microorganisms has optimum stoichiometry (1:4, with glucose as substrate) compared with facultative anaerobes (1:2), although the latter process is comparatively simpler than the former.60 Under natural conditions, fermentation is the process that generates the hydrogen used in reductive dechlorination. In the absence of externally available electron acceptors, many organisms perform internally balanced (different portions of the same substrate are oxidized and reduced) oxidation-reduction reactions of organic compounds with the release of energy; this process is called fermentation. Since only partial oxidation of the carbon atoms of the organic compound occurs, fermentation yields substantially less energy per unit of substrate compared to oxidation reactions. (Oxidation reactions are those in which external electron acceptors participate in the reaction). For instance, the fermentation of glucose to ethanol and CO2 has a theoretical energy yield of –57 k cal/mole, enough to produce about
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6 ATP. However, only 2 ATPs are produced, which implies that the organism operates at considerably less than maximum efficiency.59 In any fermentation reaction, there must be a balance between oxidation and reduction. In a number of these reactions, electron balance is maintained by the production of molecular hydrogen, H2. In H2 production, protons (H+) of the medium, derived from water, serve as electron acceptor. The energetics of hydrogen production are actually somewhat unfavorable, so that most fermentative organisms only produce a relatively small amount of hydrogen along with other fermentation products. Fermentation reactions that have pyruvate as an intermediate product have the potential of producing more H2. Conversion of pyruvate to acetyl-CoA is an oxidation process and the excess electrons generated must either be used to make a more reduced end product, or can be used in the production of H2. Fermentation by bacteria can also be important in controlling the biogeochemical environment of anaerobic aquifers. Bacterial fermentation can be divided into two categories:14,58 Primary fermentation is the fermentation of primary substrates such as sugars, amino acids, and lipids to yield acetate, formate, CO2, and H2, but also yields ethanol, lactate, succinate, propionate, and butyrate. While primary fermentation often yields H2, production of H2 is not required for these reactions to occur. Secondary fermentation or coupled fermentation is the fermentation of primary fermentation products such as ethanol, lactate, succinate, propionate, and butyrate to yield acetate, formate, H2, and CO2. Bacteria that carry out these reactions are called obligate proton reducers because the reactions must produce hydrogen in order to balance the oxidation of the carbon substrates. These secondary fermentation reactions are energetically favorable only if hydrogen concentrations are very low (10–2 to 10–4 atm or 8000 to 80 nM dissolved hydrogen, depending on the fermentation substrate). Thus these fermentation reactions occur only when the produced hydrogen is utilized by other bacteria, such as methanogens that convert H2 and CO2 into CH4 and H2O. The process by which hydrogen is produced by one strain of bacteria and utilized by another is called interspecies hydrogen transfer. It should be noted that the terminal products of anaerobic decomposition, CH4, and CO2, respectively, are the most reduced and the most oxidized carbon compounds.
There are a number of compounds besides the ones listed in Table 4.4 that can be fermented to produce hydrogen (Figure 4.10). While anaerobic degradation of BTEX compounds has been confirmed for a long time, there is still some controversy as to whether aromatic compounds (without any oxygen in the molecule) such as the BTEX compounds can be completely mineralized to CO2 without alternate electron acceptors coupled solely by fermentation with methanogenesis. Based on a number of field observations of the presence of methane, it is well known that fermentation occurs at almost all sites where BTEX compounds are present in groundwater.14,53 Since methane production requires fermentation products as methanogenic substrates, the presence of methane is clear evidence that fermentation is occurring. Metabolism of BTEX compounds to produce hydrogen probably requires the involvement of several strains of bacteria. One possible mechanism is a series of reactions, in which other electron acceptors are used by nonfermenters to break down the aromatics to simpler compounds that can be used by the fermenters.
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Figure 4.10
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Steps in the process of biodegradation of PCE by reductive dechlorination. As shown, biodegradable organic matter is required as an electron donor to initiate the process. Different types of microbes are involved at each stage. The bottom step shows that PCE must compete for electrons with sulfate, iron, and carbon dioxide, meaning that a large amount of organic electron donors may be needed to supply enough electrons. Note: CDCE = cis-dichloroethene. Source: after McCarty, 1997.
4.2.1.4.2 Competition for H2 In environments where hydrogen is the most important electron donor for dechlorination of chlorinated solvents, competition for the uptake of hydrogen between different types of microorganisms, such as methanogenic, homoacetogenic, sulfidogenic, and dechlorinating bacteria, becomes important. In several studies it has been shown that dechlorinating organisms have a higher affinity for molecular hydrogen than methanogens.27,35,55 This indicates that the dechlorinating organisms are able to survive at lower hydrogen levels, but will possibly be outcompeted by other microorganisms when elevated hydrogen levels are present. These studies suggest that a more effective dechlorination may be achieved by using an electron donor that generates low hydrogen concentrations during its fermentation, such as propionate or butyrate. The speculation is that this would then create more favorable conditions for dechlorinating bacteria than for hydrogen-consuming methanogens.27,35 Reductive dechlorination of PCE requires the addition of two electrons for each chlorine removed; for three of the seven recently identified dechlorinating organisms, H2 is one of the substrates (and in some cases, the only one) that can serve as the
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direct electron donor. Dehalobacter restrictus is another direct dechlorinator that uses only H2 as an electron donor, but dechlorinates PCE only to cis-1,2-dichloroethene (cisDCE).46,47 Dehalospirillum multivorans also dechlorinates PCE to cisDCE using H2, but has a much more widely varied biochemical repertoire: it is additionally able to use various organic substrates such as pyruvate, lactate, ethanol, formate, and glycerol as electron donors.7,37,44 Other PCE-dechlorinating organisms have been isolated that do not use H2.34,61 It was later determined that the halfvelocity constant with respect to H2 for this dechlorinator was one-tenth that of the methanogenic organisms in the culture. The threshold H2 level for dechlorination was also correspondingly lower than values typically reported for methanogens. Though confirmed thus far with only this one dechlorinator, there are thermodynamic reasons (i.e., the relatively high free energy available from dechlorination) to suspect that the threshold for H2 use by dechlorinators may generally be lower than that for hydrogenotrophic methanogens.15,27 This suggests a strategy for selective enhancement of dechlorination — managing H2 delivery so as to impart a competitive advantage to dechlorinators. Numerous microcosm and site studies have shown successful stimulation of dechlorination with substrates such as methanol, ethanol, lactate, butyrate, and benzoate.3,32,36,62,57 However, understanding the fate of the electron donors and that of the H2 evolved from their degradation, as well as the extent to which their reducing equivalents are channeled to desirable dechlorination or competing H2 sinks, has important implications for determining how best to effectively stimulate latent dechlorinating activity for in situ enhanced reductive dechlorination in an IRZ. Leading to a new school of thought, recent studies have suggested that the type of substrates and the rate of fermentation may not have an impact on reductive dechlorination. One study showed the ability of four fermentable substrates to sustain PCE dechlorination long-term (i.e., approximately four months).35 The choice of organic substrates was based upon their rates of fermentation and the H2 partial pressures that could be developed and maintained. Despite the difference in the resulting H2 partial pressures (ranging approximately 1 × 10–5 to 3 × 10–3 atm), no long-term effect on dechlorination was observed. This result may indicate either that low H2 partial pressures were not required to maintain a competitive dechlorinating community or that several isolated PCE respiring bacteria do not utilize H2 as an electron donor.43,46 H2 was not the source of PCE-reducing equivalents in all systems tested. Other laboratory and field studies have also suggested that the steady state concentration of hydrogen is controlled by the type of bacteria utilizing the hydrogen and is almost completely independent of the rate of hydrogen production. As discussed earlier, when hydrogen is produced by fermentative organisms, H2 is rapidly consumed by other bacteria. This utilization of H2 by nonfermenters is known as interspecies hydrogen transfer and is essential for fermentation reactions to proceed.59 Note, for example, that a glucose fermenter is unable to utilize glucose by itself so that both the glucose fermenter and the methanogen benefit from this symbiotic relationship. Although H2 is a waste product of fermentation, it is a highly reduced molecule, which in turn makes it an excellent, high-energy electron donor. In this symbiotic relationship, the hydrogen utilizing bacteria gain a high energy electron donor, while,
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for the fermenters, the removal of hydrogen allows continuous fermentation to be favorable energetically. In addition to methanogens, a wide variety of bacteria can utilize hydrogen as an electron donor: denitrifiers, Fe (III) reducers, sulfate reducers and halorespirators. As discussed earlier, for dechlorination to take place, halorespirators must successfully compete against all these hydrogen utilizers. It was suggested that the competition is mainly controlled by the Monod halfsaturation constant Ks (H2) (the concentration at which a specific bacterial strain can utilize hydrogen at half the maximum utilization rate).14,27,56 The measured value of Ks (H2) for halorespirators was 100 nM and for methanogens 1000 nM.14 This led to the suggestion that halorespirators would compete successfully for H2 only at low concentrations. However, a more detailed analysis of halorespiration kinetics and competition for hydrogen based on the Monod kinetic model was performed recently.14,27 Using this model, the ability of hydrogen-utilizing bacteria to compete for hydrogen can easily be predicted from substrate concentration and two properties of the bacteria, µmax (maximum specific growth rate), and Ks. Table 4.5 lists these parameters for the various hydrogen-utilizing bacteria. Table 4.5 Maximum Specific Growth Rate (µmax) and Half Saturation Coefficient (Ks) for Various H2Utilizing Bacteria (Modified from Wiedermeier et al., 1999) Bacterial Strain
µmax (hr–1)
Ks (mg/L)
Halorespirator Denitrifier Sulfate Reducer Methanogen
0.019950 0.058080 0.003936 0.003792
0.0002 — — 0.0019
Table 4.5 illustrates that, from the µmax term, halorespirators will outcompete methanogens and sulfate reducers at any hydrogen concentration (since at high substrate concentration growth rate µ ≈ µmax and at low substrate concentration µ ≈ (µmax.S)/Ks). However, denitrifiers will probably outcompete halorespirators under most conditions as their maximum specific growth rate is approximately three times faster than halorespirators’. Based on these detailed analyses and the synthesis of wide ranging data from field observations, the following probable sequence takes place at most sites undergoing halorespiration reactions:14,27 • Aerobic bacteria consume nonchlorinated organic substrates until the oxygen is depleted; to implement enhanced reductive dechlorination, oxygen depletion is forced intentionally in an IRZ. • Similarly, denitrifying bacteria will consume nonchlorinated organic substrates until the nitrate is exhausted; nitrate depletion will be forced for enhanced reductive dechlorination. • Iron reducing bacteria consume nonchlorinated organic substrates until the available Fe (III) is expended.
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• Fermentation processes consume nonchlorinated organic substrates and generate hydrogen; dechlorinating bacteria consume hydrogen for dechlorination, while sulfate-reducing bacteria consume nonchlorinated organic substrates and methanogens consume hydrogen to generate methane.
End of Remediation
Hydrogen Concentration
Recently researchers have found that steady state H2 concentrations in the field are controlled by the type of bacteria utilizing the hydrogen.14 For example, under nitrate reducing concentrations, steady state H2 concentrations were less than 0.05 nM, under Fe (III) reducing conditions they were less than 0.2 to 0.8 nM, under sulfate reducing conditions they were 1 to 4 nM, and under methanogenic conditions they were 5 to 14 nM (Figure 4.7). Thus it is clear that an increased rate of hydrogen production will result in increased halorespiration without affecting the competition between various bacteria for the available hydrogen (Figure 4.11). Attempting to stimulate halorespiration with poor fermentation substrates, as has been suggested in the past, may unnecessarily limit the amount of dechlorination taking place.
O2 Depletion
Figure 4.11
NO3 Depletion Fe (iii) Depletion
SO42- Depletion
Methanogenic Conditions
Effect of oxidative poise on H2 concentrations during field scale implementation of enhanced reductive dechlorination systems, if methanogenic conditions can be achieved and maintained (adapted from Weidermeier et al., 1999).
It is clear from this this discussion that, during field scale enhanced reductive dechlorination at contaminated sites, the oxidative poise contributed by dissolved oxygen, nitrate, Fe (III), Mn (IV) and sulfate has to be depleted as quickly as possible to achieve efficient steady state reductive dechlorination reactions.53 Thus it is prudent to use the cheapest fermentable substrate available (see Table 4.4) to overcome the oxidative poise (Figure 4.12). 4.2.1.5 Mixture of Compounds on Kinetics Because few studies have systematically investigated the effect of multiple contaminants, and not all biochemical mediators of chlorinated aliphatic transformation
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Fermentation Substrate Demand for Production of H2
SO42Fe3+, Mn4+ NO3-
O2
Figure 4.12
Pictorial depiction of oxidative poise to be overcome during implementation of an IRZ for engineered anaerobic systems.
in methanogenic cultures are known, it is difficult to predict how any two chlorinated aliphatics may influence each other’s transformation. Several possible interactions can be hypothesized wherein chlorinated aliphatic biotransformation rates may increase, decrease, or remain unaffected. Increased rates may be observed through induction of transformation pathways or growth on one chlorinated aliphatic, such as dichloromethane (DCM), which then supports transformation of other aliphatics present. Decreased rates may result from competitive inhibition, competition for reducing equivalents, or synergistic toxicity effects. There may be no observable effect if transformation processes are independent, or concentrations of chlorinated aliphatics are low. Mixtures of chlorinated aliphatics often result from reductive dechlorination of a single parent compound; discerning the effect of mixtures on the transformation of individual compounds is difficult. The most frequently cited example is the sequential reductive dechlorination of polychlorinated ethenes (e.g., perchloroethene, trichloroethene) to ethene. Rates of reductive dechlorination in this series tend to decrease as the number of chlorine substituents decrease, so compounds such as vinyl chloride and dichloroethene often accumulate.64 Since all chlorinated ethenes are transformed by the same mechanism, it is conceivable that competitive interactions also influence the distribution of products, although it has not been systematically investigated. The sequential reductive dechlorination of 3,5-dichlorobenzoate to 3-chlorobenzoate to benzoate has been reported.63 In this case, reduction of the 3-chlorobenzoate did not proceed until the 3,5-dichlorobenzoate was completely transformed. This could not be explained by competing reaction rates, but was successfully described
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with a competitive inhibition model. Similar behavior for the reduction of 4-amino3,5-dichlorobenzoate to 4-amino-3-chlorobenzoate also was observed. It is important to note that competition may be significant during degradation of chlorinated methanes and ethanes as they can be transformed by reductive dechlorination and other mechanisms.65,66 Methylene chloride (DCM) can be oxidized to CO2 or converted to acetate while serving as a growth substrate.7 Chloroform (CF) and carton tetrachloride (CT) can be hydrolyzed.66 1,1,1 Trichloroethane (TCA) can be hydrolyzed or undergo dechlorination.19 Hydrolysis processes for these compounds are of particular interest since the products are reactive in some cases decomposing to harmless end products or reacting with cell material — whereas hydrolysis of CF and TCA yields strong nucleophiles (phosgene and acid halide), which may be toxic to the cell. Subsequent hydrolysis of these intermediates yields CO2 and acetate; both are subject to complete metabolism in anaerobic culture. The interaction of dechlorination amongst CM, CF, and TCA in methanogenic acetate-enrichment cultures was investigated in another study.67 Complex interactions occurred when mixtures of these chlorinated aliphatics were present: TCA transformation rates were reduced by the presence of DCM or CF, DCM transformation was enhanced by CF and TCA, and CF transformation rates increased or decreased depending on the mixture. Acetate utilization varied depending on the mixture fed, complicating the interpretation of results. Where acetate utilization was inhibited, biomass concentrations decreased and steady-state conditions were not achieved during the study. In all cases where CF and TCA were fed together, the rate of their transformations was lower than when they were fed individually. The decrease in the rate of transformation increases with the concentration of either compound, CF being more inhibitory than TCA on mg/L basis. Results were described by a competitive inhibition model, which was more predictive for the effect of TCA on CF transformation than the effect of CF on TCA transformation.67 Competitive interactions between substrates can introduce significant limitations to bioremediation processes. Studies investigating the cometabolic transformation of chlorinated ethenes by methanotrophs,68,69 nitrifiers,70 and phenol-induced aerobes71 have identified many problems at full scale that result from similar interactions. The decrease of biodegradation rates that results from competitive inhibition may limit the applicability of bioremediation processes. This is particularly true with CF and TCA, whose growth and biotransformation rates decrease rapidly as their concentrations increase. The effect of multiple substrates on the kinetics of biotransformation reactions has not been extensively studied. The results reported in the literature demonstrate that a range of effects may be observed even with a mixture of compounds that are structurally similar. No unifying model can be constructed on the basis of the available information. These studies also demonstrate the difficulty of assessing the interactions that occur when the compounds of interest may have dissimilar degradation pathways. Unraveling the complex interactions of compounds such as those often encountered in contaminated groundwaters will require further research. It is clear from these studies that certain combinations of compounds will lead to decreased rates
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of biotransformation. This is certainly true for the combination of CF and TCA. For engineers hoping to implement anaerobic bioremediation, this is important information in the decision making and design processes. If these interactions are not considered, any existing model will significantly underestimate the time required for remediation. Additional studies should attempt to enhance our fundamental understanding of these interactions, and identify other mixtures commonly found that significantly affect biodegradation rates. Such knowledge will minimize the application of bioremediation to sites where its efficacy would be limited. 4.2.1.6 Temperature Effects Many studies have reported anaerobic reductive dechlorination of chlorinated solvents to occur within a mesophilic temperature range (20 to 37°C). However, several authors have shown dechlorination of these compounds at more ambient groundwater temperatures between 10 and 20°C,8,26,55 indicating the applicability of reductive dechlorination in many groundwater environments. Microbial dechlorination has also been demonstrated under thermophilic conditions.26 An enrichment culture, obtained from polluted harbor sediment, rapidly dechlorinated PCE to cis-DCE at an optimum temperature of 65°C. Fumarate appeared to be the best electron donor. A large number of samples from high-temperature anaerobic environments has been investigated for the presence of dechlorinating microorganisms as well, but no dechlorinating activity has been found. 4.2.1.7 Anaerobic Oxidation Microorganisms can anaerobically mineralize VC and DCE in the presence of a complex, bioavailable electron acceptor such as Fe (III) − EDTA.72,73 Studies have focused on the possibility of oxidation of VC and DCE when they are used as a primary growth substrate under anaerobic environments.72,73,74 These results show VC and DCE mineralization under methanogenic and iron reducing conditions in anaerobic streambed sediments without the accumulation of ethene or ethane and buildup of carbon dioxide.14 Decreases of VC and DCE concentrations corresponded quantitatively to the production of carbon dioxide. 4.2.1.8 Electron Acceptors and Nutrients Nutrients: In addition to proper electron donor selection, nutrient availability may be a critical factor in maintaining a healthy dechlorinating consortium. In one instance, attempts to isolate a microbial species responsible for dechlorination led to the discovery that nutritional factors probably had been supplied by other consortium members. Highly enriched dechlorinating cultures required the addition of vitamin B12 and sludge supernatant to sustain dechlorination.38 Speculation exists that acetogens may supply the unknown nutritional factors required by the dechlorinating organism(s).34 Fortunately, in situ applications support a variety of microbial species. This microbial diversity, combined with the addition of nutritional supplements, should support a healthy dechlorinating microbial community.
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Alternative Electron Acceptors: Microbial dechlorination of chlorinated aliphatic hydrocarbons has been found to occur at low REDOX potentials, mainly under methanogenic conditions, although dechlorination under sulfate-reducing conditions has also been reported.26 A recent study found that, in reactions involving polychlorinated methanes and organic reductants exhibiting mercapto groups, an alternative initial reaction step may be a halophilic dissociative two-electron transfer.75 The proposed reaction mechanisms(s) involving R-S-H or R-S-S-R groups in the complete dechlorination of polychlorinated methanes may be helpful in the (re)interpretation of microbially mediated dechlorination reactions of such compounds. Indirect microbial reductive dechlorination of PCE has also been observed under iron-reducing conditions due to magnetite formation by iron-reducing bacteria. Magnetite can chemically reduce PCE to lower chlorinated ethenes.76 Besides reduction of chlorinated solvents under iron-reducing conditions, oxidation of cis-DCE and VC has been reported to occur under conditions where Fe (III) is the final electron acceptor.73 The same authors also reported the oxidation of DCE and VC in methanogenic, organic compound-rich bed sediment, indicating that the oxidation of these compounds is coupled to the reduction of humic acid compounds.74 As discussed in the previous section, the successful application of enhanced reductive dechlorination depends upon the depletion of electron-accepting chemical species. The most environmentally relevant species include O2, NO3– , Mn (IV), Fe (III), and SO4–2 . When evaluating a site for enhanced reductive dechlorination applicability, one must investigate the relative abundance of these compounds in both the groundwater and the aquifer solids. Although aqueous-phase acceptors such as O2 and NO3– take primary consideration, it is imperative that aquifer solids be characterized because they can serve as a reservoir of relatively insoluble electron-accepting species such as Fe (OH)3 or CaSO4. Once the electron-accepting species have been quantified, the amount of electron donor required to deplete them can be estimated by evaluating the stoichiometric relationship between the selected electron donor and each electron acceptor present on site (Figure 4.12). Higher levels of electron acceptor increase the oxidative poise and thus require more electron donor, therefore raising treatment costs. A series of generic reactions is given in Table 4.6 to illustrate some of the possible reactants and products. Table 4.6 Possible Reactants and Products of Specific Terminal Electron-Accepting Processes Predicted Reaction Electron Electron Electron Electron Electron
donor donor donor donor donor
+ + + + +
O2 → CO2 + H2O NO3– → CO2 + H2O + N2 Mn4+ → Mn2+ + CO2 + H2O Fe3+ → Fe2+ + CO2 + H2O SO4–2 → H2S + CO2 + H2O
Process Aerobic respiration Denitrification Manganese reduction Iron reduction Sulfate reduction
Once an electron donor has been selected and electron acceptors have been characterized, the stoichiometric relationship between them can be determined. An equation for each electron acceptor present at the site must be balanced using the
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selected electron donor. Once balanced, the molar ratio of donor to acceptor can be determined from these equations. These molar ratios represent an ideal case where the entire electron donor dosage is used to reduce the electron acceptor present in the treatment zone. When calculating the actual electron donor dosage, a safety factor must be incorporated to account for uncharacterized electron sinks and the advective transport of electron acceptors into the treatment zone. Site-specific conditions such as groundwater flow rate, surrounding electron acceptor concentrations, depth to the water table, rainfall frequency, and level of site characterization will influence the selection of the safety factor. Because treatment alternatives and budgetary constraints are different for each site, no rule of thumb exists for screening sites based on electron acceptor concentrations. The required mass of electron donor should be estimated so its cost can be calculated. Afterwards, a site-specific, cost-benefit analysis must be undertaken to determine if the site is a good candidate for enhanced reductive dechlorination (ERD) application. 4.2.1.9 Field Implementation of IRZ for Enhanced Reductive Dechlorination The author’s success and significant experience in creating an IRZ for enhanced reductive dechlorination is based purely on biostimulation of the indigenous capacity of microorganisms present at a contaminated site for dechlorination rather than bioaugmentation. This experience is based on successful implementation of this technology at more than 100 sites. Creation of such an IRZ involves the addition of an electron donor and supplemental nutrients to the contaminated groundwater zone in order to provide the optimum biogeochemical environment conducive for reductive dechlorination.1 The author’s wide experience in this technology is mostly based on the soluble electron donor, such as molasses that must be added semicontinuously or in batch injections1 in order to sustain the microbial activity for the fermentation reactions. Recently, cheese whey has been used as a slow release substrate in less permeable geologic environments. Preference of molasses and cheese whey is based purely on economics as illustrated in Figure 4.12 and Table 4.4.1 The geologic and hydrogeologic setting in which an IRZ system is installed governs its successful application. IRZ systems rely on the delivery of dissolved reagents, such as dilute molasses, throughout a contaminant plume; administering delivery of these amendments through both the vertical and horizontal extent of contaminant plumes sounds deceptively easy, but requires careful engineering and a knowledge of the geologic parameters affecting groundwater flow and transport. Different configurations, in plan view and cross sections, used for IRZ system designs are shown in Figures 4.13a and b, and 4.14a and b. Creative engineering considerations have to be taken into account to accommodate the requirements of a smaller plume vs. a larger plume and a shallower plume vs. a deeper plume. The ultimate objective of the IRZ system design engineer should be to deliver the electron donor as fast as possible and to create a uniformly mixed reactive zone in the subsurface, as well as to maintain the optimum biogeochemical conditions for enhanced reductive dechlorination to occur.
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Groundwater Flow Direction
Injection Points
Figure 4.13a
Staggered plume-wide injection grid for an IRZ system for remediation of a small plume.
Groundwater Flow Direction
Source Zone Grid Containment Curtains
Figure 4.13b
Source area staggered grid and containment curtains at mid-plume and downgradient locations for a large-size plume.
The total treatment time for an IRZ will encompass the time it takes to overcome the oxidative poise (deplete available electron acceptors), acclimatize and stimulate a healthy population of dechlorinating microorganisms, and allow the dechlorination reactions to proceed to conclusion. Site-specific conditions will obviously influence the total time required for treatment; for instance, anaerobic and particularly methanogenic sites exhibiting a significant level of natural dechlorination will require considerably less time than sites with aerobic groundwater and no evidence of dechlorination. Other factors incluencing the time required to treat a site include aquifers with low hydraulic conductivities, which will require more time for delivery of substrate throughout the subsurface, and the presence of DNAPL, which would require considerably longer treatment times due to the rate limitation imposed by dissolution of the contaminant. The time required for oxidative poise depletion depends on the electron donor supply and utilization rate, on initial electron acceptor concentrations, and the rate at which they are replenished by groundwater flow and recharge events. The large number of variables affecting electron acceptor depletion makes it difficult to predict the time lag; field data indicate that this lag could be anywhere from ten days to about three months. When considering the time required to implement an enhanced reductive dechlorination IRZ system, one should include a minimum of six months to perform a
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Dilute Molasses
Reactive Zones
Figure 4.14a
Injection well clusters for depths between 40 feet and 100 feet of saturated zone containment.
field pilot test to obtain the design parameters to design and scale up a full scale system. This six month testing time frame assumes one to two months for creation of a reducing environment after the depletion of the oxidative poise and three to four months of evaluating treatment data. The actual time required for a pilot test may exceed six months and will depend on hydrogeologic conditions and whether the site was already reduced with partial declorination initially. The assessment of a particular site for IRZ application should include the development of a contaminant profile, a hydrogeological profile, and a biogeochemical profile. An inventory of contaminants, their concentrations, and distribution throughout the plume will be the first step in assessing the feasibility of an IRZ. The presence, relative concentration, and distribution of daughter products is particularly important when assessing sites for enhanced dechlorination potential. Co-contaminant impacts may be either beneficial or detrimental, so it is important to assess before the onset of a pilot study. The success of an IRZ application primarily depends upon the effective delivery and distribution of the electron donor and nutrients throughout the contaminated subsurface. Hence, the ability to control the movement of the injected reagents is imperative at sites with a hydraulic conductivity less than or equal to 10–5 cm/sec. These sites require the addition of a slow releasing electron donor. Faster geologic
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IRZ In Situ Reactive Zones
100'
Submersible Pump
200'
Submersible Pump Figure 4.14b
In well mixing systems with submersible pumps for creating deeper IRZ systems.
settings require the addition of a soluble and fast releasing electron donor such as molasses. Figure 4.15 illustrates the need to inject a soluble electron donor at higher concentrations to achieve a reasonable size reactive zone from each injection point to achieve the scale up of the full scale system cost effectively. Biogeochemistry influences the potential for stimulating and maintaining microbially catalyzed reductive dechlorination. These microorganisms require highly reducing conditions reflected by low REDOX potential measurements and the production of hydrogen sulfide and methane gas. Additional biogeochemical parameters such as pH, alkalinity, temperature, and dissolved organic carbon can also affect the health and stability of dechlorinating microorganisms. 4.2.1.10
Lessons Learned
Typically a pilot test or smaller-scale field test follows the initial screening process. The two primary issues to be addressed during the field testing phase are the provision of properly placed observation wells and allowing for sufficient time to demonstrate the success of ERD. The amount of time it takes to see positive
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A - Distance of the Reactive Zone for the Slow Release Electron Donor B - Distance of the Reactive Zone for the Soluble Electron Donor
Concentration at Injection Point
IRZ In Situ Reactive Zones
Soluble and Fast Degrading Electron Donor Such as Molasses
Slow Releasing and Slower Degrading Electron Donor Such as Soybean Oil
Minimum Concentration Required for Adequate Production of H2 A
B
Distance From Injection Point Figure 4.15
Effects of a soluble electron donor and a slow release electron donor in an IRZ where the hydraulic conductivity is greater than 10–5 cm/sec.
results of the IRZ implementation is related to many factors, including groundwater velocity, the time required for introduced reagents to overcome the ambient REDOX conditions in groundwater, and the locations of observation wells in relation to the injection area. It is prudent to evaluate geochemistry and achievable degradation rates from data collected from wells located at least several months of travel time apart. Consequently, the minimum duration of a typical pilot study is six months with the flexibility to extend the testing based on data collection and site-specific costs. Field tests that are shorter in duration, or are applied in too small an area, often do not provide information that is applicable to a successful or economical large-scale implementation. This minimum period of time should be sufficient to overcome the initial aquifer REDOX conditions and allow for the degradation of constituents to a degree that will be observed in the pilot test. Some observation wells should also be placed within one to two months’ groundwater travel time from the injection area. This timing/placement should allow for early observations of the IRZ development, and allow for modification of the reagent injection program (strength and frequency) early enough in the planned test duration. Reagent Delivery: The successful application of an IRZ to remediate chlorinated solvents in groundwater first and foremost relies on the timely and consistent delivery of the organic carbon reagent to the treatment zone. The author’s experience is primarily based on injecting a dissolvable sucrose solution (molasses) as a reducing reagent. This serves as a cost effective reagent (0.20– $0.30/pound) that can aggressively alter the REDOX state of groundwater (oxidative poise) in a short time period. Other reagents, or electron donor substrates, such as edible oils and semisolid forms of lactate (such as Regenesis’ HRC®) will rely more on dissolution and diffusion for delivery. On a unit cost basis, these donors
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are more expensive. However, the application of a slow diffusing reagent may be more efficient in a highly reducing, slower groundwater velocity environment. The proper reagents should be selected based on the site hydrogeology and desired treatment time frame. Cheese whey has been used by the author as the slow release substrate in less permeable, slow moving groundwater environments.1 Based on the implementation of IRZs for the application of ERD to date, reagent delivery becomes most complicated in low permeability geologic environments (10–5 centimeters per second (cm/sec) or less hydraulic conductivity) or those with low groundwater flow velocities (less than 50 ft/yr). These settings can limit the area of influence of individual reagent injection points due to the absence of sufficient reagent dispersion. Poor donor delivery can also result in other potential complications. These complications can include: • Uneven application of reagent and resulting treatment; not achieving treatment goals • Lack of sufficient or timely demonstration of the technology during pilot phase • Requirement of too many injection points for a full-scale application
In low permeability and/or low groundwater velocity environments, the reagent can also accumulate in the vicinity of the injection point. Careful monitoring of the pH, ORP, and total organic carbon (TOC) levels in the groundwater near the injection well is necessary to avoid deleterious side effects. These effects are related to fermentation and byproduct formation and are discussed later in this section. Natural Surfactant Effect: The injection of an abundant source of easily degradable organic carbon during the application of ERD typically results in a rapid and large increase in the population of microorganisms in the treatment zone. As in any microbiological system, this large population increase will also result in an increase in production of natural biosurfactants and bioemulsifiers by the microorganisms. Natural biosurfactants result in desorption of the chlorinated contaminants adsorbed to the aquifer media. To assimilate less soluble substrates, such as chlorinated solvents, microorganisms require a large contact area between themselves and the contaminant. They achieve this by emulsifying the adsorbed contaminants into an aqueous phase. Most microbes frequently synthesize and excrete chemicals that promote such emulsification. These excreted chemicals fall into two main groups: biosurfactants and bioemulsifiers (Table 4.7). Table 4.7 Microbial Surfactants Structural Type
Producing Microorganism
Carbohydrates — lipids Trehalose — lipids Amino acids — lipids Lipopeptides Fatty acids — neutral lipids
Nocardia, Mycobacterium, Coryne bacterium, Arthrobacter Bacillus, Streptomyces, Corynebacterium, Mycobacterium Pseudomonas, Acinetobacter, Mycococcus, Micrococcus, Candida Pseudomonas, Thiobacillus, Gluconobacter
Ornithine − lipids
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Biosurfactants reduce the interfacial tension between water and the chlorinated contaminant so that the chlorinated contaminant (or less water soluble compounds such as PAHs) is easily micro-emulsified into the water phase. These micro-emulsion droplets are known to be smaller than microbial cells. Some bacterial glycolipids are extremely effective surfactants. In addition to enhancing the “mobilization” of the contaminants by microemulsions, biosurfactants can also increase apparent solubilities by the partitioning the contaminants into surfactant micelles. This desorption, or natural surfactant effect, is observed in many biological treatment processes as an increase in the constituent levels in the treatment zone and, in some cases, downgradient of the treatment zone. In some cases, the constituent concentrations in the treatment zone may remain unchanged, due to increased solubilization of the contaminants, for a short period even when biodegradation endproduct data support the conclusion that sufficient mass is being degraded by the ERD processes. The production of surfactants that facilitate the partitioning of contaminants from the DNAPL to the dissolved phase (thus resulting in enhanced biodegradation) has received considerable attention recently.12,13 The success of this approach depends on enhancing and maintaining biodegradation rates faster than the rate of mass transfer from NAPL to the dissolved phase. 100
Koc = 265 (PCE)
Percent Sorbed
80
Koc = 94 (TCE)
60
40
K oc = 36 (cis-1,2-DCE) 20
0.000
0.002
0.004
0.006
0.008
0.010
Organic Carbon Fraction (foc )
Figure 4.16a
Effect of Koc on Masssorbed/Masstotal.
The magnitude and composition of soil organic carbon content combined with the distinct differences in partitioning among the chlorinated alkenes have the potential to develop additional mechanisms that cause temporary increases in constituent concentrations during enhanced reductive dechlorination applications (Figure 4.16a).76 A high TOC gradient present between the groundwater and the aquifer soil matrix, resulting from the injection of molasses, will also result in desorption of hydrophobic contaminants for the following reasons:76
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• Successive dechlorination of alkene compounds is accompanied by successive decreases in organic carbon-water partition coefficients, Koc (the solubility of aliphatic compounds rapidly increases with decreasing molecular weight). If the degradation rate for daughter compounds is equal to or lower than the rate of parent compound degradation, temporary increases in aqueous-phase concentrations will be observed. • Electron donors such as molasses, typically applied to enhance reductive dechlorination, comprise soluble and colloidal carbon compounds, creating an aqueousphase organic carbon pool that was essentially nonexistent prior to the creation of the IRZ. Creation of an aqueous-phase carbon pool is expected to result in a partial migration of chlorinated alkenes to aqueous-phase carbon sorption, resulting in increased “apparent” concentrations. Since the aqueous-phase carbon is mobile, chlorinated alkenes may also be transported from their point of origination. As the soluble organic carbon is consumed by the microbial community, a portion of the chlorinated alkenes may remain in dissolved phase for a while until eventual degradation within the IRZ.
Intuitively, the increased desorption of target constituents within the IRZ allows for greater access to the typically “untreatable” adsorbed and separate phase contaminant mass present at source areas and DNAPL locations. However, this microbial surfactant effect must be anticipated and pilot or full-scale treatment should incorporate provisions to evaluate and account for it. For example, the potential for initial increases of stable parent constituent trends can be of concern to both responsible parties and regulatory bodies as the data would tend to indicate the technology is not working and, in some cases, could be considered as actually making conditions worse. Therefore, during the full scale or pilot test planning stages the possibility of this desorption effect must be evaluated in detail and anticipated in advance. Also, the possibility for an increase of dissolved chlorinated solvent concentrations to occur in areas downgradient of the treatment zone must be addressed. Typically, an “outside-in” approach is applied, whereby a steady state IRZ is established in a downgradient portion of the plume before applying ERD to the source area. Desorbed constituents would then move into an area already undergoing treatment and capable of treating the increased level of mass flux. Fermentation and By-Product Formation: During application of ERD a highly reducing biogeochemical environment is generally created throughout the treatment zone. This zone will also contain a large excess of organic carbon in the vicinity of the injection points, particularly if the geology is less permeable. During the implementation of an IRZ, at (10–5 cm/sec or less), these conditions can result in the formation of organic acids and alcohols in the groundwater as part of the degradation process. If the formed acids and alcohols are not consumed quickly the zone around the injection zone will mimic a fermenter where additional by-products can be formed. The formation of undesirable byproducts (including acetone and thiol compounds and 2-butanone) has been observed at sites where injection was initiated without careful monitoring of altered groundwater conditions near the injection wells. The occurrences of these byproducts are generally limited in extent and often sporadic in nature. It is expected that these oxidized by-products will also be utilized
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by microbes within the IRZ. Therefore, the lessons learned regarding these potential occurrences are as follows: • Careful and regular monitoring of groundwater within the treatment zone should be provided in order to ensure that pH levels are not depressed below pH = 4.0–4.5, and TOC levels are not excessive (site specific, but generally within 2 to 3000 mg/L). • The remediation plan for application of ERD should be flexible enough to allow for modification of both frequency of delivery and mass of organic carbon delivered to prevent the build-up of organic carbon and creation of conditions amenable to formation of these byproducts. Modifications in reagent delivery should be tied to the pH, ORP, and TOC monitoring in the treatment zone.
Overcoming Oxidizing Conditions/High Groundwater Velocity: As discussed, the implementation of an IRZ for the application of ERD relies on the creation of a highly reducing biogeochemical environment through provision of excess organic carbon to the groundwater. Achieving these conditions can be problematic in groundwater flow systems in which the ambient conditions are very oxidizing (due to shallow groundwater with abundant recharge) or the groundwater velocity is very high (>1,000 ft/yr). In both situations, the amount of reagent needed to “overcome” the oxidative poise of the naturally oxidizing conditions will be cost prohibitive. In addition, the scale-up cost for the full-scale system will be uneconomical due to extremely narrow (cigar-shaped) zones of influence from each injection point. In high groundwater velocity settings the limited transverse dispersion in groundwater can limit the extent of the reactive zone created by an individual injection point. This is of particular importance in settings where drilling costs may be high (i.e., deep settings or complex geology). In such cases, these site-specific considerations need to be weighed against other treatment alternatives. Biofilm Developments: When injecting an electron donor such as molasses (and electron acceptors) into an aquifer via injection wells, biofilm development around the injection wells should be anticipated. Biofilms are large aggregations of bacteria and other microorganisms bound together in a sticky mass of tangled polysaccharide fibers that connect cells together and tie them to a surface. Aerobic and anaerobic bacteria not only can thrive side by side within biofilms when biogeochemical conditions permit, but also actually seem to collaborate to make themselves more powerful. The polysaccharide coating acts like armor, giving the microorganisms protection beyond their usual defense mechanisms. While the typical average diameter of a bacterium in established biofilms is about 0.5–1 µm, biofilm bacteria rarely adhere directly to solid surfaces. Instead, at distances shorter than 1 nm, short-range forces such as hydrogen bonding and dipole formation tend to be the dominant adhesion effects. As bacteria are held in place and fed by the organic and inorganic molecules trapped by these short-range forces, they form slime that anchors them to solid surfaces. This slime becomes a home for additional bacterial growth. If the biofilm becomes too thick to permit adequate oxygen penetration, under aerobic conditions any additional biofilm growth may actually decrease biofilm adherence due to shearing. The thickness of the biofilm
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under anaerobic conditions is significantly smaller due to the above-mentioned shearing effects and the fact that the rate of biomass growth is substantially lower under anaerobic conditions. Under unaerobic conditions, typical of an IRZ, reduction in porosity within the saturated zone due to biofilm growth will not be significant enough to impact the hydrogeologic conditions for reagent transport. However, well clogging around the injection wells is an issue to be taken into consideration. Electron donor solutions, such as dilute molasses, are injected at reasonably high concentrations of TOC before it gets diluted by mixing with the groundwater within the IRZ. As a result of the higher concentrations of TOC present around the injection wells, the amount of biomass and biofilm growth will be significant. Since the electron donor solutions are injected in a batch mode at most of the IRZ applications, resistance to injection due to clogging may be an operational issue only during the injection events. In all the sites in the author’s experience, there were only two sites where injection under pressure was difficult due to significant head buildup. Manual cleaning of the well screens will be required under those conditions. Application in Areas of Low Constituent Concentration: The application of ERD to portions of an aquifer where the constituent concentrations are low (i.e., less than 100 µg/L) can pose additional challenges. A low concentration plume will impart less microbial conditioning and, therefore, will be more difficult to stimulate the microbial community. In these environments, a longer lag time for microbial growth and conditioning should be expected. It is also difficult to observe direct evidence of degradation through the monitoring program in a low concentration plume. Application in Areas of High Constituent Concentration/DNAPL: Given the inherent problems with the use of conventional remediation techniques in areas where the constituent concentrations are very high and/or where free phase contaminant (DNAPL) may be present, ERD has been an attractive potential alternative. The benefit of applying ERD in high concentration regimes (>50 to 200 mg/L of chlorinated VOCs) is related to the microbial surfactant effect that usually accompanies this technique. The surfactant molecule is typically composed of a strongly hydrophilic (water loving) group (or moiety) and a strongly hydrophobic (water fearing) group; in fact, the entire surfactant monomer is often referred to as amphiphillic because of its dual nature. The hydrophobic portion of the surfactant monomer is typically a long hydrocarbon chain, referred to as the “tail” of the molecule. The hydrophilic “head” group often includes anions or cations. The hydrophilic group of most surfactants provides a high solubility in water; however, the hydrophobic group prefers to reside in a hydrophobic phase such as a DNAPL. These compelling effects result in the accumulation of surfactant monomers at DNAPL-water interfaces (Figure 4.16b). Physical mobilization of the residual or adsorbed DNAPL by the surfactants is undesirable and will not happen during the IRZ application due to low levels of surfactant production (compared to a surfactant flood). Enhanced solubilization of the DNAPL will take place and has to be controlled by the enhanced rate of biodegradation.
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Water Phase
DNAPL Phase
Figure 4.16b
Surfactant monomer accumulation at the DNAPL–water interface.
When the groundwater equilibrium is altered, the transfer of more constituent mass from the free or adsorbed phase into the dissolved phase should be expected. An increase in the levels of dissolved constituents in groundwater results in a more treatable portion of the total contaminant mass. This effect can be used by itself or in conjunction with other ongoing technologies (such as pump and treat) to reduce treatment life span and costs. Care needs to be taken that increased dissolution and desorption does not result in the vertical or horizontal migration of elevated dissolved concentrations from the treatment zone. The possibility of enhancing downgradient migration is more pronounced when applying ERD in a potential DNAPL environment. Therefore, prior to ERD application in these settings a clear plan to address these possibilities must be developed. This could include application of the technology in an outside-in approach in which the downgradient areas are treated initially to develop a steady state “containment IRZ” and encroach to the source area gradually. However, if properly accounted for, the possibility of concentration increases, and the impacts to overcome, an ERD can be successfully applied in these settings. The application of ERD will increase the levels of mass reduction within the IRZ and once the initial disruption in phase equilibrium is overcome the IRZ technology will provide greater control of constituent migration from the source area. 4.2.1.11
Derivation of a Completely Mixed System for Groundwater Solute Transport of Chlorinated Ethenes78
Assumptions and Definitions: The measured concentrations in a well (C1) represent conditions in a unit volume of groundwater, the volume of which is defined by the saturated aquifer thickness times the effective porosity, ηe, as follows:
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V = haq × 1 × 1 × ηe
(4.2)
This volume of water moves through the soil at a velocity computed by using Darcy’s Law (υgw), while dissolved constituents migrate at a reduced velocity proportional to the retardation factor: υcon = υgw ÷ Rf
(4.3)
Similarly, the measured concentrations (s1) in the volume characterize the dissolved mass, while the total mass of the constituent can be as follows: s10 = s0 × Rf1
(4.4)
The dissolved constituent in the groundwater is assumed to decay through first order process that can be represented in terms of half-life as follows: λ1 =
ln 2 t1
(4.5)
2
As a constituent degrades, a daughter product is formed at a rate proportional to a yield factor equal to the ratio of the molecular weights of the daughter compounds to the parent compounds: β12 =
MW2 MW1
(4.6)
Solution for a single constituent: The differential equation representing describing the concentration within this volume written as the change in mass is equal to the mass in minus the mass out minus the rate of decay; or mathematically as: V
ds = W(t ) − Qs − λVs dt
where: V = Q = s = W= λ =
Volume [L3] Flow through the system [L3/T] Concentration [M/L3] Mass loading term [M/T] First order reaction coefficient [1/T]
(4.7)
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Equation 4.7 can be written more concisely as follows: V
ds + Qs + λVs = W(t ) dt
V
ds + s(Q + λV) = W(t ), or dt
V
ds + λ ′s = W(t ) dt
(4.8)
where: λ′ = Q + λV Equation 4.8 is a nonhomogeneous ordinary differential equation. The general solution to this classification of equations can be expressed as the sum of the “complementary” or general solution when W(t) equals zero, and a particular solution when W(t) has a specific form. s = sc + sp Consider first, the solution to Equation 4.8 for a single constituent, with initial conditions s = s0, at t = 0. Dividing Equation 4.8 by V, the equation describing the complementary function is written as: ds λ ′ + s=0 dt V Separation of variables, ds λ′ = − dt s V ds
λ′
ln s = −
λ′ t + C0 V
∫ s = ∫ − V dt
(4.9)
C0 = Integration Constant If W(t) = 0, sp = 0 and the above equation is also the specific solution. The integration constants in Equation 4.9 can be solved by exponentiation and applying the initial conditions.
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λ′ s = exp − t + C 0 V λ′ s = C ′0 exp − t V At t = 0, s0 = C0′, the initial concentration in the control volume. Therefore, the equation describing the change in concentration (mass) in the control volume is as follows: λ′ s = s 0 exp − t V
(4.10)
Equation 4.10 can be applied to simple steady-state groundwater transport problems (no dispersion) by recognizing analogous processes. Consider a well in a contaminant plume, and measured concentrations have been stable through multiple groundwater sampling rounds, implying an equilibrium has been reached between the continued release of the constituent from residual source materials, and the significant transport process (advection, adsorption, and degradation). The question to be answered is how far and at what level these constituents will migrate. The control volume, V, is equivalent to the volume of active groundwater beneath the water table, i.e., V equals the saturated aquifer thickness, T, times the effective porosity, ηe. Time, t, is equal to the constituent transport time, i.e., t equals distance (x) divided by the groundwater velocity (υgw) times the retardation factor. The unit flow through the control volume, Q, is equivalent to groundwater recharge (or percolation), N. The initial dissolved concentration can also be expressed as ratio of the total mass to the retardation factor (s10 /Rf1). x λ′ s1 (t ) = s 0 exp − t = s 0 exp − λ 1′ R f1 υ gw T or
s1 (t ) =
s10 x exp − λ 1′ R f1 R f1 υ gw
where λ 1′ =
N + λ 1T T
(4.11)
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Evaluation of Daughter Products Now consider the presence of a second constituent, which could exist in the environment due to a release or be a degradation product, for example, TCE. The fate of constituent 2 can be represented mathematically as the sum of 2 separate expressions. TCE(t) = f(t) + g(t) where f(t) describes the fate of the portion of the mass that was released to the environment, and g(t) describes the fate of the TCE generated from the degradation of PCE. By inspection, f(t) can be written from Equation 4.11 as:
f (t ) =
s 20 x exp − λ ′2 R f2 R f2 υ gw
Similiarly, g(t), the mass of TCE generated by the degradation of PCE, can also be written from Equation 4.11 as:
α( t ) =
β12 x s10 − s10 exp − λ 1′ R f 1 R f1 υ gw
or
α( t ) =
s10β12 x 1 − exp − λ 1′ R f 1 R f1 υ gw
The above expression describes the change in the total mass over time of the TCE generated through the degradation of PCE. Therefore, the equation describing the change in the dissolved concentration (consistent with Equation 4.11 and f(t) above, and implicitly assuming that only the dissolved PCE degrades) is as follows:
β(t ) =
s10β12 x 1 − exp − λ 1′ R f 1 R f1 R f2 υ gw
(4.12)
The mass of TCE generated from the degradaion of PCE also degrades consistent with f(t). Therefore g(t) is written as follows:
g(t ) =
s10β12 x x 1 − exp − λ 1′ R f exp − λ ′2 R f2 1 R f1 R f2 υ gw υ gw
(4.13)
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The total change in concentration of TCE over time is therefore expressed as:
s 2 (t ) =
s 20 x s10β12 x x 1 − exp − λ 1′ R f exp − λ ′2 R f exp − λ ′2 R f2 + 1 2 R f2 υ gw R f1 R f2 υ gw υ gw (4.14)
Now, consider cis-1,2 DCE, the degradation byproduct of TCE. There are three possible fate and transport pathways for cis-1,2 DCE: a) existing cis-1,2 DCE b) cis-1,2 DCE formed by degradation of an existing source of TCE (existing) c) cis-1,2 DCE formed by degradation of TCE which originated from the degradation of PCE
a ) α( t ) =
s 30 x exp − λ ′3 R f3 R f3 υ gw
b) β(t ) =
s 20β 23 x x 1 − exp − λ ′2 R f exp − λ 3′ R f 2 3 R f2 R f3 υ gw υ gw
c) The generated total cis-1,2 DCE from the decay of dissolved TCE from dissolved PCE is s10β12β 23 x x 1 − exp − λ 1′ R f 1 − exp − λ ′2 R f 1 2 R f1 R f2 υ gw υ gw The equivalent dissolved cis-1,2 DCE is: s10β12β 23 x x 1 − exp − λ 1′ R f 1 − exp − λ ′2 R f 1 2 R f1 R f2 R f3 υ gw υ gw which changes in concentration with time:
γ (t ) =
s10β12β 23 x x x 1 − exp − λ 1′ R f 1 − exp − λ ′2 R f exp − λ 3′ R f 1 2 3 R f1 R f2 R f3 υ gw υ gw υ gw
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The equation describing the maximum down gradient concentrations of cis-1,2 DCE is DCE = s3(t) = α(t) + β(t) + γ(t) or in expanded form,
s3 (t ) =
+
s 30 x s 20β 23 x x 1 − exp − λ ′2 R f exp − λ ′3 R f exp − λ ′3 R f3 + 2 3 R f3 υ gw R f2 R f3 υ gw υ gw s10β 23β12 x x x 1 − exp − λ 1′ R f exp − λ 3′ R f − − exp λ R 1 ′ 2 f2 3 1 R f1 R f2 R f3 υ gw υ gw υ gw
x * exp − − λ 3′ R f3 υ gw (4.15) From inspection, the maximum concentrations of vinyl chloride, s4(t), are expressed as:
s 4 (t ) =
s 40 x s 30β 34 x x 1 − exp − λ ′3 R f exp − λ ′4 R f exp − λ ′4 R f4 + 3 4 R f4 υ gw R f3 R f4 υ gw υ gw
+
s 20β 23β 34 x x x 1 − exp − λ ′2 R f 1 − exp − λ 3′ R f exp − λ ′4 R f 3 4 2 R f2 R f3 R f4 υ gw υ gw υ gw
+
s10β12β 23β 34 x x 1 − exp − λ 1′ R f 1 − exp − λ ′2 R f 1 2 R f1 R f2 R f3 R f4 υ gw υ gw
x x exp − λ ′4 R f *1 − exp − λ 3′ R f3 4 υ gw υ gw (4.16) Again, from inspection, the equation for the transformation of vinyl chloride to ethane can be written. Figure 4.17 describes the transformation of PCE to the final desired end product ethene; the shapes of the individual curves will depend on the degradation rates, retardation factor and other biogeochemical parameters.
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IN SITU REACTIVE ZONES
177
Total VOC
PCE TCE DCE Vinyl Chloride
Time Figure 4.17
4.2.1.12
Aqueous-phase concentrations of chlorinated alkene compounds resulting from the successive dechlorination of PCE.
IRZ Performance Data
The author and his colleagues have implemented about 100 IRZ applications, beginning in 1993.1 During the technology evolution a lot of lessons were learned and have been described earlier. The performance data presented in the next few figures describe only the transformation or degradation of the contaminants during IRZ applications. It should be noted that the information is not presented as sitespecific case studies, due to shortage of space. Site in California This site was a former metal plating facility and was contaminated with TCE and Cr (VI). Very few daughter products were present prior to injection of molasses. A grid-ike IRZ injection system was installed throughout the entire plume (two acres in size) and injection of molasses began during the first quarter of 1996. Figure 4.18 presents the degradation and remediation of TCE and the formed daughter products in terms of average concentrations throughout the entire plume. Figure 4.19a describes the highest concentration of TCE found at the site and its decline during the implementation of the IRZ. The increased concentration of TCE at this well (17 ppm) is believed to be a result of the microbial surfactant effect. The REDOX potential within the plume was maintained at less than –250 mV via batch injections of molasses. The TOC concentrations were always maintained above 200 ppm. Figures 4.19b and 4.19c describe the reduction of Cr (VI) concentrations at the same site. Site in Northeastern U.S. At a site in northeastern U.S., PCE and its daughter products were found in a fractured bedrock environment. The plume was very long and a pump and treat system was already in place at the site. Pilot studies for IRZ implementation were performed and the primary objective was to implement a containment IRZ curtain
Sept. 1995
Dec. 1995
Mar. 1996
Sept. 1996
Dec. 1996
cis-1,2-DCE
Jun. 1996
TCE
Initial Injection
Apr. 1997
June 1997
Performance data on enhanced reductive chlorination at a site in California.
0 Apr. 1995
500
1000
1500
2000
2500
3000
3500
4000
4500
Oct. 1997
Dec. 1997
ERD Application - Observation Well COC Concentrations
Mar. 1998
VC
June 1998
Oct. 1998
Feb. 1999
178
Figure 4.18
Concentration
5000
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20
Concentration
15 TCE 10
Vinyl Chloride DCE
5
April '97
June '97
Sept. '97
Dec. '97
Mar. '98
June '98
Sept. '98
Dec. '98
Date Figure 4.19a
The reduction of TCE from about 17 ppm at the California site during an IRZ implementation. Injection of molasses begins
180
Hexavalent Chromium (mg/L)
160 140 120 100 80 60 40 20
Dec. '96
April '97
June '97
Oct. '97
Dec. '97
Date Figure 4.19b
Hexavalent chromium reduction at abandoned manufacturing facility in California.
to bifurcate the plume. Figure 4.20 describes the performance of the IRZ at a monitoring well with the decline of PCE and the formation and degradation of the daughter products. This figure also shows the formation of ethene as the final end product. Figure 4.21a and b is important to note because of the transformation of cis-1,2 DCE to the final end product ethene, and also the observed mass balance of the conversion. Figure 4.22 describes the installation locations of the containment IRZ curtain, and the bifurcation of the plume in a short time frame (nine months). At this point,
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Remediation Injection Events 70,000
Concentration (mg/L)
60,000 50,000 40,000 30,000
Total Chromium Hexavalent Chromium
20,00
Jan. '99
Jan. '99
Jan. '99
Jan. '99
Oct. '98
July '98
Apr. '98
Jan. '98
Oct. '97
July '97
Apr. '97
Jan. '97
Nov. '96
Aug. '96
May '96
Feb. '96
10,00
Date Figure 4.19c
Aerobic reduction of concentrations at all the wells.
ERD applications are taking place within the source area at this site after the establishment of the downgradient IRZ curtain. Site in Wisconsin This was a former dry cleaning location within a strip mall in Wisconsin. Figure 4.23 describes the performance of the implemented IRZ at a monitoring well located within the plume. All monitoring wells within the plume exhibited similar performance. Due to development activities at this site, some of the monitoring wells had to be replaced at identical locations after these activities were finished. That is the reason the pre-injection concentrations are shown as estimated instead of actual concentrations. In all probability, the concentrations shown as estimated are the actual pre-injection concentrations at those locations. It is important to note the decline of PCE and the formation and degradation of the daughter products. Ethylene concentrations were increasing until all the chlorinated compounds were degraded. Site in Ohio Another IRZ for ERD is being implemented at an industrial facility in Ohio. The performance of the IRZ is described by a monitoring well located about 50 feet from the injection locations (Figure 4.24). The disproportionate increase in DCE concentrations shortly after injection is believed to be due to a combination of the microbial surfactant effects and the enhanced degradation rates of the desorbed contaminants. Site in North Carolina At this site a pilot study was initiated to address contamination at very high concentrations of TCE (more than 100 ppm). The performance of the ongoing
Figure 4.20
Concentration changes in a performance monitoring well within an IRZ for ERD, 70 feet downgradient from the injection well. Note: There is a gradual increase of the final transformation product ethene and a reasonably steady level until all the chlorinated ethenes have been transformed.
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182
Figure 4.21
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Concentration changes in mg/L and µM at a monitoring well 50 feet from the injection zone.
pilot study is shown in Figure 4.25. This is the first site where the author has implemented an IRZ when the initial chlorinated contaminant concentrations were more than 100 ppm. Site in Pennsylvania An ongoing pump and treat system had reached asymplotic concentrations at this site after 13 years of operation and the desire was to accelerate the time required for closure. Once the IRZ was established, the site was closed after reaching cleanup levels in less than 12 months. (Figure 4.26).
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Figure 4.22
4.2.2
183
The effect of containment IRZs on a long plume of PCE (contamination shown are total VOCs).
In Situ Metals Precipitation
The presence of metals in the subsurface environment can be in many forms: elemental, ionic, and/or organometallic. Distribution of the commonly encountered metals in the subsurface can be categorized as follows: • Elemental form • Mercury • Lead • Gold, silver and the other noble metals • Metal alloys: brass (copper and zinc); bronze (copper, tin, and zinc); nickelcadmium • Ionic form • Arsenic: As (III) arsenite, AsO3–2 ; As (V) arsenate, AsO4–3 • Chromium: trivalent Cr (III); hexavalent Cr (VI), Cr2O7–2 and CrO4–2 • Iron: Ferrous Fe (III); Ferric Fe (III) • Copper, lead, zinc, cadmium: Cu+1, Cu2+, Pb2+, Zn2+, Cd2+ • Mercury: Hg+1, Hg+2 • Organometallic form • Dimethyl mercury: Hg (CH3)2 • Dimethyl arsenic: AS2 (CH3)4 • Tetraethyl lead: Pb (C2H5)4 • Metal cyanide complexes: Hg (CN)2; Zn (CN)4–2 ; Cu (CN)2–1 ; Fe (CN)6–4
The common range of concentrations of naturally encountered metals in the subsurface environment is shown in Table 4.8. In order to understand the fate of heavy metals in the soil-water system, it is important to understand the general characteristics of soil and the chemistry of heavy metals in an aqueous solution. In the aquatic environment, heavy metals may be classified into at least two different categories: 1) in true solution as free or complexed ions, and 2) in particulates from adsorption onto other particles, or incorporation into biomass of living organisms and inorganic precipitates such as hydroxides, carbonates, sulfides, and sulfates.
Apr. 1997
Nov. 1997
PCE
Estimated Pre-Remediation Groundwater Conditions
TCE June 1998
Dec. 1998
July 1999
cis -1,2-DCE
Initial Injection
VC Jan. 2000
Ethylene
Concentration declines in a monitoring well within an IRZ for enhanced reductive dechlorination at a site in Wisconsin.
0 Oct. 1996
500
1000
1500
2000
2500
3000
0 Aug. 2000
50
100
150
200
250
300
350
400
450
IN SITU REACTIVE ZONES
Figure 4.23
Concentration
Groundwater Contamination Concentrations vs. Time
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184
Ethylene
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Figure 4.24
185
Performance data from a monitoring well 50 feet downgradient of the injection zone. Substantial increase in mass was observed due to the microbial surfactant effects. Note the increasing concentration of the daughter products with time and the excellent correlation of mass balance on a milli molar basis. The last data point shows that the transformation is complete to ethene.
160000
140000
Concentration (ug/L)
TCE 120000
cis-1,2-DCE 100000
80000
60000 PCE 40000 Vinyl Chloride 20000
Aug. '99
Oct. '99
Dec. '99
Jan. '00
Mar. '00
May '00
June '00
Aug. '00
Oct. '00
Date
Figure 4.25
Concentrations of VOCs in the pilot observation well vs. time, in situ reactive zone pilot test.
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In-Situ Remediation Using Molasses Injection
Pump and Treat Remediation
1000
Concentration (ug/L)
Chromium TCE
100
10
May '90
Mar. '94
Aug. '95
Mar. '96
Mar. June Sept. '97 '97 '97
Date Figure 4.26
TCE and chromium concentrations vs. time.
Table 4.8 Common Concentration Range of Metals in Soils (mg/Kg)77 Element
Range
Average
Antimony (Sb) Arsenic (As) Barium (Ba) Beryllium (Be) Cadmium (Cd) Chromium (Cr) Cobalt (Co) Copper (Cu) Lead (Pb) Mercury (Hg) Nickel (Ni) Selenium (Se) Silver (Ag) Tin (Sn) Vanadium (V) Zinc (Zn)
2–10 1–50 100–3,000 0.1–40 0.01–0.7 1–1,000 1–40 2–100 2200 0.02–0.3 5–500 0.1–2 0.01–5 2–200 20–500 10–300
--5 430 6 0.06 100 8 3 10 0.03 40 0.3 0.05 1 100 50
Dec. '97
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Many metals are found as very insoluble sulfide (Zn, Ag, Hg, Cu, Cd, Pb, Ni, Co) carbonate and hydroxide (Cr, Fe) forms. Biogeochemical impacts on the groundwater concentrations of species such as sulfides (from sulfate reduction) and carbonates (via CO2 formation) enable many dissolved metallic ions to be precipitated and immobilized. In a soil-water system, the fate of heavy metals is directly related to their states of identity and the existing biogeochemical conditions. The free and complexed metal ions may be removed from solution by adsorption and precipitation mechanisms, while the particulate heavy metals may be transformed by their own dissolution and filtration mechanism of soils. In principle, the concentration of heavy metals in an aqueous system is controlled by the congruent and incongruent solubility of various oxides, carbonates, sulfates, and sulfides. Metal precipitates in soil systems represent a selective accumulation of at least two or more constituent ions into an organized solid matrix often crystalline in nature. The process by which this selective accumulation occurs to form a distinct solid phase is termed precipitation. A precipitate can be considered a particulate phase which separates from a continuous medium. The fact that solid phases form in soil-water systems means that the overall free energy of formation is negative for the combined physical-chemical processes operating during the period of formation. The actual steps leading to the formation of a separate solid phase, however, must occur at the microscale level: the joining together of the constituent ions or molecules that will eventually be recognized as a distinct separate phase.80 Under classical nucleation theory, three steps are generally considered necessary for those microscale processes to result in the formation of crystals that will persist and survive over relatively long periods of time: nuclei formation, crystallite formation, and crystal (precipitate) formation.80 Complexation reactions are also important in determining the saturation state of groundwater. A complex is an ion that forms by combining simpler cations, anions, and sometimes molecules. The cation or central atom is typically one of the metals, and the anions, often called ligands, include many of the common inorganic species found in groundwater, such as S2–, CO32– , SO42– , PO43– , NO3–, Cl–. The ligand might also be an organic molecule such as amino acid. 4.2.2.1 Principles of Heavy Metals Precipitation The mechanisms that can be used to reduce the concentrations of heavy metals dissolved in groundwater are transformation and immobilization. These mechanisms can be induced by both abiotic and biotic pathways. Abiotic pathways include oxidation, reduction, sorption, and precipitation. Examples of biotically mediated processes include: reduction, oxidation, precipitation, biosorption, bioaccumulation, organo-metal complexation and phytoremediation. In this chapter, immobilization mechanisms induced only by precipitation will be discussed. Dissolved heavy metals can be precipitated out of solution through various precipitation reactions shown below. A divalent metallic cation is used as an example in these reactions.
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Hydroxide precipitation: Me++ + 2OH– → Me (OH)2 ↓ Sulfide precipitation:
Me++ + S2– → MeS ↓
Carbonate precipitation: Me++ + CO3– – → MeCO3 ↓
(4.17) (4.18) (4.19)
Theoretical behavior of solubility of these precipitation mechanisms is shown in Figure 4.27.
log [Me++]
Carbonate Precipitate
Hydroxide Precipitate Sulfide Precipitate
pH Figure 4.27
Theoretical pathways of solubility of metals.
Hydroxide and sulfide precipitation of heavy metals have been used successfully in conventional industrial waste water systems. Lime (Ca(OH)2) or other alkaline solutions such as potash (KOH) are used as reagents for hydroxide precipitation. Sodium sulfide (Na2S) is normally used as the reagent to form extremely insoluble metallic sulfide precipitates. Injection of these chemical reagents into the contaminated aquifers to create a reactive zone will precipitate the heavy metals out of solution. However, injection of a reactive, pH altering chemical reagent into the groundwater may be objectionable from a regulatory point of view. Obtaining the required permits to implement chemical precipitation may be difficult. Furthermore, the metallic cations precipitated out as hydroxide could be resolubilized slightly as a result of any significant shift in groundwater pH.
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Under reducing conditions, heavy metal cations can be removed from solution as sulfide precipitates if sufficient sulfur is available. In systems containing a sufficient supply of sulfur, neutral to mildly alkaline pH and low REDOX conditions are most favorable for the precipitation of many heavy metals. Chromium is insoluble under reducing conditions, as Cr (III) hydroxide, but only at neutral to mildly acidic and alkaline pH values. Precipitation as sulfides is considered the dominant mechanism limiting the solubility of many heavy metals. Sulfide precipitation is particularly strong for “chalcophilic” metals exhibiting so-called “B-character,” such as Cu (I), Ag, Hg, Cd, Pb, and Zn; it also is an important mechanism for transition elements such as Cu (II), Ni (I), Co (II), Fe (II) and Mn (II).81 Two situations can be distinguished in natural systems during sulfide precipitation conditions: the existence of a certain sulfide precipitation capacity (SPC), or (when exceeding the SPC) the accumulation of free sulfide (as H2S or HS–) in the aqueous phase. At excess sulfide concentrations, solubility of some metals can be increased by the formation of thio complexes. However, the stability of these complexes is still questionable. Possible pathways of metal precipitate interactions are shown in Figure 4.28. Figure 4.29 describes the fields of dominance of the different sulfur species in groundwater. Mineral Surface
Figure 4.28
Organic Surface
Inorganic Complex
Free Ion
Organic Complex
Precipitate
Occlusion
Living Biomass
Heavy metal interactions in an aquifer matrix.
The sulfide ions necessary to mediate sulfide precipitation can be directly injected into a reactive zone in the form of sodium sulfide (Na2S). However, the sulfide ion (S2–) is one of the most reduced ion and its stability within the reactive zone is short lived. It will be converted to sulfate (SO4– –) very quickly in the presence of oxidizing
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1.40 1.20 Water Oxidized
1.00 0.80
HSO4-
Eh, in Volts
0.60 0.40
SO2-
S0
4
0.20 0.00 H2Saq
-0.20 -0.40
HS-
-0.60 Water Reduced
S2
-0.80 -1.00 0
2
4
6
8
10
12
14
pH Figure 4.29
Fields of dominance of sulfur species at equilibrium at 25˚C and 1 atmosphere (adapted from Hem, 1985).
conditions within the contaminated plume. Addition of a very easily biodegradable organic substrate, such as carbohydrates, will enhance the formation of reduced, anaerobic conditions by depleting the available oxidation potential. The presence of carbohydrates serves two purposes: microorganisms use it as their growth substrate by depleting the available oxygen, and they use it as an energy source for the reduction of sulfate to sulfide. Indirect microbial transformation of metals can occur as a result of sulfate reduction when anaerobic bacteria oxidize simple carbon substrates with sulfate serving as the electron acceptor. The net result of the process is the production of hydrogen sulfide (H2S) and alkalinity (HCO3– ). Sulfate reduction is strictly an anaerobic process and proceeds only in the absence of oxygen. The process requires a source of carbon to support microbial growth, a source of sulfate, and a population of sulfate reducing bacteria. Dilute black strap molasses solution is an ideal feed
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substrate for this purpose since typical black strap molasses contains approximately 20% sucrose, 20% reducing sugars, 10% sulfated ash, 20% organic non sugars, and 30% water.1 Whether formed biotically or abiotically, the metal sulfides result from an interaction between the metal ion and sulfide ion: Me2+ + S2– → MS ↓
(4.20)
It is the source of the sulfide that determines whether a biological agent is implicated in metal sulfide formation. If the sulfide results from bacterial sulfate reduction or from bacterial mineralization of organic compounds, it is obviously of biotic origin. If it is derived from volcanic activity, it is generally of abiotic origin. The metal sulfides, because of their relative insolubility, form readily at ordinary temperatures and pressures by interaction of metal ions and sulfide ions. Table 4.9 lists solubility products for some common simple sulfides.83
Table 4.9 Solubility Products for Some Metal Sulfides81 CdS Bi2S3 CoS2 Cu2S CuS
1.4 × 10–28 1.6 × 10–72 7 × 10–23 2.5 × 10–50 4 × 10–38
1 × 10–19 1 × 10–29 5.6 × 10–16 1 × 10–45 3 × 10–53
FeS PbS MnS Hg2S HgS
NiS Ag2S SnS ZnS H 2S HS-
3 × 10–21 1 × 10–51 8 × 10–29 4.5 × 10–24 1.1 × 10–7 1 × 10–15
The following calculation will show that relatively low concentrations are needed to form metal sulfides by reacting with H2S at typical concentrations that can be formed in an anaerobic IRZ.83 Let us examine, for instance, the case of iron. The dissociation constant for iron sulfide (FeS) is: [Fe2+][S2–] = 1 x 10–19
(4.21)
The dissociation constant for H2S is: H S [S ] = 1.1 x 10 [H ] [ ]
(4.22)
[HS ][H ] = 1.1 x 10 [H S]
(4.23)
2−
−22
2 + 2
since, −
+
s
and,
–7
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[S ][H ] = 1 x 10 [HS ] 2−
+
−15
(4.24)
−
Therefore,
[Fe
H ]= [ ]
+ 2
2+
[ ]( 2
H+ 1 x 10 −19 9.1 x 10 2 x = [H 2S] 1.1 x 10 −22 [H 2S]
)
(4.25)
About 5.08 × 10–3 mg of Fe2+ per liter (9.1 × 10–8 M) will be precipitated by 3.4 mg of hydrogen sulfide per liter (10–4 M) at pH 7. The unused H2S will ensure reducing conditions, which will keep the iron in the ferrous state. Since ferrous sulfide is one of the more-soluble sulfides, it can be seen that metals whose sulfides have even smaller solubility products will form even more readily at lower H2S concentrations. Metal sulfides have been generated in laboratory experiments utilizing H2S from bacterial sulfate reduction. It has been reported that sulfides of Sb, Bi, Co, Cd, Fe, Pb, Ni, and Zn were formed in a lactate broth culture of Desulfovibrio desulfuricans to which sulfate and salts of selected metals had been added.83 Metal toxicity to Desulfovibrio desulfuricans depends in part on the concentration of the metallic ion in question. Obviously, for the corresponding metal sulfide to be formed, the metal sulfide must be even more insoluble than the starting compound of the metal. More metals such as Cu, Ag, Cd, Pb, Zn, Ni, and Co, in addition to Fe and Mn, can also be precipitated as metallic sulfides. Precipitated metallic sulfides will remain in an insoluble, stable form, unless the subsurface REDOX conditions change dramatically. The production of alkalinity from sulfate reduction, denitrification, and other reactions causes an increase in pH, which can result in metal precipitation through the formation of insoluble metal hydroxides or oxides. This process follows the reaction: Me2+ + 2H2O → Me (OH)2 ↓ + 2H+
(4.26)
Chromium Precipitation In situ microbial reduction of dissolved hexavent chromium Cr (VI) to trivalent chromium Cr (III) yields significant remedial benefits because trivalent chromium Cr (III) is less toxic, water insoluble, and, thus, nonmobile, and precipitates out of solution. In fact, it has been stated that the natural attenuation of Cr (VI) to the reduced Cr (III) form within an aquifer is a viable groundwater remediation technique. In situ microbial reduction of Cr (VI) to Cr (III) can be promoted by injecting a carbohydrate solution, such as dilute molasses solution. The carbohydrates, which consist mostly of sucrose, are readily degraded by the heterotrophic microorganisms present in the aquifer, thus depleting all the available dissolved oxygen present in the groundwater. Depletion of the available oxygen present causes reducing conditions to develop. The mechanisms of Cr (VI) reduction to Cr (III) under induced reducing conditions can be 1) likely a microbial reduction process involving Cr (VI)
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as a terminal electron acceptor for the metabolism of carbohydrates by species such as Bacillus subtilis; 2) an extra cellular reaction with by-products of sulfate reduction such as H2S; and 3) abiotic oxidation of the organic compounds including the soil organic matter such as humic and fulvic acids.84 Cr (VI) is known to be reduced both aerobically and anaerobically in different bacterial systems. In anaerobic systems, membrane preparations reduce Cr (VI), which has been shown to serve as a terminal electron acceptor. Aerobic reduction of Cr (VI) has been found to be associated with soluble proteins. The enzymatic basis for aerobic chromate reduction is not known, but it has been proposed that chromate may be reduced by a soluble reductase enzyme with a completely unrelated primary physiological role. Based on the diversity of Cr (VI) reducing microorganisms in soil, provision of a suitable electron donor such as molasses may be sufficient and the ORP within the IRZ need not be reduced to –250 to –300 mV as is the case during ERD applications.85,86 The primary end product of Cr (VI) to Cr (III) reduction process is chromic hydroxide [Cr (OH)3], which readily precipitates out of solution under alkaline to moderately acidic and alkaline conditions.87 To ensure that this process will provide both short term and long term effectiveness in meeting groundwater cleanup objectives, the chromium precipitates must remain immobilized within the soil matrix of the aquifer, and could not be subject to Cr (OH)3 precipitate dissolution or oxidation of Cr (III) back to Cr (VI) once groundwater conditions revert back to natural conditions. Based on the results of significant research being conducted on the in situ chromium reduction process, it is readily apparent that the Cr (OH)3 precipitate is essentially an insoluble, stable precipitate, immobilized in the soil matrix of the aquifer. Contrary to the numerous natural mechanisms that cause the reduction of Cr (VI) to Cr (III), there appear to be only a few natural mechanisms for the oxidation of Cr (III). Indeed, only two constituents in the subsurface environment (dissolved oxygen and manganese dioxide) are known to oxidize Cr (III) to Cr (VI).88 The results of studies conducted on the potential reaction between dissolved oxygen and Cr (III) indicate that dissolved oxygen will not cause the oxidation of Cr (III) under normal groundwater conditions. However, studies have shown that Cr (III) can be oxidized by manganese dioxides, which may be present in the soil matrix. However, only one phase of manganese dioxides is known to oxidize appreciable amounts of Cr (III) and this process is inversely related to groundwater pH. Hence, the oxidation of Cr (III) back to Cr (VI) in a natural aquifer system is highly unlikely. The Cr (OH)3 precipitate has an extremely low solubility (solubility product, Ksp = 6.7 × 10–31) and thus, very little of the chromium hydroxide is expected to remain in solution. It has been reported that aqueous concentration of Cr (III), in equilibrium with Cr (OH)3 precipitates, is around 0.05 mg/L within the pH range of 5 to 12 (Figure 4.30). The pH range of natural aquifer systems will be within 5 to 12 and, hence, the potential for the chromic hydroxide to resolubilize is unlikely. Furthermore, the potential for co-precipitation with Ferric ions will further decrease the solubility of Cr (OH)3. Dissolved Cr (VI) can be also precipitated as Cr (OH)3 in a reactive zone by the injection of ferrous sulfate solution into a reactive zone at appropriate concentrations.
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-2
log [Cr(lll)]
Cr(OH)3(am) -4
MCL -6
-8
4
6
8
10
12
14
pH Figure 4.30
Cr (III) concentration in equilibrium with Cr (OH)3.
Cr (VI) exists as chromate, CrO42– , under neutral or alkaline conditions and dichromate, Cr2O72– , under acidic conditions. Both species react with ferrous ion: Acidic conditions:
Cr2O72– + 6Fe2+ + 14H+ → 2Cr3+ + 6Fe3+ + 7H2O (4.27)
Neutral or alkaline condition: CrO42– + 3Fe2+ + 4H2O → Cr3+ + 3Fe3+ + 8OH– (4.28) Both Cr (III) and Fe (III) ions are highly insoluble under natural conditions of groundwater (neutral pH or slightly acidic or alkaline conditions). Fe3+ + 3OH– → Fe (OH)3 ↓
(4.29)
Cr3+ + 3OH– → Cr (OH)3 ↓
(4.20)
The addition of ferrous sulfate into the reactive zone may create acidic conditions and, hence, the zone downgradient of the ferrous sulfate injection zone may have to be injected with soda ash or caustic soda to bring the pH back to neutral conditions. Arsenic Precipitation Soluble arsenic occurs in natural waters only in the pentavalent, As (V) and trivalent, As (III), oxidation states. Although both organic and inorganic forms of arsenic have been detected, organic species (such as methylated arsenic) are rarely
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present at concentrations greater than 1 ppb and are generally considered of little environmental significance compared with inorganic arsenic species. Thus, this discussion focuses exclusively on the behavior of inorganic arsenic. Thermodynamics provides useful insight into the equilibrium chemistry of inorganic arsenic species. In oxygenated waters, As (V) is dominant, existing in anionic forms of either H2AsO4– , HAsO42– , or AsO43– over the pH range of 5 to 12, which covers the range encountered in natural groundwater. Under anoxic conditions, As (III) is stable, with nonionic (H3AsO3) and anionic (H2AsO3– ) species dominant below and above pH 9.22, respectively. In the presence of sulfides, precipitation of AsS (realgar) or As2S3 (orpiment) may remove soluble As (III) and exert considerable control over trace arsenic concentrations. The thermodynamic reduction of As (V) to As (III) in the absence of oxygen could be chemically slow and may require bacterial mediation.13 As noted in the previous section, injection of dilute solution of blackstrap molasses will create the reducing conditions for As (V) to be reduced to As (III) and also provide the sulfide ions for As (III) to precipitate as As2S3. These reactions are described by the following equations:10 Reduction of As (V) to As (III) under anaerobic conditions: HAsO42– → HAsO2 In the presence of S– – under anaerobic conditions: HAsO2 + S– – → As2S3 Within oxygenated zones in the aquifer, oxidation of Ferrous ion (Fe (II)) and Mn (II) leads to formation of hydroxides that will remove soluble As (V) by coprecipitation or adsorption reactions. The production of oxidized Fe-Mn species and subsequent precipitation of hydroxides are analogous to an in situ coagulation process for removing As (V). 4.2.2.2 Aquifer Parameters and Transport Mechanisms REDOX processes can induce strong acidification or alkalinization of soils and aquifer systems. Oxidized components are more acidic (SO42– , NO3–) or less basic (Fe2O3) than their reduced counterparts (H2S, NH3). As a result, alkalinity and pH tend to increase with reduction and decrease with oxidation. Carbonates are efficient buffers in natural aquifer systems in the neutral pH range. Many events can cause changes in REDOX conditions in an aquifer. Infiltration of water with high dissolved oxygen concentration, fluctuating water table, excess organic matter, introduction of contaminants that are easily degradable, increased microbial activity, and deterioration of soil structure can impact the REDOX conditions in the subsurface. However, there is an inherent capacity to resist REDOX changes in natural aquifer systems. This inherent capacity depends on the availability of oxidized or reduced species. REDOX buffering is provided by the presence of various electron donors and electron acceptors present in the aquifer.
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An engineered in situ reactive zone has to take into consideration how the target reactions will impact the REDOX conditions within and downgradient of the reactive zone, in addition to degrading the contaminants with the available residence time. Furthermore, careful evaluation should be performed regarding the selectivity of the injected reagents towards the target contaminants and the potential to react with other compounds or aquifer materials. Careful monitoring, short term and long term, should be performed to determine whether the natural equilibrium conditions can be restored at the end of the remediation process. In some cases modified biogeochemical equilibrium conditions may have to be maintained over a long period of time to prevent the reoccurrence of contaminants. 4.2.2.3 Contaminant Removal Mechanisms As noted earlier, the mechanisms used to reduce the toxicity of dissolved contaminants can be grouped into two major categories: transformation and immobilization. Examples of some of these mechanisms have been discussed earlier. Conversion of chlorinated organic compounds to innocuous end products such as CO2, H2O, and Cl− by either biotic or abiotic reaction pathways is an example of transformation mechanisms. Precipitation of Cr (VI) as Cr (OH)3 by either abiotic or biotic reaction pathways and subsequent filtration by the soil matrix is an example of immobilization mechanisms. It can be assumed, in most cases, that the end products of transformation mechanisms will result in dissolved and gaseous species and that the impact of these end products on the natural REDOX equilibrium will be short term. If the impact is expected to be significant, it can be controlled by limiting the reaction kinetics and transport of the end products from the reaction zone. Dilution and escape of dissolved gases will also help in restoring the natural equilibrium conditions in the reaction zone. Immobilization mechanisms, which include heavy metals’ precipitation reactions, in reality transform the contaminant into a form (precipitate) which is much less soluble. In addition, transport of dissolved heavy metals in groundwater should be considered a two-phase system in which the dissolved metals partition between the soil matrix and the mobile aqueous phase. Metal precipitates resulting from an in situ reactive zone may move in association with colloidal particles or as particles themselves of colloidal dimensions. The term colloid is generally applied to particles with a size range of 0.001 to 1 µm. The transport of contaminants as colloids may result in unexpected mobility of low solubility precipitates. It is important to remember that the transport behavior of colloids is determined by the physical/chemical properties of the colloids as well as the soil matrix. Generally, when fine particles of colloid dimensions are formed, flocculation naturally occurs unless steps are taken to prevent it. Even when the primary precipitates are of colloid dimensions, if they form larger lumps a stable dispersed transport cannot take place. These larger flocs will settle on the soil matrix. Metal precipitates may be pure solids (e.g., PbS, ZnS, Cr (OH)3) or mixed solids (e.g., (Fex, Cr1–x) (OH)3, Ba(CrO4, SO4)). Mixed solids are formed when various elements co-precipitate or due to interaction with aquifer materials.
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Colloidal precipitates larger than 2 µm in the low flow conditions common in aquifer systems will be removed by sedimentation. Colloidal precipitates are more often removed mechanically in the soil matrix. Mechanical removal of particles occurs most often by straining, a process in which particles can enter the matrix, but are caught by the smaller pore spaces as they traverse it. Colloidal particles below 0.1 µm will be subjected more to adsorptive mechanisms than mechanical processes. Adsorptive interactions of colloids may be affected by the ionic strength of the groundwater, ionic composition, quantity, nature, and size of the suspended colloids, geologic composition of the soil matrix, and flow velocity of the groundwater. Higher levels of total dissolved solids (TDS) in the groundwater encourage colloid deposition. In aquifer systems with high Fe concentrations, the amorphous hydrous ferric oxide can be described as an amphoteric ion exchange media. As pH conditions change, it has the capacity to offer hydrogen ions (H+) or hydroxyl ions (OH–) for cation or ion exchange, respectively. Adsorption behavior is primarily related to pH (within the typical range of 5.0 to 8.5), and at typical average concentrations in soil, the iron in a cubic yard of soil is capable of adsorbing from 0.5 to 2 pounds of metals as cations or metallic complexes. This phenomenon is extremely useful for the removal of As and Cr. 4.2.3
In Situ Denitrification
Nitrogen can form a variety of compounds due to its different oxidation states. In the natural ecosystem, most changes from one oxidation state to another are biologically induced. The nitrogen forms in Table 4.10 are of interest in relation to the subsurface environment. Table 4.10 Nitrogen Forms Present in the Subsurface Environment Nitrogen Compound Ammonia Ammonium ion Nitrogen gas Nitrite ion Nitrate ion
Formula
Oxidation State
NH3 NH+4 N2 NO2– NO3–
–3 –3 0 +3 +5
The unionized, molecular ammonia exists in equilibrium with the ammonium ion, the distribution of which depends upon the pH and temperature of the biogeochemical system; in fact, very little ammonia exists at pH levels less than neutral. Transformation of nitrogen compounds can occur through several mechanisms, including fixation, ammonification, synthesis, nitrification, and denitrification. Ammonification refers to the change from organic nitrogen to the ammonium form. In general, ammonification occurs during decomposition of animal and plant tissue and animal fecal matter and can be expressed as follows:
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Organic Nitrogen + Ammonifying Microorganisms → NH3 /NH+4
(4.31)
Nitrification refers to the biological oxidation of ammonium ions under aerobic conditions by the chemoautotrophic organisms called nitrifiers. Two specific chemoautotrophic bacterial genera are involved, using inorganic carbon as their source of cellular carbon: NH 4+ + O 2
Nitrosomonas Nitrobacter → NO 2− + O 2 → NO 3− bacteria bacteria
(4.32)
The transformation reactions are generally coupled and proceed rapidly to the nitrate form. In situ denitrification can be accomplished by organisms belonging to the genera Micrococcus, Pseudomonas, Denitrobacillus, Spirillum, Bacillus, Achromobacter, Acinetobacter, Gluconobacter, Alcaligens, and Thiobacillus, which are present in the groundwater environment. Denitrifying organisms will utilize nitrate or nitrite in the absence of oxygen as the terminal electron acceptor for their metabolic activity. If any oxygen is present in the environment, it will probably be used preferentially. The energy for the denitrifying reactions is released by organic carbon sources that act as electron donors. The microbial pathways of denitrification include the reduction of nitrate to nitrite and the subsequent reduction of nitrite to nitrogen gas. NO3– → NO2– → N2 ↑
(4.33)
In biological wastewater treatment processes employing denitrification, a cheap, external carbon source such as methanol is added as the electron donor. It has long been known that NO3– can be converted to N2 gas in anaerobic groundwater zones in the presence of a labile carbon source. In situ microbial denitrification is based on the same principle as conventional biological wastewater treatment systems, except that it is carried out in the subsurface by injecting the appropriate organic carbon source. Since methanol could be an objectionable substrate from a regulatory point of view, sucrose or sugar solution is an optimum substrate to be injected. It should be noted that in the hierarchy of REDOX reactions, NO3– is the most favored electron acceptor after dissolved oxygen. Hence, considerable attention should be focused in maintaining the REDOX potential in the optimum range, so that Mn (IV), Fe (III), sulfate reduction conditions or methanogenic conditions are not formed in the subsurface. Furthermore, since denitrification is a reduction reaction, alkalinity and pH tend to increase in the aquifer. Since the end product N2 gas will escape into the vadose zone and, hence, the aquifer system is not a closed system, increased alkalinity will be observed in the groundwater. If the NO3– concentration is not very high, this concern will be short lived.
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199
Perchlorate Reduction
Perchlorate has been widely used as a propellant in solid rocket fuel and has recently been identified as a contaminant in both groundwater and surface waters. Perchlorate is recognized by the USEPA as a potential health risk; California has set a drinking water action level of 18 ppb. Most of the perchlorate contamination in groundwater appears to have come from the legal discharge decades ago of then unregulated waste effluents containing high levels of ammonium perchlorate. Although ammonium perchlorate was released initially, the salt is highly soluble and dissociates completely to ammonium and perchlorate ions upon dissolving in water: NH4 ClO4 Æ NH4+ + ClO4–
(4.34)
It is likely that most of the ammonium has been biodegraded and the cation is now best viewed as mostly Na+ or possibly H+, especially where perchlorate (ClO4– ) levels are below 100 ppb. At those sites where contamination dates back decades, very little (if any) ammonium has been found.89 The persistence of perchlorate in groundwater aquifers results primarily from a combination of aerobic conditions and lack of an electron donor. A number of bacteria that contain nitrate reductases are capable of dissimilatory reduction of perchlorate.89,90 Many mixed cultures have reduced perchlorate, chlorate, chlorite, nitrate, nitrite, and sulfate under the right conditions. Inhibition of perchlorate reduction also has been observed in the presence of other substrates, particularly chlorate, chlorite, and sulfate.90 Chlorate reductase has been isolated from microorganisms that also possess nitrate reductase. Although most perchlorate strains may be denitrifying facultative anaerobes, not all denitrifiers are (per)chlorate reducers. Simultaneous reduction of NO3– and ClO4– has been demonstrated in laboratory studies.90,91 The conversion of chlorine in perchlorate to chloride requires the overall transfer of eight electrons. The sequence of intermediates involved in perchlorate reduction is as follows: Æ ClO 3- Æ ClO 2- Æ O 2 + Cl ClO -4 (chloride) ( perchlorate) (chlorate) (chlorite)
(4.35)
In situ bioremediation, via an IRZ, appears to be the most economically feasible, fastest, and easiest means of dealing with perchlorate-laden groundwater at all concentrations. Microbial transformation of perchlorate to chlorite occurs in the absence of oxygen as a result of anaerobic respiration. Anaerobic respiration is an energy yielding process in which the oxidation of an electron donor, such as an easily degradable organic substrate, is coupled to the reduction of an electron acceptor, such as perchlorate and chlorate. Chlorite can be inhibitory to microbial activity, and the transformation of chlorite to chloride and O2 is believed to be a nonenergy
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yielding enzymatic detoxification mechanism that protects the cell and allows the bacterium to use perchlorate and chlorate as electron acceptors.90,91 Implementation of an IRZ with the introduction of easily biodegradable electron donors such as hexoses, acetate, or lactate (without the presence of other electron acceptors such as SO42– ) should be able to reduce the concentrations of ClO4– present in the groundwater. A tenfold reduction of perchorate was achieved in column experiments at residence times of less than 48 hours. Laboratory column experiments have demonstrated that perchlorate degradation can be achieved at influent levels ranging from 0.1 to 1000 mg/L.91 The effluent levels achieved were in the range of 0.005 mg/L which is lower than the state of California drinking water action level of 0.018 mg/L. The author and his colleagues are currently involved in initiating field scale testing to implement in situ biodegradation of perchlorate. Since most of the perchlorate plumes are decades old and also due to its high solubility, there are significantly large sized groundwater plumes of this contaminant. Hence, it is very important to select the cheapest electron donor to create an in situ reactive zone (IRZ) to achieve perchlorate degradation in situ. The persistence of perchlorate itself is enhanced by the oxidative poise available within these plumes. Hence, it is equally important to select the cheapest electron donor to overcome the oxidative poise within these large plumes.
4.3 4.3.1
ENGINEERED AEROBIC SYSTEMS
Direct Aerobic Oxidation
The majority of the compounds in petroleum products are biodegradable at significantly faster rates under aerobic conditions. The amount of oxygen required for complete aerobic mineralization of one gram of hydrocarbon ranges from three to three and a half grams. In simplistic volumetric terms, 300,000 kilograms of oxygen-saturated water must be delivered and mixed in order to mineralize one kilogram of petroleum hydrocarbons. This illustrates the need to select the technically and economically most effective method of delivering O2 into the groundwater and also to maximize the efficiency of O2 utilization by the microorganisms in the subsurface. The total cost of a pound of dissolved O2 delivered into the subsurface could range from 0.80 to $10.00, depending on the method selected and the geologic and hydrogeologic conditions encountered at a site. The cheapest method of delivering dissolved O2, if hydrogeologic conditions are conducive, is by injecting dilute hydrogen peroxide (at about 100–1000 ppm concentrations) into the contaminated zone. Other methods of oxygen delivery include various methods of air injection and expensive methods such as oxygen release compounds. In addition to the petroleum hydrocarbons, other compounds more conducive for aerobic biodegradation are: nonchlorinated phenolic compounds, alcohols, ketones, aldehydes, etc. Among the chlorinated compounds chlorobenzene, methylene chloride, and vinyl chloride are among the commonly encountered contaminants that are biodegradable faster under aerobic conditions.
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The most significant biological mechanism for the degradation of chlorinated solvents is when they are used as a primary substrate. In direct oxidation reactions, the chlorinated compound acts as an electron donor and the microorganism uses molecular oxygen as an electron acceptor. The microorganism obtains energy and organic carbon from the degraded chlorinated compound. The more chlorinated compounds, PCE, carbon tetrachloride (CT), and hexachloroethane (HCA), are neither susceptible to aerobic oxidation nor degraded under anaerobic oxidizing conditions when used as a primary substrate.14 TCE undergoes slow aerobic degradation to trichloroethanol and then to acetic acid, but the reaction is not thermodynamically favorable. Therefore, discussion of aerobic oxidation and mineralization has always been focused on DCE and vinyl chloride (VC). Rates of aerobic oxidation are more rapid for the less chlorinated organics (DCE and VC) when compared to their reductive dechlorination rates. It has been well documented in literature that VC is oxidized directly to carbon dioxide and water. Aerobic oxidation of cis-1,2 DCE has been speculated. However, it could not be ascertained whether DCE was reduced to VC and then direct oxidation of VC produced carbon dioxide or direct oxidation of DCE occurred to produce carbon dioxide. Under aerobic conditions, chlorinated aliphatic compounds with one or two carbons per molecule can be transformed by three types of microbial enzymes:13 dehalogenases, hydrolytic dehalogenases, and oxygenases. Dehalogenases, which require reduced glutathione as a cofactor, dehalogenate the substrates by means of nucleophilic substitution. The first product of this degradation pathway is an S-choloralkyl-gluthathione, which is probably nonenzymatically converted to glutathione and an aldehyde. Hydrolytic dehalogenases hydrolyze their substrates, yielding alcohols. Oxygenases use molecular oxygen as a reactant for the attack on the halogenated compounds; the products could be alcohols, aldehydes, or epoxides, depending on the structure of the compound. Numerous chlorinated short-chain aliphatic hydrocarbons have been demonstrated to undergo aerobic transformation. However, compounds that have all the available valences on their carbon atoms substituted by chlorine, such as PCE or carbon tetrachloride, have never been shown to transform through any other but reductive pathways. Generally, as the degree of chlorination increases, the likelihood of aerobic transformation decreases (Figure 4.31); the opposite is true for anaerobic (reductive) transformations. Among the methane compounds, methylene chloride (MC) and chloromethane have been found to be amenable to aerobic microbial transformation. Pure cultures of the genera Pseudomonas and Hyphomicrobium have been isolated that can grow on methylene chloride as the sole carbon and energy source.13,15 Alkylhalides (haloalkanes), such as 1,2-dichloroethane (1,2-DCA), are frequently hydrolytically dehalogenated. Xanthobacter autotrophicus utilizes 1,2-DCA as sole carbon source. Complex communities consisting of methanotrophs and heterotrophs, which inhabit groundwater aquifers, mineralize 1,2-DCA. A Pseudomonas fluorescens strain isolated from water and soil contaminated by chlorinated aliphatic hydrocarbons was shown to utilize 1,2-DCA, 1,1,2-trichloroethane (1,2,1-TCA) and TCE, but not PCE or 1,1,1-TCA.13,15
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1
Relative Rate of Oxidation
Relative Rate of Reduction
1
0 0
0 CM CA VC
Figure 4.31
MC CF 1,2DCA 1,1DCA 1,1,2TCA 1,2DCE 1,1DCE TCE Increasing Extent of Chlorination
CT TCA HCE PCE
Methanes Ethanes Ethenes
Relative rates of oxidation and reduction from a range of C1 and C2 chlorinated compounds (adapted from Semprini et al., 1992).
4.3.1.1 Aerobic Cometabolic Oxidation Chloroalkanes, such as TCE, cis- and trans-1,2-DCE, 1,1,-DCE, and VC, are also transformed by several different physiological groups of aerobes. Methanotrophic communities consisting of methanotrophs that initiate the oxidative transformation, and heterotrophs which utilize the products of oxidation and hydrolysis, are very active in this respect, and can achieve complete degradation of chlorinated alkenes. The same communities fail to transform PCE, however, because this compound is too oxidized. Pure cultures of methanotrophs such as Methylosinus trichosporium OB3b or Methylomonas sp. MM2, have been shown to partially transform TCE, trans-1,2-DCE, and cis-1,2-DCE.13,14,15 Other microorganisms capable of transforming chlorinated alkenes belong to the genera Pseudomonas, Alcaligenes, Mycobacterium, and Nitrosomonas. All of these microorganisms, except the genus Nitrosomonas, are heterotrophs which grow on various organic substrates (e.g., toluene, cresol, phenols, propane, etc.); Nitrosomonas is a chemolitotroph which derives energy from oxidation of ammonia. All of them cometabolize chlorinated compounds such as TCE or 1,2-DCE while growing on their respective growth substrates; the haloalkenes are only fortuitously transformed, not utilized for growth. However, vinyl chloride seems to be an exception: it has been demonstrated that a Mycobacterium strain isolated from soil contaminated by VC could grow on VC as a sole carbon and energy source.92 Aerobic cometabolism of chlorinated compounds at low concentrations by methane- and propane-utilizing bacteria is well documented. In comparison, butaneutilizing bacteria are less susceptible to the toxic effects of elevated chlorinated
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compound concentrations. Butane is approximately four times more soluble in groundwater than methane. Butane injection results in large radii of influence at injection wellheads. The difficulty of utilizing alkanotrophic bacteria stems from the low solubility of alkanes and the difficulty of maintaining homogeneous concentration of the dissolved alkane within the reactive zone. Methanotrophs grow on C1 compounds as sole carbon and energy sources. Their catabolic oxygenases are methane monooxygenases (MMO) that incorporate one atom of oxygen from the oxygen molecule into methane to yield methanol.12,14,92 This alcohol is further oxidized via a series of dehydrogenation steps, through formaldehyde and formic acid, to CO2 that is the final product of catabolism. MMO enzymes utilize molecular oxygen as a reactant, and require a reduced electron carrier to reduce the remaining oxygen atom to water. MMO enzymes have relaxed substrate specificity, and will oxygenate many compounds that are not growth substrates for methanotrophs. Such compounds include various alkanes, alkenes, ethers, alicycles, aromatics, nitrogen heterocycles, as well as chlorinated alkanes, alkenes, and aromatics.13,93 Two types of MMO have been suggested: a particulate (membrane-bound) and a soluble enzyme.13 The soluble MMO (purified from Methylosinus trichosporium OB3b and Methylococcus capsulatus [bath]), produced under conditions of copper limitation and increased oxygen tension, has been considered to have broader substrate specificity. It has been stated that only the soluble MMO can transform TCE. However, recent findings indicate that the particulate MMO in some methanotrophs may be as effective in the transformation of chlorinated solvents as the soluble MMO. Since the soluble MMO is not constitutively expressed, whereas the particulate MMO is, the latter methanotrophs (Methylomonas sp.) have a significant potential for in situ bioremediation. Thus TCE can be transformed (upon the induction of the oxygenase enzyme by its substrate) in the presence of the microorganismal growth substrate (cometabolism), or in its absence (resting cells transformation). However, TCE is not utilized by the bacteria as a carbon, energy, or electron source; this transformation is only fortuitous. Based on the findings with methanotrophs, it can be concluded that TCE is most likely oxygenated to TCE-epoxide (Figure 4.32).13,15,93 The epoxide is unstable and is quickly nonenzymatically rearranged in aqueous solution to yield various products including carbon monoxide, formic acid, glyoxylic acid, and a range of chlorinated acids. Recent findings with purified MMO from Methylosinus trichosporium OB3b indicate that TCE-epoxide is indeed a product of TCE oxygenation. In nature, where cooperation between the TCE oxidizers and other bacteria (most prominently heterotrophs) occurs, TCE can be completely mineralized to carbon dioxide, water, and chloride. Toluene, phenol, and cresol oxidizers, such as Pseudomonas putida or P. cepacia, express the TCE transformation activity upon induction by their aromatic substrates. These bacteria have a great potential for remediation of groundwater aquifers contaminated by mixtures of gasoline or jet fuel (or other petroleum derivatives), and chlorinated solvents, such as TCE, DCE, or VC. If the aromatic contaminants are not present, however, bacterial growth substrates need to be injected into the site in
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Methane oxidation (normal reaction) with methane monooxygenase H
OH
MMO
H C H
H C H
H
CO 2
H NADH2, O
NAD, H2O
3NAD, H2O
3NADH 2
TCE epoxidation (cometabolic dechlorination reaction) with MMO Cl
Cl C C
Cl
Cl
MMO
H
Cl NADH2, O2
NAD, H O 2
O Cl C C H
-
2CO2, 3Cl , 3H 2NAD, 3H2O
+
2NADH 2
(other microorganisms)
Figure 4.32
The top reaction shows how methanotrophs (“methane eaters”) produce the enzyme methane monooxygenase (MMO) in the process of converting methane (CH4– ) to CO2. The bottom reaction shows how MMo then causes the conversion of TCE to CO2 and HCl. NADH2 serves as the carrier of electrons released from methane and TCE. Note: NAD = nicotinamide adenine dinucleotide; NADH = reduced nicotinamide adenine dinucleotide.
order to stimulate the transformation of chlorinated solvents. In this situation, methanotrophs become more attractive agents of bioremediation because methane, their preferred substrate, is a nontoxic and inexpensive chemical. Once methane and oxygen are injected into the site, methanotrophs (if present) will start cometabolizing chlorinated solvents, as well as a great number of other contaminants (see below), and the accompanying heterotrophs will mineralize their transformation products. As mentioned earlier it is important to maintain reasonably high and uniform O2 and CH4 concentrations to achieve significant methanotrophic degradation. 4.3.1.2 MTBE Degradation There is a growing body of evidence from laboratory and field studies that MTBE can be degraded under aerobic conditions either by direct metabolism (when MTBE serves as the carbon and energy source for microbial growth) or cometabolism. Evidence on natural attenuation of MTBE is presented in Chapter 3. Microbes capable of MTBE degradation under aerobic conditions may be present at most sites, but perhaps under nonoptimum biogeochemical conditions to significantly reduce the migration of MTBE. Furthermore, these aerobic processes would be expected to be limited at many sites since the MTBE plume is migrating down a largely anaerobic path. Thus any approach to initiating or enhancing in situ aerobic biodegradation of MTBE must overcome at least two major hurdles: 1) creating steady aerobic conditions over the long term and 2) generating enough microbial biomass to accomplish the treatment at a reasonable rate. In situ chemical oxidation of MTBE has not been very successful due to the incomplete oxidation and formation of undesirable byproducts such as tertiary-butyl formate, tertiary-butyl alcohol, methyl acetate, acetone, and formic acid. Hence, enhanced biodegradation of MTBE has to be optimized and engineered based on the positive evidence found recently.
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A pure culture that degrades MTBE has been isolated;94 however, the proliferation of this organism in the natural environment may be questionable. Two recent field tests, at Pt. Hueneme, CA, and Vandenburg AFB, CA, have provided positive results to the point that the previously held notion that MTBE is not aerobically biodegradable is not true anymore. At Pt. Hueneme, controlled testing was done in plots where pure oxygen only, and pure oxygen with bioaugmentation of enriched cultures were injected in addition to a control plot where only natural conditions were allowed to exist. The bioaugmented plot showed significant reductions of MTBE at ppm range concentrations within a very short time period. The plot where only O2 was injected also showed similar reductions in MTBE concentrations — after a lag period of months, however. The question still to be answered is whether the microorganisms responsible for MTBE degradation are obligate aerobes and thus a reasonably high DO concentration has to be maintained in the groundwater. The Vandenburgh AFB field test also provided positive results and raised the possibilities of implementing engineered in situ aerobic biodegradation of MTBE. In two separate long term field tests, dissolved oxygen was released to the MTBE plume through pressurized tubing via controlled interception trenches acting like permeable walls. In both field tests, significant MTBE reductions took place in the presence of increased levels of oxygen. MTBE degradation ceased when O2 injection stopped, thus indicating that degradation was conclusively aerobic. Others have reported reductions in MTBE concentrations during in situ air sparging projects.54 Stripping may be the dominant mechanism of MTBE removal from groundwater in these projects; however, the contribution by enhanced biodegradation due to increased levels of O2 in the groundwater cannot be discounted, albeit at low levels compared to the mass removal by stripping. At the Pt. Hueneme and Vandenburgh AFB field tests the means of O2 injection was achieved by injecting pure O2. However, scaling up such a system to implement an engineered aerobic IRZ to address large MTBE plumes will be uneconomical, particularly when many of these plumes have migrated beyond property lines. Testing the injection of dilute hydrogen peroxide to sustain the reasonably higher levels of DO, which seems to be a requirement for aerobic MTBE degradation, will occur soon. Another means of providing enhanced levels of DO is through the implementation of in-well sparging. Injecting hot air into the well will also enhance the mass removal by air stripping at reasonable air to water ratios. Recirculated water saturated with oxygen will create an in situ aerobic zone around the well and thus enhance the aerobic degradation of MTBE (Figure 4.33). Based on recent field observations, enhancing MTBE degradation within engineered anaerobic zones may be a viable option. These zones have to be maintained under methanogenic conditions.
4.4
IN SITU CHEMICAL OXIDATION SYSTEMS
Chemical oxidation processes have been widely used for treatment of organic contaminants in wastewaters. Because they are aggressive and applicable to a wide variety of compounds, using these processes, coupled with delivery technologies for
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Heater Air to treatment
Air compressor
an Cle er wat
Water table
Enhanced biodegradation
Packer Air injection pipe
Contaminated groundwater
Figure 4.33
Heated in-well sparging for enhanced stripping and aerobic biodegradation of MTBE.
in situ remediation of contaminated groundwater or subsurface soils, has received increased attention. In situ chemical oxidation is an innovative technology with widely varying opinions regarding its effectiveness on a range of contaminant types. The oxidants frequently used for this purpose are hydrogen peroxide, permanganate, and ozone. In situ chemical oxidation is achieved by delivering potent chemical oxidants to contaminated media so that the contaminants are almost completely oxidized into H2O, CO2, and chloride ions or converted into innocuous compounds commonly found in nature. In situ chemical oxidation will most likely be selected to address remediation of what may be considered “difficult sites” having one or more of the following characteristics:77 low permeability soils, highly stratified soils, low-volatility target compounds, target compounds with low in situ degradation kinetic constants, and dense nonaqueous phase liquids (DNAPLS). 4.4.1
Advantages
The primary advantages of in situ chemical oxidation (ISCO) technologies is their relatively high speed destruction of contaminants. The cost of reagents is relatively high compared to biological systems, so application is generally far more costly than bioremediation systems, but significantly lower than other active source removal technologies, such as in situ thermal treatment or flushing using surfactants or cosolvents. Since the reaction is nearly immediate, treatment is far more rapid than biological techniques and can be faster than thermal or vapor recovery technologies.
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The advantages of in situ chemical oxidation can be summarized as follows: • The ability to oxidize dense nonaqueous phase liquids (DNAPLs) if targeted properly • A reduction in overall treatment time, allowing the site to reach closure relatively sooner • The elimination of capital intensive pump and treat systems • The ability to address contamination in situ without disturbing above ground structures
In situ chemical oxidation can be used as a stand alone treatment or in conjunction with other technologies such as bioremediation. The nature and location of the contamination, size of the source zone, type of soil, and hydrogeology play a significant role in choosing the most effective type of ISCO treatment system. In situations where contamination covers a vast area, economics will dictate the extent to which ISCO is used, but, in many cases, this is a cost effective pretreatment to bioremediation and natural attenuation. 4.4.2
Concerns
The primary concern is ensuring the health and safety of workers. Chemical oxidation is an exothermic reaction generating heat that can increase temperature and pressurize gases depending on loading and reaction rates. Strong oxidants are corrosive and potentially explosive. The design and operation of any ISCO system must take into account the hazards of the chemicals and the potential for vigorous, uncontrolled, exothermic reactions in the subsurface. Site conditions that would warrant particular attention in the planning stage include paved sites for which vapor pressures could build up under the pavement, sites with preferential flow paths, or utility corridors through which vapors could migrate. A significant performance concern is that the oxidation reaction is not complete, and significant DNAPL accumulations remain in untreated areas in the subsurface. Even a small percentage of the original DNAPL mass can result in a rebound in the groundwater concentrations after treatment to levels similar to those measured before treatment, or at least above levels of regulatory concern. In addition, the migration of contamination to previously uncontaminated areas due to thermal gradients caused by exothermic reactions and to trapping contaminants in gas bubbles created by the reactions should be taken into account. Another concern is the possibility of increased volatile emissions of volatile organic compounds. Oxidation can cause significant heat generation and water vapor production. As a result, in situ steam stripping is a potential mechanism for contaminant loss, particularly for highly volatile compounds like chlorinated solvents. For example, in cases where the hydrogen peroxide concentration exceeds approximately 11%, enough thermal energy can be released to cause water to boil, leading to a significant concern regarding vaporization losses.95
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
Oxidation Chemistry
The oxidation chemistry of chlorinated solvents is relatively well understood. Oxidants attack the C-C bonds in these molecules. The double bonds that characterize chlorinated ethenes are far more reactive than the single bonds of chlorinated ethanes, and hence PCE and TCE are far more susceptible to oxidation than TCA, for example. However, the chloroethanes are often claimed to be susceptible to oxidation as well.93,95,96 Current theory is that the oxidants cause formation of an unstable epoxide that then breaks down to yield ketones and aldehydes. These products may also be susceptible to further oxidation, eventually yielding carbon dioxide, water, and chloride. Several oxidants have been employed in the recent past for ISCO applications. For DNAPL sites, the most common oxidants used have been hydrogen peroxide (H2O2) and potassium permanganate (KMnO4). Permanganate is more expensive than hydrogen peroxide, but it is also more stable and effective over a broad pH range. Ozone (O3) is the strongest oxidant available, with an oxidation potential (E°) of 2.07 v. However, ozone is a gas and therefore most suitable for treating the vadose zone, or possibly LNAPL accumulations in the capillary fringe. Persulfate (S2O8–2 ) salts are also available, with an E° of 2.01v, but these oxidants are relatively expensive and require thermal activation.93,95 The relative reaction kinetics of the different oxidants are shown in Figures 4.34 and 4.35. Hydrogen peroxide apparently works through two mechanisms: free radical generation and direct oxidation. The direct oxidation has an E° of 1.76 v, and free radical formation (H2O2 ⇒ 2OH· + 2H+ + 2e–) has an E° of 2.76 v. The latter relies
Resistance to Oxidation
Oxidant Strength High
● Perchloroethylene
● Hydroxyl Radical
● Trichloroethylene
● Permanganate
● Vinyl Chloride
● Ozone
● Phenanthrene
● Hydrogen Peroxide
● Benzene
● Hypochlorite
● Hexane
● Oxygen Low
Figure 4.34
Relative strength of oxidants and relative resistance of some common contaminants to chemical oxidation.
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Figure 4.35
209
Relative reaction kinetics of various oxidants.
on so-called Fenton’s chemistry, in which iron acts as a catalyst; therefore, iron is often added with the hydrogen peroxide. In addition, pH adjustment is common because oxidation is more rapid under acidic conditions. Permanganate has an E° of 1.70 v and yields MnO2 as an insoluble precipitate under most conditions.93,95 Catalysts and pH control are not needed for permanganate oxidation. The stoichiometry of complete oxidation reactions yields the following weight ratios for permanganate (expressed as KMnO4:contaminant): PCE (1.3:1); TCE (2.4:1); DCE (4.4:1); and VC (8.5:1). Of course, this stoichiometry ignores the significant oxidant demand due to other reduced and natural organic compounds in the subsurface, which can be significant. Optimal use of the in situ chemical oxidation technology is very much dependent on understanding oxidant demand from contaminant oxidation and matrix oxidant demand. Matrix oxidant demand refers to oxidant consumption that can be attributed to background soil and groundwater conditions (Figure 4.36). Matrix demand can be derived from oxidation of natural organic matter (NOM), reduced metals, carbonates, sulfides, etc. Matrix demand can be highly variable (depending on the reductive poise of the contaminated zone), influenced by background geochemical conditions, and, since permanganate reaction rates are second order, also will depend on the permanganate solution concentration. The oxidant demand caused by the nontarget compounds can range from 10 to 100 times (or even higher) of the stoichiometric demand caused by the target contaminants. Hence, it is more important to look at the chemical oxidation demand of the system than at the total organic carbon (TOC) as an evaluation parameter for chemical oxidation. It should be noted that destruction of natural organic matter can release additional contaminants, adsorbed to the organic matter, into the dissolved phase (Figure 4.37) before being completely destroyed by the oxidant. This phenomenon is the primary factor contributing to the rebound effects of the target contaminants during chemical oxidation. The most commonly observed mobilization of metals, during ISCO, is oxidation of precipated Cr3+ to the dissolved Cr6+. The amount of Cr6+ mobilized will obviously depend on the background chromium concentration in the soils. Literature reports indicate that this dissolved Cr6+ will reattenuate within a short time frame and distance. The advantages of peroxide as an oxidant include relatively low regulatory resistance, more field experience in its use than permanganate, and a sparcity of
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Figure 4.36
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Oxidative poise of natural environment and increased potential demand of oxidants.
Chemical oxidation demand
Natural and non-target organic matter
NOM destruction effect
Initial dissolved phase target contaminant
Residual level of NOM
Progress of treatment Figure 4.37
Natural organic matter destruction releases additional dissolved phase contaminants.
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byproducts of oxidation. Disadvantages include the need for pH control in some cases and difficulties in controlling in situ heat and gas production. Permanganate is more expensive and more stable than peroxide, and is effective over a broad pH range. Oxidation also produces manganese levels, which will precipitate and potentially cause reduced porosity. Increases in dissolved manganese levels are also a potential regulatory concern depending on the geochemistry, as is the purple color of groundwater containing unreacted permanganate. Ozone has been used mostly for vadose zone treatment. It is less costly than permanganate or peroxide, but the most significant factor in choosing ozone is that it must be applied as a gas. Gases may disperse further in the unsaturated zone than liquids, but vapor recovery and possible treatment can add considerable cost. 4.4.3.1 Hydrogen Peroxide Hydrogen peroxide (H2O2) is typically used together with Fe (II) to form Fenton’s reagent. In Fenton’s reagent, H2O2 is decomposed by Fe (II) to produce highly reactive hydroxyl radicals as expressed by Equation 4.36: Fe2+ + H2O2 ⇒ Fe3+ + OH• + OH–
(4.36)
The hydroxyl radical can nonselectively attack the C-H bonds of organic molecules and is capable of degrading many solvents, chloroalkenes, esters, aromatics, and pesticides. The major advantages over other oxidation processes of using Fenton’s reagent to treat hazardous wastes can be summarized as:95 1) there are no chlorinated organic compounds formed during the oxidation process as in chlorinating; 2) both iron and hydrogen peroxide are inexpensive and nontoxic; 3) there are no mass transfer limitations because the reaction is homogeneous; 4) no light is required as a catalyst and, therefore, the design is much simpler than ultraviolet light systems; and 5) H2O2 can be electrochemically generated in situ, which may further increase the economic feasibility and effectiveness of this process for treating contaminated sites. During treatment, particulates can be generated and the pore size and continuity can, therefore, be modified with fine-grained media. As a result, the permeability can be impacted. In Fenton’s mechanism, reactions with H2O2 cycle iron between the +II and +III oxidation states, yielding OH• and other byproducts. Because OH• is a powerful indiscriminate oxidant that reacts with many compounds at near diffusion-controlled rates,97,98 H2O2 and iron have been used to generate OH• and oxidize undesirable contaminants in soils and aquifers.93,95 A wide range of organic compounds (TCE, BTEX, PCP, naphthalene, and pesticides) that are common contaminants of groundwater and soil have moderate to high reaction rate constants with OH• (108 – 1010M–1s–1). The stability of H2O2 increases with decreasing pH in Fenton systems, and oxidation efficiency is optimum under acidic conditions.95,97 Under acidic conditions and with an excess of Fe2+, the hydroxyl radical generated can further react with Fe2+ to produce Fe3+ (Figure 4.38a):76 Fe2+ + OH• ⇒ Fe3+ + OH–
(4.37)
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H2O 2
Fe2 + R'H + CO 2
RED RED
OX OX
Fe3 + H++R OH - + OH -
OX
RED
OH RH Figure 4.38a
Fenton’s reagent — idealized reactions.76
By properly controlling experimental conditions, ferric iron can be regenerated back to ferrous iron by a subsequent reaction with another molecule of H2O2: Fe3+ + H2O2 ⇒ Fe2+ + HO2• + H+
(4.38)
The HO2• radicals produced (Equation 4.38) have been shown also to participate in oxidation of some organic compounds, although they are much less reactive than OH•. Based on Equation 4.38, a low pH range of 2 to 4 is preferred to facilitate the generation of hydroxyl radicals, although the reaction is feasible up to neutral pH.99 Almost all organic compounds can be treated in situ by this technology. Limitations to Fenton-based remediation strategies arise from excessive H2O2 decomposition via nonproductive reactions (those that do not result in OH• production), reaction of OH• with nontarget species (scavenging), insufficient iron or H2O2 for radical production, and slow reaction of OH• with the target compound.93,95 For example, REDOX cycling of manganese between the +II and +IV oxidation states consumes H2O2, but does not yield OH•. Common groundwater anions (NO3– , SO42– , C1–, HPO42– , HCO3– , CO3–2 ) react with OH• and may be a source of treatment inefficiency. Furthermore, because H2O2 is generally present at high concentrations in Fenton systems and has a moderate rate constant for reaction with OH• (2.7 × 107M–1s–1), peroxide is itself a primary source of inefficiency in Fenton-driven systems.95
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Idealized reactions of Fenton’s reagent and potential reduction in efficiencies due to disproportionation are shown in Figures 4.38b and c, respectively.76 The contaminants of particular interest include chlorinated solvents (e.g., TCE, PCE), polyaromatic hydrocarbons (e.g., naphthalene), PCP, and petroleum products (e.g., BTEX). Some of these chemicals are very difficult to biodegrade or may take exceedingly long time in many subsurface settings.
H2O 2
OX
H2O + O 2
Fe2 + RED
Fe3 +
OH - + OH Figure 4.38b
Fenton’s reagent failures — Fe3+ catalyzes disproportionation.76
Major concerns for this technology are related to potential ecological effects and chemical handling. The introduction of acid solution can have potential effects on the ecosystem. During the reactions, both OH– and H+ can be produced; however, their quantities are relatively small compared with the acid introduced and thus would have no significant effect on the pH of the media. Because large quantities of chemicals are required for the treatment, it could be hazardous to handle them. In addition, special measures may be taken during the delivery process because H2O2 can easily break down into H2O vapor and O2, leading to fugitive emissions of VOCs and pressure buildup. One benefit of decomposition of H2O2 is that the released O2 can stimulate aerobic biological activity. 4.4.3.2 Potassium Permanganate Potassium permanganate (KMnO4) has been used in treatment of drinking water and wastewater for decades because it can effectively oxidize many water impurities, including phenol, Fe2+, S2–, and taste and odor-producing compounds. Only within the past few years has it been used more frequently as an alternative chemical oxidant for ISCO. KMnO4 is a dry crystalline material that turns bright purple when dissolved
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H2O 2
Fe2 +
OX
RED
Fe3 +
Precipitate Figure 4.38c
OH - + OH -
Fenton’s reagent failures — Fe3+ is lost to precipitation.76
in water. The purple color acts as a built-in indicator for unreacted chemical. Reacted KMnO4 is black or brown, indicating the presence of the MnO2 precipitate — a natural compound present in soil. Other KMnO4 oxidation byproducts include CO2, H2O and the potassium ion K+. Limitations of KMnO4 include its low solubility (65 g/l at 68°F) and its inability to oxidize petroleum compounds effectively. Sodium permanganate (NaMnO4) is an oxidant that performs very similarly to KMnO4; its attributes and limitations are much the same as KMnO4. However, NaMnO4 has a much higher solubility in water, allowing it to be used for ISCO at a much higher concentration. NaMnO4 is more expensive than KMnO4 on a pound-perpound basis and users have to be concerned about safety during handling and storage. Reaction of KMnO4 with organic compounds produces manganese dioxide (MnO2) and carbon dioxide or intermediate organic compounds. The kinetics of reaction between permanganate and contaminants are obviously an important factor in the overall treatment success achieved. It has been reported that oxidation of TCE by KMnO4 is second order, with a fast second order constant. An apparent limitation with the reactive hydroxyl radical (OH•) is that it strongly reacts with common inorganic species in groundwater such as carbonate and bicarbonate. However, permanganate, a metal-oxo reagent, does not apparently rely on generating a hydroxyl radical to oxidize chlorinated ethenes as the other oxidants do. Experience indicates that metal-oxo reagents can attack a double carbon-carbon bond powerfully through direct oxygen transfer.76 [Org] + MnO4– → MnO2 + CO2 or [Org]ox
(4.39)
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where [Org]ox is the oxidized intermediate organic compound. Permanganate ions preferentially attach carbon-carbon double bonds, in a manner similar to the attack of ozone.76 A manganate ester forms in the first stage of the reaction and rapidly decomposes to form a glycol, as shown in Figures 4.39a and b. Manganese dioxide (MnO2) precipitates from the oxidizing aqueous solution. The glycol is cleaved under high permanganate concentration or acidic conditions to form aldehydes or ketones. Aldehydes are likely to be further oxidized to carboxylic acids.76,96
O
O
O Mn
Mn
H+
O
O
C
C OX
H O
O
C
C
OH O
MnO 4
C H
HO
O
H
H
MnO 4
H H
C H
Glycol Aldehyds
H
Cyclic Hypomangnate Ester
Mn
O
OH O C
C
O
H
MnO 4 HO
Glyoxylic Acid
O H
C
OH O C
C
O OH Oxalic Acid
O H
Formaldehyde
Figure 4.39b
C
H
C
nd n bo tio C nta C- me g fra
Ethylene
O Mn O O
H
O C C
C
Permanganate oxidation of an alkene.
C
H
O
H
H
H
H+
O
RED
O
Figure 4.39a
O
HO
H
C
OH
Formic Acid
The oxidation of ethylene in a neutral to weak acidic condition.
When permanganate is used to oxidize chlorinated ethenes, chlorinated intermediates such as phosgene or formyl chloride might be produced. However, it was observed that rapid dechlorination of the manganate ester took place when
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permanganate ion was used to dechlorinate TCE and other chloroethenes.76,93,95,96 For tests run at pH ranging from 4 to 8, oxidation of the manganate ester to carbon dioxide was more rapid than the permanganate attack on the solvents. It was also noted that only the permanganate ion, MnO4– , participated in oxidation, and that manganese dioxide (MnO2) was the only manganese-bearing product of the several reactions. Several studies have been published describing permanganate oxidation of chlorinated ethenes, including reports of both field and laboratory applications.76,93-95 A common element of these studies is the focus on oxidation of the contaminant compounds, without evaluation of oxidation byproducts that may result from reaction of permanganate with naturally occurring compounds and organic species associated with solvent wastes. One particular problem in laboratory studies is that permanganate is typically applied as an excess reagent — an approach that simplifies analysis of reactions that are first-order in both reactants. But the excess permanganate oxidizes many potential reaction byproducts. In field applications, permanganate cannot be applied as an excess reagent across the entire aquifer and the appearance of ketones, aldehydes, and other reaction byproducts cannot be ruled out.95 Even when permanganate is applied as an excess reagent, byproducts such as acetone and butanone may accumulate during oxidation of contaminated aquifer soils. In an unpublished bench-scale study, a 3% potassium permanganate solution was applied to aquifer soils contaminated by an oil-solvent mixture. The application continued until permanganate depletion, during passage through the aquifer soils, to negligible levels. Newly formed acetone, 2-butanone and other oxidation products were measured in aqueous-phase samples throughout the test application.76 The compounds that can be oxidized by permanganate in addition to alkenes include aromatics, PAHs, phenols, pesticides, and organic acids. The optimum pH range is 7 to 8, but they are effective over a wide range. Because Mn is an abundant element in the Earth’s crust and MnO2 is naturally present in soils, introduction of KMnO4 to soils as well as production of MnO2 would not be an environmental concern. KMnO4 is as effective as or more effective than H2O2 in oxidizing organic compounds. Furthermore, KMnO4 is more stable and easier to handle. The potential problem is that MnO2 particles will be generated and permeability loss is possible. 4.4.3.3 Ozone Like hydrogen peroxide and permanganate, ozone is a strong oxidant that can quickly oxidize organic compounds once in contact. Compared to other technologies, in situ ozonation offers several advantages:76,93 • It is much easier to deliver ozone to the contamination zone than aqueous oxidants. • No volatilization of target chemicals is required and, therefore, mass transfer limitations associated with soil venting can be overcome. • In situ ozonation would likely be more rapid than biodegradation or soil venting processes, and thus reduce the remediation time and treatment costs.
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• • • •
Ozone can be electrically generated from air on site. In situ ozonation is conceptually similar to soil venting processes. Both vertical and horizontal wells can be used to inject ozone. Little degradation of ozone occurs during injection and on-site handling is relatively easy. • Similar to H2O2 and permanganate, ozone can be used to treat a variety of organic compounds. • Ozone is very reactive and corrosive to materials.
Ozone reacts quickly in the subsurface and does not migrate long distances from the point of delivery. Currently, ozone is used to treat chlorinated solvents, polyaromatic hydrocarbons, and petroleum products in situ. Ozone is unstable in water. The degradation of ozone involves a complex cyclic process that may be promoted or inhibited by various substances. In natural systems, the degradation of ozone may be initiated by various substances including the hydroxide ion OH–, and natural organic matter (NOM). Bicarbonate and carbonate ions and other hydroxyl radical scavengers will inhibit the degradation by ozone. Many organic compounds are able to initiate, promote, or inhibit the chain-reaction processes of ozone decomposition and degradation. Zwitterions, also known as dipolar ions, are neutrally charged but strongly polarized molecules that behave as ions.76 Many molecules exhibit a degree of dipolar behavior — zwitterions can be sufficiently dipolar to confer substantial reactivity. Another zwitterion behavior is exemplified by glycine, an amino acid:glycine acts as a base when titrated with acid, and acts as an acid when titrated with base. Ozone is a zwitterion comprised of three oxygen atoms, as shown in Figure 4.40. A resonant double bond concentrates negative charge in the terminal oxygen atom bound by the single bond.76 Although the diagram suggests a concentration of positive charge in the central portion of the molecule, the central atom exerts a pull on the electrons from the resonant double bond, transferring some of the positive charge to the double-bonded terminal oxygen.
O
O +
Figure 4.40
O
+
O
O
O
-
Zwitterion behavior.
The double-bonded terminal oxygen atom in ozone can initiate electrophilic attack on carbon-carbon double bonds, as shown in Figure 4.41. As the electron pair from the alkene migrates toward the electrophilic oxygen atom, the opposite carbon atom becomes electrophilic, attracting the singly bonded oxygen atom into a molozonide bridge. This highly unstable compound breaks and reforms as an ozonide,
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O O
O O
O
O
C C
C
C
C
O O O
O
C O Figure 4.41
C O
C
Ozonation of an alkene.
which also decomposes spontaneously. The reaction is completed by the formation of two ketones (or aldehydes) and water. The kinetics of ozone attack on chlorinated ethenes is highly influenced by the steric hindrance caused by chlorine atoms.76 The dramatic increase in reactivity to ozone from tetrachloroethene to trichloroethene is due to two factors: the reduction in steric hindrance that follows from elimination of a chlorine atom, and the reduction of the carbon atom from C (II) to C (I), making the electron pair more available to electrophilic attack (oxidation). Reaction rates of ozone and simple alkenes such as styrene are very high, while alkanes, alcohols, aldehydes, and ketones are only slightly reactive to ozone (Table 4.11). The final decomposition products of the ozonation of chlorinated ethenes are formaldehyde (CH2O), and phosgene (CCl2O). Formyl chloride (CHClO) is a theoretical product which is unreported in the chemical literature and presumable is unstable. Phosgene decomposes rapidly in water and is not expected to be observed; and formaldehyde rapidly biodegrades in the highly aerobic post-ozonation environs.76 4.4.4
Application
In general, more than a single application of oxidant is required to meet most cleanup standards. Several reinjections at periodic intervals have been used for more thorough treatment. Recently, continuous injection using recirculation of amended waters has been used to maximize the utilization efficiency of the oxidant as well
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Table 4.11 Kinetic Constants for Ozonation of Various Organic Compounds76,98 Compound
Ko3 (M–1s–1)
Tetrachloroethene Trichloroethene 1,1-Dichloroethene 1,1-Dichloroethane Cis-1,2-Dichloroethene Trans-1,2-Dichloroethene Styrene Formaldehyde Acetaldehyde
0.1 17 110 0.12 800 5700 3 × 105 0.1 1.5
as to augment the distribution rate within the reactive zone. A comparison of the properties of the three commonly used oxidants is presented in Table 4.12. Table 4.12 Comparison of Oxidants Fenton’s Reagent Physical State Molecular Composition OH• formation
Liquid OH• Yes
Oxidation Potential Reaction Times Contaminant Range Potential to Entrance
2.76 V Very Fast Many Organics Yes If the pH Conditions Are Not Low Yes
Metal Mobilization/Potential
Permanganate
Ozone
Liquid MnO4– Under Very Limited Conditions 1.70 V Slow Few Organics Unlikely
Gas O3 Under Certain Conditions 2.07 V Fast Some Organics Yes
Cr3+ → Cr6+
Yes
For single or multiple injections, permanent or temporary injection points are established, and an aqueous solution containing the oxidant and any needed catalysts is injected under pressure. The oxidant (and catalyst) concentration, the target pH, the injection well spacing (i.e., radius of influence), the number of injections, and the injection pressure are all important design parameters affecting cost and performance. The oxidation reactions occur in the aqueous phase, and NAPL and sorbed phases must be targeted and treated either by interfacial contact with or mass transfer to the aqueous phase (Figure 4.42). Successful prediction of overall rates of mass removal would require rate expressions both for nonequilibrium dissolution and oxidation. In the conceptual model (Figure 4.43a), dissolution mass transfer, driven primarily by aqueous phase chlorinated contaminant concentration gradients, is enhanced by the oxidation reaction that increases these gradients (Figure 4.43b). The efficiency of chemical oxidation for treatment of NAPLs is based on the conceptual model that attributes an
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Adsorbed DNAPL
Pooled DNAPL Figure 4.42
Chemical oxidation systems to address DNAPLs will have to target the mass of DNAPL.
Concentration Without Oxidant
CO
With Oxidant
DNAPL Phase
Oxidant Concentrations
Dissolved Phase Mass Flux
Figure 4.43a
Conceptual model describing mass removal by in situ oxidation.
increased rate of DNAPL mass transfer to chemical oxidation within the stagnant film boundary layer. As the aqueous solvent gradient is increased, the dissolution mass flux is increased. Simultaneously, the concentration gradient of the oxidant would be increased, causing an increase in oxidant mass flux towards the DNAPL/water interface.
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DNAPL
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GW / Oxidant solution
Boundary layer Figure 4.43b
Conceptual model of interface mass transfer effects of chemical oxidation.
Permanganate oxidation of a DNAPL can yield manganese oxide solids that may deposit on the interface and could result in a reduced interface mass transfer rate and DNAPL oxidation rate. This is a complex process that is not fully understood. The use of recirculation, with injection and extraction wells, is intended to increase subsurface mixing. Many investigators have tried this approach with some apparent success. The costs are likely to be higher than even multiple injections without groundwater extraction and reinjection (with possible treatment and filtration of MnO2 required). However, the degree of mixing and, therefore, contact between contaminants and oxidant, will be greater, leading to more complete treatment, especially in heterogeneous subsurfaces. Also, utilization efficiency of the oxidant will be enhanced by recirculating the unused portion of the oxidant. In some cases, mixing has been encouraged by use of injection arrays with thin screen intervals at different depths to fully saturate the target zone and limit the need for vertical migration of the oxidant (Figure 4.42). High injection pressures have also been used to create fractures in tighter subsurface materials, again to encourage migration and mixing of the reactants. Mixing has also been encouraged through the use of air injection, to “push” peroxide solutions out into the aquifer. Finally, in some cases, vapor extraction has been used in conjunction with in situ oxidation in the vadose zone to relieve off-gas pressures, to encourage oxidant migration, and/or to capture any volatile emissions. Oxygen concentrations in the soil air can reach close to 100% and thus create explosive concentrations near the points of injection.76 The presence of colloidal materials, precipitation, and gas binding can cause reduced permeability of the aquifer near injection points. If the geologic materials have excessive amounts of CaCO3 in the formation gas, binding during the injection of Fenton’s reagent could be a significant problem. There do not seem to be well-developed guidelines for the design operation and cost estimation of ISCO systems particularly when DNAPL is present (Figure 4.44). The data needs for determining well spacing, screen intervals, or oxidant mass to be injected are not clear. There is a need for guidance to estimate the ROI under different conditions (soil texture, groundwater velocity, injection pressure, etc.). The efficiency of use of oxidants is not well established, and guidance for determining the mass needed at a specific site does not seem to be available. Recommendations
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Reduction of aqueous phase concentrations to below regulated clean-up levels 100%
Degree of Mass Removal
Complete DNAPL removal
Reduction of aqueous phase mass flux
Stabilization of pooled DNAPL
Partial DNAPL removal
0%
Conceptual Remediation End-point Figure 4.44
Conceptual description of increased remediation cost due to the presence of DNAPL.
regarding operations and monitoring to prevent undesirable reactions (explosions, volatile emissions, or foaming) are also not clear. 4.4.4.1 Oxidation of 1,4-Dioxane by Ozone Ozone is a powerful oxidant that can degrade contaminants via two mechanisms. The first, commonly referred to as the direct mechanism, involves the reaction of molecular ozone with the contaminant. Secondary oxidants, particularly the hydroxyl radical, can also oxidize the contaminants present. The oxidation of compounds by hydroxyl radical or other secondary oxidants is referred to as an indirect oxidation (since ozone is not directly involved in the oxidation). OH• is a nonselective oxidant, that is, it oxidizes many substances; consequently, in natural systems it may not be a very efficient oxidant, as it will react not only with contaminants of interest, but also with other substances present, e.g., the natural organic matter. In waters that contain carbonate, hydroxyl radical scavenging is greater at higher pH. Therefore the effectiveness of ozonation systems tends to decrease at elevated pH. For substances such as 1,4-dioxane that are not very reactive with molecular ozone, the optimal pH for removal is typically around 8.76 Ozone could be used to treat groundwater extracted from the aquifer. For ex situ treatment systems a treatment time of approximately 20 minutes would be needed for 99% removal of 1,4-dioxane.76 The use of ozone for in situ treatment of groundwaters may be particularly useful for the treatment of contaminants that are strongly sorbed to the aquifer materials (e.g., PAHs) or where the aquifer materials exert relatively little ozone demand. 1,4-dioxane is infinitely soluble in water and is not strongly sorbed to solids. In this case, in situ ozonation may be desirable as it avoids the cost of removal of the
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groundwater and of the expense and problems associated with construction of an ozone contractor to treat the extracted groundwater at the site. In situ ozonation can be implemented using techniques developed for in-well air sparging. The relative cost of in situ vs. ex situ treatment will depend very much upon how effectively these systems can be implemented. Ozone is sometimes used in combination with UV light or hydrogen peroxide to treat groundwaters. Both UV light and hydrogen peroxide catalyze the decomposition of ozone to produce hydroxyl radical. The rapid decomposition of ozone can enhance the rate of degradation of compounds like 1,4-dioxane, which are not very reactive with molecular ozone. 4.4.4.2 Biodegradation Enhanced by Chemical Oxidation Pretreatment Many experimental efforts have been carried out to evaluate the enhanced biodegradation of many recalcitrant compounds such as PCBs, polychlorinated phenols, and PAHs — with limited success. With increased attention to the cleanup of sites with known DNAPLs and manufactured gas plant (MGP) sites with coal tars, pretreatment with chemical oxidation for certain compounds may be a viable technology.
4.5
NANO-SCALE FE (0) COLLOID INJECTION WITHIN AN IRZ
Considerable research during the past several years has focused on the transformation of chlorinated solvents to harmless end products by exploiting the use of zero valent elemental metals for reductive dechlorination. In addition, elemental metals can be used to reduce soluble metals such as Cr (VI) to insoluble Cr (III) or metalloids such as As (V) and Se (VI) to As (III) and Se (IV), respectively. The most common metal utilized for this purpose is elemental iron, Fe (0). Although met with initial skepticism, the transformation process is surface-based and is now widely accepted as abiotic reductive dechlorination, involving corrosion of Fe (0) by chlorinated hydrocarbon. Other metals including tin, zinc, and palladium have also been shown to be effective.100 The process can be described best as anaerobic corrosion of the metal by the chlorinated hydrocarbon. During this process, the contaminant is adsorbed directly to the metal surface where the dechlorination reactions occur. In waters contaminated with chlorinated solvents, three oxidants are available to drive corrosion of metals: water, dissolved oxygen, and the chlorinated contaminant. The corrosion reaction involving water (Equation 4.40) is slow but presumably ubiquitous, whereas corrosion of Fe (0) by reaction with dissolved oxygen (Equation 4.41) is very rapid as long as O2 is available. The reaction rates with the chlorinated contaminant (Equation 4.42) are assumed to be between the two. Under aerobic conditions, dissolved oxygen is usually the preferred electron acceptor and will compete with the chlorinated contaminant as the favored oxidant (PCE and carbon tetrachloride may be comparable). Fe (0) + 2H2O → Fe2+ + H2 + 2OH–
(4.40)
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2Fe (0) + O2 + 2H2O → 2Fe2+ + 4OH–
(4.41)
Fe (0) + RX + H+ → Fe2+ + RH + Cl–
(4.42)
When sufficient oxygen is present, the Fe2+ generated in Equation 4.40 can precipitate as ferric hydroxide or (oxy) hydroxides at an elevated pH typical of corroding Fe systems. In carbonate rich waters FeCO3 precipitation will also occur. These precipitates can exert significant additional chemical and physical effects within the surface-based Fe (0) reactive system by coating the reactive iron metal. Recent research on Fe (0) systems indicates that other mechanisms also may be involved in the reductive process. The reductive processes can be summarized as below:100 • Fe (0) can act as a reductant by supplying electrons directly from the metal surface to the adsorbed chlorinated contaminant (Figure 4.45). • Metallic Fe (0) may act as a catalyst for the reaction of H2 with the chlorinated contaminant. The hydrogen is produced on the surface of the iron metal as the result of corrosion with water (Figure 4.45).
A H
Cl
Cl
Cl
Cl -
H
H
Cl
Cl
cis-1,2-DCE
TCE
B
2e - + H +
2e - + H +
2e -
Cl
-
2Cl Chloroacetylene
H
H
Cl
Vinyl chloride
2e - + H +
2e - + 2H + Cl
Cl -
-
2e - + 2H + Acetylene
H H C C Cl
H
H
H
H
H C C H
Ethene
2e - + 2H +
H
H H Ethane
H H
Figure 4.45
H
Abiotic reductive dechlorination mechanisms by Fe (0).
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The rate of reaction of Fe (0) with chlorinated contaminants is dependent upon the reactivity of individual chemical compounds and the amount of reactive surface area on the Fe (0) particles. For degradation of a contaminant by zero-valent iron metal, the reaction model can be represented as:102 -
dC = K SA ra [C] dt
(4.43)
where C = reacting contaminant concentration KSA = specific reaction constant ra = the amount of iron surface area Based on the reported values of KSA in the literature, there seems to be an order of magnitude variability for an individual chlorinated hydrocarbon.102 It is important to note, however, that the variability in KSA for individual compounds is modest relative to the five orders of magnitude variability among the various chlorinated hydrocarbons. In addition to the primary effects of contaminant reactivity and metal surface area, several other factors influence the kinetics (KSA) of chlorinated contaminant degradation. One factor is the saturation of reactive surface area with increasing contaminant concentration. Another factor is metal “type,” which is the variable most commonly invoked to rationalize otherwise unexplained variability in degradation rates by iron. The ra term in Equation 4.43 characterizes quantity of iron surface area, but does not address differences in the reactivity of the surface. It is important to note that as the size of the metallic iron is reduced, surface area goes up as well as chemical reactivity. High surface areas can be attained either by fabricating smaller particles or clusters where the surface to volume ratio of each particle is high, or by creating materials where the void surface area (pores) is high compared to the amount of bulk material. If a metal is continually reduced in size it will eventually reach what is known as superfine particle or nano-scale particle. Such particles can be distinguished from their corresponding bulk solid form by the size of their surface areas in relation to their weight. Initial applications of this technology in the mid 1990s used iron filings. Due to size limitations (not small enough to be injected directly) of the commercially available iron filings (Table 4.13), the process had to be implemented in the subsurface as a permeable reactive barrier (PRB). In a PRB, reactive material is placed in the subsurface where a plume of contaminated groundwater must move through it as it flows, typically under its natural gradient, and treated water comes out the other side. The placement of the iron filings into the PRB was usually achieved by hydraulic fracturing (Figure 4.46), or via a funnel and gate system where the gate was filled with iron filings, or by mixing the iron filings with sand in a permeable interception trench (Figure 4.47). It is obvious that in all these methods the “peripheral” geotechnical cost for the “placement” of iron filings in the subsurface can be
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Table 4.13 Examples of the Surface Area of Different Metallic Iron [Fe (0)] Products103-105 Iron Type Iron turnings Electrolytic iron Iron granules Commercial iron filings Nano-scale iron particles
Surface Area in M2/g 0.019 0.057 0.287 0.900 33.50
M2/g M2/g M2/g M2/g M2/g
up to two orders of magnitude higher than the actual cost of the iron filings. As a result, more recent applications have used iron colloids in the micron size range to cut down on the peripheral geotechnical cost and directly inject the iron colloids into the contaminated zone.
Figure 4.46
Placement of Fe (0) filings as a reactive barrier via hydraulic fracturing.
The author and a few others have advanced metallic Fe (0) reduction technology by incorporating nano-scale particles ranging in size from 1 to 999 nanometers (.001 to .999 µm). A particle of this size has several advantages in application for in situ groundwater remediation. These advantages include: • High surface area result in greater reaction kinetics. • The increase in kinetics allows for a lower mass loading of iron in the treatment zone or reactor because the residence time required for complete dechlorination is decreased. • The small size and greater reactivity of the superfine particle allows for the application of the technology through direct in situ injection into the subsurface (Figures 4.48 and 4.49).
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Treated Groundwater
Groundwater flow Contaminated plume
Sand/ Fe o mixture
Interceptor Trench
Figure 4.47
Placement of Fe (0) filing as a permeable reactive barrier in an interceptor trench.
o
Fe / Molasses slurry
Injection well
Contaminated zone
In Situ reactive zone filled with nano scale iron Figure 4.48
Direct injection of nano-scale Fe (0) particles into the contaminated zone.
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In Situ mixing o
Mixed zone with Fe
Figure 4.49
Injection of nano-scale Fe (0) particles into the contaminated zones in the unsaturated and shallow saturated zones.
• The smaller size allows for advective particle transport. • The greater reactivity due to the small size allows for much lower overall iron mass requirements.
Conceptually, destruction of the contaminant is an interfacially controlled process, and thus the efficacy of destruction is dominated by the exposed surface area of the superfine particle. The exposed surface area is easily determined by BET nitrogen adsorption, for which the surface area can be related to an equivalent spherical diameter (desd): SSA = 6/ρ.desd
(4.44)
where SSA = specific surface area determined by BET ρ = material density In addition to the beneficial effects of increased surface area of the superfine particles, coupling of a catalyst such as palladium or platinum will lead to increased reaction rates which are multiplicative (Figure 4.50). 4.5.1
Production of Nano-Scale Iron Particles
Over the last decade, research, driven primarily by needs in the field of materials science (hi-tech electronic chips or component industry products), has contributed
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Figure 4.50
229
Nano-scale bimetallic clusters.
to general technologies designed to produce nano-scale particles. Generally, the research has been in the area of colloids composed of ceramic or other nonmetallic inorganic materials and not metal colloids. A significant part of the development effort for the technology is the adaptation of nonmetallic nano-scale production methods to the production of metallic nano-scale particles. The method for production of metal particles in the nano-scale range may be divided into two primary approaches: 1) “bottom up,” in which colloids of the appropriate size are produced by being assembled from individual atoms; and 2) “top down,” in which colloids of the appropriate size are produced by attrition of larger existing particles of the metal. The bottom up approach has a greater number of potentially applicable methods including100: • Chemical reduction using sodium borohydride; various soluble metal salts (such as ferrous or ferric chloride for iron) in suspensions of water, or various organic hydrocarbon solvents; this process may or may not be enhanced with sonification during reaction processes • Other chemical precipitation reactions in aqueous or hydrocarbon solutions capable of producing metals from soluble salts that may or may not include sonification during reaction processes • Various methods of metal volatilization and subsequent deposition, typically under vacuum
The top down approach uses two primary variations of milling or mechanical comminuation that include:
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• Using mechanical agitation of a mixture of the desired colloidal metal, a grinding media, and an organic or aqueous suspension fluid; examples include ball mills and rod mills • Systems similar to the above where the mechanical agitation is provided by high speed gas jets
There are methods available to produce superfine particles that have distinct morphology and internal crystal structure to further enhance the surface reactivity.106 In addition, it is important to recognize that, in the nano-scale range, quantum size effects begin to become apparent. For example a colloid of 10 nm diameter has about 30% of its atoms in grain boundaries (which are highly reactive and subject to quantum effects). These features have an effect on the physical/chemical behavior of the particle in use which falls into one of two broad categories reflecting on production by bottom up or top down methods.100 • A colloid produced by chemical precipitation or reduction, or through the various vapor deposition methods, will be nano-structured. This means that the colloid will have nano-scale crystal domains with sharp boundaries between crystals. The grain boundaries are typically only 1 atom thick and there is low dislocation density in the crystal structures.100 • The reactivity of a colloid of this type can be controlled primarily through the selection of an appropriate overall colloid size and resulting surface area. Smaller size means greater surface area and reactivity; larger size means lower surface area and reactivity.100 • A colloid produced by mechanical attrition will be nano-crystalline. The crystal domains in the colloid are small, relative to the overall colloid size. The individual crystal domains are separated by wide amorphous transition regions that exhibit a very high dislocation density. These transition regions may be as large as the crystal domains, but are still termed grain boundaries.100 • The amorphous transition regions will be highly reactive. The reactivity of the colloid will be dominated by the size and intensity of dislocation density of the amorphous boundary regions rather than the absolute size of the colloid. A relatively large colloid produced by this method could have reactivity the same as or greater than a much smaller colloid produced by bottom up methods.100
Control of the reactivity of the colloid is a critical feature. The iron undergoes anaerobic corrosion, reacting directly with halogenated solvents as well as with water to produce hydrogen. As the reactivity of the colloid increases, the hydrogen production rate increases as well. By controlling the rate of hydrogen production using the methods described above, it will be possible to design reactive metal colloids with reactivity that will generate hydrogen at the rate required for the desired dehalogenation processes — rather than being consumed at excessively higher rates (with just water) at which the iron colloid would be consumed (by the water) without reacting with the halogenated solvents undergoing treatment. Controlling the type of nano-scale particle produced is particularly important for in situ applications in order to maximize the rate of hydrogen production needed to achieve the remediation objectives.
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231
Injection of Nano-Scale Particles in Permeable Sediments
Injection of nano-scale particles into intragranular pore space of the geologic matric is the preferred mode of application for directly addressing microemulsions of contaminants in source areas or for treating dissolved phase contaminants. Diffusion of contaminants and hydrogen generated by Fe (0) and advection of Fe (0) particles should provide the intimate contact between contaminants and Fe (0). The mobility of colloids is governed by mechanical filtration and adsorptive processes within the porous media; it is always preferable to achieve the largest reactive zone from each injection point for economic reasons. Colloids and nano-scale particles can be mechanically removed by the soil matrix. The key parameter to this process is the pore entrance size, which is a function of grain size. In fine- to coarse-grained silts, pore entrance (throat) sizes range from 0.7 to 7 µm, in fine- to coarse-grained sands from 24 to 240 µm, and in fine- to coarse-grained gravels from 720 to 7,200 µm. It is obvious from this information that nano-scale particles will travel further from the point of injection than typical colloids, particularly within more permeable formations. Injection of nano-scale particles with shear thinning fluids also will enhance the injectability of Fe (0) particles into the porous media. In contrast to Newtonian fluids, whose viscosities are constant with shear rate, certain non-Newtonian fluids are shear thinning, that is, the viscosity of these fluids decreases with increasing shear rate.108 The primary benefit of using these fluids for this application is that they increase the viscosity of the aqueous phase without adversely decreasing the hydraulic conductivity. A suspension formulated with a shear thinning fluid will maintain a relatively high viscosity in solution near Fe (0) particles (where the shear stress is low) relative to locations near the surfaces of the porous media, where the shear stress is high. The increased viscosity decreases the rate of gravitational settling of the Fe (0) particles while maintaining a relatively high hydraulic conductivity that permits injecting the Fe (0) suspension into the porous media at greater flow rates and distances. If an easily biodegradable shear thinning fluid is selected, it will also provide an additional benefit in the form of scavenging the dissolved oxygen present within the reactive zone and ensuring that the reactive iron is consumed primarily by the degradation of the contaminant mass. Above ground engineering controls to prevent agglomeration of the Fe (0) particles in injection solution will also enable the injected particles to travel farther in the porous media. This will entail control of the ionic state of the suspension fluid to prevent agglomeration, use of surfactants, and determination of the optimum colloidal concentration for the suspension. 4.5.3
Organic Contaminants Treatable by Fe (0)
Tables 4.14 and 4.15 present a list of organic contaminants that are treatable and not treatable by Fe (0) based on the current state of science. Table 4.16 presents a list of compounds with unknown reactivity with Fe (0).
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Table 4.14 Contaminants Treatable by Zero Valent Iron, Fe (0)101 Organic Compounds
Inorganic Compounds
Methanes
tetrachloromethane trichloromethane dichloromethane
Dissolved Metals
Ethanes
hexachloroethane 1,1,1-trichloroethane 1,1,2-trichloroethane 1,1-dichloroethane tetrachloroethene trichloroethene cis-1,2-dichloroethene trans-1,1-dichloroethene 1,1-dichloroethene vinyl chloride 1,2,3-trichloropropane 1,2-dichloropropane benzene toluene ethylbenzene hexachlorobutadiene 1,2-dibromoethane freon 113 N-nitrosodimethylamine
Anion Contaminants
Ethenes
Propanes Aromatics
Other
Table 4.15 Contaminants Presently Not Treatable by Fe (0)101 Organic Compounds
Inorganic Compounds
dichloromethane 1,2-dichloroethane chloroethane chloromethane heavier PAHs
chloride perchlorate
Table 4.16 Contaminants with Unknown Treatability Organic Compounds
Inorganic Compounds
chlorobenzenes chlorophenols certain pesticides PCBs
mercury
Chromium Nickel Lead Uranium Technetium Iron Manganese Selenium Copper Cobalt Cadmium Zinc Sulphate Nitrate Phosphate Arsenic
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REFERENCES 1. Suthersan, S. S., Engineered In Situ Anaerobic Reactive Zones, U.S. Patent 6,143,177, 1998. 2. Bouwer, E. J., Rittmann, B. E., and McCarty, P. L., Anaerobic degradation of halogenated 1-, and 2- carbon organic compounds, Environ. Sci. Technol., 15, 596–599, 1987. 3. Parsons, F., Wood, P. R., and DeMarco, J., Transformations of tetrachloroethene and trichloroethene in microcosms and groundwater, J. Am. Water Works Assoc., 76, 56–59, 1984. 4. Vogel, T. M. and McCarty, P. L., Biotransformation of tetrachloroethylene to trichloroethylene, dichloroethylene, vinyl chloride, and carbon dioxide under methanogenic conditions, J. Appl. Environ. Microbiol., 49, 1080–1083, 1985. 5. Kleopfer, R. D. et al., Anaerobic degradation of trichloroethylene in soil, Environ. Sci. Technol., 19, 277–280, 1985. 6. Freedman, D. L. and Gossett, J. M., Biological reductive dehalogenation of tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions, J. Appl. Environ. Microbiol., 55, 1009–1014, 1989. 7. DeBruin, W. P. et al., Complete biological reductive transformation of tetrachloroethene to ethane, J. Appl. Environ. Microbiol., 58, 1996–2000, 1992. 8. Christiansen, N. et al., Transformation of tetrachloroethene in an upflow anaerobic sludge blanket reactor, Appl. Microbiol. Biotechnol., 47, 91–94, 1997. 9. Kastner, M., Reductive dechlorination of tichloroethylenes and tetrachloroethylenes depends on transition from aerobic to anaerobic conditions, J. Appl. Environ. Microbiol., 57, 2039–2046, 1991. 10. DiStefano, T. D. et al., Reductive dechlorination of high concentrations of tetrachloroethene to ethane by an anaerobic enrichment culture in the absence of methanogenesis, J. Appl. Environ. Microbiol., 57, 2287–2292, 1991. 11. DiStefano, T. D., Gossett, J. M., and Zinder, S. H., Hydrogen as an electron donor for dechlorination of tetrachloroethene by an anaerobic mixed culture, J. Appl. Environ. Microbiol., 58, 3622–3629, 1992. 12. Alexander, M., Biodegradation and Bioremediation, Academic Press, New York, 1999. 13. Schwarzenbach, R. P., Gschwenda, P. M., and Imboden, D. M., Environmental Organic Chemistry, John Wiley & Sons, New York, 1993. 14. Wiedemeier, T. H. et al., Natural Attenuation of Fuels and Chlorinated Solvents in the Subsurface, John Wiley & Sons, New York, 1999. 15. McCarty, P. L. and Semprini, L., Groundwater treatment for chlorinated solvents, in Handbook of Bioremediation, Norris, R. D. et al., Eds., Lewis Publishers, Boca Raton, FL, 1994. 16. Hollinger, C. and Schumacher, W., Reductive dehalogenation as a respiratory process, Antoine Leeuwenhoek, 66, 239–246, 1994. 17. Gossett, J. M. and Zinder, S. H., Microbiological aspects relevant to natural attenuation of chlorinated ethenes, in Proc. Symp. Nat. Attenuat. Chlorinat. Org. Groundwater, Dallas, TX, September 11-13, 1996, EPA/540/R-96/509. 18. Vogel, T. M., Criddle, T., and McCarty, P. L., Transformations of halogenated aliphatic compounds, Environ. Sci. Technol., 21, 722–736, 1987. 19. Neumann, A. et al., Purification and characterization of tetrachloroethene reductive dehalogenase from Dehalospirillum mutivorans, J. Biochem. Mol. Biol., 271, 16515–16519, 1996.
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20. USEPA, Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action and Underground Storage Tank Sites, Directive No. 9200.4-17P, Office of Solid Waste and Emerging Response, Washington, D.C., 1999. 21. Bradley, P. M. and Chappelle, F. H., Anaerobic mineralization of vinyl chloride in Fe (III) reducing aquifer sediments, Environ. Sci. Technol., 30, 2084–2086, 1996. 22. Luijten, M. L. et al., Effect of alternative electron acceptors on the reduction of tetrachloroethene, in Proc. 2nd Int. Conf. Remed. Chlorinat. Recalcitrant Comp., Monterey, CA, May 2000. 23. Bouwer, E. J., Bioremediation of chlorinated solvents using alternate electron acceptors, in Handbook of Bioremediation, Norris, R. D. et al., Eds., Lewis Publishers, Boca Raton, FL, 1994. 24. Gantzer, C. J. and Wackett, L. P., Reductive dechlorination catalyzed by bacterial transition metal coenzymes, Environ. Sci.Technol., 25, 715–722, 1991. 25. Middeldorp, P. J. M. et al., Anaerobic microbial reductive dehalogenation of chlorinated ethenes, Bioremed. J., 3, 151–169, 1999. 26. Belay, N. and Daniels, L., Production of ethane, ethylene and acetylene from halogenated hydrocarbons by methanogenic bacteria, Appl. Environ. Microbiol., 53, 1604–1610, 1987. 27. Ballapragada, B. S. et al., Effect of hydrogen on reductive dechlorination of chlorinated ethenes, Environ. Sci. Technol., 31, 1728–1734, 1997. 28. Carter, S. R. and Jewell, W. J., Biotransformation of tetrachloroethylene by anaerobic attached-films at low temperatures, Wat. Res., 27, 607–615, 1993. 29. Berededsamuel, Y. et al., Effect of Perchloroethylene (PCE) on methane and acetate production by a methanogenic consortium, Appl. Biochem. Biotechnol., 57-8, 915–922, 1996. 30. Fathepure, B. Z. and Tiedje, J. M., Reductive dechlorination of tetrachloroethylene by a chlorobenzoate-enriched biofilm reactor, Environ. Sci. Technol., 28, 746–752, 1994. 31. Gibson, S. A. and Sewell, G. W., Stimulation of reductive dechlorination of tetrachloroethene in anaerobic acquifer microcosms by addition of short-chain acids or alcohols, Appl. Environ. Microbiol., 58, 1392–1393, 1992. 32. Guiot, S. R. et al., Anaerobic and aerobic/anaerobic treatment for tetrachloroethylene (PCE), in 3rd Int. In Situ On-Site Bioreclam. Symp., Hinchee, R.E., Leeson, A., and Semprini, L., Eds., Batelle Press, San Diego, 1995, 297–305. 33. DiStefano, T. D. et al., Hydrogen as an electron donor for dechlorination of tetrachloroethene by an anaerobic mixed culture, Appl. Environ. Microbiol., 58, 3622–3629, 1992. 34. Wiedermeier, T. H. et al., Protocol for supporting natural attenuation of chlorinated solvents with examples, in In Situ and On-Site Bioremediation 3, Batelle Press, Columbus, OH, 1997, 147. 35. Fennell, D. E. et al., Comparison of butyric acid, ethanol, lactic acid and propionic acid as hydrogen donors for the reductive dechlorination of tetrachloroethene, Environ. Sci. Techol., 31, 918–926, 1997. 36. Maymo-Gatell, X. et al., Isolation of a bacterium that reductively dechlorinates tetrachloroethene to ethane, Science, 276, 1568–1571, 1997. 37. Maymo-Gatell, X. et al., Characterization of an H2-utilizing enrichment culture that reductively dechlorinates tetrachloroethene to vinyl chloride and ethane in the absence of methanogenesis and acetogenisis, Appl. Environ. Microbiol., 61, No. 11, 3928–3933, 1995.
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38. Hollinger, C. et al., A highly purified enrichment culture couples the reductive dechlorination of tetrachloroethene to growth, Appl. Environ. Microbiol., 59, 2991–2997, 1993. 39. Wild, A. et al., Isolation of an anaerobic bacterium which reductively dechlorinates tetrachloroethene and trichloroethene, Biodegradation, 7, 507–511, 1996. 40. Scholz-Muramatsu, H. et al., Isolation and characterization of Dehalospirillum multivorans general nov., sp. nov., a tetrachloroethene-utilizing, strictly anaerobic bacterium, Arch. Microbiol, 163, 48–56, 1995. 41. Utkin, I., Woese, C. and Wiegel, J., Isolation and characterization of Desulfitobacterium Dehalogenans general nov, sp. nov., an anaerobic bacterium which reductively dechlorinates chlorophenolic compounds, Int. J. Syst. Bacteriol., 44, 612–619, 1994. 42. Gerritse, J. et al., Desulfitobacterium sp. strain PCEI, an anaerobic bacterium that can grow by reductive dechlorination of tetrachloroethene or ortho chlorinated phenols, Arch. Microbiol., 165, 132–140, 1996. 43. Gerritse, J. et al., Complete degradation of tetrachloroethene in coupled anoxic and oxic chemostats, Appl. Microbiol. Biotechnol., 48, 553–562, 1997. 44. Miller, E. et al., Comparative studies on tetrachloroethene reductive dechlorination mediated by Desulfitobacterium sp. strain PCE-S, Arch. Microbiol., 168, 613–519, 1998. 45. Sharma, P. K. and McCarty, P. L., Isolation and characterization of a facultatively aerobic bacterium that reductively dehalogenates tetrachloroethene to cis-1, 2-dichloroethene, Appl. Environ. Microbio. 62, 761–765, 1996. 46. Krumholz, L. R., Desulfuromonas chloroethenica sp. nov. uses tetrachloroethylene and trichloroethylene as electron acceptors, Int. J. Syst. Bacteriol., 47, 1262–1263, 1997. 47. Maymo-Gatell, X. et al., Dehalococcus ethenogenes strain 195: ethane production from halogenated aliphatics, in In Situ and On-Site Bioremediation: 3, Batelle Press, Columbus, OH, 1997, 23. 48. Schumacher, W. and Holliger, C. L., The proton electron ratio of the menaquinone dependent electron transport from dihydrogen to tetrachloroethene in Dehalobacter restrictus, J. Bacteriol., 178, 2328–2333, 1996. 49. Miller, E. et al., Studies on tetrachloroethene respiration in Dehalospirillum multivorans, Arch. Microbiol., 166, 379–387, 1996. 50. Rosner, B. M. et al., In vitro studies on reductive vinyl chloride dehalogenation by an anaerobic mixed culture, Appl. Environ. Microbiol., 63, 4139–4144, 1997. 51. Harkness, M. R., Economic considerations in enhanced anaerobic biodegradation, Proc. 2nd Int. Conf. Remed. Chlorinat. Recalcitrant Compds., Wickramanayake, G. B. et al., Eds., Battelle Press, Columbus, OH, May 22–25, 2000. 52. DiStefano, T. D., PCE dechlorination with complex electron donors, Proc. 2nd Int. Conf. Remed. Chlorinat. Recalcitrant Compds., Wickramanayake, G. B. et al., Eds., Battelle Press, Columbus, OH, May 22–25, 2000. 53. Suthersan, S. S., ARCADIS G & M, Inc., personal communication, 2000. 54. Komatsu, T. et al., Biotransformation of cis-1,2-dichloroethylene to ethylene and ethane under anaerobic conditions, Water Sci. Technol., 30, 75–84, 1994. 55. Smatlak, C. R., Grossett, J. M., and Zinder, S. H., Comparative kinetics of hydrogen untilization for reductive dechlorination of tetrachloroethene and methanogenesis in an anaerobic enrichment culture, Environ. Sci. Technol., 30, 2850–2858, 1996. 56. Bouwer, E. J. and McCarty, P. L. M., Transformations of 1 and 2-carbon halogenated aliphatic organic compounds under methanogenic conditions, Appl. Environ.Microbiol., 45, 1286–1294, 1983.
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57. Fathepure, B. Z. and Vogel, T. M., Complete degradation of polychlorinated hydrocarbons by a two stage biofilm reactor, Appl. Environ. Microbiol. 57, 3418–3422, 1991. 58. Brock, T. D. and Madigan, M. T., Biology of Microorganisms, 5th ed., Prentice Hall, Englewood Cliffs, NJ. 59. Nandi, R. and Sengupta, S., Microbial production of hydrogen: an overview, CRC Rev. Microbiol., 24, 61–84, 1998. 60. Tandoi, V. et al., Environ. Sci. Technol., 28, 973–979, 1994. 61. Gossett, J. M., Environ. Sci. Technol., 21, 202–208, 1987. 62. Carr, C. S. and Hughes, J. B., Enrichment of high-rate PCE dechlorination and comparative study of lactate, methanol, and hydrogen as electron donors to sustain activity, Environ. Sci. Technol., 32, 1817–1824, 1998. 63. McCarty, P. L., An Overview of anaerobic transformation of chlorinated solvents, in Symp. Intrinsic Bioremed. Groundwater, Denver, CO, August 30 – September, 1994, 135–142. 64. Vogel, T. M. et al., Abiotic and biotic transformation of 1,1,1-trichloroethane under methanogenic conditions, Environ. Sci. Technol., 21, No. 12, 1208–1213, 1987. 65. Criddle, C. S. and McCarty, P., Electrolytic model system for reductive dehalogenation in agueous environments, Environ. Sci. Technol., 25, 973–978, 1991. 66. Hughes, J. B. and Parken, G. F., Individual biotransformation rates in chlorinated aliphatic mixtures, J. Environ. Eng., ASCE, 99–106, 1996. 67. Alvarez-Cohen, L. and McCarty, P. L., Product toxicity and cometabolic competitive inhibition of chloroform and trichloroethylene transformation by methanogenic resting cells, Appl. Environ. Microbiol., 57, 1031–1037, 1991. 68. Alvarez-Cohen, L. and McCarty, P. L., Effects of toxicity, aeration, and reductant supply on trichloroethylene transformation by a mixed methanogenic culture, Appl. Environ. Mecorbiol., 57, 228–235, 1991. 69. Rasche, E. M., Hyman, M. H., and Arp, D. J., Factors limiting aliphatic chlorocarbon degradation by Nitrosomonas europaea: cometabolic inactivation of ammonia monooxygenase and substrate specificity, Appl. Environ. Microbiol., 57, 2986–2999, 1991. 70. Folsom, B. R. and Chapman, P. J., Performance characterization of a model bioreactor for the biodegradation of trichloroethylene by Pseudomonas cepacia G4, Appl. Environ. Microbiol., 57, 1602–1608, 1991. 71. Bradley, P. M. and Chapelle, F. H., Anaerobic minerilization of vinyl chloride in Fe (III) reducing acquifer sediments, Environ. Sci. Technol., 30, 2084-2086, 1996. 72. Bradley, P. M. and Chappelle, F. H., Kinetics of DCE and VC mineralization under methanogenic and Fe (III) reducing conditions, Environ. Sci. Technol., 31, 2692–2696, 1992. 73. Bradley, P. M. and Chapelli, F. H., Effect of contaminant concentration on aerobic microbial mineralization of DCE and VC in stream-bed sediments, Environ. Sci. Technol., 30, No. 5, 553–557, 1978. 74. Bushman, A. et al., Environ. Sci. Technol., 33, 1015–1020, 1999. 75. Adrianes, P. et al., Biogeochemistry and dechlorination potential at the St. Joseph acquifer Lake Michigan interface, in Ed., In Situ and On-Site Bioremediation 3, Batelle Press, Columbus, OH, 1997, 173–178. 76. Payne, F., ARCADIS G & M, Inc., Personal Communication, 2000. 77. Pyrih, R. Z., Recognizing the natural attenuation of metals, in Proc. IBS’s 4th Ann. Conf. Nat. Attenuation, Pasadena, CA, December, 1998. 78. Potter, S., ARCADIS G & M, Inc., personal communication, 2000.
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79. Sparks, D. L., Ed., Soil Physical Chemistry, 2nd Ed., CRC/Lewis Publishers, Boca Raton, FL, 1998. 80. Hong, J., Forstner, U., and Calmano, W., Effects of REDOX processes on acid producing potential and metal mobility in sediments, in Bioavailability: Physical, Chemical and Biological Interactions, Hamlink, J. L. et al., Eds., Lewis Publishers, Boca Raton, FL, 1994. 81. Ehrlich, H. L., Geomicrobiology, Marcel Dekker, New York, 1981. 82. Melhorn, R. J., Buchanan, B. B., and Leighton, T., Bacterial chromate reduction and product characterization, in Emerging Technology for Bioremediation of Metals, Means, J. L. and Hinchee, R. E., Eds., Lewis Publishers, Boca Raton, FL, 1994. 83. Turick, C. E., Graves, C., and Apel, W. A., Bioremediation potential of Cr (VI)contaminated soil using indigenous microorganisms, Bioremed. J., 2, 1–6, 1998. 84. Bader, J. L. et al., Aerobic reduction of hexavalent chromium in soil by indigenous microorganisms, Bioremed. J., 3, 201–212, 1999. 85. Palmer, C. and Puls, R., Natural Attenuation of Hexavalent Chromium in Groundwater and Soils, USEPA, Office of Research and Development, OSWER, EPA/540/S94/505, 1994. 86. Gary, L. and Rai, D., Kinetics of chromium (III) oxidation to cromium (VI) by reaction with manganese dioxides, Environ. Sci. Technol., 21, 1187–1193, 1987. 87. Urbansky, E. T., Perchlorate chemistry: implications for analysis and remediation, Bioremed. J., 2, 89–97, 1998. 88. Logan, B. E., A review of chlorate- and perchlorate-respiring microorganisms, Bioremed. J., 2, 69–80, 1998. 89. Giblin, T., et al., Removal of perchlorate in groundwater with a flow-through bioreactor, J. Environ. Qual., 29, 578–583, 2000. 90. Hartsman, S., and de Bont, J. A. M., Aerobic vinyl chloride metabolism in Mycobacterium aunum li, Appl. Environ. Microbiol., 58, 1220–1226, 1985. 91. McCarty, P. L. and Semprini, L., Groundwater treatment for chlorinated solvents, in Handbook of Bioremediation, Norris, R. D., et al., Eds., Lewis Publishers, Boca Raton, FL, 1994. 92. Stefan, R., Environ. Corporation., personal communication, 2000. 93. Yin, Y. and Allen, H. E., In Situ Chemical Treatment, Groundwater Remediation Technologies Analysis Center, Pittsburgh, PA 1999. 94. Yan, Y. E. and Schwartz, F. W., Oxidative degradation and kinetics of chlorinated ethylenes by potassium permanganate, J. Contaminant Hydrol., 37, 343–365, 1999. 95. Haag, W. R. and Yao, C. C. D., Rate constants for reaction of hydroxyl radicals with several drinking water contaminants, Environ. Sci. Technol., 26, 1005–1013, 1992. 96. Walling, J., Fenton’s reagent revisited, Acc. Chem. Res., 8, 125–131, 1975. 97. Siegrist, R. L., In Situ chemical oxidation: technology features and applications, in Proc. Conf. Adv. Innovat. Groundwater Remed. Technol., Atlanta, December, 1998. 98. Hoigne, J. and Bader, H., Rate constants of reactions of ozone with organic and inorganic compounds in water − Nondissociating organic compounds, Water Res., 17, 173–183, 1983. 99. Basel, M.D. and Nelson, C.H., Overview of in situ chemical oxidation: status and lessons learned, paper presented at the 2nd Int. Conf. Remed. Chlorinat. Recalcitrant Compds, Monterey, CA, May, 2000. 100. Vance, D., ARCADIS G & M, Inc., Personal Communication, 2000. 101. USEPA, Permeable Reactive Barrier Technologies for Contaminant Remediation, EPA/600/R-98/125, Washington, DC, September, 1998.
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102. Tratnyek, P.G. et al., Remediating groundwater with zero valent metals: chemical considerations in barrier design, Groundwater Monitor. Rev., Fall, 1997. 103. Johnson, T.L. et al., Kinetics of halogenated organic compound degradation by iron metal, Environ. Sci. Technol., 30, 2634–2640, 1996. 104. Gillham, R.W. and O’Hannesin, S.F., Enhanced degradation halogenated aliphatics by zero valent iron, Groundwater, 32, 958–967, 1994. 105. Wang, C.B. and Zhang, W.X., Synthesizing nano-scale iron particles for rapid and complete dechlorination of TCE and PCBs, Environ. Sci. Technol., 31, 2154–2156, 1997. 106. Vance, D., Suthersan, S., and Palmer, P., Method of Making and Using Nano-Scale Metal, U.S. Patent (pending). 107. Ichinose, N., Ozaki, Y., and Kashu, S., Superfine Particle Technology, Springer-Verlag, Tokyo, 1992. 108. Cantrell, K.J., Caplan, D.I., and Gilmore, T.J., Injection of colloidal size particles of Fe (0) in porous media with shear thinning fluids as a method to emplace a permeable reactive zone, Int. Contain. Technol. Conf., St. Petersburg, FL, February, 1997.
CREDIT Figure 4.9 is adapted from Van Briesen, J.M. and Rittmann, B.E.,Natural Attentuation Consideration and Case Studies: Remediation of Chlorinated and Recalcitrant Compounds, Wickramanayake, G.B., Gavaskar, A.R., and Kelley, M.E. (Eds), Battelle Press, Columbus, OH. With permission.
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CHAPTER
5
Phytoremediation CONTENTS 5.1 5.2
Introduction ..................................................................................................240 Chemicals in the Soil–Plant System............................................................241 5.2.1 Metals..................................................................................................241 5.2.2 Organics ..............................................................................................242 5.3 Types of Phytoremediation ..........................................................................244 5.3.1 Phytoaccumulation ...........................................................................245 5.3.2 Phytodegradation..............................................................................248 5.3.3 Phytostabilization .............................................................................250 5.3.4 Phytovolatilization............................................................................251 5.3.5 Rhizodegradation..............................................................................252 5.3.6 Rhizofiltration...................................................................................256 5.3.7 Phytoremediation for Groundwater Containment ...........................259 5.3.8 Phytoremediation of Dredged Sediments ........................................260 5.4 Phytoremediation Design .............................................................................261 5.4.1 Contaminant Levels .........................................................................265 5.4.2 Plant Selection..................................................................................265 5.4.3 Treatability .......................................................................................266 5.4.4 Irrigation, Agronomic Inputs, and Maintenance .............................266 5.4.5 Groundwater Capture Zone and Transpiration Rate .......................267 References..............................................................................................................267
… many accepted agricultural techniques for cultivating, harvesting, and processing plants have now been adapted for phytoremediation. Overall, the application of phytoremediation is being driven by its technical and economic advantages over conventional approaches … .phytoremediation’s future is not a scientific issue, but rather a “scientific sociology” issue….
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5.1 INTRODUCTION Phytoremediation is defined as “the engineered use of plants in situ and ex situ for environmental remediation.” The technology involves removing or degrading organic and inorganic contaminants and metals from soil and water. The processes include all plant-influenced biological, chemical, and physical processes that aid in the uptake, sequestration, degradation, and metabolism of contaminants, either by plants or by the free living organisms that constitute a plant’s rhizosphere. Phytoremediation takes advantage of the unique and selective uptake capabilities of plant root systems, together with the translocation, bioaccumulation, and contaminant storage and degradation capabilities of the entire plant body. The concept of using plants to alter the environment has been around since plants were first used to drain swamps. What is new within the context of this new technology called phytoremediation is the systematic, scientific investigation of how plants can be used to decontaminate soil and water.1 Interest in phytoremediation has been growing in the U.S. during the past few years with potential applicaton of this technology at a wide range of sites contaminated with heavy metals, pesticides, explosives, and solvents. The potential benefits of phytoremediation seem to be as numerous as the problems it might address. One reason this technology is gaining attention is because it is potentially cheaper than conventional treatment approaches for contaminated soils and traditional pump and treat systems for contaminated groundwater, such as incineration or soil washing. Another attraction of this technology is that it may leave topsoil in usable condition, keeping soil fertility and structure intact while reducing contamination levels at the same time. Phytoremediation is well suited for applications in low permeability soils, where most currently used technologies have a low degree of feasibility or success, as well as in combination with more conventional remediation technologies. The main advantages of phytoremediation are the low capital costs, aesthetically pleasing technique, minimization of leaching of contaminants, and soil stabilization. The operational cost of phytoremediation is also substantially less than that of conventional treatments and involves mainly fertilization and watering for maintenance of plant growth. In the case of heavy metals remediation, additional operational costs include harvesting, disposal of contaminated plant mass, and repeating the plant growth cycle. It should be emphasized that there is more to phytoremediation than merely putting plants in the ground and letting them do the work. Phytoremediation also has its drawbacks, which even its ardent champions are quick to acknowledge. First of all, it is a time-consuming process that can take several growing seasons to clean a site. Vegetation that absorbs toxic heavy metals will have to be harvested and managed as a waste. This vegetation containing high concentrations of toxic metals and organics may also pose a risk to wildlife. The shutdown of plant activity during winter months and the seasonal variation of plant metabolic activity is a drawback for application of this technology in colder climates. Other limitations of phytoremediation are that contaminants present below rooting depth will not be treated or extracted and that the plant or tree may not be able to grow in soils at heavily contaminated sites due to plant toxicity.
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Phytoremediation as a technology is still in its early stages. While many scientists, engineers, and regulators are optimistic that it will eventually be used to clean up organic and metallic contaminants, at least two or three more years of field tests and analyses are necessary to validate the initial, small-scale field tests.1,2 Issues like soil characteristics and length of the growing season will need to be taken into account and scientists must also determine what sites are most amenable to phytoremediation. Other issues such as the potential impact on wildlife remain to be fully explored. Simultaneously, researchers working in the lab are trying to better understand the processes behind phytoremediation to possibly improve its performance during cleanup applications. This chapter will not do justice to this technology by claiming that it will cover the rapidly progressing state of the science and also describe how these scientific advances are being applied in the field for efficient remediation. Instead it will serve as a brief state of the science summary that will allow the reader to understand the current status of the technology and its applications, as well as activities of the research community to further enhance this technology.
5.2 5.2.1
CHEMICALS IN THE SOIL–PLANT SYSTEM
Metals
Elements occur in the soil in a variety of forms more or less available for uptake by plants. Many of the contaminants of concern at waste sites are metals or metalloids. Availability is determined by characteristics of the elements, such as behavior of the ion as a Lewis acid (electron acceptor) which determines the predominant type of strength of bond created (ionic or covalent) and, therefore, the mobility of the metal in the soil environment. Soil characteristics (e.g., pH, clay and organic matter content and type, and moisture content) also determine availability to plants by controlling speciation of the element, temporary immobilization by particle surfaces (adsorption-desorption processes), precipitation reactions, and availability in soil solution. The most general sinks for metals are iron and manganese oxides and organic matter. Although particulate soil organic matter serves to immobilize metals, soluble organic matter may act to keep metals in solution in a form absorbed and translocated by plants. Metal fractionation or sequential extraction schemes — such as toxicity characteristic leaching procedure (TCLP) — sometimes are used to describe metal behavior in soils. Most metals interact with the inorganic and organic matter that is present in the root-soil environment. Potential forms of metals include those dissolved in the soil solution, adsorbed to the vegetation’s root system, adsorbed to insoluble organic matter, bonded to ion exchange sites on inorganic soil constituents, precipitated or coprecipitated as solids, and attached to or inside the soil biomass. The final control on availability of metals and metalloids in soil to plants is the selective absorption from soil solution by the root. Metals may be bound to exterior exchange sites on the root and not actually taken up. They may enter the root passively in organic or inorganic complexes with the mass flow of water or actively
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by way of metabolically controlled membrane transport systems often meant to take up a nutrient which the “contaminant” metal mimics. At different soil solute concentrations, metals may be absorbed by both processes. Absorption mechanisms and quantity absorbed are influenced by plant species (and cultivar), growth stage, physiological state, and the presence of other elements. Once in the plant, a metal can be sequestered in the roots in vacuoles or in association with cell walls and organelles, or translocated to above ground parts in xylem as organic or inorganic complexes. Location and forms of metals in plants, as well as their toxic effects, depend on plant species, growth stage, physiological state, and presence of other metals. Mechanisms of toxicity of metals tend to be dependent on the nature of the reactivity of the metal itself and its availability in the soil and soil solution media. They may alter or inhibit enzyme activity, interfere with deoxyribonucleic acid (DNA) synthesis or electron transport, or block uptake of essential elements.2 Availability in response to toxic levels of metals by different plants is due to a number of defenses. These include exclusion from the root, translocation in nontoxic form, sequestering in nontoxic form, sequestering in nontoxic form in the root or other plant parts, and formation of unusable complexes containing metals that may otherwise be inserted into biomolecules instead of the proper element (e.g., As replacing P). 5.2.2
Organics
Organic compounds of environmental concern include nonionic compounds (such as PAHs, chlorinated benzenes, polychlorinated biphenyls (PCBs), BTEX compounds, and many pesticides), ionizable compounds (chlorophenols, carboxylic acids, surfactants, and amines), and weakly hydrophobic volatile organic compounds (trichloroethene). For the nonionic compounds, sorption in soil is mainly a function of degree of hydrophobicity and amount of sorbent hydrophobic phase (i.e., soil organic matter). Sorption of the compound by soil organic matter is reversible. The activities of these compounds in soil can be predicted by the organic matter-water coefficient, Kom, as estimated by the octanol-water coefficient, Kow.3 Absorption onto colloidal organic matter in solution may alter the availability of these nonionic compounds. Ionizable compounds contain anionic or cationic moieties or both within their structure. These charged structures interact with organic and inorganic charged surfaces in the soil in a variety of reversible reactions. The extent and nature of the associations with charged surfaces depends on characteristics of the organic compound, solution pH and ionic strength, and mineral composition of the soil particulates. Organic compounds may be degraded by microorganisms in the soil to metabolites with greater or lesser toxicity. Very stable compounds, like highly chlorinated PCBs, may persist in essentially unaltered form for many years. Plant roots are not discriminating in uptake of small organic molecules (molecular weight less than 500) except on the basis of polarity.1-4 More water-soluble molecules pass through the root epidermis and translocate throughout the plant. The less soluble compounds (like many polycyclic aromatic hydrocarbons) seem to have limited entry into the plant and minimal translocation once inside. Highly lipophilic
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compounds, such as PCBs, move into the plant root via the symplastic route (from cell to cell, as opposed to between cells) and are translocated within the plant. Within a plant the contaminant may be adsorbed on a cell surface or accumulated in the cell. Many contaminants become bound on the root surface and are not translocated. Not all organic compounds are equally accessible to plant roots in the soil environment. The inherent ability of the roots to take up organic compounds can be described by the hydrophobicity (or lipophilicity) of the target compounds. This parameter is often expressed as the log of the octanol-water partioning coefficient, Kow. Direct uptake of organics by plants is a surprisingly efficient removal mechanism for moderately hydrophobic organic compounds. There are some differences between the roots of different plants and under different soil conditions, but, generally, the higher a compound’s log Kow, the greater the root uptake. Hydrophobicity also implies an equal propensity to partition into soil organic matter and onto soil surfaces. Root absorption may become difficult with heavily textured soils and soils with high native organic matter. There are several reported values available in the literature regarding the optimum log Kow value for a compound to be a good candidate for phytoremediation (as an example, log Kow = 0.5–3.0; log Kow = 1.5–4.0).2,13 It has also been reported that compounds that are quite water soluble (log Kow < 0.5) are not sufficiently sorbed to the roots or actively transported through plant membranes. From an engineering point of view, a tree could be thought of as a shell of living tissue encasing an elaborate and massive chromatography column of twigs, branches, trunk, and roots. The analogous resin in this system is wood, the vascular tissue of the tree, and this “resin” is replenished each year by normal growth. Wood is composed of thousands of hollow tubes, like the bed of a hollow fiber chromatography column, with transpirational water serving as the moving phase. The hollow tubes are actually dead cells, whose death is carefully programmed by the tree to produce a water conducting tissue, which also functions in mechanical support. A complex, cross-linked, polymeric matrix of cellulose, pectins, and proteins embedded in lignin forms the walls of the tubes. The cell wall matrix is chemically inert, insoluble in the majority of solvents, and stable across a wide range of pH. Once an organic chemical is taken up, a plant can store (sequestration) the chemical and its fragments in new plant structures via lignification, or it can volatilize, metabolize, or mineralize the chemical all the way to carbon dioxide, water, and chlorides. Detoxification mechanisms may transform the parent chemical to nonphytotoxic metabolites, including lignin, that are stored in various places in plant cells. Many of these metabolic capacities tend to be enzymatically and chemically similar to those processes that occur in mammalian livers; one report has equated plants to” green livers” due to similarities of detoxification processes. Different plants exhibit different metabolic capacities. This is evident during the application of herbicides to weeds and crops alike. The vast majority of herbicidal compounds have been selected so that the crop species are capable of metabolizing the pesticide to nontoxic compounds, whereas the weed species either lack this capacity or perform it at too slow a rate. The result is the death of the weed species without the metabolic capacity to rid itself of the toxin.
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The shear volume and porous structure of a tree’s wood provide an enormous surface area for exchange or biochemical reactions. Some researchers are attempting to augment the inherent metabolic capacity of plants by incorporating bacterial, fungal, insect, and even mammalian genes into the plant genome.
5.3
TYPES OF PHYTOREMEDIATION
A review of where pyhtoremediation fits into the scheme of hazardous waste remediation enables us to differentiate the various types and mechanisms of phytoremediation (Figure 5.1). The scientific understanding of plant, soil and rhizosphere biochemistry, and contaminant fate and transport must be contrasted with field and pilot studies that represent the current proof of concepts. The technology is summarized below as those approaches ready for application, promising treatments expected to be tested soon, and concepts of phytoremediation requiring intensive development. Finally, the intrinsic strengths of phytoremediation as a technology and the future potential of this technology must be reviewed for regulatory acceptance in terms of hazardous waste remediation.1,2 Mechanisms for Organics
Mechanisms for Inorganics
Atmosphere
Plant Contaminant in the plant
Phytodegradation
Soil
Rhizofiltration
Contaminant in the root-zone (Rhizosphere)
Rhizodegradation Phytostabilization
Impacted Media Figure 5.1
Phytovolatilization
Phytovolatilization
Remediated Contaminant
Contaminant in the air
Phytoaccumulation
Rhyzofiltration
Phytostabilization
Impacted Media
Potential contaminant fates during phytoremediation in the soil–plant–atmosphere continuum.
Phytoremediation approaches can be summarized as follows based on current understanding of the technology: • • • • • •
Phytoaccumulation, phytoextraction, hyperaccumulation Phytodegradation or phytotransformation Phytostabilization Phytovolatilization Rhizodegradation, phytostimulation, or plant assisted bioremediation Rhizofiltration or contaminant uptake
Optimal performance of the technology is an important key to phytoremediation’s ability to gain wider acceptance as a presumptive remediation technique. With
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the possible exception of some of the above mechanisms that are already widely studied and understood, all of phytoremediation’s major applications require further basic and applied research in order to optimize field performance. Significant research and development should be carried out to 1) obtain a better understanding of mechanisms of uptake, transport, and accumulation of contaminants; 2) improve collection and genetic evaluation of hyperaccumulating plants; and 3) obtain a better understanding of interactions in the rhizosphere interactions among plant roots, microbes, and other biota. Short of true regulatory reform, phytoremediation’s ability to make further inroads will depend on how quickly federal, state, and local regulators become convinced of the technology’s efficacy. While not involved in every decision making process, the public is sometimes a key constituency as well. One can expect public interest groups to be more concerned about efficacy and safety issues than cost or other economic factors. However, phytoremediation seems to be faring well with the general public and, according to many practitioners, has already proven popular with neighbors and other interested parties at field remediation sites. 5.3.1
Phytoaccumulation
Remediation of contaminated soils using nonfood crops, called phytoaccumulation, has attracted a great deal of interest in recent years. Also called phytoextraction, phytoaccumulation, refers to the uptake and translocation of metal contaminants in the soil by plant roots into the above ground portions of plants.2 Certain plants, called hyperaccumulators, absorb unusually large amounts of metals in comparison to other plants and the ambient metals concentration (Table 5.1). Table 5.1 The Number of Taxonomic Groups of Hyperaccumulators Varies According to Which Metal is Hyperaccumulated2 Metal Ni Co Cu Zn Mn Pb Cd
Number of Taxonomic Groups of Hyper Accumulators >300 26 24 18 8 5 1
Phytoaccumulators or phytoextractors must have a high accumulation factor, that is, a high uptake of metals from the soil. The uptake should be metal specific, which diminishes the risk of impoverishing the soil of nutrient elements. The property of having a high specific uptake must be genetically stable. Since the removal of metals from the soil is actually achieved through the harvest, it is necessary that the plant have a high transport of the metal(s) from the roots to the shoots to be effective during remediation applications. In addition, a high biomass production of the
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phytoaccumulator is needed for high removal of metals per unit area. It is also an advantage if biomass production is of economic interest. Hyperaccumulators have been preferred during phytoaccumulation applications because they take up very large amounts of a specific metal. They are often endemic and of a specific population (genotypes/clones) of a species.5 However, these plants seldom have high biomass production and may also have low competitive ability in less polluted areas, probably because the plant uses its energy to tolerate such high levels of metals in the tissue instead of growth. Hyperaccumulators can accumulate ≥0.01% of Cd, ≥0.1% of Cu, or ≥1.0% Zn in leaf dry mass and may have the metal evenly distributed throughout the plant.6 There are also high accumulators that accumulate somewhat lower metal concentrations than hyperaccumulators but much more than “normal” plants. They usually have high biomass production. In these plants, there is no uniform distribution of metal throughout the plant, and thus the plant might have high accumulation either in the roots or in the shoots. These plants are selected and planted at a site based on the type of metals present and other site conditions. After they have been allowed to grow for several weeks or months, they are harvested. Landfilling, incineration, and composting are options to dispose of or recycle the metals, although this depends upon the results of TCLP and cost. Planting and harvesting of plants may be repeated as necessary to bring soil contaminant levels down to allowable limits. A plan may be required to deal with the plant biomass waste. Testing of plant tissue, leaves, roots, etc., will determine if the plant tissue is a hazardous waste. Regulators will play a role in determining the testing method and requirements for the ultimate disposal of the plant waste. The state of science in phytoaccumulation is as follows:7 • Botanical prospecting dating to the 1950s in the former USSR and U.S. is available to practitioners. • Over 400 species of hyperaccumulators worldwide have been cataloged. • Field test kits for metal hyperaccumulation have been developed. • Uptake and segregation processes using cation pumps, ion transporters, Ca blocks, metal chelating exudates and transporters, phytochelatin peptides, and metallothioneins have been evaluated and continuous research is being performed to develop further understanding.
The hyperaccumulator plants can contain toxic element levels in the leaf and stalk biomass (LSB) about 100 times more than nonaccumulator plants growing in the same soil, with some species and metal combinations exceeding conventional plant levels by a factor of more than 1000.8 Many hyperaccumulator plants, which are nonwoody (not a tree), have been identified as having the capacity to accumulate metals. Thlaspi caerulascens was found to accumulate Zn up to 2000–4000 mg/kg.9 The Indian mustard plant Brassica juncea, grown throughout the world for its oil seed, was found to accumulate significant amounts of lead.10 One planting of mustard in a hectare of contaminated land was found to soak up two metric tons of lead. If three plantings could be squeezed in per year, six tons of lead per hectare can be extracted. Both hemp dogbane (Apocynum sp.) and common ragweed also have been observed to
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accumulate significant levels of lead. Aeollanthus subcaulis var lineris and Papsalum notatus are other hyperaccumulator plants known to accumulate Cu and Cs, respectively. Hyperaccumulator plants can address contamination in shallow soils only, up to 24 inches in depth. If contamination is deeper, 6–10 feet, deep-rooted poplar trees can be used for phytoextraction of heavy metals. These trees can accumulate the heavy metals by sequestration. However, there are concerns specifically for trees that include leaf litter and associated toxic residues being blown off site. This concern may be tested in the laboratory to see whether uptake and translocation of the metals into the leaves exceed standards. Hyperaccumulators have metal accumulating characteristics that are desirable, but lack the biomass production, adaptation to current agronomic techniques, and physiological adaptations to climatic conditions required at many contaminated sites. It has been reported that harvesting at different seasons in a year had pronounced differences in accumulation levels. In the future, genetic manipulation techniques may provide better hyperaccumulator species. The success of phytoextraction depends on the use of an integrated approach to soil and plant management: the disciplines of soil chemistry, soil fertility, agronomy, plant physiology, and plant genetic engineering are currently being used to increase the rate and efficiency of heavy metal phytoextraction. Chelates have been used not only to enhance metal uptake but also to avoid metal toxicity. Metal accumulator plants have been studied extensively for organometallic complexes. It has been suggested that there is a relationship between metal tolerance and carboxylic acids. Organo-metallic complexes increase the translocation and tolerance of plants to the toxic effects of metals. For example, in Sebertia acuminata citrate seems to be a detoxifying agent as well as an agent in transporting phytotoxic Ni from root systems to the leaves until leaf fall.5,6 It has also been suggested that in copper (Cu) and cobalt (Co) accumulator plants, Co existed as an oxalate complex within the leaf. The formation of Zn–citrate complexes in Zntolerant plants was the reason for high levels of organic acid accumulation. Reports have indicated that histidine was responsible for accumulation, tolerance, and transport to shoots in nonaccumulating and hyperaccumulating (Ni) plant species.11 In Thlaspi, a Zn hyperaccumulator plant species, it has been determined that the majority of Zn in the roots was coordinated with histidine, whereas organic acids were involved in xylem transport and Zn storage in the shoots. Similarly in a Craccumulating plant, Leptospermum scoparium, it was found that soluble Cr in leaf tissue was present as the trioxalatochromium (III) ion, [Cr (C2O4)3]3–. The function of the Cr-organic acid complex was to reduce the cytoplasmic toxicity of Cr.5 Adding ethylenediaminetetraacetic (EDTA) acid, citric acid, or oxalic acid to metal contaminated soils will significantly increase the metal concentrations in plant shoots and roots.5 However, the application of these chelates during a full scale remediation application has to be carefully controlled; if not, the increased solubility of the metal chelates formed could drive these contaminants to migrate further downward by leaching when plant uptake rates are not adequate. Controlling the pH and conditioning the soils for optimum pH is an important factor when dealing with metals-contaminated soils.
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Figure 5.2
248
Process schematic describing the various processes during phytoaccumulation of heavy metals.
The schematic of the process involved in heavy metal phytoextraction is shown in Figure 5.2. Translocation from the root to the shoot must occur efficiently for ease of harvesting. After harvesting, a proper, regulartorily acceptable biomass processing step or disposal methods should be implemented. 5.3.2
Phytodegradation
Phytodegradation, also called phytotransformation, is the breakdown of contaminants taken up by plants through metabolic processes within the plant, or the breakdown of contaminants external to the plant through the effect of compounds (such as enzymes) produced by the plants. Pollutants are degraded, used as nutrients, and incorporated into the plant tissues. In some cases metabolic intermediate or end products are rereleased to the environment depending on the contaminant and plant species (phytovolatilization) (Figure 5.3). Plants synthesize a large number of enzymes as a result of primary and secondary metabolism and can quickly uptake and metabolize organic contaminants to less toxic compounds. Plant enzyme systems can be constitutive or induced and can play a role in solar driven transformations and plant adaptation and/or tolerance to adverse
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Photosynthesis O2
H2O Transpiration and Volatilization of VOCs
CO 2 Phloem Photosynthates +O2
Xylem H2O, Nutrients
Dark Respiration CO2, H O
Phytodegradation - Metabolism within the plant - Production of enzymes which help to catalyze degradation
O2
Lignification, Metabolites Sequestration
H2O, Nutients, O 2 Transpiration
Root Respiration CO2, H O O2
Contaminent Uptake Exudation O2, CH3COOH, C4H5OH Cometabolism
Figure 5.3
Contaminant
CO2, H2O, Cl Mineralization
Phytodegradation and phytovolatilization mechanisms associated with some other mechanisms essential for plant life.
growth conditions resulting from contamination of the soils. Plant-formed enzymes that are useful for phytodegradation are nitroreductases (for munitions and pesticides); dehalogenases (for chlorinated solvents and pesticides); phosphatases (for pesticides); peroxidases (for phenols); laccases (for aromatic amines); cytochrome P-450 (for pesticides and chlorinated solvents); nitrilase (for herbicides). Plant transformation pathways can be of many different types and obviously depend on plant species and tissue type. In simplistic terms, these pathways can be categorized as reduction, oxidation, conjugation, and sequestration. The “green liver model” has been proposed to describe the metabolic pathways of herbicides, pesticides, explosives, and other nitroaromatic compounds. Contaminant degradation by plant-formed enzymes can occur in an environment free of microorganisms (for example, an environment in which the microorganisms have been killed by high contaminant levels). Thus, phytodegradation potentially could occur in soils where biodegradation cannot. The current state of science in phytodegradation (phytotransformation) is summarized below:1,2 • Plant-formed enzymes that degrade organic contaminants have been isolated and metabolic pathways can be predicted. • Phytodegradation can be used for the treatment of soil, sediments, sludges, and groundwater depending on contaminant type and concentrations.
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• Mass balance and pathway analyses studies have been conducted to prove complete degradation; potential toxicity of intermediate compounds also can be predicted. • Differentiation between degradation by plant enzymes, rhizosphere microorganisms, and other breakdown processes is being performed. • Development of engineered solutions based on the use of monocultures vs. multicultures found in wetlands and terrestrial communities is being further investigated. • Organic contaminants are the main category of contaminants with the highest potential of phytodegradation. Inorganic nutrients are also consumed through plant uptake and metabolism. Phytodegradation outside the plant does not depend on log Kow and plant uptake. • Axenic plant tissue cultures of the aquatic plant Myriophyllum and the periwinkle Catharanthus are being used for elucidating plant transformation pathways.
The aquatic plant parrot feather (Myrioplillum aquaticum) and the algae Nitella have been used for the degradation of TNT. The nitroreductase enzyme has also been identified in other algae, ferns, monocots, dicots, and trees. Degradation of TCE has been detected in hybrid poplars and in poplar cell cultures, resulting in production of metabolites and in complete mineralization of a small portion of the applied TCE.12,14 Poplars have been used to remove atrazine and inorganic nutrients.2 Black willow (Salix nigra), yellow poplar (Liriodendron tulipifera), bald cypress (Taxodium diskchum), river birch (Betula nigra), cherry bark oak (Quercus falcata), and live oak (Quercus viginiana) have been known to support degradation of herbicides.13 One recent study demonstrated that poplar trees, which possess cytochrome P-450s analogous to the oxygenases responsible for transformation of compounds such as TCE in the mammalian liver, exposed to 100 mg/L of TCE did uptake and chemically alter this contaminant. TCE and its metabolites were found in the roots and tissue of the study trees, but not in control trees or in the soil used for potting the trees. In a subsequent study, poplar seedlings exposed to 14C-labeled TCE were found to generate 14C-labeled carbon dioxide. Intermediate compounds generated during oxidation are thought to be 2,2,2-trichloroethanol, and di- and trichloroacetic acid. Similar studies have shown positive results for toluene and benzene. A recent study using parrot feather showed positive results for phytotransformation of perchlorate at concentrations of up to 20 ppm.22 Based on the results of these experiments and ecological knowledge of parrot feather, this species is an excellent candidate for future research on in situ phytoremediation of contaminated water bodies. Parrot feather also is a good candidate for phytoremediation of contaminated groundwater temporarily held in artificial ponds. 5.3.3
Phytostabilization
Phytostabilization is the use of certain plant species to immobilize contaminants in the soil and groundwater through absorption and accumulation by roots, adsorption onto roots, or precipitation within the root zone and physical stabilization of soils. It is also used as a means to stabilize contaminated soil by decreasing wind
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and water erosion and to decrease water infiltration and the subsequent leaching of contaminants. This process reduces the mobility of the contaminant and prevents migration to the groundwater or air. This technique can be used to re-establish a vegetative cover at sites where natural vegetation is lacking due to high metal concentrations. Metal-tolerant species may be used to restore vegetation to such sites, thereby decreasing the potential migration of contamination through wind erosion, transport of exposed surface soils, and leaching of soil contamination to groundwater. Implementation of phytostabilization involves reduction in the mobility of heavy metals and high molecular weight organics by minimizing soil erodibility, decreasing the potential for wind blown dust, and reduction in contaminant solubility by the addition of soil amendments. Containment using plants either binds the contaminants to the soil, renders them nonavailable, or essentially immobilizes them by removing the means of transport. Erosion leads to the concentration of heavy metals because of selective sorting and deposition of different size fractions of the soil. Eroded material is often transported over long distances, thus selectively extending the effects of contamination and increasing the risk to the environment. Erosion can, therefore, cause the build up of concentrations of normally nontoxic contaminants to toxic levels at locations where transported material is deposited. Planting of vegetation at contaminated sites, particularly abandoned strip mining sites, will significantly reduce the erodibility of the soils by water and wind; density of vegetation will effectively hold the soil and provide a stable cover against erosion. An excellent example of phytostabilization is everyone’s family garden where plants help to minimize erosion and enhance the stability of the soil. Another element of phytostabilization is to supplement the system with a variety of alkalizing agents, phosphates, organic matter, and biosolids to render the metals insoluble and unavailable to leaching. Materials with a calcareous character or a high pH, such as lime and gypsum, can be added to influence the acidity. Specific binding conditions can be influenced by adding concentrated Fe, Mn or Al compounds. To maintain or raise the organic matter content in the soils, various materials such as humus or peat materials, manure, or mulch can be added. This chemical alteration should be quickly followed by establishing a plant cover and maximizing plant growth. The amendments sequester the metals into the soil matrix and plants keep the stabilized matrix in place, minimizing wind and water erosion. 5.3.4
Phytovolatilization
Phytovolatilization is the uptake and transpiration of a contaminant by a plant, with release of the contaminant or a modified form of the contaminant to the atmosphere from the plant. Phytovolatilization occurs as growing trees and other plants take up water and organic and inorganic contaminants. Some of these contaminants can pass through the plants to the leaves and volatilize into the atmosphere at comparatively low concentrations (Figure 5.3). Many organic compounds transpired by a plant are subject to phytodegradation.
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Thus far, phytovolatilization has mainly been applied to groundwater contamination. However, the potential exists for application to soil, sediments, and other contamination and needs some careful applications.2 The state of science with respect to phytovolatization can be summarized as follows:2,17 • Contaminants could be transformed to less toxic forms (e.g., elemental Hg and dimethyl selenite gas). • The contaminant or a hazardous metabolite might accumulate in vegetation. • Significant reductions of TCE, TCA, and carbon tetrachloride have been achieved in experimental studies. • Poplars, alfalfa (Medicago sativa), and black locust species have been studied to evaluate phytovolatilization. • Indian mustard and canola have been used in phytovolatilization studies of Se.2 Selenium (as selenate) was converted to less toxic dimethyl selenite gas and released to the atmosphere. Kenaf and tall fescue have also been used to take up Se, but to a lesser degree than canola. • A weed from the mustard family (Arabidopsis thaliana), genetically modified to include a gene for mercuric reductase, converted mercuric salts to metallic mercury and released it to the atmosphere.2 • Groundwater must be within the influence of plant (usually a tree) roots and soil must be able to transmit sufficient water to the plant. • Climatic factors such as temperature, precipitation, humidity, solar radiation, and wind velocity can affect transpiration rates and thus the rate of phytovolatilization. • Improved methods for measuring phytovolatilization, diurnal and seasonal variations, and precipitation vs. groundwater use need to be developed. • Significant research needs to be focused on modeling impacts of vegetation such as transpiration stream concentration factors, canopy effects, and root concentration factors.
5.3.5
Rhizodegradation
Rhizodegradation (also called phytostimulation, rhizosphere biodegradation, enhanced rhizosphere biodegradation, or plant-assisted bioremediation/degradation) is the breakdown of contaminants in the soil through microbial activity enhanced by the presence of the rhizosphere (Figure 5.4). Microorganisms (yeast, fungi, and/or bacteria) consume and degrade or transform organic substances for use as nutrient substances. Certain microorganisms can degrade organic substances such as fuels or solvents that are hazardous to humans and ecoreceptors and convert them into harmless products through biodegradation. Natural substances released by plant roots — such as sugars, alcohols, and acids — contain organic carbons that act as nutrient sources for soil microorganisms; these additional nutrients stimulate their activity. Rhizodegradation is aided by the way plants loosen the soil and transport oxygen and water to the area. Plants also enhance biodegradation by other mechanisms such as breaking apart clods and transporting atmospheric oxygen to the root zone. Soil adjacent to the root contains increased microbial numbers and populations.15 It is common knowledge that the number of bacteria in the rhizosphere is as much as 20 times that normally found in nonrhizosphere soil (Figure 5.4). Short gram negative rods (specifically Pseudomonas, Flavobacterium, and Alcaligens) are most
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Enhanced rhizosphere biodegradation - Supply of nutrients, cometabolites - Transport and retention of water - Aeration
Soil dessication Root respiration
Root intrusion
Sloughing
Enzymes dehalogenase nitroductase
Figure 5.4
Uptake
Rhizodegradation and associated processes in the root zone.
commonly found in the rhizosphere.15 The increased microbial numbers are primarily due to the presence of plant exudates and sloughed tissue that serve as sources of energy, carbon, and other growth factors. The products excreted by plants include amino acids, carboxylic acids, carbohydrates, nucleic acid derivatives, growth factors, and enzymes. The activity of microorganisms in the root zone stimulates root exudation further stimulating microbial activity.16 Several studies have evaluated the effect of plants and the associated rhizosphere on the fate of petroleum contaminants.2,4,15 For the most part, the presence of plants enhanced the degradation of contaminants. Also, in studies using 14C-labeled contaminants in closed plant chambers, mineralization was greater in rhizosphere soils than in unvegetated soils, indicating that the bioavailability of the contaminant was higher in the rhizosphere.15 Studies using deep rooted prairie grasses to remediate soils contaminated with PAH suggest that the roots of these perennial grasses may be more effective at stimulating the rhizosphere microflora due to their fibrous nature. Fibrous roots offer more root surface area for microbial colonization than other roots and result in a larger microbial population in the contaminated soil. Big bluestem (Andropogon gerardii), indian grass (Sorghastrum nutans), switch grass (Panicum virgatum), Canada wild rye (Elymus canadensis), little bluestem (Schizachyrium scoparius), side oats grama (Bouteloua curtipendula), western wheatgrass (Agropyron smithic), and blue grama (Bouteloua gracilis) are some of the species known to enhance degradation of petroleum compounds. Crested wheatgrass (Agropyron desertorum) is known to degrade PCP contaminated soils.15 Alfalfa (Meticago sativa), fescue (Festuca anundinacea), big bluestem (Andropogon gerardii), and sudan grass
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(Sorghum vulgare sudanense) are known to enhance the degradation of PAH compounds in the rhizosphere. The degradation rates among various PAHs studied correlated with the water solubility of the compound with the more soluble compound, showing the highest degradation. Cometabolic transformation of chlorinated solvents and other compounds also has been reported in the literature.2 Wherever significant cometabolic transformations took place, the following enzyme systems were present: dehalogenase, nitroreductase, peroxidase, laccase, nitrylase, and oxygenase. The rhizosphere is often divided into two general areas: the inner rhizosphere at the very root surface and the outer rhizosphere embracing the immediately adjacent soil. The microbial population is larger in the inner zone where biochemical interactions are most pronounced and root exudates are concentrated. In addition to plant exudates, the rapid decay of fine-root biomass can also become an important addition of organic carbon to soils. A recent report considers some strategies for engineering plants to improve bioremediation in the root zone. One of the simpler approaches is to make use of the organism Agrobacterium rhizogenes to induce a state called “hairy root disease.” Depending on virulence of the strain used, the extent of root production is variable, but generally, infection leads to a significant enhancement of rooting without obvious detrimental effects on the host plant. Increased root mass has the apparent advantage of increasing the surface area available for microbial colonization. Root exudation may be increased in proportion to increase in root area. Such rhizosphere enhancements could improve bioremediation potential of the plantmicrobial system. It is suggested that when water is not freely available in unlimited quantities, increased root mass could lead to greater water uptake, and hence greater contaminant mobilization and potential degradation. Different plant species often establish somewhat different subterranean floras (Figure 5.5). The differences are attributed to variations in rooting habits, tissue composition, and excretion products of the plant. The primary root population is Poplar Trees 15 ft.
Alfalfa 4-6 ft. Grasses 2 ft.
Figure 5.5
Examples of different root depths.
Indian Mustard 1 ft.
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determined by the habitat created by the plant; the secondary flora, however, depends upon the activities of the initial population. The age of the plant also alters the microbial population in the rhizosphere. Roots also harbor mycorrhizae fungi, which metabolize organic contaminants. These fungi, growing in symbiotic association with the plant, have unique enzymatic pathways, similar to white rot fungus enzymes that help to degrade organics that could not be transformed solely by bacteria. In summary, plants provide exudates that offer an excellent habitat for increased microbial populations and pump oxygen to roots, a process ensuring aerobic transformations near the root that otherwise may not occur in bulk soil. Due to the presence of certain primary substrates in the exudate system, anaerobic cometabolic transformations may also take place in the rhizosphere. Typical microbial population in the rhizosphere comprise: 5 × 106 bacteria, 9 × 105 actinomycetes, and 2 × 103 fungi per gram of air dried soil. The state of science in phytodegradation can be summarized as follows: • Contaminant degradation can be achieved in situ, which is the biggest advantage. • Translocation of the contaminant to the plant or atmosphere is less likely than with other phytoremediation techniques since degradation takes place at the source of contamination. • There are low installation and maintenance cost(s) since no harvesting and disposal are required. • Various microorganism species and enzymes have been isolated which degrade different contaminants. • Analytical methods to better quantify treatment efficiency and success are improving. • Field management techniques for nutrients, water, and plant selection are advancing. • TPH and PAHs up to hundreds of ppm have been studied in the field with varying success.2 • Degradation of various pesticides (atrazine, metolachlor, parathion, diazinon, and 2,4-D, 2,4,5-T herbicides) has been studied, again with mixed results.2 • TCE, PCP and PCB degradation have also been investigated — again with varying success. • More research needs to be done to further elucidate: microbial metabolism in the rhizosphere, toxicity towards plants, biodiversity in the rhizosphere, biogeochemical optimization in the rhizosphere, and interrelation between biological, chemical and physical characteristics of the rhizosphere.
The following plants, in addition to the ones discussed previously, have been used for successful implementation of phytodegradation at field sites:2 1) red mulberry, crabapple, spearmint, and osage orange that are capable of stimulating PCB degradation; 2) alfalfa, loblolly pine, and soybean for TCE degradation;3) alfalfa for TCA degradation; and 4) rye, St. Augustine, and white clover for TPH. Growth of hybrid poplar trees for the application of phytodegradation and rhizodegradation is shown in Figures 5.6a, b, and c.
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Figure 5.6a
Phytoremediation System, August 6, 1998.
Figure 5.6b
Phytoremediation System, September 13, 1999.
5.3.6
Rhizofiltration
Rhizofiltration is the adsorption or precipitation of contaminants onto plant roots or the absorption of contaminants into the roots when contaminants are in solution
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Figure 5.6c
257
Phytoremediation System, August 22, 2000.
surrounding the root zone. In some applications, the plants are raised in greenhouses hydroponically (with their roots in water rather than in soil). Once a large root system has been developed, contaminated water is diverted and brought in contact with the plants or the plants, are moved and floated in the contaminated water. The plants are harvested and disposed as the roots become saturated with contaminants. Plant uptake, concentration and translocation might occur, depending on the contaminant. Exudates from the plant roots might cause precipitation of some metals. Rhizofiltration first results in contaminant containment, in which the contaminants are immobilized or accumulated on or within the plant; contaminants are then removed by removing the plant. Aquatic plants and algae are known to accumulate metals and other toxic elements from solution.18 There are large differences in bioremoval rates due to species and strain differences, cultivation methodology, and process control techniques. In the past, commercial systems have used immobilized algae biomass for removing radionuclides and other heavy metals in the aqueous phase.19 Naturally immobilized, plants such as attached algae and rooted plants, and those easily separated from suspension, such as filamentus microalgae, macroalgae, and floating plants, have been found to have high adsorption capacities. In a recent study, one blue green filamentous alga of the genus Phormidium and one aquatic rooted plant, water milfoil (Myriophyllum spicatum), exhibited high specific adsorption for Cd, Zn, Ph, Ni, and Cu.18 It has been reported that porous beads containing immobilized biological materials such as sphagnum peat moss can be used for extracting metals dissolved in the aqueous phase.20 The beads designated as BIO-FIX beads readily adsorbed Cd, Pb, and other toxic metals from dilute waters. In one recent study, it was reported that
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Saccharomyces cerevisiae yeast biomass, when treated with a hot alkali, exhibited an increase in its biosorption capacity for heavy metals.21 It was also reported that caustic treated yeast immobilized in alginate gel could be reactivated and reused to remove Cu, Cd, and Zn in a manner similar to the ion exchange resin. Phytoremediation applications are summarized in Tables 5.2a and b based on contaminant fate, degradation, extraction, containment type, or a combination of these applications. In the soil–plant–atmosphere continuum, a specific contaminant can be remediated at specific points along this continuum by different phytoremediation mechanisms. Table 5.2a Types of Phytoremediation for Organic Constituents Type of Phytoremediation 1.
Phytostabilization
2.
Rhizodegradation (phytostimulation, rhizosphere bioremediation, or plant-assisted bioremediation)
3.
Rhizofiltration (contaminant uptake)
4.
Phytodegradation (phytotransformation)
5.
Phytovolatilization
Process Involved Plants control pH, soil gases, and redox conditions in soil to immobilize contaminants. Humification of some organic compounds is expected.
Contaminant Treated
Expected for phenols, chlorinated solvents (tetrachloromethane and trichloromethane) and hydrophobic organic compounds Plant exudates, root necrosis, and Polyaromatic hydrocarbons, other processes provide organic BTEX, and other carbon and nutrients to spur soil petroleum hydrocarbons, bacteria growth by two or more perchlorate, atrazine, orders of magnitude. Exudates alachlor, polychlorinated stimulate degradation by biphenyl (PCB), and other mycorrhizal fungi and microbes. organic compounds Live roots can pump oxygen to aerobes and dead roots may support anaerobes. Compounds are taken up or Hydrophobic organic sorbed by roots (or sorbed to chemicals algae and bacteria). Aquatic and terrestrial plants take Munitions (TNT, DNT, HMX, up, store, and biochemically nitrobenzene, picric acid, degrade selected organic nitrotoluene), atrazine, compounds to harmless halogenated compounds byproducts, products used to (tetrachloromethane, create new plant biomass, or trichloromethane, byproducts that are further hexachloroethane, carbon broken down by microbes and tetrachloride, TCE, other processes to less harmful tetrachloroethane, products. Reductive and dichloroethant), DDT and oxidative enzymes may be used other chlorine and in series in different parts of the phosphorus based plant. pesticides, phenols, and nitrites Volatile organic compounds are Chlorinated solvents taken up and transpired. Some (trichloroethane), organic recalcitrant organic compounds VOCs, BTEX, MTBE are more easily degraded in the atmosphere (photodegradation).
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Table 5.2b Types of Phytoremediation for Inorganic Constituents Type of Phytoremediation 1.
Phytostabilization
2.
Rhizofiltration (contaminant uptake)
3.
Phytoaccumulation (phytoextraction or hyperaccumulation)
4.
Phytovolatilization
5.3.7
Process Involved
Contaminant Treated
Plants control pH, soil gases, and redox conditions in soil to immobilize contaminants. Humification of some organic compounds is expected. Compounds are taken up or biosorbed by roots (or sorbed to algae and bacteria). Metals and organic chemicals taken up by the plant with water, or by cation pumps, sorption and other mechanisms. Volatile metals are taken up, changed in species, and transpired.
Proven for heavy metals in mine tailing ponds
Heavy metals and radionuclides
Nickel, zinc, lead, chromium, cadmium, selenium, other heavy metals radionuclides Mercury and selenium
Phytoremediation for Groundwater Containment
Phytoremediation can be applied for containment of contaminated groundwater under the right hydrogeologic conditions such as sites with shallow groundwater depths. In general, favorable economics is one factor in phytoremediation’s favor, particularly in contrast to the high cost of operation and maintenance of conventional groundwater treatment systems. Furthermore, the high pumping rates of many deep rooted trees may make them more efficient at removing water at low permeability sites. Phreatophytes (like willows, cottonwood, and hybrid poplar), which take up and “process” large volumes of soil water are good candidates for phytoremediation applications specifically for groundwater containment. For example, a single willow tree on a hot summer day transpires more than 5000 gallons of water, and a hybrid poplar can transpire about 50 to 350 gallons per day.23 Phytoremediation of groundwater plumes is preferred when the contaminants are water soluble, leachable organics, and inorganics present at concentrations that are not phytotoxic. Hydraulic control by plants can occur only within the root zone or within a depth influenced by roots; the placement depth of roots during planting can be varied. Root depth, early tree growth, and nutrient uptake were enhanced by placing poplar tree root balls closer to shallow groundwater during planting.23 The primary considerations for selecting phytoremediation for hydraulic control as the method of choice are the depth and concentration of contaminants that affect plant growth. Soil texture and degree of saturation are also influential factors. Planning technique and materials can extend the influence of plants through nonsaturated zones to water-bearing layers. As mentioned earlier, phreatophytes such as poplars are capable of extending their roots into aerobic water tables. For example, the roots of poplars growing
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alongside streams can easily be observed intertwined in the stream bottom. The degree to which poplar roots would penetrate the saturated zone cannot be easily estimated. If their access to soil moisture from precipitation is limited, poplars will draw large amounts of water from the top of saturated aquifer. Evapotranspiration will draw down the water table below the trees similar to a pump and treat system (Figures 5.7a and b). Simulations of a proposed design can be carried out based on extent of contamination, hydrogeological data, past precipitation and infiltration records, and evapotranspiration data. A big advantage of phytoremediation over conventional pump and treat systems is the ability of the roots to penetrate the microscopic scale pores in the soil matrix. Contaminants adsorbed or trapped in these micropores are impacted minimally or not at all by the pump and treat system. In the case of phytoremediation, the roots can penetrate these micropores for contaminant removal.
Above Capillary Fringe At Capillary Fringe In Capillary Fringe and Groundwater Table
Figures 5.7a
5.3.8
Placement of root ball with time due to maturation of the tree.
Phytoremediation of Dredged Sediments
Dredged material is nothing more than displaced topsoil that enters and is eventually removed from navigable waterways. Contaminant discharges into waterways over time result in contamination of bottom sediment. Dredged sediments are usually stored in confined disposal facilities (CDF).24 The application of phytoremediation to dredged material presents some challenges unique to dredged material. Dredged sediments come from an aquatic environment and are initially wet and anaerobic after placement in a CDF. Subsequent drying and oxidation depend on dewatering and management techniques. Drying and oxidation of surface layers may result in physicochemical changes that may affect plant establishment and contaminant mobility. Although the surface layer of
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31.0
Groundwater flow
Zone of tree plantation
28.0 29.0
30.5
30.0
30.0
Groundwater table elevation contours Figures 5.7b
Predicted groundwater flow conditions at maturation of tree growth.
dredged sediments in a CDF may be dry and aerobic, deeper layers may remain anaerobic. Saltwater dredged sediments provide another level of difficulty for vegetation and in most cases must be leached to reduce soluble salt levels. Dredged material management is further complicated by the potential of elevated concentrations of multiple contaminants. The selection of plant species and methods of establishment will be determined by these factors. Common contaminants present in dredged sediments are metals, PAHs, polychlorinated phenols, PCBs and other heavy molecular weight compounds. The current state of knowledge indicates that phytoremediation of dredged sediments would not be as readily effective as application to more heavily contaminated industrial sites.
5.4
PHYTOREMEDIATION DESIGN
The design of a phytoremediation system varies according to contaminants, conditions at the site, level of cleanup required, and plants used. A thorough site characterization should provide the needed data to design any type of remediation system. Clearly, phytoextraction has different design requirements from phytostabilization or rhizodegradation. Nevertheless, it is possible to specify a few design considerations that are part of most phytoremediation efforts (Figures 5.8a, b, and c). Site characterization data will provide the information required for the designer to develop a properly functioning system. The design considerations include contaminant levels; plant selection; treatability; irrigation, agronomic inputs (P N, K, salinity, zinc, etc.), and maintenance; groundwater capture zone and transpiration rate; and contaminant uptake rate and clean-up time required. Other factors to be considered during the evaluation, design, and implementation phases of phytoremediation at a contaminated site are:
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Figure 5.8a
Decision tree for phytoremediation in soil.
• Soil Water — The most crucial factor in a plant’s life is water, which links it to the soil via roots and serves as a vehicle for nutrient transport. Water also controls the exchange of gases and moderates soil temperature changes. Plant available water is held in the soil between the field capacity and permanent wilting point. Plant roots can extract water at lower potentials, depending upon the plant type and arable environment. Root growth rates are controlled by the presence of continuing supplies of water to maintain hydrostatic pressure in the elongating
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Decision Tree for Phytoremediation Groundwater
YES
NO
YES
YES
Will the climate support the proposed plants?
YES
Is time or space a constraint?
Is the contaminant physically within the range of the proposed plant (typically less than 10- 20 feet bgs for Salix species - willows, cottonwoods, poplars) ?
Will the plants be used for hydraulic control ONLY (prevent water from REACHING the contaminated zone)?
Is the contaminant at phytotoxic concentrations (this may require a greenhouse dose-response test)?
YES
YES
Will state regulations allow this type of phytoremediation?
Will the rhizosphere microbes and plant-exuded enzymes degrade the target contaminants in the rhizosphere and are the metabolic products acceptable?
YES YES
Will the plants transpire the contaminant or metabolic products?
Are the quantity and rate of transpiration acceptable for this site?
YES
YES
Can engineering controls make it acceptable? Is the final disposition of the contaminant or metabolic products acceptable?
NO
YES
Will the plant degrade the contaminant after uptake and are the metabolic products acceptable?
NO
YES
Is the level of accumulation acceptable for this site throughout the growth of the plant?
NO
Can controls be put in place to prevent the transfer of the contaminant or metabolic products from a plant to humans/animals ?
NO
YES
Can the contaminant or metabolic product be immobilized to acceptable levels ?
NO
YES
NO
NO
NO
Does the plant material constitute a waste if harvested?
Can the plant waste be economically disposed?
Phytoremediation has the potential to be effective at the site
Figure 5.8b
NO
YES
YES
YES
NO
Will the plant accumulate the contaminant or metabolic products after uptake?
NO
NO
NO
YES
Is the log Kow of the contaminant or metabolic poducts between 1 and 3.5 (will uptake occur)?
NO
NO
Will the water be mechanically pumped and applied to the phytoremediation system?
NO
YES
NO
NO
YES
NO
Phytoremediation is NOT an option at the site;consider other options
Decision tree for phytoremediation groundwater.
cells of the root, and metabolites for cell wall construction. Water flows radially into elongating root cells only when the cell’s total water potential is lower than the combined osmotic and matric potentials of the soil. Soil water content will influence plant biomass growth.
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Decision Tree for Phytoremediation Sediments Will the climate support the proposed plants?
YES
NO
YES
YES
Is time or space a constraint?
NO
Can the sediments be treated in place (wetlands)?
YES YES
Is there strong public support to treat the sediment as a soil?
Is the contaminant at phytotoxic concentrations (this may require a greenhouse dose-response test)?
YES
NO
YES
YES
YES
YES
Will the plants transpire the contaminant or metabolic products?
Are the quantity and rate of transpiration acceptable for this site?
Will the plant accumulate the contaminant or metabolic products after uptake?
YES
NO
Is the level of accumulation acceptable for this site throughout the growth of the plant?
NO
YES
YES
Can controls be put in place to prevent the transfer of the contaminant or metabolic products from a plant to humans/animals?
NO NO
NO
NO YES
Can the contaminant or metabolic products be immobilized to acceptable levels ?
Does the plant material constitute a waste if harvested?
Can the plant waste be economically disposed?
Phytoremediation has the potential to be effective at the site
Figure 5.8c
NO
YES
Is the final disposition of the contaminant or metabolic products acceptable?
YES
NO
NO
Will the plant degrade the contaminant after uptake and are the metabolic products acceptable?
YES
Can engineering controls make it acceptable?
NO
NO
Are there hotspots that can be removed or treated?
YES
YES
NO
NO
Will the rhizosphere microbes and plant-exuded enzymes degrade the target contaminants in the rhizosphere and are the metabolic products acceptable?
Is the log Kow of the contaminant or metabolic products between 1 and 3.5 (will uptake occur)?
NO
NO
Is the contaminant physically within the range of the proposed plant (typically less than 1- 2 feet bgs)?
YES
NO
Are the sediments to be dredged?
Will the regulatory statutes allow the dredged sediments to be treated as a soil?
YES
NO
NO
YES
NO
Phytoremediation is NOT an option at the site; consider other options
Decision tree for phytoremediation sediments.
• Soil Air — Plants need molecular oxygen to respire and convert carbohydrates to CO2 and H2O. This is an exothermic reaction and releases respiratory energy utilized for many plant processes. The disappearance of O2 triggers a sequence of changes in the biogeochemical properties of the soil; the absence of O2 alone is sufficient to alter plant metabolism profoundly. Suboptimal concentrations of
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•
•
•
5.4.1
265
O2 in the soil occur because of interactions among soil properties such as porosity, water content, temperature, surface water infiltration, and continuity of air filled pores with biotic activity. Soil Temperature — Temperature influences plant processes at the cellular level, such as osmotic potential, hydration of ions, stomatal activity and transpiration, Gibbs free energy available for work, membrane permeability, solute solubilities, diffusion, and enzymatic activities. Temperature and cultivar strongly influence the establishment of plants. Low temperatures also decrease metabolic activity and root growth. Physical Impedance — Physical impedance, sometimes called mechanical impedance or excessive soil strength, can severely affect normal root growth patterns. Such impedances result from increased soil bulk density, increased cohesion and friction between soil particles, reduction in soil water content, frost-heave action of soil, and presence of permafrost within the root zone. Under an excessive soil strength environment, roots enter the soil volume where pore sizes are larger than the root tip. Conversely, if pore sizes are too small for entry of the main root but not for the laterals, then laterals proliferate and produce a highly branched root system. Topography — Topography is a critical factor because it is a key factor in determining runoff velocity and erosion. In general, the amount of soil erosion increases manifold with increasing degree and length of slope. Contaminated sites with slopes greater than 10% are often not suitable for phytoremediation without surface modification because of excessive erosion. Soil pH — Plant roots are damaged at pH lower than 4.0. The roots are shortened, thickened, fewer in number, and dull brown or gray in color. Salinity is another challenge to phytoremediation applications in the field. Soluble salts reduce the total water potential of the soil solution, thus tending to reduce the potential difference between soil water and the atmosphere. Excessive soil salinity reduces root elongation and upsets hormonal balance, as well as altering soil structure that, in turn, affects plant growth.
Contaminant Levels
During the site characterization phase the concentration level of the contaminants of concern will be established. High levels of contamination may eliminate phytoremediation as a treatment option. Plants are not able to treat all contaminants. The composition of organic compounds (structure, log Kow, degree of weathering and boiling point range) and degree of adsorption are important factors in phytoremediation. It is important to understand the range of contaminants that can be treated using phytoremediation. In addition to knowing contaminants and their concentrations, the depth of the contaminants must be known. The primary consideration in this area is that the contaminant concentrations cannot be phytotoxic or cause unacceptable impacts on plant health or yield. Higher concentrations of contaminants might be tolerated more readily by plants than by soil microorganisms. 5.4.2
Plant Selection
The goal of the plant selection process is to choose a plant species with suitable characteristics for growth under site conditions that meet the objectives of
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phytoremediation. Native, crop, forage, and other types of plants that can grow under regional and climatic conditions should be preferred. Plants are selected according to the application and the contaminants of concern. For phytotransformation of organic compounds, the design requirements are that vegetation is fast growing and hardy, easy to plant and maintain, utilizes a large quantity of water by evapotranspiration, and transforms the contaminants of concern to nontoxic or less toxic products. In temperate climates, phreatophytes (e.g., hybrid poplar, willow, cottonwood, and aspen) are often selected because of fast growth, a deep rooting ability down to the level of groundwater, large transpiration rates, and the fact that they are native throughout most of the country. A screening test or knowledge from the literature of plant attributes will aid the design engineer in selection of plants. Plants used in phytoextraction include sunflowers and Indian mustard for lead; Thlaspi spp. (Pennycress) for zinc, cadmium, and nickel; and sunflowers and aquatic plants for radionuclides. Aquatic plants are used in constructed wetlands applications. The two categories of aquatic plants used are emergent and submerged species. Emergent vegetation transpires water and is easier to harvest if required. Submerged species do not transpire water but provide more biomass for the uptake and sorption of contaminants. 5.4.3
Treatability
Treatability or plant screening studies are recommended prior to designing a phytoremediation system. If the decision tree flowcharts indicate phytoremediation is an applicable technology for a site, a plant scientist should assist in the treatability studies which assure concerned parties that the phytoremediation system will achieve desired results. Treatability studies provide toxicity and transformation and assess the fate of the contaminants in plant system. Different concentrations of contaminant are tested with proposed plant species. Volatile organic compounds are often transpired to the atmosphere by plants; calculations will predict the amount and type of material transpired. 5.4.4
Irrigation, Agronomic Inputs, and Maintenance
Irrigation of plants ensures a vigorous start to the system even in drought. Hydrologic modeling may be required to estimate the rate of percolation to groundwater during irrigation conditions. Irrigation should be withdrawn if the area receives sufficient rainfall to sustain the plants. Agronomic inputs include the nutrients necessary for vigorous growth of vegetation and rhizosphere microbes. The soil must be analyzed and then items such as nitrogen, potassium, phosphorous, aged manure, sewage sludge compost, straw, and/or mulch are added as required to ensure the success of the plants. Maintenance of the phytoremediation system may include adding fertilizer, agents to bind metals to the soil, or chelates to assure plant uptake of the contaminants. Replanting may be required due to drought, disease, insects, or animals killing plants.
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267
Groundwater Capture Zone and Transpiration Rate
For applications involving groundwater remediation, a capture zone calculation can be used to estimate whether the phytoremediation pump (trees) can be effective at entraining the plume of contaminants. The goal is to create a water table depression where contaminants will flow to the vegetation for uptake and treatment. Organic contaminants are not taken up at the same concentrations in the soil or groundwater because membranes at the root surface reduce the uptake rate. Although it is possible to estimate the uptake rate of contaminants, the calculation is beyond the scope of this chapter.
REFERENCES 1. McCutcheon, S. C., USEPA, personal communications, 1999, 2000. 2. USEPA, Introduction to Phytoremediation, EPA/600/R-99/107, Washington D.C., February, 2000. 3. Schnoor, J. L. et al., Phytoremediation of organic and nutrient contaminants, Environ. Sci. Technol., 29, 1620–1631, 1995. 4. McCutcheon, S. C., Phytoremediation of organic compounds: science validation and field testing, in Workshop on Phytoremediation of Organic Wastes, Kovalick, W. W. and Olexsey, R., Eds., Ft. Worth, TX, December, 1996. 5. Shahandeh, H. and Hossner, L. R., Enhancement of Cr (III) phytoaccumulation, Int. J. Phytoremed., 2, 269–286, 2000. 6. Brooks, R. R., Plants That Hyperaccumulate Heavy Metals, CAB International, New York, NY, 1998. 7. McCutcheon, S. C., The science and practice of phytoremediation, in Phytoremediation: State of the Science Conf., Boston, MA, May, 2000. 8. Cornish, J. E. et al., Phytoremediation of soils contaminated with toxic elements and radionuclides, in Bioremedation of Inorganics, Hinchee, R. E. et al., Eds., Battelle Press, Columbus, OH, 1995. 9. Brown, S. L. et al., Zinc and cadmium uptake by hyperaccumulator Thlaspi caerulescens and metal tolerant Silene vulgaris grown on sludge amended soils, Environ. Sci. Technol., 29, 1581–1590, 1995. 10. Bishop, J. E., Pollution fighters hope a humble weed will help reclaim contaminated soil, Wall Street Journal, August 7, 1995. 11. Kramer, et al., Free histidine as a metal chelator in plants that accumulate nickel, Nature, 379, 635–638. 12. Newman, L. A. et al., Uptake and biotransformation of trichloroethylene by hybrid poplars, Environ. Sci. Technol., 31, 1062–1067, 1997. 13. Conger, R. M. and Portier, R., Phytoremediation experimentation with the herbicide bentazon, Remediation, 7, 19–37, 1997. 14. Narayanan, M., Davis, L. C., and Erickson, L. E., Fate of volentile chlorinated organic compounds in a laboratory chamber with alfafa plants, Environ. Sci. Technol., 29, 2437–2444, 1995. 15. Fiorenza, S., Oubre, C. L., and Ward, C. H., Phytoremediation of Hydrocarbon Contaminated Soil, Lewis Publishers, Boca Raton, Florida, 2000. 16. Alexander, M., Introduction to Soil Microbiology, John Wiley & Sons, New York, 1977.
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17. Stomp, A. M. et al., Genetic strategies for enhancing phytoremediation, Ann. NY Acad. Sci., 721, 481–491, 1994. 18. Want, T. C., Weissman, J. S., Ramesh, G., Varadarajan, R., and Benemann, J. R., Bioremoval of toxic elements with aquatic plants and algae, in Bioremediation of Inorganics, Hinchee, R. E. Means, J. L., and Burns, D. R., Eds., Battelle Press, Columbus, OH, 1995. 19. Feiler, H. D. and Darnall, D. W., Remediation of Groundwater Containing Radionuclides and Heavy Metals using Ion Exchange and the Alga SORD Biosorbent System, Final Report under Contract No. 02112413, DOE/CH-9212, 1991. 20. Jeffers, T. H., Bennett, P. G., and Corwin, R. R., Biosorption of metal contaminants using immobilized biomass: field studies, Report of Investigations 9461, Bureau of Mines, US Department of the Interior, 1993. 21. Lu, Yongming and Wikins, E., Heavy metal removal by caustic-treated yeast immobilized in alginate, in Bioremediation of Inorganics, Hinchee, R. E., Means, J. L., and Burris, D. R., Eds., Battelle Press, Columbus, OH, 1995. 22. Susarla, S. et al., Phytotransformation of perchlorate using parrot feather, Soil and Groundwater Cleanup, March, 1999. 23. Gatliff, E., personal communication, 2000 24. DOE, Phytoreclamation of dredged material; a working group summary, Technical Note, DOER-C9, November, 1999.
CREDIT Figures 5.1, 5.8a,b,c, and Tables 5.2a,b were reproduced from Phytoremediation Decision Tree, prepared by Interstate Technology and Regulatory Cooperation Work Group, Phytoremediation Work Team, November 1999.
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CHAPTER
6
Constructed Treatment Wetlands CONTENTS 6.1
6.2
6.3
6.4
Introduction ..................................................................................................270 6.1.1 Beyond Municipal Wastewater ........................................................272 6.1.2 Looking Inside the “Black Box” .....................................................273 6.1.3 Potential “Attractive Nuisances”......................................................274 6.1.4 Regulatory Uncertainty and Barriers ...............................................275 Types of Constructed Wetlands ...................................................................276 6.2.1 Horizontal Flow Systems.................................................................276 6.2.2 Vertical Flow Systems......................................................................277 Microbial and Plant Communities of a Wetland.........................................278 6.3.1 Bacteria and Fungi ...........................................................................278 6.3.2 Algae ................................................................................................279 6.3.3 Species of Vegetation for Treatment Wetland Systems...................279 6.3.3.1 Free-Floating Macrophyte-Based Systems.......................282 6.3.3.2 Emergent Aquatic Macrophyte-Based Systems ...............284 6.3.3.3 Emergent Macrophyte-Based Systems with Horizontal Subsurface Flow ...............................................................285 6.3.3.4 Emergent Macrophyte-Based Systems with Vertical Subsurface Flow ...............................................................285 6.3.3.5 Submerged Macrophyte-Based Systems ..........................285 6.3.3.6 Multistage Macrophyte-Based Treatment Systems..........287 Treatment-Wetland Soils..............................................................................287 6.4.1 Cation Exchange Capacity...............................................................289 6.4.2 Oxidation and Reduction Reactions ................................................290 6.4.3 pH .....................................................................................................292 6.4.4 Biological Influences on Hydric Soils.............................................292 6.4.5 Microbial Soil Processes..................................................................292 6.4.6 Treatment Wetland Soils ..................................................................293
269
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6.5
Contaminant Removal Mechanisms ............................................................294 6.5.1 Volatilization ....................................................................................294 6.5.2 Partitioning and Storage...................................................................295 6.5.3 Hydraulic Retention Time................................................................297 6.6 Treatment Wetlands for Groundwater Remediation....................................299 6.6.1 Metals-Laden Water Treatment........................................................300 6.6.1.1 A Case Study for Metals Removal ..................................302 6.6.2 Removal of Toxic Organics .............................................................306 6.6.2.1 Biodegradation ..................................................................306 6.6.3 Removal of Inorganics .....................................................................309 6.6.4 Wetland Morphology, Hydrology, and Landscape Position............309 References..............................................................................................................310
Creating or constructing a natural wetland sounds like an oxymoron, but this doesn’t mean that an “unnatural wetland” is by definition bad. It doesn’t mean we can’t mimic Mother Nature in giving natural birth to a desirable wetland. Constructed rice paddies have been responsible for feeding more people than any other enterprise on earth.
6.1
INTRODUCTION
Natural wetlands are land areas that are wet during part or all of the year because of their location in the landscape. Historically, wetlands were called swamps, marshes, bogs, fens, or sloughs, depending on existing plant and water conditions and on geographic setting. Wetlands are frequently transitional between uplands (terrestrial systems) and continuously or deeply flooded (aquatic) systems. They are also found at topographic lows (depressions) or in areas with high slopes and low permeability soils (seepage slopes). In other cases, wetlands may be found at topographic highs or between stream drainages when land is flat and poorly drained (blanket bogs). In all cases, the unifying principle is that wetlands are wet long enough to alter soil properties because of the chemical, physical, and biological changes that occur during flooding, and to exclude plant species that cannot grow in wet soils.1 The structural components of natural wetland ecosystems are shown in Figure 6.1. These components are highly variable and depend on hydrology, underlying sediment types, water quality, and climate. Starting with the unaltered sediments or bedrock below the wetlands, these typical components are1 • Underlying strata — unaltered organic, mineral, or lithic strata, typically saturated with or impervious to water and below the active rooting zone of the wetland vegetation • Hydric soils — the mineral-to-organic soil layer of the wetland, infrequently to continuously saturated with water and containing roots, rhizomes, tubers, funnels, burrows, and other active connections to the surface environment
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271
Canopy Tree
Subcanopy Tree Emergent Vegetation
Shrub
Seasonal High Water
Rhizomes Detritus Hydric Soils
Cypress Kness Buttressed Stem
Seasonally Flooded Zone Seasonal Low Water
Unaltered Sediment
Figure 6.1
Structural components of natural wetland ecosystems (adapted from Kadlec et al., 1996).
• Detritus — the accumulation of live and dead organic material in a wetland, consisting of dead emergent plant material, dead algae, living and dead animals (primarily invertebrates), and microbes (fungi and bacteria) • Seasonally flooded zone — the portion of wetland seasonally flooded by standing water and providing habitat for aquatic organisms including fish and other vertebrate animals, submerged and floating plant species that depend on water for buoyancy and support, living algae, and populations of microbes • Emergent vegetation — vascular, rooted plant species containing structural components that emerge above the water surface, including both herbaceous and woody plant species
Natural wetlands have been used as convenient wastewater discharge sites for as long as sewage has been collected (at least 100 years in some locations). Examples of old treatment wetland sites can be found in Massachusetts, Wisconsin, Florida, and Ontario. Judging by the growing number of wetlands built for wastewater treatment around the world, this “natural” technology seems to have firmly established roots. After almost 30 years of use in wastewater treatment, constructed “treatment wetlands” now number over 1000 in Europe and in North America.1 Marsh-type “surface flow” systems are most common in North America, but “subsurface flow” wetlands, where wastewater flows beneath the surface of a gravel-rock bed, predominate in Europe. This inexpensive, low-maintenance technology is reportedly in high demand in Central America, Eastern Europe, and Asia. New applications,
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from nitrate-contaminated groundwater to effluent from high-intensity livestock operations, are also increasing. In the U.S., treatment-wetland technology has not yet gained universal regulatory acceptance; projects are approved on a case-by-case basis. Some states and EPA regions are eager to endorse them, but others are wary of this nontraditional method of treating wastewater and contaminated groundwater. In part, this reluctance exists because the technology is not yet completely understood. Knowledge of how the wetland works is not far enough advanced to provide engineers with detailed predictive models. Because wetlands are natural systems, their performance is variable, subject to the vagaries of changing seasons and vegetative cycles. These treatment wetlands also pose a potential threat to wildlife attracted to this new habitat within an ecosystem exposed to potentially toxic compounds. When utilized for benign, pretreated wastewaters, wetlands do not generally pose a threat to human or wildlife health. In these circumstances, there may be significant ancillary benefits in terms of habitat creation and beneficial human use. In those situations where a potentially hazardous condition exists, the extra expense of a gravel media is warranted.1 Water and associated particulates, organisms, and sediments are then located below ground, and thus out of reach of human and wildlife contact. Subsurface wetland waters are typically anoxic or anaerobic, which is optimal for some processes such as sulfide precipitation or denitrification, but unsatisfactory for other processes, such as nitrification of ammonium nitrogen. New efforts are underway, however, to place the technology onto firmer scientific and regulatory ground. Long-term demonstration and monitoring field studies are currently probing the inner workings of wetlands and their water quality capabilities to provide better data on how to design more effective systems. Researchers are documenting the fate of toxic compounds in wetlands and the extent to which wildlife may be exposed to them. A recent study of U.S. policy and regulatory issues surrounding treatment wetlands has recommended that the federal government actively promote this technology and clear the regulatory roadblocks to enable wider use. Proponents argue that the net environmental benefits of constructed wetlands, such as restoring habitat and increasing wetland inventory, should be considered. A federal interagency work group is grappling with that recommendation, trying to balance the benefits and shortcomings of this increasingly popular technology. 6.1.1
Beyond Municipal Wastewater
Constructed wetland systems in North America have been designed predominantly for large-scale treatment of municipal wastewater, ranging from 100,000 to 15 million gallons per day.1,2 The use of treatment wetlands is well established in Europe, where the technology originated with laboratory work in Germany 30 years ago.3 Subsurface-flow systems are the norm because they provide more intensive treatment in a smaller space than marsh-type wetlands — an important design constraint in countries where open space is limited. The European thrust has been for small-scale systems primarily for domestic wastewater treatment; for example, Denmark alone has 150 systems, most in small villages handling domestic wastewater. The term “reed beds” is commonly used for treatment wetlands in Europe.
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Since the 1980s, constructed wetlands have also been built to treat other types of wastewaters, including acid mine drainage, industrial wastewater, agricultural and storm water runoff, and effluent from livestock operations.1,2 The petroleum industry is using constructed wetlands to treat a variety of wastewaters from refineries and fuel storage tanks. Food processing and pulp and paper industries are relative newcomers to treatment wetlands. Stormwater runoff also has recently become a focus of research in using constructed wetlands as a treatment method. While many of the early acid mine drainage treatment systems were marsh-like surface flow systems, the most recent projects are “passive treatment systems” that link several different types of cells — vertical limestone drain as well as vegetated cells — to sequentially treat particularly “nasty” wastewater with low pH and high metals content.2 A wetland system for the treatment of runoff from coal piles at coal-fired power plants with a pH of 2 and high levels of metals uses a series of successive alkalinity-producing systems, a rich organic layer over an anoxic limestone drain, to reduce the acidity in the wastewater before it flows into wetland cells. Landfill leachates are a subset of polluted waters requiring substantial levels of treatment. Leachates vary considerably, depending upon the materials accepted at the landfill. They may contain large concentrations of volatile and toxic organics, both as individual compounds and as COD, chlorinated organics, metals, and nitrogenous compounds.2 Wetland treatment of landfill leachates has been successfully tested at several locations. Cold climate systems are functioning properly in Norway, as well as at several locations in Canada; reed beds are used to treat leachate in the United Kingdom, Slovenia, and Poland.4 Based on current understanding of the effectiveness of wetland treatment of leachates, several U.S. projects are in planning and design phases. In addition, there are about a half-dozen other projects in various locations, such as Mississippi, Indiana, Pennsylvania, and West Virginia. Wetlands have been proposed for control of stormwater runoff from capped landfills.1,2 Continued growth in the use of treatment wetlands is expected as a result of new regulatory initiatives on nutrient management, including the Clean Water Act’s total maximum daily load (TMDL) program. Small- to medium-sized communities trying to meet new TMDLs in sensitive watersheds for phosphorus or ammonia need something that is cost-effective, and wetlands are a good option. 6.1.2
Looking Inside the “Black Box”
The rapid spread and diversification of treatment-wetland technology are running ahead of the mechanistic understanding of how they work. These complex natural systems are still, somewhat, a “black box,” according to many in the field. For example, the role of plants in transporting oxygen into the root zone to promote nitrification has been demonstrated in the laboratory but not convincingly in the field, according to many researchers. There is very little data to say whether that is an important factor or whether the plants are more or less passive. It is likely, according to some researchers, that the ratio of open water to vegetated areas is more important in creating aerobic conditions in a wetland.
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Another issue quite often debated is how important the volume of water in a wetland is to treatment performance. Is it the bottom of the wetland or the volume of water that is more important? The data coming in now are on the side of the wetland bottom:1,2 it apparently does not matter how deep the water is as long as the soil is wet. That is a surprise to civil engineers, who, for years, have designed treatment systems based on their volume and hydraulic residence time. Numerous research efforts, both broad based and focused, are currently generating a great deal of new information on treatment-wetland function.1,2,5 The extensive research activities include gathering conventional water quality data; measurements of metals, biotoxicity, and organics; bird surveys; and macroinvertebrate sampling. Expanding the species pallet of plants used in treatment wetlands is another focus of research among researchers in this field. Most constructed wetlands for treatment have been built around herbaceous species so far, and many researchers are experimenting with a greater variety of plants to see how water quality changes when multispecies systems are used. Many have found that pathogen removal is higher in a multispecies system than in a single species system. One of the things that may be important in pathogen removal is having multiple types of wetland components, for example, a duckweed system followed by a subsurface wetland.5 Looking deeper into the wetland, to the microbes in the soil and around the root systems of wetland plants, some researchers are studying the role that bacteria play in trace element removal. Researchers have found that bacteria in the root zone of bulrush increase the plants’ ability to accumulate and volatilize selenium twofold. They are now working to identify which bacteria are most responsible, and will soon move to mesocosm studies to see whether seeding the soil with those bacteria increases trace element removal. Some researchers are experimenting with an innovative wetland design — a vertical flow system — to solve the oxygen depletion problem and boost nitrification.1,2 Effluent flows over a porous surface and percolates through a vegetated sand filter, which is periodically allowed to dry to reintroduce oxygen to the system. 6.1.3
Potential “Attractive Nuisances”
Aside from research issues surrounding the design and performance of the treatment wetlands black box, another scientific issue looms large for the future of the technology: do treatment wetlands pose a threat to wildlife?1,5 This question is an important one, since many wetland projects are designed with habitat creation as one of their primary beneficial objectives. It is easier to justify the land use for a constructed wetland if it is also used for habitat restoration. Research is also being directed toward several critical issues. Some researchers are working to find out exactly where toxic trace elements from wastewater end up in a treatment wetland. They are completing laboratory studies documenting trace element uptake potential of various wetland plants and identifying where the elements go in the plants: roots, stems, leaves, or plant litter. They are also monitoring several active treatment wetlands to track trace elements in the ecosystem: sediment, water, air, plant tissues, and animal tissues.
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To address similar habitat-related issues, influent and effluent water have been analyzed for potential bioaccumulation and mutagenic activity from organic compounds.5 Toxicity tests were designed to look for physiological impacts on biota living in the system. Work also continues on the control of an unplanned threat to human health: mosquitoes. Fish have been introduced to the wetlands to consume mosquito larvae, but the density of the particular bulrush variety used may prevent the fish from reaching certain parts of the wetland. Sections of the wetlands can be reconfigured and replanted to raise the water level and give the fish greater access. 6.1.4
Regulatory Uncertainty and Barriers
Treatment wetlands do not appeal to all wastewater engineers because they lack the traditional “handles” of engineered pollution control systems, are not easy to control, and may be hard to predict. Regulators in the U.S. have similar problems with treatment wetlands because they do not fit easily into existing regulatory categories. Surface-flow treatment wetlands can be a point source discharge and a protected environment at the same time. No national guidance on the use of treatment wetlands and no uniform acceptance of them by states exist, according to researchers and consultants. In this atmosphere of regulatory uncertainty, questions abound. Concerns have been expressed that under a strict reading of the Clean Water Act, certain treatment wetlands could be considered “waters of the U.S.,” and thus discharges into them could be tightly regulated. USEPA’s environmental technology initiative (ETI) treatment wetland policy and permitting team of representatives from federal, state, and local agencies issued a report in January, 1997, that recommended “changes in regulation and/or policy that would facilitate, where appropriate, implementation of beneficial treatment wetland projects.”6,7 It also advocated that “net environmental benefits” of habitat creation, reduced use of energy and treatment chemicals, and recreational value — not just the water quality impact of a treatment wetland project — should be considered in approving it. The report catalogued numerous regulatory and policy issues. Should disinfection of effluent be done at the inlet rather than the outlet of a wetland? When should a wetland be lined to protect groundwater? Should treatment wetlands be allowed to mitigate for permitted wetland losses? Under what conditions should constructed treatment wetlands be considered “waters of the U.S.?” The report also noted that more research is needed concerning the “fate and effect of potential wastewater toxins and ecological risks in treatment wetlands.” The federal interagency work group, including representatives from USEPA wetlands and wastewater offices, the U.S. Army Corps of Engineers, the National Oceanic and Atmospheric Administration, the Bureau of Reclamation, and the U.S. Fish and Wildlife Service, was created to take up these issues.6,7 The question of where treatment wetlands should be sited has been a particularly difficult regulatory issue, and consensus must be reached on the need to handle wetland systems differently depending on whether their primary purpose is water treatment or habitat restoration. There is still some disagreement about the habitat
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value of treatment wetlands and concerns about the negative impact they could have on the environment. USEPA currently is not developing the type of specific guidance documents and formal agency actions recommended in the ETI study to promote the use of treatment wetlands. Nevertheless, wetlands experts are encouraged because the issues are now being discussed at the national level.
6.2 6.2.1
TYPES OF CONSTRUCTED WETLANDS
Horizontal Flow Systems
The purposeful construction of treatment wetland ecosystems is a relatively new technology. Constructed wetlands for pollution control, wastewater treatment, and, recently, for contaminated groundwater treatment are divided into two basic types: free water surface (FWS) and subsurface flow (SSF) wetlands. Both types consist of a channel or a basin with some sort of barrier to prevent seepage and utilize emergent aquatic vegetation as part of the treatment system. The difference between FWS and SSF wetlands is that SSF uses some kind of media as a major component (Figures 6.2a and b). In an FWS treatment wetland, soil supports the roots of the emergent vegetation; water at a relatively shallow depth of 6 to 24 inches flows through the system with the water surface exposed to the atmosphere. Oxygen is provided by diffusion through the water surface.
Inlet
Outlet Weir
Figure 6.2a
Free water surface (FWS) wetland.
An SSF treatment wetland bed contains a suitable depth (1.5 – 3.0 feet) of permeable media, such as coarse sand or crushed stone, through which the water flows. The media also support the root structure of the emergent vegetation. The surface of the flowing water is beneath the surface of the top layer of the medium, determined by proper hydraulic design and appropriate flow control structures. In both systems the polluted water undergoes physical, biological, and chemical treatment processes as it flows through the wetlands. The rate at which organic contaminants move through wetlands can be determined by several transport mechanisms. These mechanisms often act simultaneously on the organics and may include such processes as convection, diffusion, dispersion, and zero- or first-order production or decay.
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Inlet Effluent Porous Media
Figure 6.2b
Subsurface flow (SSF) wetland.
Currently, constructed wetlands for municipal wastewater treatment are designed based on the assumptions of plug-flow hydrodynamics and first-order biochemical oxygen demand (BOD) removal kinetics. The first assumption implies that dispersion in the system is negligible and all the fluid particles have a uniform detention time traveling through the system. The plug-flow model seems to give a reasonably accurate estimate of the performance of SSF-constructed wetlands. However, some designers have recognized the limitation of using the plug-flow model for constructed wetlands design. Three types of hydraulic inefficiencies may occur in treatment wetlands: one caused by internal islands and topographical features, a second caused by preferential flow channels on a large-distance scale, and a third caused by mixing effects, such as water delays in litter layers and transverse mixing. 6.2.2
Vertical Flow Systems
Vertical flow constructed wetlands are vegetated systems in which the flow of water is vertical rather than horizontal as in FWS and SSF wetlands (Figure 6.3).
Figure 6.3
Vertical flow constructed wetland.
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Polluted water is applied at time intervals over the entire surface of the wetland. The water flows through a permeable medium and is collected at the bottom. The intermittent application allows the cell to drain completely before the next application. This type of operation allows for much more oxygen transfer than typical SSF systems and thus may be a good option for treatment of wastewaters with a relatively high oxygen demand. This type of system has been recommended for removal of high levels of ammonia through nitrification. High BOD levels may cause clogging due to biomass buildup; mineral buildup may also cause clogging. Intermittent application gives the advantage of greatly increasing the oxygen available for microbial reactions, but also greatly increases the mechanical and operational requirements of the system over the more traditional wetland treatment processes.
6.3
MICROBIAL AND PLANT COMMUNITIES OF A WETLAND
Because of the presence of ample water, wetlands are typically home to a variety of microbial and plant species. This biological diversity, from the smallest virus to the largest tree, creates interspecies interactions resulting in greater diversity, more complete utilization of energy inflows, and ultimately, emergent properties of the wetland ecosystem.1,6,7 The treatment wetland system designer should not expect to maintain a system with just a few known species. The successful wetland designer creates the gross environmental conditions suitable for groups or guilds of species, seeding the wetland with diversity by planting multiple species, using soil seed banks, and inoculating from other similar wetlands, and then using minimum external control to guide the wetland development.1 This form of ecological engineering results in lower initial cost, lower operation and maintenance costs, and the most consistent system performance. 6.3.1
Bacteria and Fungi
Wetland and aquatic habitats provide suitable environmental conditions for the growth and reproduction of microorganisms, two important groups of which are bacteria and fungi. These organisms are important in treatment wetland systems primarily because of their role in the assimilation, transformation, and recycling of chemical constituents present in contaminated waters. Bacteria and fungi are typically the first organisms to colonize and begin the sequential decomposition of contaminants and wastes. Also, microbes typically have first access to dissolved constituents in the wastewater or contaminated groundwater. Some bacteria are sessile, while others are motile by use of flagella. In wetlands, most bacteria are associated with solid surfaces of plants, decaying organic matter, and soils. Bacteria also play a significant role in altering the biogeochemical environment because they are responsible for processes such as nitrification, denitrification, sulfate reduction and methanogenesis, etc. Fungi represent a separate kingdom of eucaryotic organisms and include yeasts, molds, and fleshy fungi. Most fungal nutrition is saprophytic, which means it is
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based on the degradation of dead organic matter. Fungi are abundant in wetland environments and play an important role in treatment. 6.3.2
Algae
Algae are unicellular or multicellular photosynthetic bacteria and plants that lack the variety of tissues and organs of higher plants. Algae are a highly diverse assemblage of species that can live in a wide range of aquatic and wetland habitats. Major algae life forms typical in wetland environments are unicellular, colonial, filamentous, and macroscopic forms. For the most part, algae depend on light for their metabolism and growth and serve as the basis for an autochthonous food chain in wetland habitats. Organic compounds created by algae photosynthesis contain stored energy which is used for microbial respiration or which enters the aquatic food chain and provides food to a variety of microbes. Alternately, this reduced carbon may be directly deposited as detritus to form organic peat sediments in wetlands. When light and nutrients are plentiful, algae can create massive populations and contribute significantly to the overall food web and nutrient cycling of a treatment wetland ecosystem. When shaded by the growth of macrophytes, algae frequently play a less important role in wetland energy flows and treatment (Figure 6.4). Filamentous algae mats are sometimes a dominant component of the plant biomass in wetland systems. The mats are made of a few dominant species of green or blue-green filamentous algae in which individual filaments may include thousands of cells. Filamentous algae mats first develop below the water surface on the substrate of wetland in areas with little emergent vegetation. During the day, entrained gas bubbles (primarily pure oxygen resulting from photosynthesis) may cause the mats to move up through the water column and float at the surface. During the night, the mats sink again to the wetland substrate.1 Filamentous algae that occur in wetlands as periphyton or mats may dominate the overall productivity of the wetland, controlling DO and CO2 concentrations within the treatment wetland water column. Wetland water column DO can fluctuate diurnally from near zero during the early morning following a night of high respiration to well over saturation (>15 mg/L) in high algae growth areas during a sunny day. Dissolved carbon dioxide and consequently the pH of the water varies proportionally to DO because of the corresponding use of CO2 by plants during photosynthesis and release at night during respiration. As CO2 is stripped from the water column by algae during the day, pH may rise by 2 to 3 pH units (a 100- to 1000fold increase in H+ concentration). These daytime pH changes are reversible, and the production of CO2 at night by algae respiration frequently returns the pH to the previous day’s value by early morning (Figure 6.5.)1 6.3.3
Species of Vegetation for Treatment Wetland Systems
The term macrophyte includes vascular plants with tissues that are easily visible. Vascular plants differ from algae through their internal organization into tissues resulting from specialized cells (Figure 6.6). The U.S. Fish and Wildlife Service has
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Figure 6.4
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Major energy sources and ecological niches affecting the occurrences of algae in wetlands (adapted from Kadlec and Knight, 1996).
10
pH 10
pH
Dissolved Oxygen (mg/L)
20
DO
0
0 12 MN
6 AM
12 NOON
6 PM
12 MN
TIME Figure 6.5
Typical diurnal plots of DO concentration and pH in a wetland dominated by filamentous algae.
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Cattail Duck Potato
Emergent Herbaceous
a. Buttonbush Shrub
Emergent Woody
b. Water Lilly
c.
Floating Leaved
Hydrilla
d. Figure 6.6
Submerged Growth forms of rooted wetland and aquatic vascular plants (adapted from Kadlec and Knight, 1996).
more than 6700 plant species on their list of obligate and facultative wetland plant species in the U.S. Obligate wetland plant species are defined as those which are found exclusively in wetland habitats, while facultative species are those that may be found in upland or in wetland areas.1 Wetland macrophytes are the dominant structural component of most wetland treatment systems. The vascular macrophytes can be categorized morphologically
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by descriptors such as woody, herbaceous, annual, perennial, emergent, floating, and submerged. Woody species have stems or branches that do not contain chlorophyll. Since these tissues are adapted to survive for more than one year, they are typically more durable or woody in texture. Herbaceous species have aboveground tissues that are leafy and filled with chlorophyll-bearing cells that typically survive only one growing season. Woody species include shrubs that attain heights of up to six to ten feet and trees that are generally more than ten feet in height when mature.1 The terms emergent, floating, and submerged refer to the predominant growth form of a plant species. In emergent plant species, most of the aboveground part of the plant emerges above the waterline and into the air. These emergent structures may be self-supporting or may be supported by other physical structures. Emergent plant species are important because they provide surface area for microbial growth important in many of the contaminant assimilation processes in treatment wetland systems.1,2 Floating species have leaves and stems buoyant enough to float on the water surface. Submerged species have buoyant stems and leaves that fill the niche between sediment surface and the top of the water column. Floating and submerged species may appear in treatment wetlands when water depths exceed the tolerance range for rooted, emergent species. Aquatic macrophyte-based wetlands treatment systems may be classified according to the life form of the dominating macrophyte into 1) free-floating macrophyte-based treatment systems; 2) rooted emergent macrophyte-based wastewater treatment systems; 3) submerged macrophyte-based wastewater treatment systems; and 4) multistage systems consisting of a combination of the above-mentioned concepts and other kinds of low-technology systems (e.g., oxidation ponds and sanitary filtration systems). 6.3.3.1 Free-Floating Macrophyte-Based Systems Free-floating macrophytes are highly diverse in form and habit, ranging from large plants with rosettes of aerial and/or floating leaves and well-developed submerged roots (e.g., water hyacinth, Eichhornia crassipes) to minute surface-floating plants with few or no roots (e.g., duckweeds, Lemna, Spirodella, Wolffia sp.) (Figure 6.7a).2 Influent Effluent
Figure 6.7a
Schematic description of a free-floating water hyacinth-(Eichhornia crassipes) based treatment wetland system.
Water Hyacinth-Based Systems: The water hyacinth is one of the most prolific and productive plants in the world. This high productivity is exploited in wetland treatment facilities. Two different concepts are applied in water hyacinth-based
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wastewater treatment systems: 1) tertiary treatment systems (i.e. nutrient removal) in which nitrogen and phosphorus are removed by incorporation into the water hyacinth biomass, which is harvested frequently to sustain maximum productivity and to remove incorporated nutrients. Nitrogen may also be removed as a consequence of microbial denitrification; and 2) integrated secondary and tertiary treatment systems (i.e., BOD and nutrient removal) in which degradation of organic matter and microbial transformations of nitrogen (nitrification-denitrification) proceed simultaneously in the water hyacinth ecosystem. Harvesting of water hyacinth biomass is only carried out for maintenance purposes. The latter system should include aerators, that is, areas with a free water surface where oxygen can be transferred to the water from the atmosphere by diffusion and where algal oxygen production can occur. The retention time in the systems varies according to wastewater characteristics and effluent requirements, but is generally on the order of from 5 to 15 days.2 The role of water hyacinths in the process of suspended solids removal is well documented. Most suspended solids are removed by sedimentation and subsequent degradation within the basins, although some sludge might accumulate on the sediment surface. The dense cover of water hyacinths effectively reduces the effects of wind mixing and also minimizes thermal mixing. The shading provided by the plant cover restricts algal growth, and hyacinth roots impede the horizontal movement of particulate matter. Furthermore, electrical charges associated with hyacinth roots are reported to react with opposite charges on colloidal particles such as suspended solids, causing them to adhere to plant roots, where they are removed from the wastewater stream and slowly digested and assimilated by the plant and microorganisms. The efficiency of water hyacinths in removing BOD and in providing good conditions for microbial nitrification is related to their capability of transporting oxygen from the foliage to the rhizosphere. The extensive root system of the water hyacinth provides a huge surface area for attached microorganisms, thus increasing the potential for decomposition of organic matter.1,2 Water hyacinth-based wetland treatment systems are sufficiently developed to be applied successfully in tropics and subtropics. Water hyacinths are severely affected by frost; the growth rate is greatly reduced at temperatures below 10°C. Consequently, in temperate regions, water hyacinth-based systems can only be used in greenhouses or outdoors during summer. Pennywort (Hydrocotyle umbellate), on the other hand, has a high growth rate and a high nutrient uptake capacity even during relatively cold periods in subtropical areas.2 It has been suggested that water hyacinths and pennywort can be alternately cultured, winter and summer, in order to maintain performance at a high level year-round. Duckweed-Based Systems: Duckweeds (Lemna, Spirodella, and Wolffia sp.) have not been investigated as much as water hyacinths for use in wetlands treatment. Duckweeds, have a much wider geographic range than water hyacinths, however, as they are able to grow at temperatures as low as 1 to 3°C. Compared to water hyacinths, duckweeds, play a less direct role in the treatment process because they lack extensive root systems and therefore provide a smaller surface area for attached microbial growth.2 The main use of duckweeds is therefore in recovering nutrients from secondary treated wastewater. A dense cover of duckweed on the surface of water inhibits both oxygen entering the water by diffusion and the photosynthetic
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production of oxygen by phytoplankton because of poor light penetration. The water consequently becomes largely anaerobic, which in turn favors denitrification. The light absorption of duckweed cover restricts growth of phytoplankton and therefore the production of suspended solids. Duckweed-based systems may be plagued by problems, as high winds can pile the duckweed into thick mats and eventually completely sweep the plants from the water. Therefore, in large systems, it is necessary to construct some kind of barrier on the water surface to prevent this. The retention time in duckweed-based wetland treatment systems depends on wastewater quality, effluent requirements, harvesting rate, and climate, but it varies typically from 30 days during summer to several months during winter. 6.3.3.2 Emergent Aquatic Macrophyte-Based Systems Rooted emergent aquatic macrophytes are the dominant life form in wetlands and marshes, growing within a water table ranging from 18 inches below the soil surface to a water depth of 60 inches or more. In general, they produce aerial stems and leaves, and an extensive root and rhizome system. The depth penetration of the root system, and thereby the exploitation of sediment volume, is different for different species. Typical species of emergent aquatic macrophytes are the common reed (Phragmites australis), cattail (Typha latifolia), and bulrush (Scirpus lacustris).2 All species are morphologically adapted to growing in a waterlogged sediment by virtue of large internal air spaces for transportation of oxygen to the roots and rhizomes. Most species of emergent aquatic macrophytes possess an extensive internal lacunal system that may occupy 50 to 70% of the total plant volume. Oxygen is transported through the gas spaces to the roots and rhizomes by diffusion and/or by convective flow of air. Part of the oxygen may leak from the root system into the surrounding rhizosphere, creating oxidized conditions in the otherwise anoxic sediment and stimulating both decomposition of organic matter and growth of nitrifying bacteria. Emergent macrophyte-based wetland treatment systems can be constructed with different designs; see Figure 6.7b for an example. These types of systems are also currently applied for the precipitation and removal of dissolved heavy metals under anaerobic conditions as a sulfide or carbonate precipitate.
Figure 6.7b
Emergent macrophyte treatment wetland system with surface flow (adapted from Mohiri, 1993).
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6.3.3.3 Emergent Macrophyte-Based Systems with Horizontal Subsurface Flow Design typically consists of a bed planted with the common reed Phragmites australis and underlain by an impermeable membrane to prevent seepage if required. The medium in the bed may be soil or gravel. During the passage of wastewater or contaminated groundwater through the rhizosphere of the reeds, organic matter is decomposed microbiologically, nitrogen may be denitrified, and phosphorus and heavy metals may be fixed in the soil. The reeds have two important functions in the process: 1) to supply oxygen to the heterotrophic microorganisms in the rhizosphere, and 2) to increase and stabilize the hydraulic conductivity of the soil. The quantitative significance of the uptake of nutrients in the plant tissue is negligible, as the amount of nutrients taken up during a growing season constitutes only a few percent of the total content introduced with the wastewater. Moreover, nutrients bound in the plant tissue are recycled in the system upon decay of the plant material. Surface runoff is a general problem in soil-based treatment facilities because it prevents the wastewater from coming into contact with the rhizoshere. Furthermore, the oxygen transport capacity of the reeds seems to be insufficient to ensure aerobic decomposition in the rhizosphere and deliver the oxygen needed for quantitatively significant nitrification. 6.3.3.4 Emergent Macrophyte-Based Systems with Vertical Subsurface Flow In a vertical flow system the requirements for sufficient hydraulic conductivity in the bed medium and improved rhizosphere oxygenation can be established. A design consisting of several beds laid out in parallel with percolation flow and intermittent loading will increase soil oxygenation several-fold compared to horizontal subsurface flow systems. During the loading period, air is forced out of the soil; during the drying period, atmospheric air is drawn into the porespaces of the soil, thus increasing soil oxygenation. Furthermore, diffusive oxygen transport to the soil is enhanced during the drying period, as the diffusion of oxygen is approximately 10,000 times faster in air than in water. This design and operational regime provides alternating oxidizing and reducing conditions in the substrate, thereby stimulating sequential nitrification–denitrification and phosphorus adsorption (Figure 6.7c). The limited information available on the treatment performance of such systems indicates good performance with respect to suspended solids and aerobically biodegradable organics, ammonia, and phosphorus. 6.3.3.5 Submerged Macrophyte-Based Systems Submerged aquatic macrophytes have their photosynthetic tissue entirely submerged (Figure 6.7d). The morphology and ecology of the species vary from small, rosette-type, low-productivity species growing only in oligotrophic waters (e.g., Isoetes lacustris and Lobelia dortmanna) to larger eloedid-type, high-productivity species growing in eutrophic waters (e.g., Elodea canadensis).
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Figure 6.7c
Emergent macrophyte based treatment wetland system with vertical percolation.
Figure 6.7d
Schematic description of a submerged macrophyte-based treatment wetland system.
Submerged aquatic plants are able to assimilate nutrients from polluted waters. However, they only grow well in oxygenated water and therefore cannot be used in wastewater with a high content of readily biodegradable organic matter because the microbial decomposition of the organic matter will create anoxic conditions. The prime potential use of submerged macrophyte-based wastewater treatment systems is therefore for “polishing” secondarily treated wastewaters, although good treatment of primary domestic effluent has been obtained in an Elodea nuttallii-based system. The presence of submerged macrophytes depletes dissolved inorganic carbon in the water and increases the content of dissolved oxygen during the periods of high photosynthetic activity. This results in increased pH, creating optimal conditions for volatilization of ammonia and chemical precipitation of phosphorus. High oxygen concentrations also create favorable conditions for the mineralization of organic matter in the water. The nutrients assimilated by the macrophytes are largely retained within the rooting tissues of the plants and by the attached microflora. Losses from the foliage of plant nutrients upon senescence of the macrophyte tissues are readily taken up by the periphytic community so that very little leaves the littoral detritus and macrophyte-epiphyte complexes. Much of the detrital matter produced in these systems will be accumulated and retained in the sediments. The use of submerged macrophytes for wastewater treatment is still in the experimental stage, with species like egeria (Egeria densa), elodea (Elodea canadensis and Elodea nuttallii), hornwort (Ceratophyllum demersum), and hydrilla
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(Hydrilla verticillata) being the most promising. Present knowledge suggests that their prime area of application will be as a final step in multistage systems. 6.3.3.6 Multistage Macrophyte-Based Treatment Systems Different types of these macrophyte-based wastewater treatment systems may be combined with each other or with conventional treatment technologies. For example, a multistage system could consist of 1) a mechanical clarification step for primary treatment; 2) a floating or emergent macrophyte-based treatment system for secondary treatment; and 3) a floating, emergent, or submerged macrophyte-based step for tertiary treatment. The types of secondary and tertiary treatment step will, among other factors, depend on wastewater characteristics, treatment requirements, climate, and amount of available land.
6.4
TREATMENT WETLAND SOILS
Several individual component wetland processes combine to provide the observed overall treatments. Sedimentation and filtration remove solids. Chemical precipitation (abiotic or microbially induced), ion exchange, and plant uptake remove metals. Nutrients are utilized by plants and algae, and cycled to newly formed sediments. The definition of hydric soils indicates that any upland soil utilized for construction of a wetland treatment system will become a hydric soil following a short to long period of flooding and continuous anaerobiosis.1 Hydric soils are defined as soils that, in their undrained condition, are saturated, flooded, or ponded long enough during the growing season to develop anaerobic conditions favoring the growth and regeneration of hydrophylic vegetation.8 Since most wetlands are constructed in former uplands, most constructed wetlands are initially dominated by mineral soils. As constructed wetland treatment systems mature, the percent of organic matter in the soil generally increases, and in some systems, soils might eventually cross the arbitrary line between mineral and organic (Figures 6.8a and b). Mineral soils are classified by particle size distributions, color, depth, and a number of other factors. The three major mineral soil classes are clays, silts, and sands. Clays are soils with very fine particles packed closely together. Because of their very fine texture and low hydraulic conductivity, clays may function as aquitards. The existence of many natural wetlands depends on impermeable clay lenses in sedimentary or wind-blown (loess) deposits. Clays typically have the highest adsorption potential of any soils because of their high surface area to volume ratio resulting from their small particle size distribution. When water in a wetland is in contact with underlying clays or when water percolates through the bottom of a clay-lined wetland, the presence of clays may greatly increase treatment potential for conservative ions such as phosphorus and metals. Organic soils, called peat, muck, or mucky peat may be classified by their extent of decomposition. Those soils with the least amount of decomposition (less than one third decomposed) are called peat. Fibric peats have more than two thirds of
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Young Plants
Water
Saturated Mineral Soils
Aerobic
Mildly Anaerobic (Positive REDOX)
Figure 6.8a
The types of soils present in a newly planted treatment wetland system (adapted from Kadlec, 1996).
Figure 6.8b
Types of soil layers developed after a period of maturation in a treatment wetland system (adapted from Kadlec, 1996).
the plant fibers still identifiable. Saprists or mucks have greater than two thirds of the original plant materials decomposed. Hemists (mucky peat or peaty mucks) are between saprists and fibrists. Due to their fibrous nature, organic soils may shrink, oxidize, and subside when they are drained. Fire may also accelerate this oxidation process, and agricultural
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practices (drainage, cropping, harrowing, and burning) are known to result in soil subsidence in highly organic soils such as those in the Everglades agricultural area where subsidence rates have been estimated at about 3 cm/yr. Drying organic soils promotes oxidation and gasifies carbon, but not the mineral nutrients associated with those soils. Although the available nutrient content of a peat or muck is often quite low, there are large amounts of nitrogen, phosphorus, sulfur, and other mineral constituents organically bound in unavailable forms. Oxidation destroys these recalcitrant organics and releases the associated substances. Upon reflooding, those substances can dissolve and provide relatively high concentrations of nutrients and other dissolved minerals. Organic soils cannot easily be characterized by grain size because the necessary act of drying destroys the physical-chemical structure. The general range of hydraulic conductivity for soils found in sedge, reed, and alder wetlands is 0.1 to 10 m/d, placing these materials in the range of other mineral soils. However, this is true only for fully saturated soils; even a slight degree of unsaturation lowers the hydraulic conductivity by two orders of magnitude, due to the extremely large capillary suction pressure created in the micropores. This means that organic soils and sediments are virtually undrainable; they retain a very high percentage of water. Organic soils are typically dark in color, ranging from black mucks to brown peats. Soil chemical properties are primarily related to chemical reactivity of soil particles and the surface area available for chemical reactions. Chemical reactivity is related to the surface electrical charge of the soil particles, is typically highest in clays and organic soil particles. 6.4.1
Cation Exchange Capacity
Wetland soils have a high trapping efficiency for a variety of chemical constituents; they are retained within the hydrated soil matrix by forces ranging from chemical bonding to physical dissolution within the water of hydration. The combined phenomena are referred to as sorption. A significant portion of chemical binding is cation exchange, which is replacement of one positively charged ion, attached to the soil or sediment, with another positively charged ion. The humics substances found in wetlands contain large numbers of hydroxyl and carboxylic functional groups, which are hydrophilic and serve as cation binding sites. Other portions of these molecules are nonpolar and hydrophobic in character. The result is the formation of micelles, groups of humic molecules with their nonpolar sections combined in the center and their negatively charged polar portions exposed on the surface of the micelle. Protons or other positively charged ions may then associate with these negatively charged sites to create electrical neutrality. Micelles are one form of ligand that can bind metal ions. The cumulative process of binding a metal ion to a ligand (L) to form a complex may be described by a chemical equation; here, it is illustrated for the binding of a divalent metal ion (M):1 2HL + M2+ ⇔ ML2 + 2H+
(6.1)
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The number of ligands per gram of dry solid is determined from the number of metal ions that can be sorbed by a fully protonated sample. This is referred to as the cation exchange capacity (CEC) of the material, usually measured in milliequivalents per gram. Peats have CEC values of approximately 1.0 to 1.5 meq/g. For a heavy metal such as copper, this can translate to a large binding capacity, on the order of a few percent by weight. Clearly, the pH of the soil or sediment has a large influence on the partitioning of a metal to the ligand because excess hydrogen ions drive Equation 6.1 toward the ionic form of the metal. Drying of the organic material will destroy some of the character of the highly hydrated micellular chemical-physical structures, therefore destroying some of the sorption capacity of the material. The sorption capacity of dried, harvested peats is not as large as that of wet, living peats. 6.4.2
Oxidation and Reduction Reactions
Wetlands are ideal environments for chemical transformations because of the range of oxidation states that naturally occur in wetland soils. Free oxygen decreases rapidly with depth in most flooded soils because of the metabolism of microbes that consume organic matter in the soil and through chemical oxidation of reduced substances. This decline in free oxygen, in other words the depletion of oxidizing potential, is measured as an increasingly negative electric potential between a standard platinum electrode and the concentration of oxygen in the soil. This measure of electric potential is called reduction-oxidation or REDOX potential (ORP) and provides an estimate of soil oxidation or reduction potential (Figure 6.9).
Figure 6.9
Typical depth profile for potential oxidation-reduction reactions taking place in a treatment wetland system.
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When ORP >100 mV, conditions are termed aerobic because dissolved oxygen is available. When ORP <–100 mV, conditions are termed anaerobic because there is no dissolved oxygen. Some authors refer to intermediate conditions (near-zero DO) as transient or anoxic. The reduction of ferric iron (Fe3+) to ferrous iron (Fe2+) is a REDOX half reaction typical of anaerobic sediments: Fe3+ + e– = Fe2+
(6.2)
The typical complementary half reaction for this reduction of ferric to ferrous iron is the oxidation of reduced sulfur (hydrogen sulfide, H2S) to sulfate (SO42– ) by the half reaction: H2S + 4H2O = SO42– + 10H+ + 8e–
(6.3)
for an overall oxidation-reduction reaction of: 8Fe3+ + H2S + 4H2O = 8Fe2+ + SO42– + 10H+
(6.4)
However, the real situation in treatment wetlands is far more complicated because there are more species present and more REDOX routes than those used here for illustration. For instance, iron reduction can be accompanied by sulfate reduction — not sulfate production — and the ferrous ion, Fe (II), will be precipitated out as FeS. The REDOX potential of many wetland soils decreases with vertical depth into the sediments because the only source of free oxygen is from atmospheric diffusion at the top of the sediment layer (Figure 6.9). The typical oxygen gradient in wetland sediments includes a thin (less than a few centimeters) oxidized surface horizon at the sediment-water interface, underlain by increasingly reduced conditions with depth based on the amount of biological and chemical reducing activity in the sediments. Vertical REDOX gradients in treatment wetland soils will vary in response to distance from the point of wastewater loading. Northern wetlands are sometimes sealed by an ice cap in winter that prevents supply of oxygen to the water and/or soils. This then shifts the REDOX profile to much lower Eh values, causing sulfate reduction and methanogenesis to dominate even the upper soil horizons. Gaseous sulfur compounds cannot escape and may reach lethal levels for wetland biota. Treatment wetlands are often subjected to waters with higher oxygen demands exerted by both carbonaceous and nitrogenous compounds. This causes a greater depletion of electron acceptors such as oxygen, nitrate, sulfate, and iron in both the water column and the underlying soils. The REDOX potential in treatment wetlands is therefore typically lower than for natural wetlands, ranging from the denitrifying regime downward to the methanogenesis regime.
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pH
Prior to flooding, soils may have widely varying pH conditions ranging from about 3 to 10 units. Following flooding, pH in wetland soils may initially decline due to aerobic decomposition liberating carbon dioxide into the interstitial water. However, this initial pH swing is generally transient and is followed by a typical trend in both acidic and alkaline soils toward pH neutrality (pH 6.7 to 7.2 units) over time. This typical trend is considered to be the result of ferric iron reduction under flooded soil conditions. In some highly organic soils, pH may remain very low, even following long periods of flooding. This result is likely due to the slow oxidation of organic sulfur compounds resulting in production of sulfuric acid and to the presence of humic acids. 6.4.4
Biological Influences on Hydric Soils
Biological, chemical, and physical properties of wetland soils are interdependent. Microbial wetland fauna make up a significant fraction of the organic carbon occurring in hydric soils. These tiny organisms are competing for sometimes limited and rapidly shifting supplies of energy containing compounds and nutrients, and their growth and death have a very significant effect on the fate and transport of the majority of soil chemical constituents. In addition to the microbial populations, macrophytic plants diversify soil structure through the growth and death of roots and creation of decaying plant litter, and wetland animals dig, burrow, scrape, and otherwise cause bioturbation of wetland sediments on an almost continuous basis. Some of the major interactions between wetlands biology and sediments are described below. 6.4.5
Microbial Soil Processes
Soil microbial populations have significant influence on the chemistry of most wetland soils. Important transformations of nitrogen, iron, sulfur, and carbon result from microbial processes. These microbial processes are typically affected by the concentrations of reactants as well as the REDOX potential and pH of the soil. Organic nitrogen is biologically transformed to ammonia nitrogen through the process of mineralization, which results as a consequence of organic matter decomposition that results from actions of both aerobic and anaerobic microbes. Ammonia is in turn converted to nitrite and nitrate nitrogen through an aerobic microbial process called nitrification. Nitrate nitrogen can be further transformed to nitrous oxide or nitrogen gas in anaerobic wetland soils by the action of another group of microbes (denitrifiers). Nitrogen gas can also be transformed to organic nitrogen by bacterial nitrogen fixation in some aerobic and some anaerobic wetland soils. Bacteria can transform reduced iron and possibly manganese to oxidized forms. These chemosynthetic processes utilize oxygen as an electron acceptor and usually are accelerated by acidic conditions typical of acidic coal mine drainage waters. Sulfate can be reduced to sulfide by anaerobic bacteria in wetlands. The sulfate serves as an electron acceptor in the absence of free oxygen at low REDOX
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potentials. Sulfides can provide a source of energy for chemautotrophic and photosynthetic bacteria in aerobic wetlands, resulting in the formation of elemental sulfur and sulfate. The formation of sulfide under anaerobic conditions enables the precipitation of dissolved heavy metals in treatment wetlands. Organic carbon is microbially degraded to carbon dioxide by aerobic respiration when oxygen is available and by fermentation under anaerobic conditions. Greater energy is released under aerobic respiration, resulting in more efficient assimilation of organic matter into microbial cellular material. In fermentation, organic matter serves as the terminal electron acceptor, forming acids and alcohols. Methane can be formed in wetlands due to the action of bacteria using carbon dioxide as an electron acceptor at very low REDOX potentials. 6.4.6
Treatment Wetland Soils
The sediments that form in treatment wetlands are often different from those that form in natural wetlands, for a number of reasons. First, the enhanced activity of various microbes, fungi, algae, and soft-bodied invertebrates leads to a greater proportion of fine detritus compared to leaf, root, and stem fragments. There is significant formation of low-density biosolids (sludge). Second, there may be a precipitation of metal hydroxides, carbonates, or sulfides, which add mineral flocs to the sediments. Finally, there is often a high ionic strength associated with treatment of effluents reflected in high dissolved salt content. The effect of high ionic strength is to alter the structure of the highly hydrated organic materials comprising wetland sediments and soils. The considerations listed previously show that wetland soils are a vital and integral link in processes that govern water quality. Therefore, it seems reasonable to expect consideration of wetland soils to be an important part of treatment wetland design; however, that is not the case. Antecedent soils are altered and replaced by new organic soils. The sorption capacity of the antecedent soils is re-equilibrated with the new water quality of the incoming water, perhaps along a gradient from inlet to outlet. If there are leachable chemicals, they are depleted and exit the wetland. Roots and rhizomes of new plantings (constructed wetlands) or replacement species (natural wetlands) repopulate the top 30 cm of the wetland and set up a new biogeochemical cycle. The long hydroperiods of treatment wetlands are conducive to the buildup of organics: first litter and microdetritus, then the sediments formed from their decomposition, and, finally, the organic soils generated from those sediments and deposited mineral solids. In short, the wetland rearranges itself to accommodate the environment created by the designer. The functioning of the wetland after such adaptation is no longer dependent upon the previous condition and type of soils, hydrology, and biota; it is now totally dependent upon the new soils, hydrology, and biota. It is this new sustainable mode of wetland operation that is the target of most designs. Available data indicate that the final state of a treatment wetland, and the accompanying suite of water quality functions, are largely independent of the initial condition of the real estate upon which it is built. During the interim period of
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adaptation, antecedent conditions are important because they dictate the short-term performance of the wetland. That period of adaptation appears to extend for up to 2 years for newly constructed wetlands and longer for alteration of natural wetlands to a treatment function.1,2 If soils are unsatisfactory for plant or microbial growth, the wetland treatment system is liable to have inadequate plant cover. A knowledge of the physical and chemical composition of site soils is essential to predict accurately some internal chemical and biological processes in treatment wetlands. The rate of soil accretion in wetland treatment systems affects the potential removal of conservative elements such as phosphorus and heavy metals and also is an important consideration during design of berm height above the wetland substrate. Some sorbed materials can be released if exposed to lower concentrations in incoming water. Violent hydrologic events are capable of resuspending particulate deposits. It is therefore incumbent on the designer to minimize or prevent such occurrences. The main environmental factor that influences the nature of wetlands soils is dissolved oxygen concentration. Vertical oxygen gradients are typically established in wetland soils due to bacterial respiration and chemical oxidation demand and due to the greatly reduced rate of oxygen diffusion in saturated soils compared to unsaturated soils. These oxidation gradients result in a chain of oxidation-reduction reactions that provide many wetlands with their typical profile of declining REDOX potentials with depth (Figures 6.10a and b). REDOX, in turn, affects the microbial processes important in most aspects of wetland use for water quality improvement, especially including removal of organic carbon and nitrogen. Wetland soils are as dynamic in character as all other aspects of the wetland ecosystem. Soils in constructed wetlands built on upland sites undergo gradual transformation, resulting in accumulations of organic carbon and reduced elements such as iron and sulfur typical of natural wetland soils. Many of the changes that occur during wetland development and succession are the result of biological factors that occur in wetlands such as growth of bacteria and fungi, algae and macrophytic plants, micro- and macroinvertebrates, and larger animals. While many of these natural biological processes are not within the control of the wetland treatment system designer and operator, their effects should be considered when trying to maximize chances for success with a wetland project.
6.5 6.5.1
CONTAMINANT REMOVAL MECHANISMS
Volatilization
A shallow wetland water body provides the opportunity for air stripping of volatiles. The efficiency is not as great as for mechanical devices, but the difference is more than compensated for by the long detention times and large surface areas in the wetland. Half-lives (time to volatilize one half the substance) range from 2–4 days for compounds such as benzene, toluene, and naphthalene, and are projected to be less for more volatile compounds, such as vinyl chloride and chloromethane.9,10
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Figure 6.10a
295
Idealized macrophyte root structure in a wetland soil illustrating development of an oxidized rhizophere (adapted from Kadlec and Knight, 1996).
More soluble, less volatile organics are less likely to volatilize, but then have an opportunity for enhanced biodegradation to occur. As a result, half-lives for substances such as phenol, tetrahydrofuran, and methanol are also fairly short; these have been measured to be in the range of 10 to 40 hours in wetlands.11 The half-lives of low molecular weight alkanes (
Partitioning and Storage
Many organics are known to sorb strongly to the organic sediments and substrates in wetlands. In addition, they may be complexed with dissolved organic matter
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Root
CH 4 N2
Organic C
O2 CH 4 Root Aerobic NH4+ NO3
Organic N NH4+
NO3-
Anaerobic Figure 6.10b
N2
Pathways of nitrogen transformations in the immediate vicinity of a wetland plant root (adapted from Reddy and Graetz, 1988).
(DOM). For example, the partition coefficient of hexachlorobenzene (HCB) to wetland sediments has been investigated and it has been determined that the association of dosed HCB with DOM is a fast reaction that equilibrates in one hour.9 The role of wetland vegetation in buffering air emissions from waste sites and the role of partitioning between the plant–air interface and resulting impact cycling of contaminants have also been investigated.13 The impact of the quality of organic matter on the magnitude of desorption–resistance has been studied. Most studies indicate the presence of a sizeable desorption–resistant-phase in wetland soils and an increase in the size of the desorption–resistance in “older” organic matter. Volatile substances are gasified. Many materials undergo microbial transformations. These processes all lead to the transformation and transfer of a “removed” pollutant either to the atmosphere or to the wetland sediments and soils. The vegetation is extremely important for nutrient transformations and transfers because it plays a key role in the cycling and temporary storage of many substances.
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6.5.3
297
Hydraulic Retention Time
Removals proceed over the time water is held in the wetland. Therefore, detention time (or the equivalent hydraulic loading rate) becomes the key design variable. The longer the water is held in the wetland, the better the treatment — so long as it is not at the expense of added depth, which contributes little to the active transfers and storages. However, wetlands possess irreducible background concentrations of some substances: about 5 mg/L of BOD and TSS, for instance. Typical detention times range from one to ten days for existing treatment wetlands, which corresponds to hydraulic loading rates (rainfall equivalent) of 1–10 cm/day. This technology requires land instead of mechanical devices to accomplish treatment. If the necessary land is available, it typically offers capital savings over competitive processes. The passive nature of wetlands technology typically offers a very large advantage in operating costs because operation is simple and maintenance is very low. It has been assumed by some people that wetlands are easily conceived and built and are low technology devices; therefore, they should be easily described in terms of simple equations. In fact, they are exceedingly complex “ecoreactors” that require complex descriptions if they are to meet designed expectations. A design model must first do an adequate job of predicting wetland hydraulilcs. The basic tool is the interior water budget. Stream inflows and outflows are typically measurable and controllable in design. The wetland often will not communicate with groundwater due to existing or constructed clay seals or liners. The uncontrollable elements of the water budget are precipitation and evapotranspiration. The fundamental hydraulics question for treatment wetlands is conveyance capacity and the related issues of depths and slopes. These issues are answered by the interior water budget, together with data on the flow resistance of the wetland. FSF and SSF wetlands are a bit different with respect to frictional characteristics and should be treated separately. The hydraulic sizing of the wetland is then set so that operation will occur within the desired parameters of depth and flow.1 Other tools necessary to determine the size of the treatment wetland are the parameters required to achieve the biogeochemical and physical processes for contaminant mass reduction. These are a description of internal flow paths, the interior chemical mass balance, and the reaction rate constants. Detailed description of the designed process of a treatment wetland system is out of the scope of this chapter and book, but a few lessons from expert practitioners are described below.1,2 A pictorial description of the energy balance in a wetland system is shown in Figure 6.11. In addition, Figure 6.12 portrays mass balance objectives; meeting this objective is the primary function of a treatment wetland system. For the sake of brevity, only parameters of importance for the design are listed below:1 • Evapotranspiration, water losses, and gains • Area requirements, number of cells, and cell shape • Overland flow and geohydrology
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Incoming Sunlight (Short Wave)
Clouds Absorb and Reflect Surface Reflects
Evapotranspiration
Evaporation
Transpiration
Radiative Heat Loss (Long Wave)
Convective Transfer to/from Air Outgoing Water Heat Content
Incoming Water Heat Content Change in Storage
Conductive Transfer to/from Ground
Figure 6.11
Components of the wetland energy balance (adapted from Kadlec and Knight, 1996).
Volumetric Flow = Qi Concentration = Ci Mass Inflow = Qi Ci Surface Area = A
Volumetric Outflow = Q e Concentration = Ce Mass Outflow = Qe Ce
Mass Removal
Figure 6.12
Components of a chemical budget and associated terminology (adapted from Kadlec and Knight, 1996).
• Water depth, subsurface wetland hydraulics, bed friction, hydraulic conductivity and changes with deposition, and clogging • Surface water elevation profiles, dynamic responses with precipitation, shadow and dead zones • Nonideal flow patterns, vertical and transverse • Mixing flow velocity and residence time • Biochemical reaction rates, residence time variations based on flow paths, substrate additions to achieve desired treatment reactions, plug flow vs. CSTR reactor models • Mass transfer mechanisms • Temperature, dissolved oxygen, and pH profiles • Suspended solids
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• • • •
6.6
299
Nutrients demand and cycling Flood protection Earth moving, dikes, berms, and levees Plant selection and planting, which includes factors such as geographic location, physiographic features, regional hydrology, climatic conditions, soil characteristics, water treatment objectives, type and intensity of adjacent land use impacts, wildlife, and aesthetic preferences
TREATMENT WETLANDS FOR GROUNDWATER REMEDIATION
Utilization of wetlands for the purpose of remediating contaminated groundwaters is a relatively new concept. The evolution of ideas has been from an independent point of view, not based upon other pre-existing aspects of the technology. Constructed wetlands in this context could be a subset of options under the general heading of phytoremediation in the previous chapter. Several processes are envisioned as effective in pollutant reduction: phytoextraction, phytostabilization, transpiration, stripping, and rhizofiltration. Phytoextraction refers to plant uptake of toxicants, which is known to occur and has been studied in the stormwater and mine water treatment wetland context. However, in many cases the contaminant is selectively bound up in below ground tissues, roots, and rhizomes, and is not readily harvested. Phytostabilization refers to the use of plants as a physical means of holding soils and treated matrices in place. This process is also one of the chief underpinnings of treatment wetlands as it relates to sediment trapping and erosion prevention in those systems. Wetland plants possess the ability to transfer significant quantities of gases to and from their root zone and the atmosphere (Figure 6.10). This ability is part of their adaptation required for survival in flooded environments. Stems and leaves of wetland plants contain airways (aerenchyma) that transport oxygen to the roots, and vent water vapor, methane, and carbon dioxide to the atmosphere. There may also be transport of other gaseous constituents, such as dinitrogen and nitrogen oxides, and volatile hydrocarbons. The dominant gas outflow is water vapor, creating a transpiration flux upward through the plant. Rhizofiltration refers to a set of processes that occur in the root zone, resulting in the transformation and immobilization of some contaminants. Wetlands are vertically stratified with respect to REDOX potential and other important chemical attributes (Figure 6.9). These vertical gradients provide for interzonal processes with both oxic and anoxic character. The microenvironments in the root zone of wetland plants also have a great deal of structure, leading to extremely different biogeochemical conditions in zones in close proximity. Typically, oxidized conditions are found adjacent to roots, while anaerobic conditions may exist only millimeters away. This varied environment makes possible diffusional transfers between chemically different regions. A prime example is the precipitation of dissolved heavy metal ions as sulfides in the anaerobic zones of wetlands. Despite this wealth of scientific knowledge about individual contaminant removal processes, there have been few attempts to implement wetlands for groundwater
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remediation. Therefore, the potential of the technology must currently be assessed from other applications. The next sections briefly address the state of knowledge in closely related applications. 6.6.1
Metals-Laden Water Treatment
A very large application area for constructed wetlands is the treatment of acid coal mine drainage. Hundreds of wetlands are now in operation serving this function. The contaminants of interest are typically pH, iron, and manganese. Despite the large number of such wetlands, there is not yet a clearly stated design methodology available for acid mine drainage treatment. Individual metals have been targeted at mining sites that involve them. For example, a good deal is known about wetland treatment of copper,14 aluminum,15 and zinc. There are other reports on removal efficiencies of a number of wetlands receiving a variety of mine effluents. Moderately high efficiencies, usually in the 60–90% range are reported from full and pilot scale projects.16,17 Many treatment wetland field studies have investigated removal efficiencies for multimetal wastewaters, mostly at low to moderate concentrations in domestic/industrial combinations and urban stormwaters.18-21 Moderately high efficiencies, usually in the 60–90% range are also reported in these studies. Laboratory and mesocosm scale studies bear out these results under more controlled, but less realistic conditions; for example, fast and high reductions in copper and chromium (VI) in mesocosms.11 Metals in wastewater and contaminated groundwater must be removed prior to final discharge to protect the environment form toxic effects, but the use of wetlands to accomplish this goal must be examined cautiously. The potential for economical treatment is nevertheless quite attractive, and the concept of metal removal in wetlands has undergone considerable investigation. Metals are removed by cation exchange to wetland sediments, precipitation as sulfides, carbonates, and other insoluble salts, and plant uptake. The anaerobic sediments provide sulfate reduction to sulfide and facilitate chemical precipitation. As a result, good removals of metals are reported for operating wetland facilities. Removal of heavy metals such as Ni, Nz, Cu, Fe, Mn, Co, and other metals has been reported at varying removal efficiencies. Under optimum operating conditions, some of these metals have been removed at efficiencies of up to 96–98%. At most of these treatment wetland systems designed for heavy metal removal, a majority of the mass reduction has taken place at deeper depth, indicating that anaerobic conditions are a must for the precipitation mechanisms. Researchers have shown that spent mushroom or other organic substrates enhance the formation of anaerobic conditions in both SSF and FWS systems.1,2 A pretreatment system with limestone and peat mixture beds has been used at a few treatment wetland systems for the treatment of acid mine drainage or coal stockpile drainage waters. Limestone is used to neutralize the pH of the input drainage water and provide for increased alkalinity and precipitations as metal hydroxides. It has been shown that these drainage waters (with a pH of 2–3) enter the actual treatment zone with a pH greater than 6.0 from the limestone pretreatment beds (Figures 6.13 and 6.14).
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Limestone Bed Flow Control Culvert
Peat Mixture Open Water Pool
Open Water Pool
Open Water Pool
Water Flow Direction Limestone Bed Peat Mixture Treatment Open Water Pool
Intermittent Creek Present Extent of Wetland Control Culvert Berm
Figure 6.13
Typical peat/wetland treatment system, designed for heavy metal removal from stockpile drainage (adapted from Moshiri, 1993).
Limestone Drain Deep Pond
Aeration
+Alkalinity Fe Removal Mn Coprecipitation TSS Removal
Figure 6.14
Deep Marsh Fe Removal
Shallow Marsh Mn Removal
Rock Filter Mn Removal
Alkaline Polishing Cell Discharge pH Increase Stormwater Control Chemicals TSS Removal Compliance
General schematic of staged aerobic constructed wetlands.
The peat mixture beds are placed downstream of the limestone, with the flow dispersed laterally to increase the contact area and encourage subsurface flow (SSF). The subsurface flow through the peat mixture encourages the development of anaerobic conditions for sulfate reduction activities. In addition, the dynamics of the physical/chemical and biological components of peat/wetland systems utilize other metal removal mechanisms such as adsorption, cation exchange, and complexation of soil organic matter (SOM). Some practitioners have added spent mushrooms with peat to increase the SOM. Removal efficiencies of around 95–98% have been reported for similar systems. It is important to pay attention to the selection of the
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right type of peat for these systems. The most important factors are high hydraulic conductivity and available organic content. Similar removals are obtained for other metals; for instance, chromium is reduced by more than 70% in about 70 hours in surface flow wetlands.1,2 The area design approach selects the wetland area required to remove a given amount of metal. Recommended areas for iron and manganese reduction are in the range of 100–500 square meters per kilogram of metal removed per day. Longer detention times produce even greater reductions. Less is known about some elements and compounds, such as boron, arsenic, cyanide, and selenium. However, enough is known that the correct conditions for removal may be intentionally designed into the wetland.22 In summary, metal removal in wetlands stems from a variety of biogeochemical processes, including aerobic and anaerobic processes in the water column, on the surface of living and decaying plants, and in sediments. The most significant mechanisms include: • • • • • • •
Sorption and/or exchange onto organic matter (detritus) Filtration of solids and colloids Formation of insoluble metal sulfides Formation of carbonates Association with iron and manganese oxides Metal hydrolysis (catalyzed by bacteria under acidic conditions) Reduction to nonsoluble, nonmobile forms (also catalyzed by bacteria)
These processes are involved to various degrees depending on specific circumstances, and it is critical to identify their relative importance in specific treatment wetlands. Identifying and quantifying these processes provides a basis for the rational design of treatment wetlands and informs on the stability and biological availability of the contaminants they retain. 6.6.1.1 A Case Study for Metals Removal* The site is a former fibers manufacturing site located in the coastal plain of Virginia. A technical solution was needed to reduce zinc loadings to meet NPDES or Publicly Owned Treatment Works (POTW) zinc limits for discharge from the impacted area. After a feasibility study and pilot test, a constructed treatment wetlands was retrofitted into an impoundment area holding zinc containing sediments. The area also receives a discharge of zinc and iron laden landfill leachate from an adjacent industrial 36-acre landfill. A passive treatment system was preferred to a mechanical treatment system as a cost-effective and long-term solution to rehabilitate the impacted area and provide treatment of the water in the impounded area. At the start of the restoration project, the wetlands contained no vegetation as a result of the acidic water. Early efforts included pH adjustment by adding limestone gravel and slurry to raise pH. Attempts were made to replant the area with several * Courtesy of Joseph McKeon, BASF Corporation, Remediation Manager, Ecology and Safety Department.
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species to determine which would survive. Phragmites australis volunteered and populated the entire wetlands, crowding out most of the replanted species. The Phragmites community continues to thrive and has spread to areas with hospitable moisture conditions, as well as repopulating areas where vegetation was removed for construction. The wetlands was designed as a two stage aerobic/anaerobic treatment system to remove the iron in an oxic environment and the zinc as sulfide in an anaerobic environment. A partitioning dike was built across the wetlands to separate the aerobic one third of the 16.2-acre wetlands. The characteristics of the leachate coming from the adjacent landfill approximates high strength acid mine drainage with an acidic pH and high iron concentrations. The leachate is passively intercepted in a series of anoxic limestone drains (ALDs), which are mitigation chambers installed below the water table. Each ALD contains a specified volume of high calcium carbonate limestone that dissolves upon contact with the acidic leachate, thus raising the pH and increasing the alkalinity. At the point of ALD discharge into the wetlands, the pH is about 7.0 and contains both iron (II) and zinc. The iron oxidizes to Iron (III) Oxide and precipitates to the substrate of the wetlands. Some zinc coprecipitates with the iron and some is transported in solution to the anaerobic treatment component of the wetlands. Sulfate naturally occurring in the leachate and supplementally added in the compost cell is reduced to hydrogen sulfide and forms a relatively insoluble and nonbiologically available precipitate of zinc sulfide. Spent mushroom compost is used to drive the wetlands anaerobic at the aerobic/anaerobic boundary and to add alkalinity and sulfate to the water. One of the principle components of the compost is manure which serves as a carbon (food) source for the biologically mediated reduction process. Limestone in the compost adds alkalinity, providing additional buffering capacity to stabilize pH. Other nutrients in the compost contribute to the health of the vegetation in the wetlands. A compost cell is also installed in the area receiving the largest amount of leachate and zinc loadings to further condition the water before it enters the anaerobic treatment area. Attenuation of zinc is dependent on the reduction of sulfate to produce hydrogen sulfide. The rhizosphere of the Phragmites provides the substrate to support the sulfate reducing bacteria. Phragmites also transpires large volumes of water during the warm months to control water levels in the impounded wetlands area. This species of plant is not an important part of the food chain so uptake of metals is not an issue for managing the passive treatment system. Phragmites does provide habitat for insects, reptiles, and amphibians and is used by deer as a bedding area. Zinc concentrations as high as 400 mg/l are introduced into the wetlands in the landfill leachate and are reduced to as low as 0.3 mg/l enabling discharge to the local POTW without further treatment. Treatment efficiency is higher in the warmer months when biological activity is at maximum and lower during the cold months. Water level management helps to minimize discharge during the colder months, thus enabling more contact time with the treatment area and thus assisting zinc attenuation. The treatment system was designed with excess treatment capacity to manage other zinc containing streams, if necessary. Various sections of the treatment wetland system is shown in Figures 6.15a-d.
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Figure 6.15a
Headwaters of the wetlands. (Photo courtesy of Joseph McKeon.)
Figure 6.15b
The aerobic or oxic area. (Photo courtesy of Joseph McKeon.)
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Figure 6.15c
Spent mushroom compost at the beginning of the anoxic or anaerobic area. (Photo courtesy of Joseph McKeon.)
Figure 6.15d
Vegetation in the anaerobic area. (Photo courtesy of Joseph McKeon.)
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Removal of Toxic Organics
Organic contaminants partition strongly to solid organic substrates. Wetland sediments are therefore excellent sinks for organics. Subsequent to such partitioning, the organic chemical may diffuse downward, or undergo biodegradation. The wetland environment is complicated by the existence of biofilms on submerged plant parts and litter. Although small in terms of mass per unit volume, these biofilms are very active in biodegradation, and consequently serve as important sinks for organics.23 In this aspect, wetlands resemble conventional attached growth bio-treatment processes. The number of toxic organics that can potentially create environmental problems is very large. Not all of these have been studied with respect to wetland technology. Those that have been investigated show removals due to volatilization, storage, and biological degradation in the wetland environment. 6.6.2.1 Biodegradation Anthropogenic organic chemical additions to wetlands for treatments are modifications to an exceedingly complex background array of hydrocarbons and organic chemical reactions. Hydrocarbon molecules are susceptible to fragmentation and chemical conversion in the wetland environment, predominantly via microbially mediated pathways. Partial conversion may occur via hydrolysis, dealkylation and ring cleavage, or the removal of amino, nitro, chlorine, hydroxyl, acid, or thio groups from the parent molecule. Oxidative processes ultimately produce carbon dioxide and water, while anaerobic processes will enhance reductive dechlorination and may terminally result in the evolution of methane. It is therefore important to identify the byproducts of degradation. Trichloroethene (TCE) and 1,1,2,2-Tetrachloroethane (PCA), at concentration ranges of 100 to 1000 ppb, were naturally attenuated when contaminated groundwater from an aerobic, sandy aquifer discharged through anaerobic sediments in a freshwater wetland at a U.S. army base in Maryland.24 Microcosm experiments confirmed field observations that Cis 1,2-DCE, and VC are the dominant daughter products from anaerobic biodegradation of both TCE and PCA in the wetland sediments. Cis 1,2-DCE was produced by reductive dechlorination of TCE and dihaloelimination of PCA. VC was produced by reductive dechlorination of Cis 1,2DCE and dihalaelimination of the 1,1,2-TCA that was produced through reductive dechlorination of PCA. Parent and daughter concentrations in the microcosms decreased to less than 5 ppb in less than 35 days, showing extremely rapid degradation rates in these organic rich sediments. Wetlands have all the basic elements needed for the attenuation of chlorinated alkenes and alkanes, which include high organic carbon content in the sediments to bind the contaminants; high microbial density and diversity in the sediments to biodegrade contaminants; and both anaerobic and aerobic conditions to ensure that contaminants can be fully degraded without the accumulation of potentially toxic intermediates such as VC. Treatment wetlands can be designed and constructed specifically to enhance the natural physical and biochemical processes to degrade chlorinated organic compounds (Figure 6.16). Some basic steps to be taken into
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Biofilm formation
Figure 6.16
Biofilms dominate the sediment–water interface and the surfaces of litter and standing dead material.
consideration on such a project include site suitability assessment, risk management, modeling and conceptual design;,permitting, design, construction, and operation and maintenance. Though anaerobic environments, such as high organic carbon wetland sediments, are often sufficiently reducing to make reductive dechlorination reactions thermodynamically favorable, the transfer of electrons from reduced species to a chlorinated solvent is often kinetically constrained. The reduction rate may be enhanced in the presence of compounds capable of facilitating transfer of electrons from the bulk reductant to the chlorinated solvent of interest. Nickel and copper complexes with Aldrich humic acid have recently been shown to be effective electron mediators for the reductive dechlorination of TCE.25 Porewaters from the high organic carbon sediment zone of the West Branch Canal Creek Wetland at Aberdeen Proving Grounds, Maryland, were used. Using titanium (III) citrate as the bulk reductant, Ni and Cu complexes (of DOC) present at the site sediments rapidly reduced TCE. The reaction was pseudo-first order with a half-life of approximately one hour for both DOC-metal complexes. Reaction rates were comparable to the Aldrich humic acid-transition metal systems. Dechlorination was complete with the dominant products being ethene and ethane, the mass balance was near 100%, and chlorinated intermediates were either absent or at extremely low concentrations. The partitioning of compounds from the aqueous phase into biota is not the only significant process that occurs after the initial discharge of xenobiotics into wetland systems. Partitioning of xenobiotics from the aqueous phase into the sediment phase may be of equal significance, attested to by the structural range of organic compounds (PAHs, PCBs, PCDDs, Chloro Phenols, etc.) recovered from contaminated sediments. It is well established that many heavy molecular weight and high Kd compounds after introduction into the wetlands environment are not readily accessible to chemical recovery. This does not necessarily imply, however, that they are of no environmental significance: the degree to which they are desorbed and therefore become accessible to biota is a central issue that has implicaitons both for toxicity of
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xenobiotics and for their resistance to microbial attack. Results that indicate decreasing recoverability and irreversible sorption from the sediment phase with increasing time from deposition may be accommodated under the general description of aging.26 Details of the mechanisms by which xenobiotics are “bound” to components of the sediment phase have not been fully established, although several plausible hypotheses have been put forward. Proposed mechanisms of interaction include ionic and covalent binding, long range (Van der Waals) forces, or sorption by undefined mechanisms. A recent study27 indicated 30–62% of chlorobenzene and 47–83% of phenathrene of the adsorbed mass residing in the desorption-resistant fraction. In another recent study partitioning of hexachlorobenzene (HCB), lower chlorinated benzenes, and hexachlorbutadiene (HCBD) between the plant–air interface was measured in field and laboratory studies using bark, leaf, and leaf litter of bald cypress (Taxodium distichum) and black willow (Salix nigra).28 A first order kinetic model was applied to the experimental results to determine a plant–air partition coefficient (KPA). The role of wetland vegetation in buffering air emissions from waste sites was established. With high microbial populations compared to terrestial environments, wetland soils are highly effective at consuming and processing labile, complex organics. The supply of low molecular weight organic substrates from decomposition of organic matter for dechlorination in organic rich wetland sediments is important to understanding the potential of using sediment microorganisms for natural degradation of chlorinated organic compounds. A recent study29 indicated that 2,3,5,6-Tetrachloro Phenol can be dechlorinated in wetland sediments and that 2,4,5-Trichloro Phenol, and 2,5-and 3,4-dichlorophenols are major intermediates before the eventual dechlorination to 3-chlorophenol is achieved. In another study30 intrinsic biodegradation of PAHs was evaluated using two microbiological analyses: heterotrophic bacterial production using 3H-Leucine incorporation to estimate the growth rate of the entire community and PAH mineralization rates using degradation of 14C-naphthalene, phenanthrene, and fluoranthene to 14CO2. Bacterial growth was generally reduced at locations that had PAH concentrations above 100 ppm. However, PAH mineralization rates were significant across the site and PAH mineralization accounted for up to 15% of the heterotrophic bacterial growth substrate demand. Additional investigations revealed that complete mineralization of hexachlorobenzene (HCB) is possible in heavily anaerobic, methanogenic wetland sediment environments.31 HCB was reductively dechlorinated to isomes of dichlorobenzene (DCB) and monochlorobenzene. After further reductive dechlorination of DCB, all the monochlorobenzene was degraded to CO2 under methangenic conditions. This study also found that uncomplexed Cd and Pb in the sediment porewater was inhibitory to the reductive dechlorination reactions. Research focused on the transformations of alachlor and atrazine (members of the chloroacetamide and chloro-s-triazine herbicide classes, respectively)32 revealed that degradation is promoted by naturally occurring inorganic sulfur nucleophiles in anoxic salt marsh porewaters. Bisulfide (HS) and polysulfides (Sn2– ) have been reported in salt marsh porewaters as high as 5.5 mm and 0.43 mm, respectively. The rates of sulfhydrolysis in these environments may greatly exceed rates of other
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removal processes (biotic or aerobiotic) for the above compounds and may represent an important sink for these herbicides in salt marsh environments. Studies have examined the relationship between irreversible sorption (aging) and bioavailability in the vegetated wetland environment using plant uptake as an endpoint. In a recent study using black willow (Salix nigra), and Scirpus olney, it was found that plant uptake of about 4–6% of the phenanthrene was observed in sediment with phenanthrene in a fully desorption resistant state.33 Both plants appeared to access the desorption resistant phase to some extent. Two mechanisms appear to be important: the direct uptake of the porewater in equilibrium with the desorption resistant phase, and the sorption of the organic onto the root followed by uptake. Other studies have reported that phenolic compounds at high concentrations can be remediated in treatment wetlands using plants such as Schoenoplectus and Phragmites. Trinitrotoluene (TNT), cyclotrimethaylenetrinitramine (RDX), and cyclotetramethylenetetranitramine (HMX) at concentrations of 4, 4, and 0.10 ppm, respectively, were treated in a treatment wetland planted with canary grass, woolgrass, sweet flag, parrot feather, sago pond weed, water stargrass, and elodea. The treatment efficiencies were greater than 95%.34 6.6.3
Removal of Inorganics
Reduction of phosphorus and nitrogen compounds requires the longest detention of any pollutants. Approximately 90% of the incoming inorganic load can be eliminated by 14 days detention in a surface flow wetland.1 These substances are required, at low loading rates, to sustain a healthy wetland. Nitrogen removal may require active or passive reaeration to promote nitrification; denitrification requires a carbon source to fuel the denitrifer population. In a recent study using native wetland plants such as bulrushes (Scirus spp.), cattails (Typha spp.), and sedges (Carex spp.), an influent water containing 62 mg/L of NO3– was treated to consistently nondetectable levels of effluent quality.7 This test is continuing to see whether perchlorate also can be treated by the same system. It is not possible to forecast the types of substances that will be contained in leachates over the entire history or anticipated discharge. Treatment wetlands have capabilities for removing a wide variety of substances, and hence operate as “broad spectrum” treatment technology. They have the further property of pulse averaging because of long detention times. A brief “spike” of a given substance will, at a minimum, be diluted by the relatively large volume of water in the wetland. Averaging is approximately over the detention time of the wetland. 6.6.4
Wetland Morphology, Hydrology, and Landscape Position
Wetlands are often constructed to remove pollutants from ground or surface waters. The intention is that constructed wetlands will degrade or sequester pollutants so they do not pose a risk to humans and wildlife downstream of pollutant sources. Pollutants retained in constructed wetlands may be converted to less toxic forms or sequestered in the sediments or biota of wetlands. Ultimately, pollutants may be
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removed from wetlands by harvesting soils and biota and transferring them to longterm storage sites such as landfills. While wetlands may remove pollutants from ground and surface waters, pollutants may leave wetlands through aquatic-terrestrial food chain links. Many birds and mammals feed on larger aquatic invertebrates and vertebrates that may accumulate significant amounts of pollutants. Thus, factors that promote the accumulation of pollutants in larger aquatic organisms can lead to risks to wildlife and, ultimately, human health. By managing the factors that promote accumulation of pollutants in organisms inhabiting constructed wetlands, a manager can reduce the ecological risk these prey items represent to terrestrial organisms. The relationships among wetland landscape position, hydrology, morphology, and toxic accumulation in fish in natural depression wetlands can be manipulated in constructed wetlands to reduce risk to wildlife from accumulated pollutants. The distribution of wetlands in the landscape and the amount of time they hold water during the annual hydrological cycle determines the amount of time fish will be present; wetlands close to other aquatic habitats and holding water for much of the year are more likely to have fish populations. Fish in wetlands with shallow maximum depth and greater water level fluctuations accumulate significantly more toxins. These relationships suggest: 1) constructed wetlands should be located in the landscape such that the chance of colonization by fish is minimized; 2) when possible, stable water levels should be maintained in constructed wetlands; and 3) wetlands should be constructed with steep sides and flat bottoms.
REFERENCES 1. Kadlec, R. H. and R. L. Knight, Treatment Wetlands, CRC Press, Boca Raton, FL, 1996. 2. Moshiri, G. A., Constructed Wetlands for Water Quality Management, Lewis Publishers, Chelsea, MI, 1993. 3. Birkbeck, A. E., D. Reil, and R. Hunter, Application of natural and engineered wetlands for treatment of low-strength leachate, in Constructed Wetlands in Water Pollution Control, P. F. Cooper and B. C. Findlater, Eds., Pergamon Press, Oxford, UK, 1990, 441–418. 4. Maehlum, T., Treatment of landfill leachate in on-site lagoons and constructed wetlands, Proc. 4th Int. Conf. Wetland Syst. Water Pollut. Control, Guangzhou, China, 553–559, 1994. 5. USEPA, Constructed Wetlands for Wastewater Treatment and Wildlife Habitat; 17 Case Studies, EPA 832-R-93-005, 1995. 6. CH2M-Hill and Payne Engineering, Constructed Wetlands for Livestock Wastewater Management, EPA Gulf of Mexico Program, Nutrient Enrichment Committee: Stennis Space Center, January, 1997. 7. Cole, S., The emergence of treatment wetlands, Environ. Sci. Technol., 218–222, 1998. 8. U.S. Soil Conservation Service (USSCS), Hydric Soils of the United States, National Technical Committee for Hydric Soils, Washington, D.C., 1987. 9. Mackay, D. and P. J. Leinonen, Rate of evaporation of low-solubility contaminants from water bodies to the atmosphere, Environ. Sci. Technol., 9, 1178–1180, 1975.
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10. Shugai, D. et al., Removal of priority organic pollutants in stabilization ponds, Water Res., 28, No. 3, 681–685, 1994. 11. Srinivasan, K. R. and R. H. Kadlec, Wetland Treatment of Oil and Gas Well Wastewaters, Report to U.S. Dept. of Energy on Contract DE-AC22-92MT92010, 55, 1995. 12. Peng, J., J. K. Bewtra and N. Biswas, Effect of turbulence on volatilization of selected organic compounds from water, Wat. Env. Res., 67, No. 1, 101–107, 1995. 13. Boethling, R. S. and D. Mackay, Handbook of Property Estimation Methods of Chemicals, Lewis Publishers, Boca Raton, FL, 2000. 14. Eger, P. et al., The use of wetland treatment to remove trace metals from mine drainage, Constructed Wetlands for Water Quality Improvement, G. A. Moshiri, Ed., Lewis Publishers, Boca Raton, FL, 1993, 171–178. 15. Reily, J. M. and H. A. Wojnar, Treating and reusing industrial wastewater, Water Environ. Technol., 52–53, 1992. 16. Haffner, W. M., Palmerton zinc superfund site constructed wetlands, paper presented at American Society for Surface Mining and Reclamation, Duluth, MN, 1992. 17. Noller, B. N., P. H. Woods and B. J. Ross, Case studies of wetland filtration of mine waste water in constructed and naturally occurring systems in Northern Austrailia, Water Sci. Technol., 29, No. 4, 257–265, 1994. 18. Crites, R. W., R. C. Watson, and C. R. Williams, Removal of metals in constructed wetlands, conference preprint, Water Environ., 68th Ann. Conf., Miami, FL, 1995. 19. Delgado, M., M. Biggeriego and E. Guardiola, Uptake of Zn, Cr, and Cd by Water Hyacinths, Water Res., 27, No. 2, 269–272, 1993. 20. Strecker, E. W., R. R. Horner, and T. E. Davenport, The Use of Wetlands for Controlling Stormwater Pollution, The Terrene Institute, Washington, D.C., 1992, 66. 21. Zhang, T., J. B. Ellis, D. M. Revitt, and R. B. E. Shutes, Metal uptake and associated pollution control by Typha latifolia in urban wetlands, in Constructed Wetlands in Water Pollution Control, P. F. Cooper and B. C. Findlater, Eds., Pergamon Press, Oxford, UK, 1990, 451-459. 22. Masscheleyn, P. H., R. D. Delaune, and W. H. Patrick, Jr., Arsenic and selenium chemistry as affected by sediment redox potential and pH, J. Environ. Qual., 20, 522–527, 1991. 23. Alvord, H. H. and R. H. Kadlec, The interaction of atrazine with wetland sorbents, Ecol. Eng., 5, No. 4, 469–479, 1995. 24. Lorah, M.M. and L. D. Olsen, Natural attenuation of chlorinated volatile organic compounds in a freshwater tidal wetland, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 25. Burris, D. R. and E. J. O’Loughlin, Reductive dehalogenation of trichloroethene mediated by wetland DOC-transition metal complexes, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 26. Neilson, A. H., Organic Chemicals: An Environmental Perspective, CRC/Lewis Publishers, Boca Raton, Florida, 1999. 27. Pardue, J. H. and W. S. Shin, Desorption resistance of organic compounds in wetland sould, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 28. Leppich, J., J. H. Pardue, and W. A. Jackson, Plant – Air partitioning of chlorobenzenes in wetland vegetation at a superfund site, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 29. Chiang, S. V. and F. M. Saunders, Intrinsic microbial dechlorination of 2,3,4,6tetrachloro phenol in anaerobic wetland sediment slurry, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.
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30. Montgomery, M. T. et al., Measuring intrinsic bacterial degradation of PAHs in a salt marsh, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 31. Jackson, W. A. and J. H. Pardue, Natural attenuation case study for chlorogenzenes in a forrested wetland, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 32. Roberts, A. L. and K. A. Lippa, Reactions of herbicides with reduced sulfur species in salt marshes, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 33. Gomez-Hermosillo, C., Bioavailability of desorption resistant phenanthrene to wetland plants, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999. 34. L. L. Behrends et al., Phytoremediation of explosives contaminated groundwater using constructed wetlands, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.
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7
Engineered Vegetative Landfill Covers CONTENTS 7.1 7.2 7.3 7.4 7.5 7.6
Historical Perspective on Landfill Practices................................................314 The Role of Caps in the Containment of Wastes........................................315 Conventional Landfill Covers ......................................................................316 Landfill Dynamics........................................................................................317 Alternative Landfill Cover Technology .......................................................321 Phyto-Cover Technology..............................................................................321 7.6.1 Benefits of Phyto-Covers over Traditional RCRA Caps.................326 7.6.2 Enhancing In Situ Biodegradation...................................................326 7.6.3 Gas Permeability ......................................................................327 7.6.4 Ecological and Aesthetic Advantages .........................................327 7.6.5 Maintenance, Economic, and Public Safety Advantages ..............329 7.7 Phyto-Cover Design .....................................................................................329 7.7.1 Vegetative Cover Soils .....................................................................330 7.7.2 Nonsoil Amendment ........................................................................331 7.7.3 Plants and Trees ...............................................................................331 7.8 Cover System Performance..........................................................................332 7.8.1 Hydrologic Water Balance ...............................................................332 7.8.2 Precipitation .....................................................................................335 7.8.3 Runoff...............................................................................................335 7.8.4 Potential Evapotranspiration — Measured Data .............................337 7.8.5 Potential Evapotranspiration — Empirical Data .............................339 7.8.6 Effective Evapotranspiration ......................................................340 7.8.7 Water Balance Model ................................................................343 7.9 Example Application....................................................................................344 7.10 Summary of Phyto-Cover Water Balance....................................................347 7.11 General Phyto-Cover Maintenance Activities .............................................348 7.11.1 Site Inspections ................................................................................348 7.11.2 Soil Moisture Monitoring ................................................................349
313
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7.11.2.1 Drainage Measurement.....................................................350 7.11.3 General Irrigation Guidelines ..........................................................352 7.11.4 Tree Evaluation ................................................................................356 7.11.4.1 Stem ..................................................................................356 7.11.4.2 Leaves ...............................................................................357 7.11.5 Agronomic Chemistry Sampling .....................................................358 7.11.6 Safety and Preventative Maintenance..............................................359 7.11.7 Repairs and Maintenance.................................................................359 7.12 Operation and Maintenance (O&M) Schedule............................................359 7.12.1 Year 1 — Establishment ..................................................................360 7.12.2 Years 2 and 3 — Active Maintenance.............................................360 7.12.3 Year 4 — Passive Maintenance .......................................................361 7.13 Specific Operational Issues..........................................................................362 7.13.1 Irrigation System Requirements ......................................................362 7.13.2 Tree Replacement.............................................................................362 References..............................................................................................................362
Maintaining and enhancing the closed landfill as a bioreactor requires modification of design and operational criteria normally associated with traditional landfill closure…
7.1
HISTORICAL PERSPECTIVE ON LANDFILL PRACTICES
The practice of using shallow earth excavations, or landfills, for disposal of liquid and solid waste has a very long history. Landfill practices basically followed the design philosophy of “out of sight, out of mind” in that a pit or trench was excavated into the ground, waste was placed into the excavation, and, when it was full, the excavation was covered with soil and abandoned. If thought was ever given to the matter, it was likely assumed that the soil surrounding the waste effectively prevented contaminant migration from the burial zone. It was not until 1976, with the passage of the Resource Conservation and Recovery Act (RCRA), and 1980, with the passage of Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) that federal and state regulations mandated much improved methods for disposal of waste in landfills. Today there are a plethora of federal and state regulations controlling all aspects of landfill disposal of municipal, radioactive, and hazardous waste. The problem in the U.S., however, is that hundreds of thousands of landfills were operated and then decommissioned prior to the requirements of current regulations. Many of these old landfills now come under the closure requirements of RCRA or CERCLA, depending on the agreements between the responsible parties. In 1989, U.S. Environmental Protection Agency (USEPA) stated that there are 226,000 sanitary landfills in the U.S. requiring evaluation for potential risks to human and environmental receptors.1 Regardless of the corrective action imposed on these old sites, almost all of them will require installation of a new cover as a final step in the closure process. The
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design of most landfill covers in the U.S. has been based on criteria developed by EPA for use in closing either RCRA subtitle C (hazardous waste) or subtitle D (municipal solid waste) landfills. Two major themes emerge in reviewing recent work in landfill cover design:2 1) there has been an overemphasis on regulatory compliance, thus inhibiting innovative and creative design that looks at the entire landfill system as a holistic biogeochemical environment, and 2) there are few published data on field performance of constructed cover systems and their impacts on the biogeochemistry of the groundwater within the footprint of the landfill.
7.2
THE ROLE OF CAPS IN THE CONTAINMENT OF WASTES
Because of the expense and risk associated with treating or removing large volumes of landfill wastes, remediation usually relies upon containment, which requires the construction of a suitable cover. Both regulators and the public usually accept covers as part of the presumptive remedy for final landfill remediation; therefore, covers are likely to be included in the optimal remedial actions for closure of most landfills. The intent of landfill remediation is to protect the public health and the environment. In keeping with this intent, a modern philosophy has evolved requiring contaminants in the waste to be isolated from receptors and contained within the landfill. As a result, landfills have become warehouses in which wastes are stored for an indefinite time, possibly centuries. There are fundamental scientific and technical reasons for placing a cover on landfill sites. Although regulations are often the most apparent influence governing the selection and design of landfill covers today, these regulations were drafted because of specific environmental concerns and were based upon scientific and technical understandings. The three primary requirements for landfill covers are to: • Minimize infiltration: water that percolates through the waste may dissolve contaminants and form leachate, which can pollute both soil and groundwater as it travels from the site. • Isolate wastes: a cover over the wastes prevents direct contact with potential receptors at the surface and prevents movement by wind or water. • Control landfill gas: landfills may produce explosive or toxic gases, which, if allowed to accumulate or to escape without control, can be hazardous.
Landfills have been covered by barriers for years, usually built with little regard for the monetary and environmental costs associated with constructing and maintaining them. A typical landfill cover design consists of a sequence of layered materials to control landfill gas infiltration and promote internal lateral drainage. The uppermost layer of a landfill cover consists of a vegetative soil layer to prevent erosion, promote runoff, and insulate deeper layers from temperature changes. The landfill cover is not a single element but a series of components functioning together.3
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Landfill covers are designed to minimize infiltration of rainfall and melting snow into the landfill in order to minimize postclosure leachate production. This objective is achieved by converting rainfall into surface runoff and infiltration into evapotranspiration and lateral drainage without compromising cover integrity. Secondary performance objectives of landfill cover design include the following:3 minimize postclosure maintenance; return the site to beneficial use as soon as possible; make the site aesthetically acceptable to adjacent property owners; accommodate post-closure settlement of the waste; address gas and vapor issues; provide stability against slumping, cracking, and slope failure; provide resistance to disruption by animals and plants; and comply with landfill closure regulations. The design features of a landfill cover are varied to affect changes in the overall water balance within the landfill to meet primary landfill cover objectives. The design adopted must take into account numerous other considerations, including costs, long term maintenance implications, and construction risks. The relatively large areas that landfill covers protect, and the thickness and number of individual layers within them, make covers a cost-intensive component of landfill facility design.
7.3
CONVENTIONAL LANDFILL COVERS
Nearly all conventional landfill covers in current use incorporate a barrier within the cover. The “impermeable” barrier layer is intended to prevent water from moving downward in response to the force of gravity. In effect, these covers are designed to oppose the forces of nature. Barrier-type covers commonly include five layers above the waste (Figure 7.1).1 The top layer consists of cover soil typically two feet thick and supports a grass cover that provides erosion control. The barrier layer consists of either a single low-permeability barrier or two or more barriers in combination. The fourth layer is designed to remove landfill gases as they accumulate underneath the barrier layer. The bottom layer is a foundation layer of variable thickness and material; its purpose is to separate the waste from the cover and to establish sufficient gradient to promote rapid and complete surface drainage from the finished cover. The barrier layer is the defining characteristic of conventional landfill covers. It may be composed of compacted clay, a geomembrane, a clay blanket, or two or more layers of materials in combination. A compacted clay layer is frequently specified to have a maximum saturated hydraulic conductivity (K) ≤ 1 × 10–7 cm/sec. In contrast, both the drainage and gas collection layers are constructed to enhance flow and commonly contain washed and selectively sieved sand, gravel, or specially designed synthetic materials. The soil in the top layer of barrier-type covers is usually too thin or has inadequate water holding capacity to store infiltrating precipitation during a large storm. These covers rely on barrier layers and rapid drainage through lateral drainage layers to prevent precipitation from reaching the waste. Barrier-type covers must accommodate specific site conditions, and supplemental components are sometimes added. For example, gravel may be added to the surface soil in desert regions to control
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Vegetation Topsoil 0'-6"
Vegetative Layer Common Borrow Material
1'-18"
Protective Cover Layer
Geocomposite (Textile-Net-Textile) 40-mil LLDPE
2'-30" (min.)
Foundation Layer
Waste
Figure 7.1
Typical single barrier cover system.
wind erosion, or a layer of cobbles may be used with the cover to discourage animal burrowing into the waste.
7.4
LANDFILL DYNAMICS
Landfills that contain a large amount of organic, putrescible materials (such as municipal solid waste) literally function as bioreactors. Most “landfill bioreacters” in general contain anaerobic and/or facultative microorganisms. Landfill leachate is generated as a result of the percolation of water or other liquids through the waste and also due to the accumulation of moisture generated as a result of microbial degradation of waste. Leachate is a concentrated fluid containing a number of dissolved and suspended materials, specifically, high concentrations of organic compounds (organic acids, hydrocarbons, etc.) and certain inorganic compounds (ammonia, sulfates, dissolved metals, etc. characteristic of the parent waste materials) from which it is derived. In addition, natural microbial activity in landfills also results in the generation of gases such as methane, carbon dioxide, ammonia, and hydrogen sulfide, a fraction of which will be dissolved in the leachate and may be introduced into the groundwater. Numerous landfill investigation studies4 have suggested that the stabilization of waste proceeds in sequential and distinct phases. The rate and characteristics of leachate produced and biogas generated from a landfill vary from one phase to another and reflect the processes taking place inside the landfill. These changes are depicted in Figure 7.2.
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Figure 7.2
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Description of leachate and gas concentration changes during landfill lifecycle.
The initial phase is associated with initial placement of waste and accumulation of moisture within landfills. Favorable biochemical conditions are created for the decomposition of waste. During the next phase, transformation from an aerobic to anaerobic environment occurs, as evidenced by the depletion of oxygen trapped within and introduced to landfill media and continuous consumption of nitrates and sulfates. Subsequent phases involve the formation of organic acids and methane gas. During maturation phase, the final state of landfill stabilization, available organic carbon and nutrients become limiting, and microbial activity shifts to very low levels of activity. Gas production dramatically drops and leachate strength remains constant at much lower concentrations than in earlier phases. Biochemical decomposition of putrescible solid waste is shown below by an example (Equation 7.1). Typical landfill gas composition during peak activity as a bioreactor is: 60% methane, 40% carbon dioxide, 5–10% other gases, and 0.3–1.0% VOCs and non-monitored organic compounds. Gas generation rates during peak activity typically fall within the ranges of 5–15 ft3 per pound of refuse per year.9 C 6 H10 O 5 H 2 O
Anaerobic → 3CH 4 + 3CO 2 Bacteria
(7.1)
Due to very high gas pressures generated at the source areas within the landfill (up to 4 atmospheres), migration of dissolved contaminants into the gaseous phase could be a serious concern. Contaminants transferred into the gas phase could be readsorbed in the waste above the water table, or dissolve in the moisture, condense in the waste zone, or migrate away from the landfill. The potential for contaminant migration from the dissolved phase into the landfill gas can be evaluated as shown in Equation 7.2, and Figure 7.3.
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Figure 7.3
319
Equilibrium mass transfer conditions of contaminants into landfill gas.
Under non-equilibrium conditions:
(
)
dm = K C g − HC w A dt
(7.2)
where dm dt K H Cg Cw A
= transfer rate from gas to water = = = = =
phase transfer coefficient Henry’s Law Constant of VOC gas phase concentration of VOC water phase concentration of VOC gas/liquid contact area
The progress toward final stabilization of any landfill and the organic waste in it is subject to physical, chemical, and biological factors within the landfill environment, age and characteristics of the waste, operation and management controls applied, as well as site-specific external conditions. Although barrier layers are sometimes referred to as “impermeable” layers, in practice this is seldom true. An extensive review of failures and failure mechanisms for compacted soil covers in landfills was performed and emphasized that “…natural physical and biological processes can be expected to cause [clay] barriers to fail in the long term.”5 Another study discussed a field test conducted in Germany in which,
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Precipitation
Runoff 0.15 m Topsoil
0.47 m
Soil Barrier K ≤1 x 10-5 cm/sec
Foundation Gravel
Waste
Figure 7.4
Subtitle D cover for municipal solid waste landfills.
seven years after construction, percolation through the compacted clay was almost 200 mm/yr and increasing. Geomembrane barriers are also prone to leak.6 Others have traced most leaks in geomembranes to holes left by construction.7,8 A modification of the typical barrier cover is the subtitle D cover (Figure 7.4) that relies on compaction to create a layer of soil with reduced K value. Used primarily for municipal landfills in dry regions, its use and components are specified in subtitle D of RCRA (40 CFR, Part 258.60), hence the name. From the surface downward, the cover includes an erosion control layer and a layer of compacted soil. A major advantage of the subtitle D cover is that its construction cost is lower than for an RCRA subtitle C cover. Even though it has gained regulatory and public acceptance, the subtitle D cover cannot ensure long-term protections against infiltration of water into the waste, even in dry regions, because 1) the topsoil layer has limited water holding capacity, 2) there is no drainage layer, 3) few roots can grow in the barrier layer to remove water, and 4) soil freezing and root activity are likely to increase the K value of the barrier soil layer over time.
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7.5
321
ALTERNATIVE LANDFILL COVER TECHNOLOGY
Alternative covers to the RCRA subtitle C or D design include evapotranspiration (ET) covers and capillary barriers. The ET cover uses no barrier or horizontal drainage layers; it is designed to work with the forces of nature rather than attempting to control them. An ET cover in its simplest form is a vegetated soil cover with a sufficiently deep soil profile so that infiltrated water is stored until removal by evaporative losses from the soil surface and by plant roots at depth in the profile. A capillary barrier also relies on water removal by ET, but is designed such that water storage near the surface is enhanced to promote the efficient removal of infiltrated water by the ET process. Optimization of material types and thicknesses for capillary barriers is critical to their effective performance. The use of sands or clays as the fine-soil component in the capillary barrier has proven to be less effective in storing water than silt loams. Capillary barriers can be thought of as enhanced ET covers — alternative cover systems that work best in semi- and/or arid environments where high ET rates and low precipitation make it possible to remove all infiltrated water by ET. However, even in arid environments there are situations where ET covers and capillary barriers can allow excessive percolation, particularly where the soil used in the cover design has insufficient storage capacity to accommodate winter snow melt events.
7.6
PHYTO-COVER TECHNOLOGY
The phyto-cover is the most popular application of the ET cover and is an engineered agronomic system that harnesses the natural transpiration process of plants to limit percolation to the groundwater. A phyto-cover relies on shallow- and deep-rooted plants to create a thick root zone from which the plants can extract available moisture. In effect, the plants serve as natural, solar-powered “pumps” to withdraw soil moisture and either convert it into biomass or evaporate it through their leaves. The withdrawal rate of the botanical pumps is limited by the available energy (sunlight), rate of growth, and available soil moisture; withdrawal virtually ceases during winter dormancy. Accordingly, the depth and composition of the root zone must be sufficient to store accumulated water like a sponge and hold it until the plants remove it. Properly designed, this “sponge and pump” water removal system (Figure 7.5) can limit water from percolating below the root zone and can be equally protective of groundwater as a RCRA cap. Thus, a phyto-cover serves as a functional alternative to natural clay, geocomposite, or geosynthetic membrane cap, yet offers several advantages over those technologies. The effectiveness of poplars in maintaining low soil moisture levels was first documented by data collected from a phyto-cover application in Iowa.10 The phytocover consistently maintained soil moisture levels substantially below the soil’s field capacity (i.e., the amount of water that soil can retain without allowing percolation) of 40–45%. Soil dryness was maintained by the trees’ prodigious water extracting ability. The capacity of certain trees such as hybrid poplar and willow trees to extract soil moisture has been demonstrated by monitoring data from landfill at many sites.
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Precipitation
Poplars Evapo-transpiration
Evaporation Infiltration
Surface Runoff 10'-0"
Root Depth
Topsoil Storage Native Soil Storage Daily Cover Waste
Figure 7.5
Conceptual phyto-cover diagram.
The poplars are employed at this site not as a cover, but to treat collected landfill leachate, which is applied to the poplars during the growing season. The total amount of water extracted from the soil in one growing season by these two- and three-yearold poplars was equivalent to about 62 inches of precipitation.10 One of the most important design considerations for a phyto-cover is choosing appropriate tree species and varieties. Selected trees must be capable of achieving the desired treatment objective and adapt to the irrigation water, soils, and climate of the site. Typically, achieving the highest possible rate of evapotranspiration is an important goal. Critical site conditions for plant selection include soil chemistry, irrigation water or groundwater chemistry, and adaptation to pests and diseases of the area. Any factor that compromises tree health and growth will reduce performance. For example, hybrid poplar clones that include either trichocarpa or maximowiczii parents are quite susceptible to Septoria canker if used in the U.S. midwest.10 Especially for the commonly used Salicaceae, a number of different types of plant materials may be used. These include stem cuttings, whips or poles, and bare root or potted material. Use of larger or rooted plant material will result in more rapid establishment and reduced weed competition, but plant material and planting costs are much higher than for smaller material. Whips and poles are commonly used for deep planting applications. Economics, especially planting costs, drive most larger installations (>5 ha) toward short stem cuttings. Certain varieties may result in a more valuable final wood product because of straighter stems or better paper processing properties. Significant differences in damage from voles has been observed among hybrid poplar trees at phytoremediation sites. Salt tolerance is a very important selection criterion, as differences between species and varieties can be significant. Only limited data for Salicaceae are currently available to guide design, but a number of relevant research programs are ongoing. For an increasing number of sites, use of non-native species is unacceptable for local community groups and sometimes for regulators. Use of native material will generally ensure resistance to local pests and disease, but may not afford the greatest efficiency.
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Once the tree system forms a complete canopy, spacing has little effect on evapotranspiration or nutrient requirements. The impact of spacing on hydraulic and nutrient loading is primarily an early establishment phase concern. Establishing dense initial plantings with the intention of thinning may provide small increases in early capacity, but thinning operations are often neglected and the resulting mature tree stands are excessively dense. Enough space must be left between tree rows to allow planned maintenance activities such as mowing or spraying. The engineered phyto-cover system consists of densely planted, deep-rooted trees and understory vegetation (perennial rye grass and clover). Photographs of hybrid poplar tree stands of varying ages are shown in Figure 7.6. The water-holding root zone (“sponge”) includes the existing topsoil and fill at the site (including intermediate and daily cover soil) supplemented with additional soil or soil amendments as dictated by design calculations. A phyto-cover will provide a protective, living “skin” for a landfill that permanently heals the wound to the landscape originally created by anthropogenic activities. This skin can equal the percolationblocking performance of a “rain coat” RCRA cap while being substantially more cost effective and providing additional benefits. The final design of a phyto-cover often includes provisions for monitoring soil moisture levels to ensure that performance criteria are met. Two-year Old Stand of Poplars
Four-year Old Poplar Trees
Figure 7.6
Phyto-covers: comparison of two-year-old and four-year-old growth of a phytocover (courtesy of Licht, 1998).
Engineered phyto-cover systems have been applied to contain spilled petrochemicals and cover landfills, as well as buffers to remove nitrogen from industrial and municipal wastewater. Sites where phyto-covers have been installed and recent research and demonstration sites for phyto-cover systems include the following:2,10-13 • A 15-acre construction debris landfill in Beaverton, OR was covered with trees in 1990 as an alternative to excavation of the fill, the installation of a liner, and then recovering with a geomembrane. The phyto-cover is serving to protect groundwater cost-effectively. The owner has continued to expand the cover as new areas are closed.
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• From 1992 to 1993, the Riverbend Landfill in McMinneville, OR planted a 17acre phyto-cover to manage landfill leachate water and soluble compounds. All nutrient and water cycling results indicate the cap is achieving all regulatory requirements for ammonia treatment and ground protection. • From 1993 to 1995, a 15-acre perimeter buffer was planted to reduce infiltration from upgradient runoff, grow a visual and sound barrier, and intercept downgradient leachate seepage. Data collected at the site indicate the cap has been successful in stopping all leachate. • At the Grundy county Landfill in Grundy Center, IA, a two-acre cap and perimeter buffer was planted from 1993 to 1994 to reduce leachate formation by installing a phyto-cap over a completed subtitle D cap. The cap also provides a visual and sound barrier, and intercepts downgradient leachate drainage. • A three-acre poplar cap was planted in 1994 at the Bluestem 1 Landfill in Cedar Rapids, IA over a pre-subtitle D cap. The cap serves to reduce leachate formation vertically and intercepts downgradient leachate drainage. The data collected at this site have been used in writing specifications for soil and compost cover requirements and use of MSW waste as a planting media. • A five-acre cap and perimeter boundary were planted in 1994 over a pre-subtitle D cap at the Bluestem 2 Landfill in Marion, IA to reduce leachate formation. Moisture management data from this cap have been used in subtitle D equivalence comparison between a soil/clay cover and the “sponge and pump” concept for deep rooted trees in four feet of rootable soil. • In 1995, a ten-acre area was planted with poplar trees and a clover/grass understory over a subtitle D cell filled with MSW and industrial waste. The Department of Environment Quality and governor’s office were interested in future phytoclosures for many funded pre-RCRA landfills in Virginia that have been abandoned and are creating potential environmental risk. The trees are growing well and are being maintained by the owner. A soil moisture measurement system using time domain reflectometry (TDR) is used to monitor the impact of tree roots on vadose zone water content. A drip irrigation system using collected storm water can control the water stress during periods when moisture in the root zone has been exhausted. • At a railroad RCRA site in Oneida, TN, a one-acre area impacted by coal and creosote from past manufacturing activities was covered with poplar trees and grass in 1997. The site soils were amended with compost and mineral fertilizer, then trenched in the root zone. The trees and grass managed to accelerate biomass growth with resulting water uptake and in situ constituent removal. The site groundwater is monitored by a university research grant to measure groundwater elevation and the containment of constituents. The concept is similar to landfill capping where the phytosystem pumps water at high rates during the growing season and minimizes groundwater movement during the dormant season. • The Woodburn WasteWater Treatment Plant in Woodburn, OR has been a demonstration site since 1995; a full-scale installation took place in 1998. This site is the first full-scale tertiary treatment of secondary municipal wastewater effluent and is being designed for no leakage through the root zone in the summer months. • The Mid-Lakes Co-op site in Bonduel, WI used an aesthetically pleasing poplar cover over a spill plume to contain pollutant migration, make use of all available precipitation, protect public exposure, and remove constituents from the groundwater . Closure requirements for this site included planting trees, monitoring the depth to groundwater, and monitoring groundwater quality over a three-yearperiod.
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• In Staten Island, NY a phyto-cover consisting of poplars, willows, paulowia, and grasses is being used to prevent constituent migration and formation of leachate. Enhancement of existing vegetation is expected to establish hydraulic control of groundwater by reducing water infiltration through the landfill materials. • Evidence collected at a closed landfill in Elmore, OH indicates naturally occurring trees have created a treatment barrier for leachate seeps. An evaluation of on-site box elder and osage orange trees yielded evidence of TCE uptake. An evaluation of the existing cover for supplemental enhancement for additional groundwater remediation and restoration was then conducted. • A poplar tree phyto-cover was installed in 1996 at a landfill in Acme, NC. The trees were planted in the most downgradient area of the landfill to stop leachate migration. Groundwater constituent concentrations have dropped substantially in the area of the poplar trees but not in areas where trees were not planted. • From 1992 to 1993, over 2000 poplar trees were planted at a site in Anderson, SC to be used for processing waste from mining ore material. The waste was used to fill low areas over six acres of the site. Data collected at the site indicate infilitration and leachate formation is being controlled. • In 1991, a succession of trees (willow and black locust), legumes, and grasses were planted to dewater slurry waste at a site in Baton Rouge, LA. The waste material was in a slurry state from a depth of 6 inches to 30 feet below ground surface. The planted vegetation reduced the hydrated state of the waste and the occurrence of leachate through the impoundment. • A process waste was placed as a slurry into an impoundment at a site in Texas City, TX. Naturally occurring trees (osage, orange, and mulberry) and vegetation have reduced the hydrated state of the top ten feet of the waste. Research on the site has found that dewatering and net water removal are directly correlated to the size of the trees. • Ongoing research, funded by the USEPA Great Plain and Rocky Mountain Hazardous Substance Research Center involves planting trees at CERCLA sites to control erosion and leaching of zinc, arsenic, lead, and cadmium. • A grass/soil cover system is one of five alternative covers being evaluated by Sandia National Labs in NM as part of an alternative landfill cover demonstration study. Similar phyto-cover systems are being considered as potential demonstration sites by USEPA ORD at sites in Tulsa, OK; Beatty, NV; and Hill Air Force Base in CA. • A phyto-cover system has been proposed and designed at the 95% completion level for the F.E. Warren Air Force Base Superfund Site in Cheyenne, WY. This site is currently being considered as a technology demonstration candidate by USEPA-Region VIII. • Pfitzer junipers have been used in a landfill cover field demonstration at Beltsville, MD. The juniper phyto-cover was installed over a clay layer to add to the “robust” cover development, but not as a replacement of the low-permeability layer. The objective of the demonstration study was to document the influence of junipers as water scavengers, yet maintain the water runoff performance of the lowpermeability cap. Compared to a reference soil, the “bioengineered” juniper cover reduced infiltration; it was demonstrated that the mature plant system improved the system’s resilience to failure. • Research regarding the establishment of sufficient vegetation to provide adequate biomass growth with resultant evapotranspiration is being conducted at Idaho National Engineering Lab, Idaho Falls, ID. This research focuses on four plant
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species to deplete soil moisture, and the configuration of a capillary barrier and root zone to prevent deep percolation during wet periods. The use of such phyto-covers has been demonstrated to be applicable to landfill sites in the semi-arid west. • In Ljubljana, Slovenia, a ten-acre cover was planted with poplar trees in 1993–1994 with the primary goal of protecting groundwater by reducing leachate formation through municipal and industrial wastes. Installation of the cover has greatly improved the aesthetics of the area and increased the value of the wildlife habitat. The design concept is being considered as a model that will become national policy. • The author also knows of many other phyto-cover applications in MA, OH, MD, NC, MI, PA, NY, NJ, and IL.
In 1998, USEPA began an effort to establish a design database and improve numerical prediction methods for alternative landfill covers. The initial task of the Alternative Cover Assessment Project (ACAP) was to catalog past and existing research efforts into measurement of cover performance and to describe the current state of numerical prediction methods. The primary criterion was to measure percolation directly. Several research sites operated by branches of the federal government were included in this study. These sites include the national laboratories at Hanford, Sandia, Los Alamos, Savannah River, and Idaho Falls, and DOE locations at Monticello, UT, Nevada Test Site, DoD locations at California, Hawaii, Colorado, Utah, and others. 7.6.1
Benefits of Phyto-Covers over Traditional RCRA Caps
In addition to satisfying the critical antileaching requirement, phyto-covers provide a number of significant pollution control, ecological, and economic benefits when compared to traditional RCRA caps: • A phyto-cover actually enhances natural biodegradation processes, instead of interfering with them, as a RCRA cap would. • A gas-permeable phyto-cover allows for passive venting of gaseous byproducts of biodegradation and allows oxygen to move into the fill to facilitate additional biodegradation. • A phyto-cover provides a forest ecosystem and an attractive alternative to an RCRA cap. • A phyto-cover can be installed with less cost and less risk to public safety than a RCRA cap and, once the cover is established, the system has a natural stability that minimizes long-term maintenance requirements.
7.6.2
Enhancing In Situ Biodegradation
Installing a phyto-cover at a site has the potential to enhance biodegradation of waste materials, including organic waste and contaminants, in the root zone. In natural ecosystems, high concentrations of indigenous soil microorganisms are found in association with plant roots, because the roots exude a wide variety of compounds such as sugars, amino/acids, carbohydrates, and essential vitamins. These compounds not only sustain the microbial consortia, which can degrade many organic compounds directly, but also enhance and accelerate cometabolic degradation of
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other pollutants resistant to direct degradation. In addition, the plants themselves will take up and metabolize or volatilize some of the organic contaminants in the fill. Finally, exuded organic acids also help in sequestering and immobilizing any metals present in the root zone. By contrast, a RCRA cap provides no stimulation to natural biodegradation and would be expected to substantially change biogeochemical conditions in the fill by trapping the gaseous byproducts of biodegradation (e.g., methane, carbon dioxide, and ammonia), thereby affecting factors critical to natural attenuation mechanisms, such as pH and REDOX potential. The main reason for the enhanced in situ biodegradation in landfills with phytocovers is the ability of the atmospheric oxygen to transfer into the landfill. The primary mechanism transferring oxygen into the landfill is diffusion into the soil from the atmosphere, based on an excellent summary shown below:14 The exchange of gases between the soil and the atmosphere … is facilitated by two mechanisms: mass flow and diffusion. Mass flow of air, which is due to pressure differences between the atmosphere and the soil air, is less important than diffusion in determining the total exchange that occurs. It is enhanced, however, by fluctuations in soil moisture content. As water moves into the soil during a rain … air will be forced out. Likewise, when soil water is lost by evaporation from the surface or is taken up by plants, air is drawn into the soil. Mass flow is also modified slightly by other factors such as temperature, barometric pressure, and wind movement. Most of the gaseous interchange in soils occurs by diffusion.
The minimum rate of oxygen diffusion at the bottom of the root zone was estimated to be 5 × 10–8 grams per centimeter, squared per minute, or 2340 pounds per year per acre.14 The maximum rate could be up to 9200 pounds per year per acre. Over the surface of a 30-acre landfill, this translates into at least 70,000 pounds of oxygen per year into the landfill, which facilitates stabilization of the waste. By contrast, the single-barrier cap would admit only an estimated 75 pounds of oxygen or about one tenth of one percent of the influx that could support the aerobic natural attenuation mechanisms (Figures 7.7a and b).15 7.6.3
Gas Permeability
Unlike RCRA caps, which are essentially impermeable to gases and therefore require elaborate gas venting systems to deal with gases and vapors generated by biodegradation of the fill, a phyto-cover is porous and permeable to gas. A phytocover can thus eliminate the need for a gas collection system at many sites. Equally important, a phyto-cover will allow oxygen to migrate into the fill, which will help to support additional aerobic biodegradation and thereby hasten the completion of the waste life cycle. 7.6.4
Ecological and Aesthetic Advantages
Both a phyto-cover and an RCRA cap are designed to be vegetated on the surface, but vegetation on a phyto-cover has the appearance of a tree farm and, eventually,
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Figure 7.7a
Biogeochemical conditions and mass balance for a presumptive remedy.
Figure 7.7b
Biogeochemical conditions and mass balance for a holistic remedy.
a forest, and serves the same ecological function as a forest while the RCRA cap is covered with grass and, in order to protect the impermeable liner, must be prevented from functioning like local natural ecosystems. Specifically, maintenance of the integrity of the RCRA cap’s impermeable layer dictates that deep-rooted plant species, such as trees and shrubs, not be allowed to colonize the site through natural succession. Moreover, protection of the impermeable liner also requires that small burrowing mammals, such as those normally associated with a meadow, must be perpetually monitored for and eliminated when found. By contrast, the trees of a phyto-cover provide nest sites for birds and other arboreal species and readily accept in-fill by shrubs and native tree species, as deemed appropriate under site management criteria. Because no animal is likely to excavate below the deep root zone, it is not necessary to prevent native fauna from inhabiting the phyto-cover. Besides offering a preferred natural ambiance, the phyto-cover forest would also serve the community as a noise pollution buffer and assist incrementally with global climate issues by fixing substantially more carbon dioxide from the atmosphere than a grass RCRA cap.
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7.6.5
329
Maintenance, Economic, and Public Safety Advantages
The ongoing maintenance requirements for an established phyto-cover are minimal. Although relatively intensive monitoring for disease and pests is needed during the three growing seasons that the trees need to become fully established, maintenance activities thereafter are expected to be minimal because of the self-healing nature of the phyto-cover. Like a natural forest, the phyto-cover is expected to be resistant to wind and water erosion. Unlike a RCRA cap, which can suffer cracks, rips, and tears due to factors such as differential settling or physical intrusion, the phyto-cover maintains its integrity and actually heals itself with new root growth in response to physical disturbances. Thinning of trees may be undertaken in the future to avoid crowding as the trees reach their mature size. However, the trees cut in a thinning operation represent a valuable forestry crop, so revenue from their sale should compensate for the operation’s costs. The lower economic cost of the phyto-cover compared to the RCRA cap is accompanied by lower noneconomic social costs in the form of safety risks. Studies have shown that remedy implementation imposes risks of injury and death to site workers, neighbors, and the public using transportation routes. These risks are more certain and typically substantially greater in magnitude than risks to the public from exposure to site contaminants. For example, assuming that bulk construction materials can be found at an average distance of 15 miles (i.e., 30 miles round trip) and using the U.S. truck fatality rate of 4.7 × 10–8/mile, construction of RCRA “C” cap at a 30-acre landfill site could lead to an estimate of transportation fatalities risk of 0.033.16 This estimate will be further increased if the nontruck driver fatalities estimate is combined with this. Since phyto-covers require less site work and fewer truckloads of imported material, such as borrow soil and gas collection layer sand, constructing a phyto-cover instead of an RCRA cap would involve less risk of an accidental injury or fatality to a construction worker and lower risks of traffic incidents associated with truckloads of construction materials carried over local roads Finally, unlike a RCRA cap, which locks the site into an “impermeable barrier” strategy, the phyto-cover system can be operated in a flexible way to take into account the results of ongoing performance monitoring data. For example, if performance data show that native species perform as well as hybrid poplars in preventing infiltration, then the natural transition to native species can be accelerated, to enhance the ecological service provided by the area. By the same token, in the unlikely event that performance data show that a part of the cover is not performing to expectation, then a supplementary measure such as additional “sponge” or denser planting would be available to improve performance.
7.7
PHYTO-COVER DESIGN
The typical components of an engineered phyto-cover system consist of vegetative cover soils (existing and supplementary), soil amendments, nonsoil amendments, understory grasses and plants, and trees. An irrigation system is an optional component to ensure sufficient water for tree growth in case of drought. Irrigated
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trees grow more rapidly, thus meeting closure objectives in less time; however, there is often lack of a convenient water source or on-site operation and maintenance personnel to make an irrigation system feasible at a site. The need for on-site irrigation should be based upon the expected water consumption of the trees. 7.7.1
Vegetative Cover Soils
The existing cover soil at many sites is sufficient to support an adequate root system for healthy tree growth. This is evidenced by the vigorous growth of trees often seen at abandoned landfills (Figure 7.8); however, the ability to grow trees is not evidence that percolation and leachate production are controlled. Typically, natural stands of vegetation are not effective at controlling percolation. Therefore, sufficient soil and nonsoil amendments may need to be added to meet the requirements for tree growth, and to achieve minimum land surface slopes to promote surface drainage and to provide sufficient soil water holding capacity for storage to function as an adequate “sponge.” The amount of soil and nonsoil amendments necessary must be determined by site-specific information, often collected in the later stages of design.
Figure 7.8
Existing root penetration of a tree at a landfill site.
Any supplemental cover soil added to achieve the required grades, as well as sufficient water storage capacity, will comprise common borrow soils. Supplemental soil should be placed in 6-inch thick and loose lifts, and be lightly compacted to achieve the minimum slope and thickness. This material is typically available from several sources in the vicinity of most sites; the specific local source usually depends upon availability during the construction period. The surficial lift of supplemental soil and existing cover, depending upon which is exposed at the final grade surface,
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must be ripped in two directions following final grading to assure noncompaction and to prepare the surface to receive the nonsoil amendments. 7.7.2
Nonsoil Amendment
The addition of nonsoil amendments will increase the water-holding capacity and nutrient transfer properties of the common borrow soils. Typical nonsoil amendments include compost, chipped wood, digested sewage biosolids, lime-stabilized sludge, manure, and other organic biomass. The incorporation of this type of organic matter into the existing and supplemented soils will greatly increase the tilth, fertility, and water-holding capacity of the soil, and further reduce percolation through the cover. Biosolids compost and lime-stabilized sludge are readily available through a compost contractor. Typically a minimum 6-inch thick layer of organic amendments needs to be applied to the soil surface after achieving final grade. This material is spread evenly in a six-inch thick layer on the area to be planted with the engineered phyto-cover system and is ripped into the surficial soils to a depth of 14 to 18 inches. Ripping is performed in both an east-west and north-south orientation in order to achieve a uniform mixing within the soil profile. Finally, the site is tilled in preparation for planting. If the organic materials used for the nonsoil amendment have a high carbon to nitrogen ratio, fertilizer is added along with organic amendments to aid in stabilizing these amendments and to provide sufficient nutrients to the rooting plants. This organic amendment is expected to supply all micronutrients required by the plants. A mineral fertilizer is also added as needed, based on nutrient analyses of the applied compost, to supplement the macronutrient reserves of nitrogen, phosphorous, and potassium. All amendment addition, ripping, and tilling is completed prior to understory planting in the fall and before the trees are installed in the following early spring. 7.7.3
Plants and Trees
The area to be planted will generally exhibit a minimum 2% or greater grade; therefore, stabilization of the site cover material remains necessary to prevent erosion. Understory planting will be established for early erosion control and water uptake during the first year. Understory establishment is a combination of annual and perennial grasses, such as varieties of rye, oats, wheat, barley, and fescue, applied at a rate of 20 to 40 pounds per acre. This mixture of seed is designed to meet the short- and long-term objectives of the understory. Annual species will be fast growing to control near-term erosion; perennial grasses will be deep-rooted species selected as the primary long-term understory for the site. The long-term effectiveness of the overstory is dependent upon establishment and long-term maintenance of the understory, which understory depletes shallow soil moisture, turn causing tree roots to grow deeper to meet water requirements. As discussed earlier, the success of a phytocover is dependent upon establishment of deep-rooted trees to create a sufficient sponge to store soil moisture in the dormant season.
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The trees normally selected for construction of a phyto-cover are hybrid poplars of the variety Deltoides x nigra.10,13 The candidate varieties, DN-21, DN-34, OP 367 and others, are planted based on demonstrated growth ability and hardiness in the environment. The poplars are installed as either rooted plants or whips at a density of approximately 1200 trees per acre.15 The rows are located by measurement and flagged, and the trees installed by a tractor and mechanical planter. These trees are typically planted with an in-row spacing of 3 feet and a row spacing of 10 to 13 feet. They are planted in rows positioned along the land elevation contours, perpendicular to slopes to aid in reducing sheet flow velocities and surface erosion.
7.8
COVER SYSTEM PERFORMANCE
The engineered phyto-cover system should be designed to meet the post-closure and remediation objectives established for any landfill site as specified below: • Minimize infiltration of precipitation through the cover system into the waste to protect groundwater quality at the site. • Resist surface soil erosion by wind and precipitation. • Minimize long-term maintenance. • Protect human health and the environment. • Offer post-closure and future beneficial use.
The achievement of these objectives is outlined in this and subsequent sections. 7.8.1
Hydrologic Water Balance
The engineered phyto-cover system is designed to mature into a remedial system that exceeds the hydrologic performance of more conventional cover systems. However, instead of acting as a constructed barrier layer, which diverts precipitation from the cover area as surface runoff or internal drainage, this system intercepts and uses the water for plant growth. In other words, the engineered phyto-cover functions as a sponge and pump system, with the root zone acting as the sponge, and trees acting as the solar-driven pumps. In contrast to restrictive permeability barrier design, the engineered phyto-cover design involves the storage of free water in soil pores and the extraction of stored water by the tree roots. The effectiveness of engineered phyto-cover systems as landfill closure systems has been demonstrated at sites in the U.S. At sites in various climates with engineered and agronomically optimized growing conditions, rapidly growing poplar trees are capable of transpiring all natural precipitation that infiltrates into a site. While the performance of engineered phyto-cover systems has been demonstrated, a proven tool to design phyto-covers is not available. Therefore, to support the design and demonstrate the effective performance of phyto-cover systems, this section discusses some fundamental scientific methods of water balance analysis. As discussed previously, the phyto-cover system utilizes specially selected trees with a grass understory to optimize evapotranspiration and achieve the equivalent
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performance of a conventional barrier cover system. This alternative landfill cover system has been designed to minimize percolation to the waste by incorporating a landfill soil cover with sufficient evapotranspirative and water holding capacity to store precipitation temporarily in the nongrowing season for subsequent evapotranspiration by vegetation in the growing season. The two key design elements in engineering a phyto-cover system are 1) determining the thickness and material composition of the soil cover system required to provide sufficient water storage capacity; and 2) incorporating a supportive phyto-cover system to access water stored in the soil cover system for evapotranspiration to the atmosphere. Moisture flow and moisture content in a landfill are extremely important to the dynamic processes of decomposition and potential leachate generation. The fundamental means to assess the moisture conditions is through an evaluation of various processes comprising a water mass balance. A water mass balance analysis is an “accounting” procedure for tracking the moisture inputs to storage and the moisture outputs that influence the potential flux of water through the cover into the waste. The primary elements of a water mass balance include precipitation, surface runoff (R/O), potential evapotranspiration (PET), infiltration (I), soil moisture storage (ST), actual evapotranspiration (AET), and flux (or percolation) of water through the system. The water shedding efficiency of a cap is then derived by calculating the percentage of flux relative to total precipitation. The phyto-cover system design concept involves maximizing efficiency by optimizing ET, runoff, and soil moisture storage to minimize infiltration, flux, and potential leachate generation. The water balance accounting for a phyto-cover can be summarized by the following equation and Figure 7.9: Percolation = Precipitation – Runoff – Evapotranspiration – Moisture Storage (7.3) The water mass balance processes within a landfill are typically evaluated using the hydrologic evaluation of landfill performance (HELP) model, developed by the Waterways Experiment Station.17 The applicability of this model to design and evaluation of an engineered phyto-cover system has been reviewed, and it has been determined that the HELP model is inappropriate for this analysis because of several computational deficiencies.18,19 The HELP model was developed based on assumptions pertaining to water management through low permeability soil covers with vegetative covers comprising short-rooted grasses. No opportunity exists for user input of higher ET values more representative of plant species with significantly higher potential water uptake than the short grasses assumed by the HELP model. Therefore, the model significantly underestimates evapotranspiration from trees and other deeply rooted vegetation that are key elements of a phyto-cover system. This application limitation of the HELP model results in an overestimation of infiltration and coincident underestimation of efficiency (overestimation of drainage) if the model were to be applied to an evaluation of a phyto-cover system. A detailed assessment of various computer models used for landfill cover designs during the early phases of the alternative cover assessment program (ACAP) came to similar conclusions.11,12 Of the four codes tested, HELP was the most widely used
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Precipitation
Leaf Transpiration Canopy Interception Surface Evaporation
Soil Evaporation Surface Evaporation
Surface Cover Interception
1) Surface Litter or Compost
Infiltration Storage
Root Depth
2) Soil and Nonsoil Amendments
3) Waste Rootable Upper Layer (Contributes toward Storage)
Potential Infiltration
Depth of Capillary Zone Draw
Thickness of the Arrow is Proportional to the Volume of Water Figure 7.9
Diagram of a phyto-cover (modified from Licht, 1998).
for landfill design, and the most user-friendly and highly dependable. HELP predictions consistently provided the highest estimates of drainage. Three concerns with HELP were 1) a nonrealistic response of increased drainage as available water capacity increased, 2) insensitivity of drainage to thickness of the cover surface layer, and 3) consistent overprediction of drainage. EPIC was also relatively easy to use, but consistently underpredicted drainage in comparison to other codes. The study suggested that Richards’ equation-based codes (HYDRUS–2D, UNSAT–H) were better able to capture the behavior of alternative landfill covers than simple water balance codes such as HELP and EPIC. Although the HELP model itself cannot accurately simulate the hydraulic effects of an engineered phyto-cover system, the water balance method that is the fundamental principle applied within the HELP model has been employed to evaluate the performance of vegetative cover systems.20 These same scientific principles are employed to design and evaluate the performance of an engineered phyto-cover system with a new software tool called PHYTOSOLV.15,21 In using the water balance method, the first step is to acquire accurate precipitation records applicable to the site and encompassing various extreme wet and dry periods. The second step is to determine the quantity of surface water runoff and infiltration (which are functions
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of the site soils, slope, and surface texture). Infiltration is computed as the difference between precipitation rates for the site and surface-water runoff from the soil cover. The third step is to apply PHYTOSOLV, assuming a variety of soil cover depths, to generate a range of annual hydrologic water balances using daily precipitation data. Finally, a supporting phyto-cover system is designed that would access infiltrated soil water throughout the entire extent of root growth (the “sponge”), and the necessary evapotranspiration rate (the “pump”) required to deplete soil moisture during the growing season is computed. This iterative water balance analysis is used to select the appropriate soil cover design to best achieve desired hydraulic performance, thereby minimizing generation of leachate. The measure of performance for the designed phyto-cover is compared to the water-shedding efficiency of traditional barrier cover systems. Presented below is a discussion of each of these steps and the basis for the general engineered phyto-cover system design. 7.8.2
Precipitation
Long-term precipitation data need to be assembled from the closest weather station to evaluate local hydrologic conditions. There are no established regulatorily approved procedures or protocols to evaluate the hydrologic performance of a phytocover design. Therefore, long-term data are needed in order to characterize the longterm precipitation trends and extremes. Typically, precipitation can vary widely from site to site for a given year, season, or month. To demonstrate this variability, data should be assembled summarizing average monthly and annual precipitation for decades at weather stations near any given site. For example, during a long period at a site in Maryland, the average annual precipitation varied from a minimum of 26.29 inches in 1965 to a maximum of 62.36 inches in 1996. Similar variability can also be observed in monthly precipitation totals. To the extent practical, these dynamics must be accounted for in the design of the phyto-cover system to demonstrate adequate performance under extreme weather conditions. The application of these data to evaluate the phyto-cover design assumes that daily precipitation totals are the result of individual storm events. 7.8.3
Runoff
Runoff from the designed phyto-cover is computed using the USDA Soil Conservation Service (SCS) curve number model.22 The model computes direct runoff from an individual storm event as a portion of total precipitation (Figure 7.10). The method was developed from field studies by measuring runoff from various soil cover, land slope, and soil type combinations. Curve numbers were developed based upon each of the combined hydrologic effects of these factors and enable the model to be applied to any area within the U.S. The curve number model is widely used and is incorporated into the HELP model and other agronomic models to compute rainfall runoff and other elements comprising a water balance. The major deficiency in this model is that it underestimates runoff from small precipitation events. This discrepancy results in overestimates of infiltration and the amount of water that must be managed by the cover system.3 Consequently, the resultant engineered
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8
Direct Runoff (Q), inches
7 6
Curves on this sheet are for the case Is = 0.2S, so that (P - 0.2s)2 Q= P + 0.8S
0
r be
5
=
10
90
m Nu
ve
80
r Cu
4
70 60
3
50
2
40
1 0 0
1
2
3
4
5
6
7
8
9
10
11
12
Rainfall (P), inches
Figure 7.10
Runoff — SCS Method.
phyto-cover is overdesigned and conservative: the engineered phyto-cover has the ability to control more infiltration than it is designed to manage.15
(P − I a ) Q= (P − I a ) + S 2
(7.4)
where Q P S Ia
= = = =
runoff (in) precipitation (in) potential maximum retention after runoff begins (in) initial abstraction (in)
The initial abstraction is all water loss before runoff begins. It includes water detained in surface depressions, as well as water intercepted by vegetation, evaporation, and infiltration. The initial abstraction is highly variable but from data collected from small agricultural watersheds, Ia was approximated using the following equation: Ia = 0.2 S
(7.5)
By eliminating Ia as an independent parameter, this approximation allows use of a combination of retention storage (S) and precipitation (P) to predict a unique runoff amount. Substituting Equation 7.5 into Equation 7.4 gives 2 P − 0.2 S) ( Q=
P + 0.8 S
(7.6)
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where S is related to the soil and cover conditions of the watershed through the curve number CN. CN has a range from 30 to 100, and is related to S by the following equation: S=
1000 − 10 CN
(7.7)
The use of the SCS runoff equation for this analysis assumes that the difference between precipitation and runoff is infiltration.15 The curve number can be estimated by either using the HELP model or other computations. The HELP model computes a curve number based upon final grade, soil type, and vegetative cover. Using this model is recommended because it objectively estimates a curve number based upon the final design. In addition, the methods in the HELP model were developed and approved by the USEPA. When using the model, the minimum final grade should be specified for the land surface slope and a good vegetative cover be assumed for the understory. These two assumptions ensure that the selected curve number is conservative (minimize runoff/maximize infiltration). 7.8.4
Potential Evapotranspiration — Measured Data
Potential evapotranspiration (PET) is a measure of the maximum rate at which evapotranspiration can occur when adequate soil moisture is available for utilization by vegetation. These data are measured in the field utilizing lysimeters planted with single species covers (usually perennial grasses). Soil moisture levels are maintained at optimum levels and evapotranspiration is measured by weighing the lysimeter. Data collected through these methods are the most reliable and most defendable estimates of potential evapotranspiration; however, measured site-specific data are not readily available for most sites. The monthly potential evapotranspiration rates measured for grasses are adjusted to best represent the supplemental evapotranspiration available from the trees. This step is performed by incorporating a consumptive-use coefficient (Kc) applicable to the trees utilized in the phyto-cover design. A consumptive-use coefficient of 1.35 was measured for areas with cottonwood trees, willows, and grass.15,23 This value is near the low end of the range for consumptive-use coefficient values derived for densely planted trees in other parts of the U.S.; it has been reported that consumptiveuse coefficient values for densely planted trees vary from 1.3 to 1.6.24 The cottonwood is one species used to develop the hybrid poplar trees selected for the phytocover system. Hybrid poplar trees have been developed specifically to achieve a high consumptive use coefficient, in addition to disease resistance and high growth rates; therefore, the selection of a consumptive-use coefficient of 1.35 is conservative for this engineered phyto-cover system. The consumptive-use coefficient is used to adjust the measured potential evapotranspiration for grasses only during the growing season for trees — April through October (Figures 7.11a and b). The growing season begins with approximately 10%
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Potential Evapotranspiration (inches)
10 9 8 7 6 5 4 3 2 1 0 Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
Total 46.18 Inches Figure 7.11a
Measured potential ET for a site with cool wet climatic conditions.
Potential Evapotranspiration (inches)
10 9 8 7 6 5 4 3 2 1 0 Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
Total 55.62 Inches Figure 7.11b
Potential ET of phyto-cover at the same site adjusted for consumptive use.
leafing in April, 50% leafing in May, and 100% leafing in June, July, and August. Thereafter, leaf falloff occurs at a level of 70% in September and 20% in October.15 No leafing is assumed during the dormant season. Factoring these together, a combined consumptive-use coefficient for grasses and trees can be developed and used to compute the monthly potential evapotranspiration rate for trees and grasses for the phyto-cover design. It should be noted again, though, that these PET rates cannot be achieved unless adequate soil moisture is available.
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7.8.5
339
Potential Evapotranspiration — Empirical Data
Modern equations to compute potential ET normalize estimated PET to a reference crop evapotranspiration rate (Erc; mm/day). The reference crop ET rate is defined as the rate of ET from an idealized grass crop with a fixed crop height of 0.12 m, albedo of 0.23, and a surface resistance of 69 m/s. The reference crop closely resembles an extensive surface of short green grass cover of uniform height that is actively growing, completely shading the ground, and not short of water. The normalization of PET rates facilitates the comparison of computed rates from different methods and equations. This also extends the application of crop specific consumptive use coefficients for estimating PET between methods. The potential evapotranspiration from a site can be estimated empirically utilizing a radiation-based equation with the general form Erc = α
∆ (R − G) ∆+γ n
(7.8)
where: Erc α ∆ γ Rn G
= potential ET [mm/day] = constant = gradient of the saturation vapor curve as a function of temperture (kPa/˚C) = psychrometric constant (kPa/˚C) = net radiation exchange for the crop cover (mm/day) = soil heat flux (mm/day)
Each of the constants in Equation 7.5 can be evaluated using average daily temperature, mean elevation of the site above sea level, and atmospheric constants. Substantial evidence supports the application of Equation 7.8 for determining PET for areas with uniform vegetation cover. The following radiation-based equations for estimation of potential evapotranspiration are recommended:15,25 E r c = 1.74
∆ (R − G) ∆+γ n
(7.9)
for arid locations with relative humidity less than 60% in the month having peak evaporation, and E r c = 1.76 for all other humid locations.
∆ (R − G) ∆+γ n
(7.10)
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The variables in Equations 7.9 and 7.10 are evaluated using the following expressions: ∆=
4098e s (237.3 + T)2
(7.11)
where T es
= temperature (˚C) = saturation vapor pressure (kPa) 17.27 T e s = 0.6108 exp (237.3 + T)
γ P
(7.12)
P λ = atmospheric pressure (kPa) estimated from the site elevation using the relationship = 0.0016286
P = 101.0 – 0.011 5E + 5.44 χ10–7 E2 E λ
(7.13)
= land surface elevation (meters) = latent heat of evaporation of water (MJ/kg) λ = 2.501 – 0.002361 T
(7.14)
G = 0.38(Tday2 – Tday1)
(7.15)
(
)
n n R n = 0.25 + 0.50 S0 − 0.9 + 0.1 0.34 − 0.14 e d σ T 4 N N n N S0 ed σ 7.8.6
(7.16)
= = = =
bright sunshine hours per day total day length in hours extraterrestrial radiation (MJ/m2/day) vapor pressure (kPa) = relative humidity (fraction) times the es (computed above) = Stefan-Boltzmann Constant (4.903 × 10–9 MJ/m2/˚K4/day) Effective Evapotranspiration
The actual or effective ET is calculated by adjusting the PET value to account for the reduction in the ET rates as soil moisture is depleted. This adjustment is performed using a standard model of ET as a function of soil moisture.15,26 The model is shown graphically in Figure 7.12. The effective ET rate occurs at the PET
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Figure 7.12
341
Actual vs. potential evapotranspiration.
rate until soil moisture content is at a percentage of field capacity, then declines linearly at soil moisture levels drier than this value until it is approximately zero at the wilting point of the plants. Mathematically, the relationship between soil moisture is expressed as follows: ∂Θ(t ) = −E r ∂t
(7.17)
where Er = PET Er =
Θ( t ) PET Θi
when
Θi ≤ Θ ≤ Θ s
when Θ w ≤ Θ ≤ Θ i
(7.18)
(7.19)
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The application of Equations 7.17 through 7.19 to compute the change in soil moisture between two time intervals requires the integration of these equations. Consider the time interval from 0 to a time t. The soil moisture at time t expressed as a function of the soil moisture at time 0 is as follows: t
Θ(t ) = Θ(0) −
∫ E dt r
(7.20)
0
If the soil is wet (moisture content greater than Θi) the moisture content at time t (Θ2) is simply the initial moisture content (Θ1) minus the PET rate multiplied by the time interval. Θ2 = Θ1 – PET t
(7.21)
If the soil is dry (moisture content less than Θi) the moisture content at time t is the initial moisture content minus the integrated ET rate over the time interval. t
Θ(t ) = Θ(0) −
∫ 0
Θ(t ) PET dt Θi
(7.22)
Equation 7.22 cannot be solved directly because there is no relationship for soil moisture as a function of time. The equation to characterize the change in moisture content when the soil is dry can be derived by substituting Equation 7.19 into Equation 7.17. ∂Θ(t ) Θ( t ) =− PET ∂t Θi
(7.23)
Separating variables and integrating Equation 7.23 between limits, Θ2
∫
Θ1
t
dΘ(t ) PET =− dt Θ(t ) Θi
∫
(7.24)
0
t
Θ
lnΘ Θ2 == 1
PET t Θi 0
(7.25)
Substituting and exponentiation of both sides of Equation 7.22 yields PET Θ2 = exp − t Θ1 Θi
(7.26)
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The moisture content at time t (Θ2) can therefore be computed from the initial moisture content (Θ1) using the formula: PET Θ 2 = Θ1 exp − t Θi
(7.27)
Equations 7.21 and 7.27 describe the change in moisture content with time when the soil is dry or wet, respectively. A third case that needs to be considered is when conditions change between wet and dry during the time period being evaluated (one day for this model). Before Equation 7.21 is evaluated, the time necessary for conditions to change from wet to dry at current PET rates is computed with the following equation: td =
(Θ 1 − Θ i )
PET
(7.28)
If the time is greater than one day (the time period), Equation 7.21 is applied. If the time is less than one day, Equation 7.21 is applied for the computed time (td), and Equation 7.27 is applied for the remaining portion of the time period (1 – td) 7.8.7
Water Balance Model
Water balance models of the soil profile are based upon fundamental principles of the behavior of water in the soil. During a storm event, surficial soils are saturated by precipitation. Initially, water percolates vertically in the profile, redistributing moisture until the remaining water held by surface tension on the soil particles is in equilibrium with gravitational forces causing drainage. The moisture content at which this equilibrium occurs is termed field capacity. Water uptake by plants continues to drive the drying process with little soil moisture restrictions until moisture contents reach 25 to 80% of field capacity. Transpiration rates decrease as the soil continues to dry until the wilting point of the plants is reached. Further declines in soil moisture levels are controlled by the hydraulic conductivity of the soil and occur as a result of evaporative and diffusion processes.25,27 These principles are used to develop a water balance model of an engineered phyto-cover system. The water balance for a phyto-cover system begins with precipitation. It is assumed for this analysis that all precipitation is in the form of rain; this assumption causes conservative design considerations related to cover thickness and total water storage requirements. The soil surface separates precipitation into runoff and infiltration. Runoff is estimated using the curve number model discussed previously. This procedure was selected because it consistently underpredicts runoff volumes and adds an additional degree of conservatism in that infiltration is overestimated. The runoff analysis for water balance assumes that daily precipitation totals correspond to individual storms. After precipitation infiltrates the soil, water is either stored, removed through ET, or, if moisture content is in excess of field capacity, percolated through root zone and into the waste.
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7.9
EXAMPLE APPLICATION
An example application was developed for a hypothetical site in Central Maryland. Precipitation data were assembled from the Conowingo Dam weather station to evaluate local climatic conditions. Average annual precipitation for the 37-year record evaluated is 45.2 inches. A more detailed look at long-term precipitation trends and totals indicates that precipitation can vary widely for a given year, season, or month. From 1960 through 1996 the average annual precipitation varied from a minimum of 26.29 inches in 1965 to a maximum of 62.36 inches in 1996. Similar variability can also be observed in the monthly precipitation totals. To the extent practical, this variability must be accounted for in the design of the phyto-cover system to demonstrate adequate performance under extreme weather conditions. Therefore, the available daily precipitation data for the 37-year period between 1960 and 1996 were assembled in order to design the phyto-cover system. Figure 7.13 shows the daily precipitation data for a three-year period of the total years analyzed. Daily precipitation records for all 37 years are utilized in this analysis in order to account for rainfall runoff and extreme precipitation conditions.
Daily Precipitation (in)
3.00
2.00
1.00
2
62 10 /6/
6/1 6/6
/61
4/6 2 2/2
11 /4
1
7/1 5/6 1
60
3/2 5/6
12 /3/
3/6 0 8/1
3/6 0 4/2
1/1 /
60
0.00
Date Figure 7.13
Daily precipitation data for a three-year period for the example application; this analysis was performed for 37 years of data looked at for the site.
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9.00
0.90
8.00
0.80
7.00
0.70
0.00 0.50 1.00 1.50 2.00 2.50 3.00
5.00 0.50 4.00 0.40 3.00
0.30
Daily ET
62
Daily Storage
/6/
2
Legend
10
2
6/6 6/1
4/6 2/2
1
11
/4/
5/6 7/1
5/6 3/2
/3/ 12
3/6
3/6
8/1
1/1
61
0.00 1
0.00 60
1.00
0
0.10
0
2.00
/60
0.20
Soil Water Storage (in)
6.00
0.60
Daily Precipitation
345
1.00
4/2
Daily Evapotranspiration (in)
ENGINEERED VEGETATIVE LANDFILL COVERS
Date
Figure 7.14
Predicted daily performance of the final engineered phyto-cover design.
Potential ET rates are known for this area of the U.S. from data collected during a 12-year lysimeter study in Seabrook, NJ.15,28 Actual ET rates are computed based upon potential ET rates and soil moisture levels. For this example, the actual and potential ET rates are assumed to be equal until moisture contents fall to less than 25% of soil moisture storage between field capacity and wilting point. At lower moisture contents, ET is assumed to decline linearly to zero at the wilting point. Other soil moisture models support the assumption that ET rates begin to decline when the moisture content is 25% of the difference between field capacity and wilting point (according to EPIC27 and CREAMS29) developed by the U.S.D.A. Therefore, the methodology and assumptions of this water balance analysis are technically defensible and comparable to the EPIC and CREAMS models developed by the USDA. An example of the expected daily performance of the final design of the engineered phyto-cover is shown in Figure 7.14. Figures that simulate daily soil water storage, daily evaportranspiration, and observed precipitation from 1960 through 1996 can be obtained through this analysis. If the total water holding capacity of the designed phyto-cover is more than the highest daily storage predicted, then the phyto-cover can be theoretically expected to prevent percolation. The water mass balance analysis employing the data presented above was used to determine the site-specific performance of various engineered phyto-cover systems with different water holding capacities. This analysis is summarized in Figure 7.15 and shows that the average annual percolate is a function of the total water holding capacity of the cover design. The greater the total water holding capacity of the phyto-cover the less the average annual percolate. Table 7.1 shows the designed water holding capacity of a phyto-cover utilizing existing soil cover, supplemental imported soils, and municipal solid waste. The existing soil cover at the site is assumed to have a thickness of 1.0 foot. The water
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10.00
Existing Conditions (8.5 in/yr)
Standard Single-Barrier Cover Compacted Clay (1.09 in/yr)
1.00
0.10
10.00 inches
Phyto-Cover, Total Water Storage 10 inches (0.24 in/yr)
5.95 inches
Average Annual Flux through Cover (in/yr)
Decreased Leachate Production
100.00
0.01 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 Total Water Storage Capacity (in) Increased Cover Thickness
Figure 7.15
Long-term performance as a function of total water holding capacity.
Table 7.1 Analysis of the Water Holding Capacity (WHC) of the Phyto-Cover “Sponge” Material Additional Imported Soil Cover Existing Landfill Soil Cover Root Growth in Landfill Matrix Total
Thickness Field Wilting (feet) Capacity Point 2.0 1.0 5.0
0.284 0.244 0.292
0.135 0.136 0.077
WHC (in/foot)
Total WHC (inches)
1.79 1.30 2.58
3.58 1.30 12.90 17.58
holding capacity (WHC) for the soil type at the site (silt clay loam) is 1.3 inches per foot.15,17 The supplemental soil cover has been assumed to be a silt loam with a water holding capacity of 1.8 inches per foot for this water balance analysis. Figure 7.16 shows the water holding capacity of different types of soils. The water holding capacity of waste in a mature landfill is between 0.5 and 1.7 inches of water per foot depending on the percentage of municipal solid waste.3 Water holding capacity as high as 4.85 inches of water per foot of waste has been reported. The landfill is assumed to consist primarily of municipal solid waste. The HELP model reports that this type of waste has a water holding capacity of 2.58 in/foot. Poplar trees grow vigorously over landfills, as evidenced by conditions at
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347
Water Content in Inches per Foot of Soil
4 Gravitational Water ts
y
3
cit
lan
a ap
eld
C
dily
r
2
ble
ssi
Fi
te Wa
cce
P to
te
A
a Re
Ra
ed
tR
a nly
d uce
O ible
ss cce
rA
te Wa
1
t
oin gP
tin Wil
Limited Availability
Figure 7.16
Clay
Hv. Clay Loam
Clay Loam
Silt Loam Lt. Clay Loam
Loam
Fine Sandy Loam
Sandy Loam
Fine Sand
Sand
0
Water holding characteristics of soils.
abandoned landfills, and routinely develop roots deeper than 8 feet below the soil surface. Accordingly, the engineered phyto-cover system is designed to root into the waste to capture additional water holding capacity. According to Table 7.1, 17.78 inches is the total water holding capacity with an additional 2 feet of soil cover, the existing average 1.0 feet of soil cover, and root growth 5 feet into the landfill matrix. Therefore, the performance efficiency of the designed phyto-cover is greater than 98%; this level of efficiency favorably compares with the performance efficiency of the RCRA cap over a long period of time. 7.10
SUMMARY OF PHYTO-COVER WATER BALANCE
After the phyto-cover is in place and functioning, the moisture levels in the waste and soils below the root zone and above the water table will fall well below field capacity as gravity and capillary action pull moisture out. These now-drier unsaturated materials will provide a water absorption capacity in addition to that within the root zone to absorb potential infiltration below the root zone during occasional extremely wet periods. As the overlying soils again fall below field capacity with additional evapotranspiration, capillary action will act to remove moisture again from below the root zone and above the water table. Thus relatively small amounts of water moving through the root zone during infrequent periods of extreme precipitation will not proceed directly to the water table, but will be held by the waste and unsaturated soils, and acted on by capillary action and osmotic potential to move back up and into the atmosphere. In contrast, there is no upward gradient for movement of potential leakage from a membrane cap.
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This design for the phyto-cover system incorporates a margin of safety because the water-balance calculations were based on the following conservative input values: • Underestimation of surface-water runoff by the curve number method (overestimation of infiltration) • Assumption of a consistently short growing season • Consumptive-use coefficient lower than known measured values for the type of trees in the design • Assumption of root growth depth shallower than that known to occur with the type of trees in the design • Not accounting for additional storage capacity provided by the very large volume of landfill matrix which will remain below field capacity most of the time, and the subsequent capillary removal of moisture which may, on occasion, extend into this mass
In addition to the use of these conservative design parameter inputs, the engineered phyto-cover system design, unlike the design of RCRA barrier caps, can include the actual measurement of soil moisture as an element of operations and maintenance (O&M), thus ensuring continuous evaluation of the system’s hydraulic performance. This analysis demonstrates that the designed phyto-cover could perform at least as effectively as a conventional landfill cover over a long period of time. In addition, once the phyto-cover is installed, the waste below the root zone will continuously dry up and will encounter a significant loss in moisture content. This continuous drying will provide additional water holding capacity to handle the flux associated with an extreme precipitation event. In summary, the analysis effectively demonstrates that the engineered phyto-cover system presented has the ability to achieve the objectives of conventional landfill covers and to protect human health and the environment, by protecting groundwater quality through minimizing the generation of landfill leachate. In addition, if coupled with monitored natural attenuation of groundwater, the engineered phyto-cover technology provides the optimum balance for achieving all of the stated goals for landfill closure.
7.11
GENERAL PHYTO-COVER MAINTENANCE ACTIVITIES
7.11.1 Site Inspections To ensure the continued proper functioning and integrity of the engineered phytocover system, regularly scheduled inspections should be conducted by on-site personnel and the local expert. In addition, during the early part of the growing season or upon discovery of any unusual operational conditions, inspections by members of the maintenance team should be completed. The local expert should have general knowledge of erosion control, irrigation systems, soil moisture monitoring, and vegetation care. Inspections should be conducted during the growing season on a weekly schedule during the first year of the O&M period, biweekly during the second
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year, and monthly or as needed thereafter throughout the O&M period. Detailed descriptions of the activities specific to each time period are provided in later sections. During the first three years of the maintenance period, additional inspections of the cover system should be conducted following storm events. Inspection observations should be recorded on an inspection log and stored on-site in a threering log binder. Photographic documentation of the growth and maturation of phytocover should also be collected and included with the inspection log.30 The log from each inspection should be reviewed by the maintenance team to evaluate phyto-cover system maturation and recommend any required corrective action. Expert personnel should be consulted, as necessary, to assist in the identification of diseases and pests that may affect the overall maintenance and health of the cover system and, in particular, the hybrid popular trees planted at the site. To determine the condition of the phyto-cover system during routine inspections, the site operator should walk the perimeter and traverse the cover along a significant number of randomly selected, noncontiguous tree rows. Different rows should be selected each week, except when reinspection is necessary. Conditions that the site operator will inspect include, but are not limited to 1) surface disturbances or rutting; 2) gullies, washouts, or other cover disturbances caused by water erosion; 3) settlement/subsidence; 4) wear, such as burrows due to animals; 5) vegetation condition; 6) evidence of malfunctions within the irrigation system; 7) saturated soils and precipitation ponding; 8) insects or pests; 9) tensiometer readings; 10) irrigation totalizer readings; and 11) structural integrity of the perimeter fencing. Immediate restorative actions are required if any of the following conditions are observed during a visual inspection: • Evidence of stressed vegetation, or evidence of animal, insect, disease, or other damage to vegetation • Evidence of saturated soils, ponding water, or excessively dry soils • Erosion that could compromise the integrity of the phyto-cover system • Other conditions that may interfere with the proper performance of the phytocover system
The site operator should document the inspection results and any observed deficiencies on an inspection log, and corrective action should be undertaken and documented. 7.11.2 Soil Moisture Monitoring As the phyto-cover matures, trees will develop roots through the surficial vegetative soil zone and into the underlying waste.30 Initially, the majority of root growth will be laterally along the surficial soil zone (the upper 24 inches). The cover uses root uptake to “de-water” the root zone and underlying waste. By monitoring the soil-moisture content in this zone, the site operator can qualitatively estimate water removal by plant uptake and moisture replenishment through irrigation and natural precipitation. Measurements will be made using two tensiometers placed at selected locations within the phyto-cover. Tensiometers will be placed at the beginning of
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the growth season between trees in the rows, in areas of contrasting moisture content based on visual observation of the site after significant precipitation events. A tensiometer is a simple device that measures the relative moisture of the soil by comparing uptake or loss of water through a ceramic membrane and recording the change in vacuum (Figure 7.17). The tensiometer data will be used along with site-specific precipitation data to determine the rate and frequency of irrigation applications as discussed below. Tensiometers must be removed and stored in accordance with the manufacturer’s instructions prior to the onset of freezing conditions.
Figure 7.17
7.11.2.1
Photograph of a field tensiometer.
Drainage Measurement
While many studies have documented parameters related to phyto-cover performance (soil-moisture content, precipitation, runoff), these measurements by themselves do not address the central issue, namely the actual deep percolation through the cover. In most cases, the collection of soil moisture, runoff, and precipitation data has been performed to meet regulatory performance requirements. Methods utilizing these data to estimate the ability of a cover design to limit the flux of water have to rely on predictive methods and thus have inherent uncertainties.11,12 Methods of determining deep percolation include those based on fixed fractions of annual precipitation, groundwater quality and level changes within the footprint of the landfill, water balance models, environmental tracer models, and lysimetry.11,12 Water-balance lysimetry is the most direct, quantitative method utilized to directly measure deep percolation through engineered phyto-covers. It has not been used universally due to the high cost of installation.
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Water-balance lysimeters are typically soil-filled containers buried flush with the soil surface that allow for some method of collecting drainage. Often these lysimeters can also be used to measure water storage either directly through weighing or by using independent methods (e.g., neutron probe or TDR) to monitor soil water storage. Lysimeters are semipermanent structures that are often expensive to construct and must be designed to account for soil physical properties and site-specific environmental variables. Despite these concerns, it was concluded that lysimeters provide the most reliable means of measuring deep percolation, given adequate surface area and long-term monitoring, with precision in drainage “often better than 1 mm/year.”11,12 Measurement of percolation by drainage lysimetry is made possible by the presence of an impermeable geomembrane forming the bottom boundary. The interruption of downward movement of moisture by the geomembrane causes increased moisture content in the soil layer immediately above the membrane. Drainage occurs when the soil above the membrane liner reaches near-saturation point. This requirement presents a design problem if roots from surface vegetation are allowed to penetrate to the bottom of the lysimeter. When the downward flux of moisture is impeded by an impermeable membrane and plants are allowed access to the trapped moisture, percolation is reduced, which can result in false negatives. This factor can be addressed if the lysimeter is relatively deep and roots are restrained from approaching the bottom liner of the lysimeter. Of all currently available methods, only the use of water-balance lysimeters (over several years, combined with climatic observations, plant community activities, and soil parameters) can provide the data necessary to quantify the performance of phyto-covers and validate and calibrate numerical models used for design purposes.19 Many different configurations of water-balance lysimeters are shown in Figures 7.18a through h. Surface Flow Diversion
10 m
Manhole
Cover Materials: Variable Depth Electronic Measurement of Runoff and Drainage
20
m
tu Na
pe Slo ral
Interim Cover: Variable Depth
Geosynthetic Root Barrier
3 to 5% Slope
60-mil HDPE Liner Geocomposite Drainage Layer French Drain, Sump Pump
Figures 7.18a
Detail of proposed lysimeter test facility design (courtesy of Rock and USEPA, 1999).
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Access Trench
Berm
Drainage
Measurements Drainage Precipitation Soil Moisture Content Soil Moisture Potential
Figures 7.18b
Lysimeters: 2 rows of 5 each Dimensions: 3m wide 3m long 3m deep Construction: concrete walls and floor
Lysimeter facility at INEEL EBTF (courtesy of Rock and USEPA, 1999).
Lysimeter Dimension: 12 m x 18 Diversion Berms
37% slope
Runoff Collection
60-mil HDPE Liner Measurements
Drainage
Soil Moisture Drainage Precipitation Air Temperature Relative Humidity Solar Radiation Wind Speed Runoff Dew Point
Figures 7.18c
Live oak (Atlanta) lysimeter (courtesy of Rock and USEPA, 1999).
7.11.3 General Irrigation Guidelines The phyto-cover system will require intermittent irrigation from approximately May through September, depending on observed weather conditions and year of operation. An existing site precipitation gauge should be utilized to measure precipitation received at the site on a weekly basis at a minimum. If precipitation exceeds
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33% Slope
Cover Materials
Runoff Collection
Drainage Grid and Hypalon Membrane Sand Bedding Measurements
Collection Sump
Soil Moisture Drainage Precipitation Soil Temperature Runoff Vegetation Activities
Figures 7.18d
Drainage Collection
Omega Hills lysimeter facility (courtesy of Rock and USEPA, 1999).
TDR
Cover Soil 40-mil HDPE
Berm
3% Slope Collection Pipe
Drainage Collection
Measurements Soil Moisture
Figures 7.18e
Twentynine Palms lysimeter facility (courtesy of Rock and USEPA, 1999).
0.25 inches in a 24-hour period, or a total of 1 inch in any 1-week period, the irrigation system should be turned off at the backflow prevention valve to prevent overwatering of the phyto-cover system and saturation of the soils. The precipitation gauge information should be augmented by the tensiometer data collected at the phyto-cover. The tensiometer readings should be used along with the tree indicator parameters described later to determine if the precipitation events have provided sufficient soil moisture or if augmentation through irrigation is required. The cover system soils must also be visually inspected prior to reactivation of the irrigation system to determine soil moisture conditions. If either saturated soils or standing water is observed, irrigation of the engineered phyto-cover system should not be conducted until the affected surface is dry, standing water is absent, and the tensiometer readings indicate a reduction in soil moisture to less than the assumed
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3% Slope Cover Materials
Drainage Geocomposite
60-mil VFPE Geomembrane Measurements
Overall Dimensions: 15.2 m long by 9.1 m wide
Drainage Collection
Drainage Precipitation Soil Moisture Irrigation Runoff
Figures 7.18f
Rocky Mountain Arsenal lysimeter facility (courtesy of Rock and USEPA, 1999).
Cover Soils Size: 4.6 m wide x 11.0 m long Instrument Access
20-mil Reinforced Geomembrane
Fiberglass _______
Drainage
15 cm PVC Pipe
Measurements Drainage Precipitation Soil Moisture Soil Temperature Figures 7.18g
Air Temperature Soil Heat Flux Net Radiation Snow Depth
Hill AFB lysimeter facility (courtesy of Rock and USEPA, 1999).
field capacity (+/– 30 centibars). Again, indicator parameters should be the primary method used to determine when to resume irrigation. The irrigation system zone capabilities make it possible to irrigate discrete areas of the cover in the event that significant areal variability in soil moisture is observed. An example of an irrigation system is shown in Figure 7.19.
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Minirhizotron Access Tube
355
Datalogger
Cover Materials
Interim Cover
Soil Water Content Figures 7.18h
Figure 7.19
Soil Water Pressure
Soil Temperature
Detail of a vadose zone monitoring station (courtesy of Rock and USEPA, 1999).
An example of an irrigation system for a phyto-cover.
For the first 3 years, hybrid polar trees planted at the site will require approximately 1 inch of water per week to promote rapid growth and development of a healthy root system. This water can be from natural precipitation or through irrigation. The ideal distribution of any required irrigation water would be in daily events, irrigating at night to reduce evaporative losses. The drip irrigation system will be adjusted by the site operator to assure that trees receive the appropriate amount of water. This is very important because insufficient water will cause trees to wilt and excess watering will stunt growth and possibly cause trees to die from lack of oxygen at the root zone. Should natural precipitation occur coincident with irrigation, the irrigation should continue until either 1 inch of precipitation is received within 7 days, or if any single
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daily precipitation event exceeds 0.25 inches.30 In the event that irrigation is stopped due to the 0.25-inch standard being met, then irrigation should continue again within 2 days of the event, unless soils are visually saturated or tensiometer readings indicate less than 30 centibars vacuum. The irrigation system must be turned off and winterized during the dormant season for trees; generally, this occurs in mid- to late September. Soil-moisture monitoring and inspection of the engineering phyto-cover system should be conducted at the beginning of the next growing year to verify soil moisture content before the drip irrigation system is reactivated. 7.11.4 Tree Evaluation The most important and effective indication of maturation and health of the tree population is the presence of abundant, large leaves and an increase in both stem height and girth. Most woody plants grow intermittently, and it is not unusual to see growth in what are known as “flushes” when environmental factors are favorable, and slower growth when unfavorable conditions exist. The following field evaluation techniques will assist in the evaluation of tree performance. 7.11.4.1
Stem
The woody stems of one-year-old rooted stock poplar trees at planting have an average diameter of less than one half inch at planting. During the first year of growth, it is not unusual for the diameter to double; This increase in girth is referred to as secondary growth. In addition to the increase in girth, the height of the tree may increase by an average of 25% or more. The increase in length is known as primary growth (Figure 7.20). In addition to increases in primary and secondary growth, the stem will form branches to expose as many leaves as possible for the production of energy. It is not unusual for poplar species present at the site to form branches all along the stem during the first year of growth. As primary growth increases in later years, these branches are generally lost through the process of abscission, and replaced by branches higher up the stem at the tree crown. Abscission is a change in plant hormonal conditions that causes the plant to shed leaves and small limbs, while at the same time generating protective materials (cork) to cover the wound or abscission site. A few simple field tests can be completed during the early years to assess stem condition. The simplest method involves bending the top of the stem, which should be supple and flexible and not crack or break. A second test involves scratching bark from the stem to reveal the cambium. This test should be performed if the bending test indicates problems with the meristematic tissue. If the cambium is green, then the meristematic tissue is healthy, and the stem is healthy. To assess the cambium, the stem should be scratched beginning at the top and working down to the base of the tree, as die-back in the stem generally occurs from the extremities back to the roots.
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Figure 7.20
7.11.4.2
357
Healthy stem of a growing poplar tree.
Leaves
The leaves of trees serve to convert raw materials (sunlight, water, CO2, etc.) into sugars and control the loss of water by evapotranspiration. The leaves are composed of several different cell types, but the most important structures affecting performance of the phyto-cover system are the cuticle-covered epidermis and the stomata located on the underside of the leaves. The cuticle is a waxy coating that limits secondary or cuticle transpiration. The stomata are the primary means of transpiration, and account for more than 90% of transpirative losses. Indicators to check include uniformity in leaf color and growth. Poplar tree leaves form from buds on the stem and branches. Healthy poplar leaves form large delta-shaped leaves with a diameter that can grow in excess of six inches at the widest part, with lengths in excess of six inches as well (Figure 7.21a). The leaves are a brilliant green color with variations of darker green also present; They should not contain black or brown areas of discoloration. The leaves are attached to the stem by a flat-shaped petiole. Leaves are the primary indication of stress or disease (Figure 7.21b). They indicate stress from a variety of environmental conditions including drought, overwatering, lack of oxygen or excessive CO2 at the roots, and animal foraging and infestation. Drought conditions are evidenced by withering and, if severe enough, by early abscission, whereas other stresses can cause the leaves to change color and wither on the stem without falling. Infestation and foraging are more readily diagnosed by physical evidence associated with discoloration or the loss of part of an otherwise healthy leaf. Visible evidence of leaf stress must be immediately addressed, as certain conditions can cause rapid declines in health and survivability of the tree.
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Figure 7.21
NATURAL AND ENHANCED REMEDIATION SYSTEMS
(a) Healthy and (b) diseased leaves of a poplar tree.
7.11.5 Agronomic Chemistry Sampling Soil and foliage samples must be collected in August during the first 3 years of operations, and in subsequent years as the need may arise to diagnose growth problems or apparent disease. Nutrient analyses of soils and leaves must be used to determine the fertilizer addition necessary, including organic content analyses (American Society for Testing and Materials [ASTM] D-2974) and pH (ASTM D4972), among others (e.g., nitrogen, phosphorous, etc.). Additional testing must be performed as necessary to diagnose performance problems identified during inspections, and assess the results of any corrective action taken.
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The phyto-cover system must be divided into representative sections by assigning transects perpendicular to the tree rows. A reasonable number of soil subsamples across each transect must be collected and combined into one sample. Soils can be collected within 18 inches of trees, and must be collected from a depth of approximately 12 inches below grade. The samples must be analyzed for pH and essential micronutrients as discussed above. Results should be reviewed and compared to suggested concentrations provided in the environmental soil analysis report supplied by the agronomic testing laboratory to determine necessary fertilizer addition or pH adjustments at the site. Laboratory recommendations for fertilizing deciduous trees should be used to determine the appropriate fertilization rate. In addition to soil samples, samples of leaf tissue must be collected from the same transects. For each transect, a total of 20 leaves (or the minimum number required by the laboratory, if higher) must be collected from 20 trees. It is suggested that samples be collected from healthy, recently matured leaves. These samples must be composited and analyzed for 13 nutrient elements. 7.11.6 Safety and Preventative Maintenance During maintenance, all fertilizers, herbicides, and insecticides must be brought on-site in ready-to-apply concentrations. No mixing of undiluted applications should be permitted, and all chemicals brought on-site must be approved by the site operator and accompanied by appropriate material safety data sheets. All chemicals brought on-site must be documented to have no deleterious effect on the trees. If required by local statute, all herbicides and insecticides must be applied by a licensed contractor. 7.11.7 Repairs and Maintenance Based upon the outcome of the visual observations, the inspection logs must address, as needed, repair and maintenance of security control devices, vegetation, erosion of the phyto-cover system, surficial settlement, and irrigation system. If a problem is identified during the course of site inspections, the situation should be evaluated by the site operator and necessary responses must be initiated in a time frame appropriate to the condition. After the repair, a reinspection must be made and documented on an inspection log. If, upon reinspection, it is determined that additional corrective maintenance is necessary, prompt attention must be given to the deficiency.
7.12
OPERATION AND MAINTENANCE (O&M) SCHEDULE
The O&M associated with the phyto-cover is divided into three distinct phases: year 1 — establishment; years 2 and 3 — active maintenance; and year 4 and succeeding years — passive maintenance. The site operator or assignee must perform maintenance activities, and record observations on the inspection log form. Each maintenance period is discussed in detail in the following sections.
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7.12.1 Year 1 — Establishment Management tasks for year 1— Establishment include the following activities: 1. The inspector(s) will conduct weekly site inspections of the entire cover system during the growing season and after major storm events. Biweekly inspections will be conducted during the dormant season. 2. Periodic mowing will be conducted to maintain the understory health and reduce understory competition with the trees. It is anticipated that monthly mowing between the tree rows during the growing season will be sufficient. Care must be exercised so that irrigation drip lines are not damaged during mowing. Mowing between trees within the rows is not necessary and will result in damage to the drip piping. 3. Herbicides will be applied by an appropriate contractor, as required, to prevent the propagation of undesirable species. The selected herbicide must not have any effect on the poplar trees. 4. Insecticide will be applied by an appropriate contractor, as required, based on inspections for infestations of cottonwood beetles, eastern tent caterpillars, or other pests that pose potential problems. The selected herbicide will have no effect on the trees. 5. Monitoring will include foliar and soil sampling during August of the first year; results will be analyzed for year 2 fertilization requirements. Foliar and soil sampling will be conducted in accordance with the methods of the Agricultural Research Service, USDA, to determine soil fertility status as it affects foliation and plant rooting patterns. 6. Repair or replacement of erosion control and security features will be implemented, as required, by an appropriate contractor. This will include surface application of compost mulch, straw, or shredded yard debris. 7. Irrigation system inspection and maintenance will be conducted, including: • inspecting and maintaining the backflow prevention valve, supply manifold, zone valves, and drip lines; and, • winterization of the drip system, including turning off water supply valving, blowing out the lines with air, and removal of batteries from the zone valves at the end of the irrigation season.
7.12.2
Years 2 and 3 — Active Maintenance
The following tasks will be performed for years 2 and 3 — Active Maintenance for the phyto-cover system: 1. The inspector will conduct biweekly site inspections of the entire cover system, year round and after major storm events. 2. The inspector will contract to apply fertilizer in April, as recommended by agronomic analyses from the year 1 August foliar and soil sampling. 3. The inspector will supply weed control on an as-needed basis. Less emphasis should be required in year 2, and weed control should only be applied if specific stress indicators that can be attributed to weeds are present on the trees. 4. Mowing will take place as required to maintain the understory health. 5. Dead or diseased trees must be removed and replaced with a healthy tree. 6. Insecticides may be required. Spraying leaf surfaces is especially important for an infestation of cottonwood beetle or eastern tent caterpillar, should it occur.
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361
Such insect infestations are normally a more significant problem during the first 2 years of establishment. 7. Monitoring will include August foliar and soil sampling as in year 1 for nutrient content. Results will be analyzed to determine the following year’s spring fertilization requirements. 8. Repair or replacement of erosion control structures will be implemented as required. 9. Irrigation system inspection and maintenance will be conducted, including: • inspecting and maintaining the backflow prevention valve, supply manifold, zone valves, and drip lines; and, • winterization of the drip system, including turning off water supply valving, blowing out the lines with air, and removal of batteries from the zone valves at the end of the irrigation season.
7.12.3 Year 4 — Passive Maintenance Passive maintenance of the phyto-cover system will be conducted beginning in year 4 and continuing for succeeding years. These tasks will include the following activities: 1. Monthly site inspections will be conducted. More frequent inspections will be completed during the spring leafing and fall bud setting to evaluate the tree growth, at the discretion of the site operator. 2. Fertilization will be applied, as required, based on the previous year’s agronomic analyses. 3. Weed control will not be required unless undesirable species present a threat to the trees. 4. Dead or diseased trees must be removed and replaced with a healthy tree. 5. Insecticides will be applied only if necessary. 6. August foliar and soil sampling may not be required, but nutrients may be applied periodically in following years to maintain the best growth at the discretion of the site operator. 7. Repair or replacement of erosion control features will be implemented as required. 8. Irrigation system operations should only be required during periods of unusually dry weather, when precipitation amounts are lower than normal. Inspection and maintenance of the irrigation system will be conducted on an annual basis, and before the system is used after a long period of inactivity, including: • clearing of potential sediment buildup in the drip lines; and, • maintaining the connection of the backflow prevention valve, manifold, zone valves, and drip lines.
Over the life of the phyto-cover, indigenous species will be allowed to invade in several stages of forest community succession, leading to climax culture of a mature forest predominant in the region. This is inherent in the design of the system, leading to a self-sustaining biotic community and providing an alternative to harvesting and replanting. Furthermore, it allows the slower growing, longer lived trees already populating the perimeter of the site to colonize into the poplar stand as the poplar trees reach the end of their life expectancy of 50 or more years.
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
The protection of human health and the environment provided by the phytocover system, as designed, will be equivalent in performance to that of a singlebarrier cover system. If it is determined that the invasion of indigenous species does not maintain the performance at an equal or better level of protection, these other species will not be allowed to establish themselves and supplant the design species (poplars).
7.13
SPECIFIC OPERATIONAL ISSUES
7.13.1 Irrigation System Requirements The irrigation system comprises several components that require routine evaluation and maintenance. The backflow prevention valve will require seasonal frost proofing. Prior to operating the valve in the spring, the bleed valves must be closed, and the unit checked for leaks. The irrigation manifold piping will also require seasonal frost proofing. After closing and frost proofing the backflow prevention valve, the manifold should be evacuated using compressed air. In addition to frost-proofing requirements, the zone control valves require routine maintenance. The valves should be monitored during weekly inspections to ensure proper operation. This can be accomplished by opening and closing the valves with the programmable controller. The valves can be operated by nine-volt batteries. More frequent battery changeouts may be required depending upon the programmed irrigation cycle frequency. 7.13.2 Tree Replacement Should the replacement of trees be required, the optimal planting time would be the end of March or beginning of April. Trees can also be planted in the fall; however, this is not recommended for bare-root stock. One-year-old rooted stock of the same species and preferably from the same supplier should be obtained and planted immediately upon receipt. Trees should be planted in the same location as the tree being replaced. To plant the trees, dig or drill a 6-inch diameter or larger, 18- to 24inch-deep hole and place the root mass directly into the hole. Carefully backfill the hole with the soils removed during excavation or with a good quality organic humus. Compact the soils around the tree carefully by hand, and water each tree with at least ten gallons of water. Place a bamboo stake and protective cover around the tree, and replace the drip irrigation pipe so that the drip emitter is within 6 inches of the tree stem.
REFERENCES 1. USEPA, Final Covers on Hazardous Waste Landfills and Surface Impoundments, Technical Guidance Document, Office of Solid Waste and Emergency Response, Washington, D.C., 1989.
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2. Rock, S. A. and P. G. Soyre, Phytoremediation of hazardous wastes: potential regulatory acceptability, Remed. J., Autumn, 1998. 3. McBean, E. A., F. A. Rovers, and G. J. Farquhar, Solid Waste Landfill Engineering and Design, Prentice Hall, Englewood Cliffs, NJ, 1995. 4. Reinhart, D. R. and T. G. Townsend, Landfill Bioreactor Design and Operation, Lewis Publishers, Boca Raton, FL, 1997. 5. Suter, G. W., R. J. Luxmoore, and E. D. Smith, Compacted soil barriers at abandoned landfill sites are likely to fail in the long term, J. Environ. Qual., 22, 217–226, 1993. 6. Koerner, R. M. and D. E. Daniel, Final Covers for Solid Waste Landfills and Abandoned Dumps, ASCE Press, Reston, VA, 1997. 7. Board, M. and D. Laine, Corralling liner nightmares, MSW Manage., 5, 48–51, 1995. 8. Crozier, F. and T. Walker, How much does your liner leak?, Wastes Manage., 24–26, 1995. 9. Finn, S., Golder Associates, personal communication, 1998. 10. Licht, L., Ecolotree, Inc., personal communications, 1996, 1997, 1998. 11. USEPA, Alternative Cover Assessment Project (ACAP), Phase I Report, prepared by Water Resources Center, Desert Research Institute, August, 1999. 12. Rock, S., personal communications, 1997, 1998, 1999, 2000. 13. Gatliff, E., personal communication, 2000. 14. Brady, N. C., The Nature and Properties of Soils, MacMillan Publishing Company, New York, 1992. 15. Potter, S., ARCADIS G & M, Inc., personal communications, 1997, 1998, 1999, 2000. 16. Bent, Tim, Bridgestone-Firestone, Inc., personal communication, 1997. 17. Schroeder, D. R. et al., The Hydrologic Evaluation of Landfill Performance (HELP) Model: User’s Guide for Version 3, EPA/600/9-94/1689, USEPA Risk Reduction Engineering Laboratory, Cincinnati, OH, 1994. 18. Ecolotree, Inc. and CH2H-Hill, Inc., Ecolotree Cap System; Focused Feasibility Study for Selected Virginia Landfills, April 22, 1996. 19. Khire, M. V., C. H. Benson, and P. J. Bosschur, Water balance modeling of earthen final covers, J. Geotech. Geoenviron. Eng., 123, 744–754, 1997. 20. Anderson, J. E., Soil Plant Cover Systems for Final Closure of Solid Waste Landfills in Arid Regions, in Landfill Capping in the Semi-Arid West: Problems, Perspectives, and Solutions, Environmental Science and Research Foundation, Idaho Falls, ID, 1997, 27–38. 21. ARCADIS G & M, Inc., Proprietary Software PHYTOSOLVE for Phyto Cover Evaluation, 1998. 22. Soil Conservation Service, Urban Hydrology for Small Watersheds, Technical Release 55, 1986. 23. Schultz, E. F., Problems in Applied Hydrology, Water Resource Publications, Fort Collins, CO, 1973. 24. Stewart, B. A. and D. R. Nielson, Irrigation of Agricultural Crops, Agronomy No. 30, American Society of Agronomy and Soul Science Society of America, Madison, WI, 1990. 25. Maidment, D. R., Ed., Handbook of Hydrology, McGraw-Hill, Inc., New York, 1993. 26. Fedder, R. A., P. J. Kowalrik, and H. Zaradyn, Simulation of Field Water Use and Crop Yield, Centre for Agricultural Publishing and Documentation, Wageningen, Netherlands, 1978. 27. Sharpley, A. N. and J. R. Williams, Eds., Erosion/Productivity Impact Calculator: 1. Model Documentation, Technical Bulletin, 1768, U.S. Dept. Agric., Washington, D.C., 1990.
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28. American Society of Civil Engineers (ASCE), Evapotranspiration and Irrigation Water Requirements, ASCE Press, New York, 1990. 29. Knisel, W. G. Ed., CREAMS: A Field Scale Model for Chemicals Runoff and Erosion from Agricultural Management Systems, U.S. Dept. Agric., Conservation Research Report, No. 26, 345, 1980. 30. Hannum, E., ARCADIS G & M, Inc., personal communications, 1999, 2000.
L1282/Appendix A/frame Page 365 Monday, June 18, 2001 9:16 AM
APPENDIX
A
Physical Properties of Some Common Environmental Contaminants
365
A
Compound
154.21 152.20 58.08 56.06 53.06 364.92 178.24 44.05 60.05 102.09 41.05 223.27 71.08 58.08 76.53 114.14 169.23 94.12 17.04 130.19 130.19 93.13 123.15 23.15 202.27
Molecular Weight
7.92 × 10–5 (25°C) 2.8 × 10–4 3.97 × 10–5 (25°C) 4.4 × 10–6 (25°C) 1.10 × 10–4 (25°C) 4.96 × 10–4 6.51 × 10–5 (25°C) 6.61 × 10–5 (25°C) 1.23 × 10–3 (25°C) 3.92 × 10–6 (20°C) 3.46 × 10–6 (25°C) — 3.03 × 10–3 (20°C) 5.00 × 10–6 (25°C) 1.08 × 10–2 (25°C) 3.83 × 10–6 (20°C) 3.89 × 10–10 (25°C) — 2.91 × 10–4 (20°C) 3.88 × 10–4 (25°C) 4.87 × 10–4 (20°C) 0.136 (25°C) 1.25 × 10–6 (25°C) — —
Henry’s Law Constant atm·m3/mol
Physical Properties of Some Common Environmental Contaminants
0.00155 (25°C) 0.0290 (20°C) 266 (25°C) 265 (25°C) 110-115 (25°C) 6 × 10–6 (25°C) 1.95 x 10–4 (25°C) 760 (20.2°C) 11.4 (20°C) 5 (25°C) 73 (20°) — 7 × 10–3 (20°C) 20 (20°C) 360 (25°C) 3.6 (20°C) 6 x 10–5 (20–30°C) Low 10 atm (25.7°C) 4.1 (25°C) 10 (35.2°C) 0.6 (20°C) 0.1 (30°C) — ≈ 0 (20°C)
Vapor Pressure mm Hg.
3.47 (25°C) 3.93 (25°C) Miscible 200,000 (25°C) 80,000 (25°C) 0.011 (25°C) 0.075 (25°C) Miscible Miscible 12% by wt. (20°C) Miscible — 2.155 g/L (30°C) Miscible — 141 g/L 842 (20–30°C) 100 wt. % at (20°C) 531g/L (20°C) 1.8 g/L (20°C) 0.2 wt. % (20°C — 1.3 wt. % (20°C) 3.3 (Room Temp) 600 (20°C)
Solubility mg/L
1.25 3.68 –0.43 –0.28 –1.13 2.61 4.41 Unavailable Unavailable Unavailable 0.34 3.20 Unavailable 0.51 1.68 Unavailable 2.03 Unavailable 0.49 Unavailable Unavailable 1.41 Unavailable Unavailable —
Log Koc
366
Acenaphthene Acenaphthylene Acetone Acrolein Acrylonitrile Aldrin Anthracene Acetaldehyde Acetic Acid Acetic Anhydride Acetonitrile 2-Acetylaminofluorene Acrylamide Allyl Alcohol Allyl Chloride Allyl Glycidyl Ethe 4-Aminobiphenyl 2-Aminopyridine Ammonia n-Amyl Acetate sec-Amyl Acetate Aniline o-Anisidine p-Anisidine Antu
Table A1
L1282/Appendix A/frame Page 366 Monday, June 18, 2001 9:16 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
78.11 84.24 228.30 252.32 252.32 122.12 276.34 252.32 108.14
312.37 290.83 290.83 290.83 173.04 143.01 171.07 390.57 163.83 252.73 249.20 72.11 252.32 126.59 154.21 157.01 129.39 148.91 54.09 58.12
Benzene Benzidine Benzo[a]anthracene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzoic Acid Benzo[ghi]perlene Benzo[a]pyrene Benzyl Alcohol
Benzyl Butyl Phthalate α-Bhc β-Bhc δ-Bhc Bis(2-chloroethoxy) Methane Bis(2-chloroethyl) Ether Bis(2-chloroisopropyl) Ether Bis(2-Ethylhexyl) Phthalate Bromodichloromethane Bromoform 4-Bromophenyl Phenyl Ether 2-Butanone Benzo[e]pyrene Benzyl Chloride Biphenyl Bromobenzene Bromochloromethane Bromotrifluoromethane 1, 3-Butadiene n-Butane
B 0.00548 (25°C) 3.88 × 10–11 (25°C) 8.0 × 10–6 1.2 × 10–5 (20-25°C) 0.00104 7.02 × 10–8 1.4 × 10–7 (25°C) < 2.4 × 10–6 Insufficient vapor pressure data for calculation at 25°C 1.3 × 10–6 (25°C) 5.3 × 10–6 (20°C) 2.3 × 10–7 (20°C) 2.5 × 10–7 (20-25°C) 3.78 × 10–7 1.3 × 10–5 1.1 × 10–4 1.1 × 10–5 (25°C) 2.12 × 10–4 5.6 × 10–4 1.0 × 10–4 4.66 × 10–5 (25°C) 4.84 × 10–7 (25°C) 3.04 × 10–4 (20°C) 4.15 × 10–4 (25°C) 2.4 × 10–3 (25°C) 1.44 × 10–3 (24–25°C) 5.00 × 10–1 (25°C) 6.3 × 10–2 (25°C) 9.30 × 10–1 (25°C)
1800 (25°C) 500 (25°C) 0.014 (25°C) 0.0012 (25°C) 0.00055 (25°C) 3400 (25°C) 0.00026 (25°C) 0.0038 (25°C) 42,900 (25°C)
42.2 (25°C) 2.0 (25°C) 0.24 (25°C) 31.4 (25°C) 81,000 (25°C) 10,200 (25°C) 1700 (20°C) 0.4 (25°C) 4500 (0°C) 3.130 (25°C) No data found 25.57 wt. % (25°C) 0.0038 (25°C) 493 (20°C) 7.5 (25°C) 409 (25°C) 0.129 M (25.0°C) 0.03 wt. % (20°C) 735 (20°C) 61 (20°C)
95.2 (25°C) 0.83 (20°C) 1.1 × 10–7 (25°C) 5 × 10–7 (20°C) 9.59 × 10–11 (25°C) 0.0045 (25°C) 1.01 × 10–10 (25°C) 5.6 × 10–9 (25°C) 1 (58°C) 8.6 × 10–6 (20°C) 2.5 × 10–5 (20°C) 2.8 × 10–7 (20°C) .7 × 10–5 (20°C) 1 (53°C) 1.55 (25°C) 0.85 (20°C) 6.2 × 10–8 (25°C) 50 (20°C) 5.6 (25°C) 0.0015 (20°C) 100 (25.0°C) 5.54 × 10–9 (25°C) 1 (22.0°C) 10–2 (25°C) 4.14 (25°C) 141.07 (24.05°C) 149 (20°C) 2105 (25°C) 1820 (25°C)
.83–2.54 3.279 3.553 3.279 2.06 1.15 1.79 5.0 1.79 2.45 4.94 0.09 5.6 2.28 3.71 2.33 1.43 2.44 2.08 Unavailable
1.92 .60 6.14 5.74 6.64 1.48–2.70 6.89 5.60–6.29 1.98
L1282/Appendix A/frame Page 367 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 367
Compound
76.13 153.82 409.78 409.78
409.78
127.57 112.56 142.59 64.52 106.55 119.38
Carbon Dissulfide Carbon Tetrachloride Chlordane cis-Chlordane
trans-Chlordane
4-Chloroaniline Chlorobenzene p-Chloro-m-cresol Chloroethane 2-Chloroethyl Vinyl Ether Chloroform
0.0133 0.024 (20°C) 4.8 × 10–5 Insufficient vapor pressure data for calculation at 25°C Insufficient vapor pressure data for calculation at 25°C 1.07 × 10–5 (25°C) 0.00445 (25°C) 1.78 × 10–6 0.0085 (25°C) 2.5 × 10–4 0.0032 (25°C)
–1
2.5 × 10 (25°C) 2.36 × 10–6 3.3 × 10–4 (25°C) 1.91 × 10–4 (20°C) — 8.81 × 10–6 (25°C) 1.02 × 10–5 (25°C) 1.20 × 10–5 (25°C) 1.25 × 10–2 (25°C) 1.14 × 10–2 (25°C) 1.17 × 10–2 (25°C) 7.04 × 10–3 (20–22°C)
Henry’s Law Constant atm·m3/mol
0.025(25°C) 11.8 (25°C) No data found 1064 (20°C) 26.75 (20°C) 198 (25°C)
No data found
360 (25°C) 113 (25°C) 1 × 10–5 (25°C) No data found
2230 (25°C) 0.76 (20°C) 15 (25°C) 10 (20°C) — 7.0 (25°C) 13 (20°C) 42 (25°C) 1.03 (25°C) 1.81 (25°C) 2.14 (25°C) 55.5 (25°C)
Vapor Pressure mm Hg.
3.9 g/L (20–25°C) 502 (25°C) 3850 (25°C) 5,740 (20°C) 15,000 (20°C) 9300 (25°C)
No data found
2300 (22°C) 1.160 (25°C) 1.85 (25°C) 0.051 (20–25°C)
222 (25°C) Miscible 5000 (25°C) 0.8 wt. % (20°C) — 74,700 (25°C) 201,000 (20°C) Miscible 1.26 (25.0°C) 309 (25.0°C) 34 (25.0°C) 590 (22°C)
Solubility mg/L
2.42 1.68 2.89 0.51 0.82 1.64
6.0
2.38–2.55 2.35 5.57 6.0
Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable 3.40 2.95 2.83 Unavailable
Log Koc
368
C
56.11 118.18 116.16 116.16 116.16 74.12 74.12 74.12 134.22 134.22 134.22 90.18
Molecular Weight
Physical Properties of Some Common Environmental Contaminants (Continued)
1-Butene Butoxyethano n-butyl Acetate sec-Butyl Acetate tert-Butyl Acetate n-Butyl Alcohol sec-Butyl Alcohol tert-Butyl Alcohol n-Butylbenzene sec-Butylbenzene tert-Butylbenzene n-Butyl Mercaptan
Table A1
L1282/Appendix A/frame Page 368 Monday, June 18, 2001 9:16 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
p, p′ DDD p, p′ DDE p, p′ DDT
D
157.56 123.54 164.38 88.54 350.59 70.09 98.19 84.16 100.16 98.14 82.15 66.10
p-Chloronitrobenzene l-Chloro-1-nitropropane Chloropicrin Chloroprene Chloropyrifos Crotonaldehyde Cycloheptane Cyclohexane Cyclohexanol Cyclohexanone Cyclohexene Cyclopentadiene
320.05 319.03 354.49
70.13 68.12
154.60 188.61
α-Chloroacetophenone o-Chlorobenzylidenemalonitrile
Cyclopentane Cyclopentene
162.62 128.56 204.66 228.30 152.24 201.22 221.26 78.50
2-Chloronaphthalene 2-Chlorophenol 4-Chlorophenyl Phenyl Ether Chrysene Camphor Carbaryl Carbofuran Chloroacetaldehyde
400 (31.0°C) —
1.86 × 10–1 (25°C) 6.3 × 10–2 (25°C)
1.02 × 10–6 (30°C) 6.49 × 10–6 (30°C) 1.9 x 10–7 (25°C)
<1 (20°C) 5.8 (25°C) 23.8 (25°C) 200 (20°C) 1.87 × 10–5 (25°C) 30 (20°C) — 95 (20°C) 1 (20°C) 4 (20°C) 67 (20°C) —
<6.91 × 10–3 (20°C) 1.57 × 10–1 (20–25°C) 8.4 × 10–2 3.20 × 10–2 4.16 × 10–6 (25°C) 1.96 × 10–5 — 1.94 × 10–1 (25°C) 5.74 × 10–6 (25°C) 1.2 × 10–5 (25°C) 4.6 × 10–2 (25°C) —
2.16 × 10–5 2.34 × 10–5 5.2 × 10–5
0.012 (20°C) 3.4 x 10–5 (20°C)
0.017 (25°C) 1.42 (25°C) 0.0027 (25°C) 6.3 x 10-9 (25°C) 0.18 (20°C) 6.578 × 10–6 (25°C) 2 × 10–5 (33°C) 100 (20°C)
— Not applicable reacts with water
6.12 × 10–4 5.6 × 10–7 (25°C) 2.2 × 10–4 7.26 × 10–20 3.00 × 10–5 (20°C) 1.27 × 10–5 (20°C) 3.88 × 10–8 (30–33°C) —
0.160 (24°C) 0.0013 (25°C) 0.0004 (25°C)
0.003 wt. % (20°C) <0.8 wt. % (20°C) 1.621 g/L (25°C) — 2 (25°C) 18.1 wt. % (20°C) 30 (25°C) 58.4 (25°C) 36,000 (20°C) 23,000 (20°C) 213 (25°C) 0.0103 mol/L at room temperature 164 (25°C) 535 (25°C)
6.74 (25°C) 28,000 (25°C) 3.3 (25°C) 0.006 (25°C) 0.12% (20°C) 0.4105 (25°C) 700 (25°C) about 50 wt. %, forms a hemihydrate Miscible Not applicable reacts with water
4.64 6 6.26
Unavailable Unavailable
Unavailable Not applicable reacts with water 2.68 3.34 0.82 — 3.86 Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable
3.93 2.56 3.6 5.39 Unavailable 2.42 2.2 Unavailable
L1282/Appendix A/frame Page 369 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 369
208.28 278.35 147.00 147.00 147.00 253.13 120.91 98.96 98.96 96.94 96.94 163.00 112.99 110.97 110.97 380.91 222.24 122.17 194.19 198.14 184.11 182.14 182.14 390.57
Dibromochloromethane Di-n-butyl Phthalate 1, 2-Dichlorobenzene 1, 3-Dichlorobenzene 1, 4-Dichclorobenzene 3, 3′-Dichlorobenzidine Dichlorodifluoromethane 1, 1-Dichloroethane 1, 2-Dichloroethane 1, 1-Dichloroethylene trans-1, 2-Dichloroethylene 2, 4-Dichloropheno 1, 2-Dichlorophenol cis-1, 3-Dichloropropylene trans-1, 3-Dichloropropylene Dieldrin Diethyl Phthalate
2, 4-Dimethylphenol Dimethyl Phthalate 4, 6-Dinitro-o-cresol 2, 4-Dinitrophenol 2, 4-Dinitrotoluene 2, 6-Dinitrotoluene Di-n-octyl Phthalate
Molecular Weight 278.36 168.20
Compound
76 (20°C) 1.4 × 10-5 (25°C) 1.5 (25°C) 2.3 (25°C) 0.4 (25°C) 1 × 10–5 m/L (22°C) 4.887 (25°C) 234 (25°C) 87 (25°C) 591 (25°C) 410 (30°C) 0.089 (25°C) 50 (25°C) 43 (25°C) 34 (25°C) 1.8 × 10–7 (25°C) 0.22 (±0.7) Pa (25°C) 0.098 (25°C) 0.22 ± 0.7 Pa (25°C) 5.2 × 10–5 (25°C) 0.00039 (20°C) 1.1 × 10–4 (20°C) 3.5 × 10–4 (20°C) 0.0014 mm (25°C)
≈ 10 (20°C) No data found –10
Vapor Pressure mm Hg.
7868 (25°C) 4320 (25°C) 250 (25°C) 6000 (25°C) 270 (22°C) ≈ 300 3 (25°C)
4000 (20°C) 400 (25°C) 145 (25°C) 143 (25°C) 74 (25°C) 3.11 (25°C) 280 (25°C) 5060 (25°C) 8300 (25°C) 5000 (25°C) 6300 (25°C) 4500 (25°C) 2800 (25°C) 2700 (25°C) 2800 (25°C) 0.20 (25°C) 1000 (25°C)
0.00249 (25°C) 10 (25°C)
Solubility mg/L
2.07 1.63 2.64 1.25 1.79 1.79 8.99
1.92 3.14 3.23 3.23 2.2 3.3 2.56 1.48 1.15 1.81 1.77 2.94 1.71 1.68 1.68 4.55 1.84
6.22 3.91–4.10
Log Koc
370
6.55 × 10–6 (25°C) 4.2 × 10–7 1.4 × 10–6 1.57 × 10–8 (18–20°C) 8.67 × 10–7 2.17 × 10–7 1.41 × 10–12 (25°C)
–9
7.33 × 10 Insufficient vapor pressure data for calculation at (25°C) 9.9 × 10–4 6.3 × 10–5 0.0024 (25°C) 0.0047 (25°C) 0.00445 (25°C) 4.5 × 10–8 (25°C) 0.425 (25°C) 0.00587 (25°C) 9.8 × 10–4 (25°C) 0.021 0.00674 (25°C) 6.66 × 10–6 0.00294 (25°C) 0.00355 0.00355 2 × 10–7 8.46 × 10–7
Henry’s Law Constant atm·m3/mol
Physical Properties of Some Common Environmental Contaminants (Continued)
Dibenz[a,h]anthracene Dibenzofuran
Table A1
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
184.24 221.04 138.25 142.28 116.16 235.91 236.36 209.82 197.03
120.91 114.96
220.98 73.14 117.19 203.83 203.83 142.24 101.19 115.18 45.08 225.30 121.18 86.18 86.18 112.22 112.22 73.09 60.10
1, 2-Diphenylhydrazine 2, 4-D Decahydronaphthalene
n-Decane Diacetone Alcohol 1, 4-Dibromobenzene 1, 2-Dibromo-3-chloropropane
Dibromodifluoromethane 1-3-Dichloro-5, 5 Dimethylhydantoin
Dichlorofluoromethane sym-Dichloromethyl Ether
Dichlorvos Diethylamine 2-Diethylaminoethanol 1, 1-Difluorotetrachloroethane 1, 2-Difluorotetrachloroethane Diisobutyl Ketone Diisopropylamine N, N-Dimethylacetamide Dimethylamine p-Dimethylaminoazobenzene Dimethylaniline 2, 2-Dimethylbutane 2, 3 -Dimethylbutane cis-1, 2-Dimethylcyclohexane trans-1, 4-Dimethylcyclohexane Dimethylformamide 1, 1-Dimethylhydrazine
688 (20°C) —
760 (8.9°C) —
0.0527 (25°C) 195 (20°C) 1 (20°C) 40 (19.8°C) 40 (19.8°C) 1.7 (20°C) 60 (20°C) 1.3 (25°C) 1520 (10°C) — 1 (29.5°C) 319.1 (25°C) 234.6 (25°C) 14.5 (25°C) 22.65 (25°C) 3.7 mm (25°C) 157 (25°C)
≈ 2.42 × 10–2 (20–30°C) Not applicable reacts with water 5.0 × 10–3 2.56 × 10–5 (25°C) — — 1.07 × 10–1 (20°C) 6.36 × 10–4 (20°C) — — 1.77 × 10–5 (25°C) — 4.98 × 10–6 (20°C) 1.943 (25°C) 1.18 (25°C) 3.54 × 10–1 (25°C) 8.70 × 10–1 (25°C) — 2.45 × 10–9 (25°C)
1.35 (25°C) 1 (22.0°C) 0.161 (25°C) 0.8 (21°C)
1.87 × 10–1 (25°C) — 5.0 × 10–4 (25°C) 2.49 × 10–4 (20°C) — Not applicable reacts with water
2.6 × 10–5 (25°C) 0.0047 (20°C) 1 (22.5°C)
4.11 × 10–11 (25°C) 1.95 × 10–2 (20°C) 39.2 (25°C)
≈ 1 wt. % (20°C) 815,000 (14°C) Miscible — 0.01 wt. % (20°C) 0.05 wt. % (20°C) Miscible Miscible Miscible 13.6 (20–30°C) 1105.2 (25°C) 21.2 (25°C) 19.1 (25°C) 6.0 (25°C) 3.84 ppm (25°C) Miscible Miscible
1 wt. % (20°C) Decomposes
0.022 (25°C) Miscible 16.5 (25°C) 1000 at room temperature — 0.21 wt. % (25°C)
221 (25°C) 890 ppm (25°C) 0.889 ppm (25°C)
— Not applicable reacts with water 1.57 Not applicable reacts with water 9.57 Unavailable Unavailable — 2.78 Unavailable Unavailable Unavailable Unavailable 3 Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable –0.7
Unavailable Unavailable 3.2 2.11
2.82 1.68 Unavailable
L1282/Appendix A/frame Page 371 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 371
Compound
406.92 406.92 422.92
380.92 380.92 06.17 92.53 323.31 61.08 90.12 132.18 88.11 100.12
α-Endosulfan β-Endosulfan Endosulfan Sulfate
Endrin Endrin Aldehyde Ethylbenzene Epichlorohydrin EPN Ethanolamine 2-Ethoxyethanol 2-Ethoxyethyl Acetate Ethyl Acetate Ethyl Acrylate
E
100.20 100.20 100.20 72.15 157.22 126.13 168.11 168.11 168.11 88.11 233.11 174.34
Molecular Weight
1.01 × 10–4 (25°C) 1.91 × 10–5 (25°C) Insufficient vapor pressure data for calculation 5.0 × 10–7 3.86 × 10–7 (25°C) 0.00868 (25°C) 2.38–2.54 × 10–5 (20°C) — — — 9.07 × 10–7 (20°C) 1.34 × 10–4 (25°C) 1.94–2.59 × 10–3 (20°C)
1.73 (25°C) 3.152 (25°C) 1.84 (25°C) 2.18 (25°C) — 2.96 × 10–6 (20°C) <1.47 × 10–3 (20°C) 2.75 × 10–7 (35°C) 4.79 × 10–7 (35°C) 4.88 × 10–6 (25°C) 1.46 × 10–9 (25–30°C) 24.2 (25°C)
Henry’s Law Constant atm·m3/mol
0.530 (25°C) 0.280 (25°C) 0.117
0.26 (25°C) 0.26 (25°C) 152 (25°C) 60,000 (20°C) — Miscible Miscible 23 wt. % (20°C) 100 ml/L (25°C) .5 wt. % (20°C)
7 × 10–7 (25°C) 2 × 10–7 (25°C) 10 (25.9°C) 13 (20°C) 0.0003 (100°C) <1 (20°C) 4 (20°C) 2 (20°C) 94.5 (25°C) 29.5 (20°C)
5.25 (25°C) 5.50 (25°C) 5.94 (25°C) 33.2 (25°C) 1.795 (25°C) 2.8 wt. % (20°C) 0.015 wt. % (20°C) 0.05 wt. % (20°C) 0.01 wt. % (20°C) Miscible 42 (25°C) 0.008 (25°C)
Solubility mg/L
10–5 (25°C) 10–5 (25°C) No data found
100 (33.3°C) 98.4 (25°C) 82.8 (25°C) 1.287 (25°C) — 0.5 (20°C) <1 (20°C) 8.15 × 10–4 (35°C) 2.25 × 10–4 (35°C) 37 (25°C) 2 × 10–7 (30°C) 0.057 (25°C)
Vapor Pressure mm Hg.
Physical Properties of Some Common Environmental Contaminants (Continued)
3.92 4.43 1.98 1 3.12 Unavailable Unavailable Unavailable Unavailable Unavailable
3.31 3.37 3.37
Unavailable Unavailable Unavailable Unavailable — 0.61 Unavailable 2.18 Unavailable 0.54 2.51 Unavailable
Log Koc
372
2, 3-Dimethylpentane 2, 4 Dimethylpentane 3, 3-Dimethylpentane 2, 2-Dimethylpropane 2, 7-Dimethylquinoline Dimethyl Sulfate 1, 2-Dinitrobenzene 1, 3-Dinitrobenzene 1, 4-Dinitrobenzene Dioxane Diuron n-Dodecane
Table A1
L1282/Appendix A/frame Page 372 Monday, June 18, 2001 9:16 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
H
G
Heptachlor Heptachlor Epoxide Hexachlorobenzene Hexachchlorobutadiene Hexachlorocyclopentadiene Hexachloroethane 2-Hexanone
Glycidol
Fluoranthene Fluorene Formaldehyde Formic Acid Furfural Furfuryl Alcohol
F
Ethylamine Ethyl Bromide Ethylcyclopentane Ethylene Chlorohydrin Ethylenediamine Ethylene Dibromide Ethylenimine Ethyl Ether Ethyl Formate Ethyl Mecaptan 4-Ethylmorpholine 2-Ethylthiophene
373.32 389.32 284.78 260.76 272.77 236.74 100.16
74.08
202.26 166.22 30.03 46.03 96.09 98.10
45.08 108.97 98.19 80.51 60.10 187.86 43.07 74.12 74.08 62.13 115.18 112.19
0.0023 3.2 × 10–5 0.0017 0.026 0.016 0.0025 0.00175 (25°C)
—
0.0169 (25°C) 2.1 × 10–4 3.27 × 10–7 1.67 × 10–7 at pH 4 1.52–3.05 × 10–6 (20°C) —
1.07 × 10–5 25°C) 7.56 × 10–3 (25°C) 2.10 × 10–2 (25°C) — 1.73 × 10–9 (25°C) 7.06 × 10–4 (25°C) 1.33 × 10–7 (25°C) 1.28 × 10–3 (25°C) 2.23 × 10–4 (25°C) 2.74 × 10–3 (25°C) — —
4 × 10–4 (25°C) 2.6 × 10–6 (20°C) 1.089 × 10–5 (20°C) 0.15 (20°C) 0.081 (25°C) 0.8 (30°C) 3.8 (25°C)
0.9 (25°C)
5.0 × 10-6 (25°C) 10 (146°C) 400 (–33°C) 35 (20°C) 2 (20°C) 0.4 (20°C)
400 (2.0°C) 386 (20°C) 40 (25.0°C) 8 (25°C) 10 (21.5°C) 11 (25°C) 250 (30°C) 442 (20°C) 194 (20°C) 527.2 (25°C) 6.1 (20°C) 60.9 (60.3°C)
180 ppb (25°C) 0.350 (25°C) 0.006 (25°C) 3.23 (25°C) 1.8 (25°C) 27.2 (25°C) 35,000 (25°C)
Miscible
0.265 (25°C) 1.98 (25°C) Miscible Miscible 8.3 wt. % (20°C) Miscible
Miscible 0.9 wt. % (20°C) 245 (25°C) Miscible Miscible 3370 Miscible 6.05 wt. % (25°C) 118,000 (25°C) 1.3 wt. % (20°C) Miscible 292 (25°C)
4.34 4.32 3.59 3.67 3.63 3.34 2.13
Unavailable
4.62 3.7 0.56 Unavailable Unavailable Unavailable
Unavailable 2.67 Unavailable Unavailable Unavailable 1.64 0.11 Unavailable Unavailable Unavailable Unavailable Unavailable
L1282/Appendix A/frame Page 373 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 373
Compound
Indeno[1, 2, 3-cd]pyrene Isophorone Indan Indole Indoline 1-Iodopropane Isoamyl Acetate Isoamyl Alcohol Isobutyl Acetate Isobutyl Alcohol Isobutyl Benzene Isopropyl Acetate Isopropylamine Isopropylbenzene Isopropyl Ether
276.34 138.21 118.18 117.15 — 169.99 130.19 88.15 116.16 74.12 134.22 102.13 59.11 120.19 102.18
2.96 × 10–20 (25°C) 5.8 × 10–6 — — — 9.09 × 10–3 5.87 × 10–2 (25°C) 8.89 × 10–6 (20°C) 4.85 × 10–4 (25°C) 9.25 × 10–6 (20°C) 1.09 × 10–2 (25°C) 2.81 × 10–4 (25°C) — 1.47 × 10–2 (25°C) 9.97 × 10–3 (25°C)
2.035 (25°C) 1.44 × 10–4 (25°C) 4.20 × 10–5 (20°C) 4.13 × 10–1 (20°C) 4.22 × 10–1 (25°C) 1.184 (25°C) 4.35 × 10–1 (25°C) 4.38-5.84 × 10–3 (20°C) <2.07 × 10–9 (20–25°C)
Henry’s Law Constant atm·m3/mol
0–10 (25°C) 0.38 (20°C) — — — 43.1 (25°C) 4 (20°C) 2.3 (20°C) 20 (25°C) 10.0 (20°C) 2.06 (25°C) 73 (25°C) 478 (20°C) 4.6 (25°C) 150 (25°C)
45.85 (25°C) 2.6 (20°C) 1.4 (25°C) 48 (25°C) 49 (25°C) 151.5 (25°C) 186.0 (25°C) 4 (20°C) 1 (132.4)
Vapor Pressure mm Hg.
0.062 12,000 (25°C) 88.9 (25°C) 3558 (25°C) 10,800 (25°C) 0.1065 wt. % (23.5°C) 0.2 wt. % (20°C) 26,720 (22°C) 6300 (25°C) 8.7 wt. % (20°C) 33.71 (25°C) 18,000 (20°C) Miscible 48.3 (25°C) 0.65 wt. % (25°C)
2.24 (25°C) 0.43 wt. % (25°C) 14,300 (20°C) 15 (25°C) 15 (25°C) 9.47 (25°C) 50 (25°C) 0.013 wt. % (20°C) 70,000 (25°C)
Solubility mg/L
7.49 1.49 2.48 1.69 1.42 2.16 1.95 Unavailable Unavailable Unavailable 3.9 Unavailable Unavailable 3.45 Unavailable
Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable 0.98
Log Koc
374
I
100.20 114.19 114.19 98.19 98.19 86.18 84.16 144.21 110.11
Molecular Weight
Physical Properties of Some Common Environmental Contaminants (Continued)
n-Heptane 2-Heptanone 3-Heptanone cis-2-Heptene trans-2-Heptene n-Hexane l-Hexene sec-Hexyl Acetate Hydroguinone
Table A1
L1282/Appendix A/frame Page 374 Monday, June 18, 2001 9:16 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
94.94 50.48 84.93 142.20
100.16 108.14 108.14 330.36 98.06
98.14 74.08 86.09 76.10 32.04 31.06 107.16 192.96
Methyl Bromide Methyl Chloride Methylene Chloride 2-Methylnaphthalene
4-Methyl-2-Pentanone 2-Methylphenol 4-Methylphenol Malathion Maleic Anhydride
Mesityl Oxide Methyl Acetate Methyl Acrylate Methylal Methyl Alcohol Methylamine Methylaniline 2-Methylanthracene
290.83
345.66
M
L
490.68
Methoxychlor
Lindane
Kepone
K
4.01 × 10–6 (20°C) 9.09 × 10–5 (25°C) 1.23–1.44 × 10–4 (20°C) 1.73 x 10–4 (25°C) 4.66 × 10–6 (25°C) 1.81 × 10–2 (25°C) 1.19 × 10–5 (25°C) —
Insufficient vapor pressure data for calculation at (25°C) 0.2 0.010 (25°C) 0.00269 (25°C) Insufficient vapor pressure data for calculation 1.49 × 10–5 (25°C) 1.23 × 10–6 (25°C) 7.92 × 10–7 (25°C) 4.89 x 10–9 (25°C) Not applicable reacts with water
4.8 × 10–7
3.11 × 10–2 (25°C)
8.7 (20°C) 235 (25°C) 70 (20°C) 400 (25°C) 127.2 (25°C) 3.1 atm (20°C) <1.0 (20°C) —
15 (20°C) 0.24 (25°C) 0.108 (25°C) 7.95 × 10–6 (25°C) 5 × 10–5 (20°C)
1633 (25°C) 3789 (20°C) 455 (25°C) No data found
No data found
6.7 × 10–5 (25°C)
2.25 (25°C)
3 wt. % (20°C) 240,000 (20°C) 52,000 33 wt. % (20°C) Miscible 9.590 (25°C) 5.624 g/L (25°C) 0.039 (25°C)
1.91 wt. % (25°C) 25,000 (25°C) 23,000 (25°C) 330 (30°C) —
13,000 (25°C) 7400 (25°C) 13,000 (25°C) 25.4 (25°C)
0.1 (25°C)
7.52 (25°C)
2.7 (20–25°C)
0.79 1.34 1.69 2.46 Not applicable reacts with water Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable 5.12
1.92 1.4 0.94 3.93
4.9
3.03
4.74
L1282/Appendix A/frame Page 375 Tuesday, June 19, 2001 1:13 PM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 375
Compound 68.12 72.15 70.13 76.10 118.13 98.19 112.17 96.17 84.16 60.05 114.23 128.21 100.20 100.20 46.07 141.94 57.05 48.10 100.12 128.26 86.18 86.18 84.16 84.16 192.26 58.12 56.11 118.18
Molecular Weight –2
7.7 × 10 (25°C) 1.35 (25°C) 5.35 × 10–1 (25°C) — — 4.35 × 10–1 (25°C) — — 3.62 × 10–1 (25°C) 2.23 × 10–4 (25°C) 3.70 (25°C) 1.30 × 10–4 (20°C) 3.42 (25°C) 1.55–1.64 (25°C) — 5.87 × 10–3 (25°C) 3.89 × 10–4 (20°C) 3.01 × 10–3 (25°C) 2.46 × 10–4 (20°C) 10.27 (25°C) 1.732 (25°C) 1.693 (25°C) 2.77 × 10–1 (25°C) 6.15 × 10–1 (25°C) — 1.171 (25°C) 2.1 × 10–1 (25°C) —
Henry’s Law Constant atm·m3/mol 550.1 (25°C) 687.4 (25°C) 902.1 (25°C) 6 (20°C) 7 (20°C) 46.3 (25°C) ≈1 (20°C) — 137.5 (25°C) 625 (25°C) 19.5 (25°C) 2 (25°C) 65.9 (25°C) 61.6 (25°C) 49.6 (25°C) 405 (25°C) 348 (20°C) 1516 (25°C) 40 (26°C) 7 (25°C) 211.8 (25°C) 189.8 (25°C) 195.4 (25°C) 270.8 (25°C) — 10 atm (66.8°C) 2.270 (25°C) 1.9 (20°C)
Vapor Pressure mm Hg.
Physical Properties of Some Common Environmental Contaminants (Continued)
2-Methyl-1, 3-Butadiene 2-Methylbutane 3-Methyl-1-Butene Methyl Cellosolve Methyl Cellosolve Acetate Methylcyclohexane o-Methylcyclohexanone l-Methylcyclohexene Methylcyclopentane Methyl Formate 3-Methylheptane 5-Methyl-3-Heptanone 2-Methylhexane 3-Methylhexane Methylhydrazine Methyl Iodide Methyl Isocyanate Methyl Mercaptan Methyl Methacrylate 4-Methyloctane 2-Methylpentane 3-Methylpentane 2-Methyl-1-Pentene 4-Methyl-1-Pentene 1-Methylphenanthrene 2-Methylpropane 2-Methylpropene α-Methylstyrene
Table A1
642 (25°C) 49.6 (25°C) 130 (25°C) Miscible Miscible 16.0 (25°C) — 52 (2°C) 41.8 (25°C) 30 wt. % (20°C) 0.792 (25°C) 0.26 wt. % (20°C) 2.54 (25°C) 4.95 (25°C) Miscible 2 wt. % (20°C) 6.7 wt. % (20°C) 23.30 g/L (20°C) 1.5 wt. % (20°C) 0.115 (25°C) 13.8 (25°C) 17.9 (25°C) 78 (25°C) 48 (25°C) 269 ppb (25°C) 48.9 (25°C) 263 (25°C) —
Solubility mg/L
Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable — Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable 1.36 Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable 4.56 Unavailable Unavailable —
Log Koc
L1282/Appendix A/frame Page 376 Monday, June 18, 2001 9:16 AM
376 NATURAL AND ENHANCED REMEDIATION SYSTEMS
128.18 138.13 138.13
138.13 123.11 139.11 139.11 74.09 198.22 130.19
380.79
143.19 143.19 230.90 199.21 75.07 61.04 89.09 89.09 137.14 137.14
4-Nitroaniline Nitrobenzene 2-Nitrophenol 4-Nitrophenol N-Nitrosodimethylamine N-Nitrosodiphenylamine N-Nitrosodi-n-Propylamine
Naled
1-Naphthylamine
2-Naphthylamine
Nitrapyrin 4-Nitrobiphenyl Nitroethane Nitromethane 1-Nitropropane 2-Nitropropane 2-Nitrotoluene 3-Nitrotoluene
N
224.16 87.12
Naphthalene 2-Nitroaniline 3-Nitroaniline
Mevinphos Morpholine
4.66 2.86 8.68 1.23 4.51 5.41
× × × × × ×
— 10–5 (25°C) 10–5 10–5 (25°C) 10–4 (25°C) 10–5 (20°C) 10–5 (20°C)
2.13 × 10–3
2.01 × 10–9 (25°C)
1.27 × 10–10 (25°C)
4.6 × 10–4 9.72 × 10–5 (25°C) Insufficient vapor pressure data for calculation 1.14 × 10–8 (25°C) 2.45 × 10–5 3.5 × 10–6 3.0 × 10–5 (20°C) 0.143 (25°C) 2.33 × 10–8 (25°C) Insufficient vapor pressure data to calculate —
— —
— 45 ml/L (20°C) 22 ml/L (20°C) 1.4 wt. % (20°C) 1.7 wt. % (20°C) 0.06 wt. % (20°C) 0.05 wt. % (20°C)
40
586 (20–30°C)
1700
—
2 × 10–4 (20°C) 6.5 × 10–5 (20–30°C) 2.56 × 10–4 (20–30°C) 0.0028 (20°C) — 15.6 (20°C) 27.8 (20°C) 7.5 (20°C) 12.9 (20°C) 0.15 (20°C) 0.25 (25°C)
800 (18.5°C) 2,000 (25°C) 2,000 (25°C) 16,000 (25°C) Miscible 35.1 (25°C) 9900 (25°C)
30 (25°C) 1260 (25°C) 890 (25°C)
Miscible Miscible
0.0015 (20°C) 0.28 (25°C) 0.20 (25°C) 10-4 (20°C) 8.1 (25°C) No data found No data found
0.23 (25°C) 8.1 (25°C) 1 (119.3°C)
0.003 (20°C) 13.4 (25°C)
2.64 — Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable
2.11
Not applicable reacts with water 3.51
1.08 2.36 1.57 2.33 1.41 2.76 1.01
2.74 1.23–1.62 1.26
Unavailable Unavailable
L1282/Appendix A/frame Page 377 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 377
Pentachlorophenol Phenanthrene Phenol Pyrene Parathion Pentachlorobenzene
PCB-1254 PCB-1260
PCB-1248
P 257.90 192.00 221.00 154–358 with an average value of 261 222–358 with an average value of 288 327 (average) 324–460 370 (average) 266.34 178.24 94.11 202.26 291.27 250.34
403.73 114.23 112.22 90.04
137.14 128.26
Molecular Weight
0.012 (25°C) 0.080 (24°C)
7.71 × 10–5 (25°C) 4.05 × 10–5 (25°C) 1.7 × 10–4 (20°C) 6.80 × 10–4 (25°C) 0.34 (25°C) 6.85 × 10–7 (25°C) 9.8 × 10–6 (25°C) 6.0 × 10–3 (20–30°C)
0.0027 0.0071 3.4 × 10–6 2.56 × 10–5 (25°C) 3.97 × 10–7 (25°C) 1.87 × 10–5 8.56 × 10–8 (25°C) 0.0071 (20°C)
20–25 (25°C) 1.18(25°C) 93,000 (25°C) 0.148 (25°C) 24 (25°C) 2.24 × 10–6 M (25°C)
0.054
0.0035
4.94 × 10–4 (25°C)
— 0.431 (25°C) 2.7 (25°C) 9.81 wt. % (25°C)
0.005 wt. % (20°C) 0.122 (25°C)
Solubility mg/L
0.22–0.25 1.5 (25°C) 1.45 (25°C) 0.24 (25°C)
<1 (20°C) 14.14 (25°C) 17.4 (25°C) <0.001 (20°C)
5.484 (26.0°C) 4.3 (25°C)
Vapor Pressure mm Hg.
4 × 10–4 (25°C) 0.0067 (25°C) 0.0046 (25°C) 4.06 × 10–4 (25°C)
750 3.24 × 10–4 8.64 × 10–4 5.6 × 10–4
— 3.225 (25°C 9.52 × 10–1 (25°C) 1.43 × 10–10 (pH 4)
–5
5.0 × 10 (25°C) 5.95 (25°C)
Henry’s Law Constant atm·m3/mol
2.96 3.72 1.43 4.66 3.68 6.3
5.61 6.42
5.64
4.7 2.44 2.83 3.71
— Unavailable Unavailable 0.89
Unavailable Unavailable
Log Koc
378
PCB-1016 PCB-1221 PCB-1232 PCB-1242
Octachloronaphthalene n-Octane l-Octene Oxalic Acid
O
Compound
Physical Properties of Some Common Environmental Contaminants (Continued)
4-Nitrotoluene n-Nonane
Table A1
L1282/Appendix A/frame Page 378 Monday, June 18, 2001 9:16 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
108.10
105.09 40.06 79.10
n-Propyl Nitrate Propyne Pyridine
p-Quinone
68.12 72.15 86.13 70.13 70.13 70.13 140.28 108.14 170.21 108.14 148.12 229.11 230.25 44.10 72.06 102.12 60.10 120.19 112.22 58.08
l, 4-Pentadiene n-Pentane 2-Pentanone l-Pentene cis-2-Pentene trans-2-Pentene Pentycyclopentane p-Phenylenediamine Phenyl Ether Phenylhydrazine Phthalic Anhydride Picric Acid Pindone Propane β-Propiolactone n-Propyl Acetate n-Propyl Alcohol n-Propylbenzene Propylcyclopentane Propylene Oxide
Q
202.28
Pentachloroethane
9.48 × 10–7 (20°C)
0.1 (20°C)
18 (20°C) 4310 (25°C) 20 (25°C)
734.6 (25°C) 512.8 (25°C) 16 (25°C) 637.7 (25°C) 494.6 (25°C) 505.5 (25°C) — — 0.12 (30°C) <0.1 (20°C) 2 × 10–4 (20°C) <1 (20°C) — 8.6 atm (20°C) 3.4 (25°C) 35 (25°C) 20.8 (25°C) 3.43 (25°C) 12.3 (25°C) 445 (20°C)
1.20 × 10–1 (25°C) 1.255 (25°C) 6.44 × 10–5 (25°C) 4.06 × 10–1 (25°C) 2.25 × 10–1 (25°C) 2.34 × 10–1 (25°C) — — 2.13 × 10–4 (20°C) — 6.29 × 10–9 (20°C) <2.15 × 10–5 (20°C) — 7.06 × 10–1 (25°C) 7.63 × 10–7 (25°C) 1.99 × 10–4 (25°C) 6.74 × 10–6 (25°C) 1.0 × 10–2 (25°C) 8.90 × 10–1 (25°C) 8.34 × 10–5 (20°C) — 1.1 × 10–1 (25°C) 8.88 × 10–6 (25°C)
4.5 (25°C)
2.45 × 10–3 (25°C)
1.5 wt. % (20°C)
— 3640 (20°C) Miscible
7.69 and 500 were reported at 25°C and 20°C 558 (25°C) 39.5 (25°C) 5.51 wt. % (25°C) 148 (25°C) 203 (25°C) 203 (25°C) 0.115 (25°C) 38,000 (24°C) 21 (25°C) — 0.62 wt. % (20°C) 1.4 wt. % (20°C) 18 (25°C) 62.4 (25°C) 37 vol. % (25°C) 18,900 (20°C) Miscible 55 (25°C) 2.04 (25°C) 41 wt. % (20°C)
Unavailable
Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable Unavailable 1.9 — 2.95 Unavailable Unavailable Unavailable Unavailable 2.87 Unavailable Not applicable reacts with water Unavailable Unavailable Unavailable
3.28
L1282/Appendix A/frame Page 379 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 379
215.89
1, 2, 3, 5-Tetrachlorobenzene
104.15 334.42 322.30
321.57
Molecular Weight
321.98 167.85 165.83 92.14 413.82 181.45 133.40 133.40 131.39 137.37 197.45 197.45 255.48 393.70 345.65 215.89
T
S
R
Compound
7.2 × 10–10 (25°C) 6 (25°C) 20 (25°C) 22 (20°C) 0.2–0.4 (25°C) 0.29 (25°C) 124 (25°C) 19 (20°C) 72.6 (25°C) 792 (25°C) 0.022 (25°C) 0.017 (25°C) 6.46 × 10–6 (25°C) — 0.1 (20°C) 2.6 × 10–2 (25°C) 1 (58.2°C)
1.58 × 10–3 (25°C)
6.45 (25°C) — 0.00017 (20°C)
8 × 10–4 (25°C)
Vapor Pressure mm Hg.
5.40 × 10–23 (18–22°C) 4.56 × 10–4 (25°C) 0.0153 0.00674 (25°C) 0.063 0.00232 0.0162 (25°C) 9.09 × 10–4 (25°C) 0.0091 1.73 (25°C) 1.76 × 10–7 (25°C) 9.07 × 10–8 (25°C) 4.87 × 10–8 (25°C) — 6.40 × 10–5 (20°C) 6.9 × 10–3 (20°C)
— 2.88 × 10–6 (20°C)
0.00261
8.46 × 10–6 (25°C)
Henry’s Law Constant atm·m3/mol
Physical Properties of Some Common Environmental Contaminants (Continued)
5.19 (25°C)
0.0193 ppb (22°C) 2970 (25°C) 150 (25°C) 490 (25°C) 0.2–0.4 (25°C) 31.3 (25°C) 950 (25°C) 4,500 (20°C) 1100 (25°C) 1240 (25°C) 1.2 g/L (25°C) 800 (25°C) 278 (25°C) 0.040 0.07 wt. % (20°C) 5.92 (25°C)
0.031 wt. % (25°C) 0.02 wt. % (20°C) 25
40 (25°C)
Solubility mg/L
6.66 2.07 2.42 2.06 3.18 2.7 2.18 1.75 1.81 2.2 2.85 3.03 1.72 4.82 2.45 5.4 average value 6.0 average
2.87 2.45 2.87
2.76
Log Koc
380
TCDD 1, 1, 2, 2-Tetrachloroethane Tetrachloroethylene Toluene Toxaphene 1, 2, 4-Trichlorobenzene 1, 1, 1-Trichloroethane 1, 1, 2-Trichloroethane Trichloroethylene Trichlorofluoromethane 2, 4, 5-Trichlorophenol 2, 4, 6-Trichloropheno 2, 4, 5-T 1, 2, 4, 5-Tetrabromobenzene 1, 1, 2, 2-Tetrabromoethane 1, 2, 3, 4-Tetrachlorobenzene
Styrene Strychnine Sulfotepp
Ronnel
Table A1
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
215.89 290.20
72.11 134.22 196.03 287.15 84.14 269.35 174.15
107.16 314.80 266.32 181.45 181.45 147.43 187.38 368.37 101.19 335.29 120.19 120.19 120.19 126.24 112.22 128.26 114.23 114.23 227.13 326.29
1, 2, 4, 5-Tetrachlorobenzene Tetraethyl Pyrophosphate
Tetrahydrofuran 1, 2, 4, 5-Tetramethylbenzene Tetranitromethane Tetryl Thiophene Thiram 2, 4-Toluene Disocyanate
o-Toluidine 1, 3, 5-Tribromobenzene Tributyl Phosphate 1, 2, 3-Trichlorobenzene 1, 3, 5 Trichlorobenzene 1, 2, 3-Trichloropropane 1, 1, 2-Trichlorotrifluoroethane Tri-o-Cresyl Phosphate Triethylamine Trifluralin 1, 2, 3-Trimethylbenzene 1, 2, 4-Trimethylbenzene 1, 3, 5-Trimethylbenzene 1, 1, 3-Trimethylcyclohexane 1, 1, 3-Trimethylcyclopentane 2, 2, 5-Trimethylhexane 2, 2, 4-Trimethylpentane 2, 3, 4-Trimethylpentane 2, 4, 6-Trinitrotoluene Triphenyl Phosphate
<0.1 (25°C) 1.55 × 10–4 (20°C) 145 (20°C) 0.49 (25°C) 13 (25°C) <1 (20°C) 79.7 (25°C) — 0.01 (20°C)
0.1 (20°C) — — 1 (40°C) 0.58 (25°C) 3.4 (20°C) 270 (20°C) — 54 (20°C) 1.1 × 10–4 (25°C) .51 (25°C) 2.03 (25°C) 2.42 (25°C) — 39.7 (25°C) 16.5 (25°C) 49.3 (25°C) 27.0 (25°C) 4.26 × 10–3 (54.8°C) <0.1 (20°C)
1.0 × 10–2 (20°C) — 7.06 × 10–5 (25°C) 2.49 × 10–2 (25°C) — <1.89 × 10–3 (20°C) 2.93 × 10–3 (25°C) — — 1.88 × 10–6 (25°C) — — 8.9 × 10–3 (20°C) 1.9 × 10–3 (20°C) 3.18 × 10–4 (25°C) 3.33 × 10–1 (20°C) — 4.79 × 10–4 (20°C) 4.84 x 10–5 (23°C) 3.18 x 10–3 (25°C) 5.7 × 10–3 (25°C) 3.93 × 10–3 (25°C) — 1.57 (25°C) 2.42 (25°C) 3.01 (25°C) 2.98 (25°C) — 5.88 × 10–2 (20–25°C)
15,0000 (25°C) 2.51 × 10–6 (25°C) 0.1 wt. % (20°C) 18.0 (25°C) 6.01 (25°C) — 0.02 wt. % (20°C) 3.1 (25°C) 15,000 (20°C) 240 75.2 (25°C) 51.9 (25°C) 48.2 (25°C) 1.77 (25°C) 3.73 (25°C) 1.15 (25°C) 2.05 (25°C) 1.36 (25°C) 0.013 wt. % (20°C) 0.001 wt. % (20°C)
Miscible 3.48 (25°C) — 0.02 wt. % (20°C) 3015 (25°C) 30 Not applicable reacts with water
0.465 (25°C) Miscible
6.1 average Not applicable reacts with water Unavailable 3.79 — 2.37 1.73 — Not applicable reacts with water 2.61 4.05 2.29 3.87 5.7 (average) — 2.59 3.37 Unavailable 3.73 3.34 3.57 3.21 Unavailable Unavailable Unavailable Unavailable Unavailable 2.48 3.72
L1282/Appendix A/frame Page 381 Monday, June 18, 2001 9:16 AM
PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 381
106.17 106.17 106.17
308.33
—
0.00535 (25°C) 0.0063 (25°C) 0.0063 (25°C)
4.81 × 10–4 2.78
Henry’s Law Constant atm·m3/mol
6.6 (25°C) 8.287 (25°C) 8.763 (25°C)
—
115 (25°C) 2660 (25°C)
Vapor Pressure mm Hg.
213 (25°C) 173 (25°C) 200 (25°C)
17 (20°C)
25,000 (25°C) 1100 (25°C)
Solubility mg/L
Sources: Montgomery, J. H. and L. M. Welkom, 1990, Groundwater Chemicals Desk Reference, Lewis Publishers, Chelsea, MI. Montgomery, J. H., 1991, Groundwater Chemicals Desk Reference, Vol. 2, Lewis Publishers, Chelsea, MI.
X
W
86.09 62.50
Molecular Weight
2.11 3.2 2.31
2.96
0.45 0.39
Log Koc
382
o-Xylene m-Xylene p-Xylene
Warfarin
V
Compound
Physical Properties of Some Common Environmental Contaminants (Continued)
Vinyl Acetate Vinyl Chloride
Table A1
L1282/Appendix A/frame Page 382 Monday, June 18, 2001 9:16 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
L1282/Appendix B/frame Page 383 Monday, June 18, 2001 11:29 AM
APPENDIX
B
Useful Information for Biogeochemical Sampling
Table B1
Factors Used to Correct Atmospheric Pressures Adjusted to Sea Level
Elevation of Weather Station (feet above sea level)
Value to Subtract (millimeters of mercury)
0 1000 2000 3000 4000 5000 6000
0 27 53 79 104 128 151
Source: USGS Groundwater Sampling Field Manual (Chapter 6).
383
795
15.3 15.1 14.8 14.6 14.4
14.2 14.1 13.9 13.7 13.5
13.3 13.2 13.0 12.8 12.7
12.5 12.4 12.2 12.1 11.9
0.0 0.5 1.0 1.5 2.0
2.5 3.0 3.5 4.0 4.5
5.0 5.5 6.0 6.5 7.0
7.5 8.0 8.5 9.0 9.5
12.4 12.3 12.1 12.0 11.9
13.3 13.1 12.9 12.8 12.6 12.4 12.2 12.1 11.9 11.8
13.2 13.0 12.8 12.7 12.5
14.1 13.9 13.7 13.5 13.3
15.1 14.9 14.7 14.5 14.3
785
12.3 12.1 12.0 11.8 11.7
13.1 12.9 12.8 12.6 12.4
14.0 13.8 13.6 13.4 13.3
15.0 14.8 14.6 14.4 14.2
780
12.2 12.1 11.9 11.8 11.6
13.0 12.8 12.7 12.5 12.4
13.9 13.4 13.5 13.3 13.2
14.9 14.7 14.5 14.3 14.1
775
12.1 12.0 11.8 11.7 11.6
12.9 12.7 12.6 12.4 12.3
13.8 13.6 13.4 13.3 13.1
14.8 14.6 14.4 14.2 14.0
12.0 11.9 11.8 11.6 11.5
12.8 12.7 12.5 12.3 12.2
13.7 13.5 13.3 13.2 13.0
14.7 14.5 14.3 14.1 13.9
12.0 11.8 11.7 11.5 11.4
12.7 12.6 12.4 12.3 12.1
13.6 13.4 13.3 13.1 12.9
14.6 14.4 14.2 14.0 13.8
11.9 11.7 11.6 11.5 11.3
12.7 12.5 12.3 12.2 12.0
13.5 13.3 13.2 13.0 12.8
14.5 14.3 14.1 13.9 13.7
11.8 11.7 11.5 11.4 11.2
12.6 12.4 12.3 12.1 12.0
13.4 13.3 13.1 12.9 12.7
14.4 14.2 14.0 13.8 13.6
11.7 11.6 11.4 11.3 11.2
12.5 12.3 12.2 12.0 11.9
13.3 13.2 13.0 12.8 12.7
14.3 14.1 13.9 13.7 13.5
11.6 11.5 11.4 11.2 11.1
12.4 12.2 12.1 11.9 11.8
13.3 13.1 12.9 12.7 12.6
14.2 14.0 13.8 13.6 13.4
11.6 11.4 11.3 11.2 11.0
12.3 12.2 12.0 11.9 11.7
13.2 13.0 12.8 12.6 12.5
14.1 13.9 13.7 13.5 13.3
11.5 11.3 11.2 11.1 10.9
12.2 12.1 11.9 11.8 11.6
13.1 12.9 12.7 12.6 12.4
14.0 13.8 13.6 13.4 13.3
11.4 11.3 11.1 11.0 10.9
12.2 12.0 11.8 11.7 11.6
13.0 12.8 12.6 12.5 12.3
13.9 13.7 13.5 13.3 13.2
Atmospheric pressure, in millimeters of mercury 770 765 760 755 750 745 740 735 730 725 720
11.3 11.2 11.1 10.9 10.8
12.1 11.9 11.8 11.6 11.5
12.9 12.7 12.6 12.4 12.2
13.8 13.6 13.4 13.2 13.1
715
11.3 11.1 11.0 10.8 10.7
12.0 11.8 11.7 11.5 11.4
12.8 12.6 12.5 12.3 12.1
13.7 13.5 13.3 13.2 13.0
710
11.2 11.0 10.9 10.8 10.6
11.9 11.7 11.6 11.5 11.3
12.7 12.5 12.4 12.2 12.1
13.6 13.4 13.2 13.1 12.9
705
11.1 11.0 10.8 10.7 10.6
11.8 11.7 11.5 11.4 11.2
12.6 12.5 12.3 12.1 12.0
13.5 13.3 13.2 13.0 12.8
700
11.0 10.9 10.7 10.6 10.5
11.7 11.6 11.4 11.3 11.1
12.5 12.4 12.2 12.0 11.9
13.4 13.2 13.1 12.9 12.7
384
14.2 14.0 13.8 13.6 13.4
15.2 15.0 14.7 14.5 14.3
790
Solubility of Oxygen in Water at Various Temperatures and Pressures (Temp ˚C, temperature in degrees Celsius; atmospheric pressures from 695 to 600 millimeters mercury begin after 40˚C)
Temp ˚C
Table B2
L1282/Appendix B/frame Page 384 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
11.8 11.7 11.5 11.4 11.3
11.1 11.0 10.9 10.8 10.6
10.5 10.4 10.3 10.2 10.1
10.0 9.9 9.8 9.7 9.6 9.5 9.4 9.3 9.2 9.1
10.0 10.5 11.0 11.5 12.0
12.5 13.0 13.5 14.0 14.5
15.0 15.5 16.0 16.5 17.0
17.5 18.0 18.5 19.0 19.5
20.0 20.5 21.0 21.5 22.0
9.4 9.3 9.2 9.2 9.1
9.9 9.8 9.7 9.6 9.5
10.5 10.4 10.2 10.1 10.0
11.1 10.9 10.8 10.7 10.6
11.7 11.6 11.4 11.3 11.2
9.4 9.3 9.2 9.1 9.0
9.9 9.8 9.7 9.6 9.5
10.4 10.3 10.2 10.1 10.0
11.0 10.9 10.7 10.6 10.5
11.6 11.5 11.4 11.2 11.1 10.8 10.7 10.6 10.5 10.4
11.5 11.4 11.2 11.1 11.0 10.8 10.7 10.5 10.4 10.3
11.4 11.3 11.2 11.0 10.9 10.7 10.6 10.5 10.4 10.2
11.3 11.2 11.1 11.0 10.8 10.6 10.5 10.4 10.3 10.2
11.3 11.1 11.0 10.9 10.8 10.6 10.4 10.3 10.2 10.1
11.2 11.1 10.9 10.8 10.7
9.3 9.2 9.1 9.0 9.0
9.8 9.7 9.6 9.5 9.4 9.3 9.2 9.1 9.0 8.9
9.7 9.6 9.5 9.4 9.3 9.2 9.1 9.0 8.9 8.8
9.7 9.6 9.5 9.4 9.3 9.1 9.0 8.9 8.9 8.8
9.6 9.5 9.4 9.3 9.2 9.1 9.0 8.9 8.8 8.7
9.5 9.4 9.3 9.3 9.2 9.0 8.9 8.8 8.7 8.7
9.5 9.4 9.3 9.2 9.1
10.3 10.3 10.2 10.1 10.1 10.0 10.2 10.2 10.1 10.0 10.0 9.9 10.1 10.0 10.0 9.9 9.8 9.8 10.0 9.9 9.9 9.8 9.7 9.7 9.9 9.8 9.8 9.7 9.6 9.6
10.9 10.8 10.7 10.6 10.4
11.6 11.4 11.3 11.2 11.0
8.9 8.9 8.8 8.7 8.6
9.4 9.3 9.2 9.1 9.0
9.9 9.8 9.7 9.6 9.5
10.5 10.4 10.3 10.1 10.0
11.1 11.0 10.9 10.7 10.6
8.9 8.8 8.7 8.6 8.5
9.3 9.3 9.2 9.1 9.0
9.9 9.8 9.7 9.5 9.4
10.4 10.3 10.2 10.1 10.0
11.0 10.9 10.8 10.7 10.5
10.9 10.8 10.6 10.5 10.4
10.8 10.7 10.6 10.4 10.3
10.7 10.6 10.5 10.4 10.3
10.7 10.5 10.4 10.3 10.2
10.6 10.5 10.3 10.2 10.1
8.8 8.7 8.6 8.6 8.5
9.3 9.2 9.1 9.0 8.9
9.8 9.7 9.6 9.5 9.4
8.8 8.7 8.6 8.5 8.4
9.2 9.1 9.0 8.9 8.9
9.7 9.6 9.5 9.4 9.3
8.7 8.6 8.5 8.4 8.4
9.2 9.1 9.0 8.9 8.8
9.7 9.6 9.5 9.4 9.3
8.6 8.6 8.5 8.4 8.3
9.1 9.0 8.9 8.8 8.7
9.6 9.5 9.4 9.3 9.2
8.6 8.5 8.4 8.3 8.2
9.0 8.9 8.8 8.8 8.7
9.5 9.4 9.3 9.2 9.1
8.5 8.4 8.4 8.3 8.2
9.0 8.9 8.8 8.7 8.6
9.5 9.4 9.3 9.2 9.1
10.4 10.3 10.2 10.1 10.1 10.0 10.2 10.2 10.1 10.0 10.0 9.9 10.1 10.1 10.0 9.9 9.8 9.8 10.0 9.9 9.9 9.8 9.7 9.7 9.9 9.8 9.8 9.7 9.6 9.6
11.0 10.8 10.7 10.6 10.5
8.5 8.4 8.3 8.2 8.1
8.9 8.8 8.7 8.6 8.5
9.4 9.3 9.2 9.1 9.0
9.9 9.8 9.7 9.6 9.5
10.5 10.4 10.3 10.2 10.0
8.4 8.3 8.2 8.1 8.1
8.8 8.7 8.7 8.6 8.5
9.3 9.2 9.1 9.0 8.9
9.9 9.7 9.6 9.5 9.4
10.4 10.3 10.2 10.1 10.0
8.3 8.3 8.2 8.1 8.0
8.8 8.7 8.6 8.5 8.4
9.3 9.2 9.1 9.0 8.9
9.8 9.7 9.6 9.5 9.4
10.4 10.2 10.1 10.0 9.9
L1282/Appendix B/frame Page 385 Monday, June 18, 2001 11:29 AM
USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 385
795 9.0 9.0 8.9 8.8 8.7 8.6 8.5 8.5 8.4 8.3 8.2 8.2 8.1 8.0 8.0 7.9 7.8 7.8 7.7 7.6
22.5 23.0 23.5 24.0 24.5
25.0 25.5 26.0 26.5 27.0
27.5 28.0 28.5 29.0 29.5
30.0 30.5 31.0 31.5 32.0
7.8 7.8 7.7 7.6 7.6
8.2 8.1 8.0 8.0 7.9 7.8 7.7 7.7 7.6 7.5
8.1 8.1 8.0 7.9 7.9
8.5 8.4 8.4 8.3 8.2
8.9 8.8 8.8 8.7 8.6
785
7.7 7.7 7.6 7.5 7.5
8.1 8.0 7.9 7.9 7.8
8.5 8.4 8.3 8.2 8.2
8.9 8.8 8.7 8.6 8.5
780
7.7 7.6 7.6 7.5 7.4
8.0 8.0 7.9 7.8 7.8
8.4 8.3 8.3 8.2 8.1
8.8 8.7 8.6 8.6 8.5
775
7.6 7.6 7.5 7.5 7.4
8.0 7.9 7.8 7.8 7.7
8.3 8.3 8.2 8.1 8.0
8.8 8.7 8.6 8.5 8.4
7.6 7.5 7.5 7.4 7.3
7.9 7.9 7.8 7.7 7.6
8.3 8.2 8.1 8.1 8.0
8.7 8.6 8.5 8.4 8.4
7.5 7.5 7.4 7.4 7.3
7.9 7.8 7.7 7.7 7.6
8.2 8.2 8.1 8.0 7.9
8.6 8.6 8.5 8.4 8.3
7.5 7.4 7.4 7.3 7.2
7.8 7.7 7.7 7.6 7.5
8.2 8.1 8.0 8.0 7.9
8.6 8.5 8.4 8.3 8.3
7.4 7.4 7.3 7.3 7.2
7.8 7.7 7.6 7.6 7.5
8.1 8.0 8.0 7.9 7.8
8.5 8.4 8.4 8.3 8.2
7.4 7.3 7.3 7.2 7.1
7.7 7.6 7.6 7.5 7.4
8.1 8.0 7.9 7.8 7.8
8.5 8.4 8.3 8.2 8.1
7.3 7.3 7.2 7.1 7.1
7.7 7.6 7.5 7.5 7.4
8.0 7.9 7.9 7.8 7.7
8.4 8.3 8.2 8.2 8.1
7.3 7.2 7.1 7.1 7.0
7.6 7.5 7.5 7.4 7.3
8.0 7.9 7.8 7.7 7.7
8.3 8.3 8.2 8.1 8.0
7.2 7.2 7.1 7.0 7.0
7.5 7.5 7.4 7.3 7.3
7.9 7.8 7.8 7.7 7.6
8.3 8.2 8.1 8.0 8.0
7.2 7.1 7.0 7.0 6.9
7.5 7.4 7.4 7.3 7.2
7.8 7.8 7.7 7.6 7.6
8.2 8.1 8.1 8.0 7.9
Atmospheric pressure, in millimeters of mercury 770 765 760 755 750 745 740 735 730 725
7.1 7.1 7.0 6.9 6.9
7.4 7.4 7.3 7.2 7.2
7.8 7.7 7.6 7.6 7.5
8.2 8.1 8.0 7.9 7.9
720
7.1 7.0 6.9 6.9 6.8
7.4 7.3 7.3 7.2 7.1
7.7 7.7 7.6 7.5 7.5
8.1 8.0 8.0 7.9 7.8
715
7.0 7.0 6.9 6.8 6.8
7.3 7.3 7.2 7.1 7.1
7.7 7.6 7.5 7.5 7.4
8.0 8.0 7.9 7.8 7.7
710
7.0 6.9 6.8 6.8 6.7
7.3 7.2 7.1 7.1 7.0
7.6 7.6 7.5 7.4 7.3
8.0 7.9 7.8 7.8 7.7
705
6.9 6.9 6.8 6.7 6.7
7.2 7.2 7.1 7.0 7.0
7.6 7.5 7.4 7.4 7.3
7.9 7.9 7.8 7.7 7.6
700
386
8.6 8.5 8.4 8.3 8.3
9.0 8.9 8.8 8.7 8.7
790
Solubility of Oxygen in Water at Various Temperatures and Pressures (Temp ˚C, temperature in degrees Celsius; atmospheric pressures from 695 to 600 millimeters mercury begin after 40˚C) (Continued)
Temp ˚C
Table B2
L1282/Appendix B/frame Page 386 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
7.6 7.5 7.4 7.4 7.3 7.3 7.2 7.2 7.1 7.0 7.0 6.9 6.9 6.8 6.8 6.7
32.5 33.0 33.5 34.0 34.5
35.0 35.5 36.0 36.5 37.0
37.5 38.0 38.5 39.0 39.5
40.0
6.7
6.9 6.9 6.8 6.8 6.7
7.2 7.2 7.1 7.0 7.0
7.5 7.5 7.4 7.3 7.3
6.6
6.9 6.8 6.8 6.7 6.7
7.2 7.1 7.1 7.0 6.9
7.5 7.4 7.3 7.3 7.2
6.6
6.8 6.8 6.7 6.7 6.6
7.1 7.1 7.0 7.0 6.9
7.4 7.4 7.3 7.2 7.2
6.5
6.8 6.7 6.7 6.6 6.6
7.1 7.0 7.0 6.9 6.9
7.4 7.3 7.2 7.2 7.1
6.5
6.8 6.7 6.6 6.6 6.5
7.0 7.0 6.9 6.9 6.8
7.3 7.3 7.2 7.1 7.1
6.4
6.7 6.7 6.6 6.5 6.5
7.0 6.9 6.9 6.8 6.8
7.3 7.2 7.1 7.1 7.0
6.4
6.7 6.6 6.6 6.5 6.5
6.9 6.9 6.8 6.8 6.8
7.2 7.2 7.1 7.0 7.0
6.4
6.6 6.6 6.5 6.5 6.4
6.9 6.8 6.8 6.7 6.7
7.2 7.1 7.1 7.0 6.9
6.3
6.6 6.5 6.5 6.4 6.4
6.8 6.8 6.7 6.7 6.6
7.1 7.1 7.0 6.9 6.9
6.3
6.5 6.5 6.4 6.4 6.3
6.8 6.7 6.7 6.6 6.6
7.1 7.0 7.0 6.9 6.8
6.2
6.5 6.4 6.4 6.3 6.3
6.7 6.7 6.6 6.6 6.5
7.0 7.0 6.9 6.8 6.8
6.2
6.4 6.4 6.3 6.3 6.2
6.7 6.6 6.6 6.5 6.5
7.0 6.9 6.9 6.8 6.7
6.1
6.4 6.3 6.3 6.2 6.2
6.6 6.6 6.5 6.5 6.4
6.9 6.9 6.8 6.7 6.7
6.1
6.3 6.3 6.2 6.2 6.1
6.6 6.5 6.5 6.4 6.4
6.9 6.8 6.8 6.7 6.6
6.0
6.3 6.2 6.2 6.1 6.1
6.5 6.5 6.4 6.4 6.3
6.8 6.8 6.7 6.7 6.6
6.0
6.2 6.2 6.1 6.1 6.0
6.5 6.4 6.4 6.3 6.3
6.8 6.7 6.7 6.6 6.5
5.9
6.2 6.1 6.1 6.0 6.0
6.4 6.4 6.3 6.3 6.2
6.7 6.7 6.6 6.6 6.5
5.9
6.1 6.1 6.0 6.0 6.0
6.4 6.3 6.3 6.2 6.2
6.7 6.6 6.6 6.5 6.5
5.9
6.1 6.0 6.0 6.0 5.9
6.3 6.3 6.2 6.2 6.1
6.6 6.6 6.5 6.5 6.4
L1282/Appendix B/frame Page 387 Monday, June 18, 2001 11:29 AM
USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 387
695
13.3 13.1 13.0 12.8 12.6
12.4 12.3 12.1 12.0 11.8
11.6 11.5 11.4 11.2 11.1
10.9 10.8 10.7 10.5 10.4
0.0 0.5 1.0 1.5 2.0
2.5 3.0 3.5 4.0 4.5
5.0 5.5 6.0 6.5 7.0
7.5 8.0 8.5 9.0 9.5
10.9 10.7 10.6 10.5 10.3
11.6 11.4 11.3 11.1 11.0 10.8 10.6 10.5 10.4 10.3
11.5 11.3 11.2 11.0 10.9
12.3 12.1 11.9 11.8 11.9
13.1 13.0 12.8 12.6 12.4
685
10.7 10.6 10.4 10.3 10.2
11.4 11.2 11.1 11.0 10.8
12.2 12.0 11.8 11.7 11.5
13.0 12.9 12.7 12.5 12.3
680
10.6 10.5 10.4 10.2 10.1
11.3 11.2 11.0 10.9 10.7
12.1 11.9 11.8 11.6 11.5
12.9 12.8 12.6 12.4 12.2
675
10.5 10.4 10.3 10.2 10.0
11.2 11.1 10.9 10.8 10.7
12.0 11.8 11.7 11.5 11.4
12.8 12.7 12.5 12.3 12.2
10.5 10.3 10.2 10.1 10.0
11.1 11.0 10.9 10.7 10.6
11.9 11.7 11.6 11.4 11.3
12.8 12.6 12.4 12.2 12.1
11.0 10.8 10.7 10.6 10.4
11.7 11.6 11.4 11.3 11.1
12.6 12.4 12.2 12.0 11.9
10.9 10.7 10.6 10.5 10.3
11.6 11.5 11.3 11.2 11.0
12.5 12.3 12.1 12.0 11.8
10.8 10.7 10.5 10.4 10.3
11.5 11.4 11.2 11.1 10.9
12.4 12.2 12.0 11.9 11.7
10.7 10.6 10.4 10.3 10.2
11.4 11.3 11.1 11.0 10.9
12.3 12.1 11.9 11.8 11.6
10.6 10.5 10.4 10.2 10.1
11.4 11.2 11.1 10.9 10.8
12.2 12.0 11.8 11.7 11.5
10.4 10.3 10.2 10.1 10.1 10.0 10.2 10.2 10.1 10.0 9.9 9.9 10.1 10.0 10.0 9.9 9.8 9.7 10.0 9.9 9.8 9.8 9.7 9.6 9.9 9.8 9.7 9.7 9.6 9.5
11.1 10.9 10.8 10.6 10.5
11.8 11.7 11.5 11.3 11.2
12.7 12.5 12.3 12.1 12.0
9.9 9.8 9.7 9.5 9.4
10.5 10.4 10.3 10.1 10.0
11.3 11.1 11.0 10.8 10.7
12.1 11.9 11.7 11.6 11.4
9.8 9.7 9.6 9.5 9.4
10.5 10.3 10.2 10.1 9.9
11.2 11.0 10.9 10.7 10.6
12.0 11.8 11.6 11.5 11.3
Atmospheric pressure, in millimeters of mercury 670 665 660 655 650 645 640 635 630 625 620
615
11.0 10.9 10.7 10.6 10.4
11.8 11.6 11.5 11.3 11.1
610
10.9 10.8 10.6 10.5 10.3
11.7 11.5 11.4 11.2 11.1
605
10.8 10.7 10.5 10.4 10.3
11.6 11.4 11.3 11.1 11.0
600
10.7 10.6 10.4 10.3 10.2
11.5 11.3 11.2 11.0 10.9
9.7 9.6 9.5 9.4 9.3
9.7 9.5 9.4 9.3 9.2
9.6 9.5 9.3 9.2 9.1
9.5 9.4 9.3 9.2 9.0
9.4 9.3 9.2 9.1 9.0
10.4 10.3 10.2 10.1 10.0 10.2 10.2 10.1 10.0 9.9 10.1 10.0 9.9 9.9 9.8 10.0 9.9 9.8 9.7 9.7 9.9 9.8 9.7 9.6 9.5
11.1 10.9 10.8 10.7 10.5
11.9 11.7 11.6 11.4 11.2
388
12.4 12.2 12.0 11.9 11.7
13.2 13.1 12.9 12.7 12.5
690
Solubility of Oxygen in Water at Various Temperatures and Pressures (Temp ˚C, temperature in degrees Celsius; atmospheric pressures from 695 to 600 millimeters mercury begin after 40˚C) (Continued)
Temp ˚C
Table B2
L1282/Appendix B/frame Page 388 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
10.3 10.2 10.1 9.9 9.8 9.7 9.6 9.5 9.4 9.3 9.2 9.1 9.0 8.9 8.8 8.7 8.6 8.5 8.4 8.4 8.3 8.2 8.1 8.0 8.0
10.0 10.5 11.0 11.5 12.0
12.5 13.0 13.5 14.0 14.5
15.0 15.5 16.0 16.5 17.0
17.5 18.0 18.5 19.0 19.5
20.0 20.5 21.0 21.5 22.0
8.2 8.1 8.1 8.0 7.9
8.6 8.6 8.5 8.4 8.3
9.1 9.0 8.9 8.8 8.7
9.6 9.5 9.4 9.3 9.2
8.2 8.1 8.0 7.9 7.8
8.6 8.5 8.4 8.3 8.2
9.1 9.0 8.9 8.8 8.7
9.6 9.5 9.4 9.3 9.2
8.1 8.0 7.9 7.9 7.8
8.5 8.4 8.3 8.3 8.2
9.0 8.9 8.8 8.7 8.6
9.5 9.4 9.3 9.2 9.1
8.0 7.9 7.9 7.8 7.7
8.5 8.4 8.3 8.2 8.1
8.9 8.8 8.7 8.6 8.5
9.4 9.3 9.2 9.1 9.0
10.2 10.1 10.1 10.0 10.1 10.0 9.9 9.9 10.0 9.9 9.8 9.8 9.9 9.8 9.7 9.6 9.8 9.7 9.6 9.5
8.0 7.9 7.8 7.7 7.7
8.4 8.3 8.2 8.1 8.0
8.8 8.8 8.7 8.6 8.5
9.4 9.3 9.1 9.0 8.9
9.9 9.8 9.7 9.6 9.5
7.9 7.8 7.8 7.7 7.6
8.3 8.2 8.2 8.1 8.0
8.8 8.7 8.6 8.5 8.4
9.3 9.2 9.1 9.0 8.9
9.8 9.7 9.6 9.5 9.4
7.8 7.8 7.7 7.6 7.5
8.3 8.2 8.1 8.0 7.9
8.7 8.6 8.5 8.4 8.3
9.2 9.1 9.0 8.9 8.8
9.8 9.7 9.5 9.4 9.3
7.8 7.7 7.6 7.6 7.5
8.2 8.1 8.0 7.9 7.9
8.6 8.6 8.5 8.4 8.3
9.1 9.0 8.9 8.8 8.7
9.7 9.6 9.5 9.4 9.2
7.7 7.6 7.6 7.5 7.4
8.1 8.0 8.0 7.9 7.8
8.6 8.5 8.4 8.3 8.2
9.1 9.0 8.9 8.8 8.7
9.6 9.5 9.4 9.3 9.2
7.7 7.6 7.5 7.4 7.4
8.1 8.0 7.9 7.8 7.7
8.5 8.4 8.3 8.2 8.2
9.0 8.9 8.8 8.7 8.6
9.5 9.4 9.3 9.2 9.1
7.6 7.5 7.5 7.4 7.3
8.0 7.9 7.8 7.8 7.7
8.4 8.4 8.3 8.2 8.1
8.9 8.8 8.7 8.6 8.5
9.5 9.4 9.2 9.1 9.0
7.5 7.5 7.4 7.3 7.2
7.9 7.9 7.8 7.7 7.6
8.4 8.3 8.2 8.1 8.0
8.9 8.8 8.7 8.6 8.5
9.4 9.3 9.2 9.1 9.0
7.5 7.4 7.3 7.3 7.2
7.9 7.8 7.7 7.6 7.6
8.3 8.2 8.1 8.0 8.0
8.8 8.7 8.6 8.5 8.4
9.3 9.2 9.1 9.0 8.9
7.4 7.3 7.3 7.2 7.1
7.8 7.7 7.7 7.6 7.5
8.2 8.2 8.1 8.0 7.9
8.7 8.6 8.5 8.4 8.3
9.2 9.1 9.0 8.9 8.8
7.4 7.3 7.2 7.1 7.1
7.7 7.7 7.6 7.5 7.4
8.2 8.1 8.0 7.9 7.8
8.6 8.5 8.5 8.4 8.3
9.2 9.1 9.0 8.8 8.7
7.3 7.2 7.2 7.1 7.0
7.7 7.6 7.5 7.4 7.4
8.1 8.0 7.9 7.8 7.8
8.6 8.5 8.4 8.3 8.2
9.1 9.0 8.9 8.8 8.7
7.2 7.2 7.1 7.0 7.0
7.6 7.5 7.5 7.4 7.3
8.0 8.0 7.9 7.8 7.7
8.5 8.4 8.3 8.2 8.1
9.0 8.9 8.8 8.7 8.6
7.2 7.1 7.0 7.0 6.9
7.6 7.5 7.4 7.3 7.2
8.0 7.9 7.8 7.7 7.6
8.4 8.3 8.2 8.2 8.1
8.9 8.8 8.7 8.6 8.5
7.1 7.0 7.0 6.9 6.8
7.5 7.4 7.3 7.3 7.2
7.9 7.8 7.7 7.7 7.6
8.4 8.3 8.2 8.1 8.0
8.9 8.8 8.7 8.6 8.5
L1282/Appendix B/frame Page 389 Monday, June 18, 2001 11:29 AM
USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 389
695 7.9 7.8 7.7 7.7 7.6 7.5 7.4 7.4 7.3 7.2 7.2 7.1 7.0 7.0 6.9 6.9 6.8 6.7 6.7 6.6
22.5 23.0 23.5 24.0 24.5
25.0 25.5 26.0 26.5 27.0
27.5 28.0 28.5 29.0 29.5
30.0 30.5 31.0 31.5 32.0
6.8 6.7 6.7 6.6 6.6
7.1 7.1 7.0 6.9 6.9 6.8 6.7 6.6 6.6 6.5
7.1 7.0 6.9 6.9 6.8
7.4 7.3 7.3 7.2 7.1
7.8 7.7 7.6 7.5 7.5
685
6.7 6.6 6.6 6.5 6.5
7.0 6.9 6.9 6.8 6.8
7.3 7.3 7.2 7.1 7.1
7.7 7.6 7.6 7.5 7.4
680
6.7 6.6 6.5 6.5 6.4
7.0 6.9 6.8 6.8 6.7
7.3 7.2 7.2 7.1 7.0
7.6 7.6 7.5 7.4 7.4
675
6.6 6.5 6.5 6.4 6.4
6.9 6.8 6.8 6.7 6.7
7.2 7.2 7.1 7.0 7.0
7.6 7.5 7.4 7.4 7.3
6.5 6.5 6.4 6.4 6.3
6.8 6.8 6.7 6.7 6.6
7.2 7.1 7.0 7.0 6.9
7.5 7.5 7.4 7.3 7.2
6.5 6.4 6.4 6.3 6.3
6.8 6.7 6.7 6.6 6.6
7.1 7.1 7.0 6.9 6.9
7.5 7.4 7.3 7.3 7.2
6.4 6.4 6.3 6.3 6.2
6.7 6.7 6.6 6.6 6.5
7.1 7.0 6.9 6.9 6.8
7.4 7.3 7.3 7.2 7.1
6.4 6.3 6.3 6.2 6.2
6.7 6.6 6.6 6.5 6.5
7.0 6.9 6.9 6.8 6.7
7.3 7.3 7.2 7.1 7.1
6.3 6.3 6.2 6.2 6.1
6.6 6.6 6.5 6.5 6.4
6.9 6.9 6.8 6.8 6.7
7.3 7.2 7.2 7.1 7.0
6.3 6.2 6.2 6.1 6.1
6.6 6.5 6.5 6.4 6.3
6.9 6.8 6.8 6.7 6.6
7.2 7.2 7.1 7.0 7.0
6.2 6.2 6.1 6.1 6.0
6.5 6.5 6.4 6.4 6.3
6.8 6.8 6.7 6.6 6.6
7.2 7.1 7.0 7.0 6.9
6.2 6.1 6.1 6.0 6.0
6.5 6.4 6.4 6.3 6.2
6.8 6.7 6.7 6.6 6.5
7.1 7.0 7.0 6.9 6.8
6.1 6.1 6.0 6.0 5.9
6.4 6.4 6.3 6.2 6.2
6.7 6.7 6.6 6.5 6.5
7.1 7.0 6.9 6.9 6.8
Atmospheric pressure, in millimeters of mercury 670 665 660 655 650 645 640 635 630 625
6.1 6.0 6.0 5.9 5.9
6.4 6.3 6.2 6.2 6.1
6.7 6.6 6.5 6.5 6.4
7.0 6.9 6.9 6.8 6.7
620
6.0 6.0 5.9 5.9 5.8
6.3 6.3 6.2 6.1 6.1
6.6 6.6 6.5 6.4 6.4
6.9 6.9 6.8 6.7 6.7
615
6.0 5.9 5.9 5.8 5.8
6.3 6.2 6.1 6.1 6.0
6.6 6.5 6.4 6.4 6.3
6.9 6.8 6.7 6.7 6.6
610
5.9 5.9 5.8 5.8 5.7
6.2 6.1 6.1 6.0 6.0
6.5 6.4 6.4 6.3 6.3
6.8 6.8 6.7 6.6 6.6
605
5.9 5.8 5.8 5.7 5.7
6.2 6.1 6.0 6.0 5.9
6.4 6.4 6.3 6.3 6.2
6.8 6.7 6.6 6.6 6.5
600
390
7.5 7.4 7.3 7.2 7.2
7.8 7.7 7.7 7.6 7.5
690
Solubility of Oxygen in Water at Various Temperatures and Pressures (Temp ˚C, temperature in degrees Celsius; atmospheric pressures from 695 to 600 millimeters mercury begin after 40˚C) (Continued)
Temp ˚C
Table B2
L1282/Appendix B/frame Page 390 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
6.3 6.2 6.2 6.1 6.1 6.0 6.0 6.0 5.9 5.9 5.8
35.0 35.5 36.0 36.5 37.0
37.5 38.0 38.5 39.0 39.5
40.0
5.8
6.0 6.0 5.9 5.9 5.8
6.3 6.2 6.1 6.1 6.1
6.5 6.5 6.4 6.4 6.3
5.7
6.0 5.9 5.9 5.8 5.8
6.2 6.2 6.1 6.1 6.0
6.5 6.4 6.4 6.3 6.3
5.7
5.9 5.9 5.8 5.8 5.7
6.2 6.1 6.1 6.0 6.0
6.4 6.4 6.3 6.3 6.2
5.6
5.9 5.8 5.8 5.7 5.7
6.1 6.1 6.0 6.0 5.9
6.4 6.3 6.3 6.2 6.2
5.6
5.8 5.8 5.7 5.7 5.6
6.1 6.0 6.0 5.9 5.9
6.3 6.3 6.2 6.2 6.1
5.5
5.8 5.7 5.7 5.6 5.6
6.0 6.0 5.9 5.9 5.8
6.3 6.2 6.2 6.1 6.1
5.5
5.7 5.7 5.6 5.6 5.5
6.0 5.9 5.9 5.8 5.8
6.2 6.2 6.1 6.1 6.0
5.4
5.7 5.6 5.6 5.5 5.5
5.9 5.9 5.8 5.8 5.7
6.2 6.1 6.1 6.0 6.0
Source: USGS Groundwater Sampling Field Manual (Chapter 6).
6.6 6.5 6.5 6.4 6.4
32.5 33.0 33.5 34.0 34.5
5.4
5.6 5.6 5.5 5.5 5.4
5.9 5.8 5.8 5.7 5.7
6.1 6.1 6.0 6.0 5.9
5.4
5.6 5.5 5.5 5.4 5.4
5.8 5.8 5.7 5.7 5.6
6.1 6.0 6.0 5.9 5.9
5.3
5.5 5.5 5.4 5.4 5.4
5.8 5.7 5.7 5.6 5.6
6.0 6.0 5.9 5.9 5.8
5.3
5.5 5.4 5.4 5.4 5.3
5.7 5.7 5.6 5.6 5.5
6.0 5.9 5.9 5.8 5.8
5.2
5.4 5.4 5.4 5.3 5.3
5.7 5.6 5.6 5.5 5.5
5.9 5.9 5.8 5.8 5.7
5.2
5.4 5.3 5.3 5.3 5.2
5.6 5.6 5.5 5.5 5.4
5.9 5.8 5.8 5.7 5.7
5.1
5.3 5.3 5.3 5.2 5.2
5.6 5.5 5.5 5.4 5.4
5.8 5.8 5.7 5.7 5.6
5.1
5.3 5.3 5.2 5.2 5.1
5.5 5.5 5.4 5.4 5.3
5.8 5.7 5.7 5.6 5.6
5.0
5.3 5.2 5.2 5.1 5.1
5.5 5.4 5.4 5.3 5.3
5.7 5.7 5.6 5.6 5.5
5.0
5.2 5.2 5.1 5.1 5.0
5.4 5.4 5.3 5.3 5.3
5.7 5.6 5.6 5.5 5.5
5.0
5.2 5.1 5.1 5.0 5.0
5.4 5.3 5.3 5.2 5.2
5.6 5.6 5.5 5.5 5.4
L1282/Appendix B/frame Page 391 Monday, June 18, 2001 11:29 AM
USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 391
0
1.000 1.000 1.000 1.000 1.000
1.000 1.000 1.000 1.000 1.000
1.000 1.000 1.000 1.000 1.000
1.000 1.000 1.000 1.000 1.000
0.0 1.0 2.0 3.0 4.0
5.0 6.0 7.0 8.0 9.0
10.0 11.0 12.0 13.0 14.0
15.0 16.0 17.0 18.0 19.0
0.997 0.997 0.997 0.997 0.997
0.996 0.996 0.997 0.997 0.997 0.993 0.993 0.993 0.993 0.993
0.993 0.993 0.993 0.993 0.993
0.993 0.993 0.993 0.993 0.993
0.992 0.992 0.992 0.993 0.993
2000
0.990 0.990 0.990 0.990 0.990
0.989 0.989 0.989 0.990 0.990
0.989 0.989 0.989 0.989 0.989
0.989 0.989 0.989 0.989 0.989
3000
0.986 0.986 0.986 0.987 0.987
0.986 0.986 0.986 0.986 0.986
0.985 0.985 0.985 0.986 0.986
0.985 0.985 0.985 0.985 0.985
0.983 0.983 0.983 0.983 0.983
0.982 0.982 0.982 0.983 0.983
0.981 0.982 0.982 0.982 0.982
0.981 0.981 0.981 0.981 0.981
0.979 0.979 0.980 0.980 0.980
0.979 0.979 0.979 0.979 0.979
0.978 0.978 0.978 0.978 0.978
0.977 0.977 0.977 0.977 0.978
0.976 0.976 0.976 0.976 0.976
0.975 0.975 0.975 0.975 0.976
0.974 0.974 0.974 0.975 0.975
0.973 0.973 0.973 0.974 0.974
0.972 0.972 0.973 0.973 0.973
0.971 0.971 0.972 0.972 0.972
0.970 0.970 0.971 0.971 0.971
0.969 0.969 0.970 0.970 0.970
0.969 0.969 0.969 0.969 0.970
0.968 0.968 0.968 0.968 0.969
0.966 0.967 0.967 0.967 0.967
0.965 0.965 0.966 0.966 0.966
0.965 0.966 0.966 0.966 0.966
0.964 0.964 0.965 0.965 0.965
0.963 0.963 0.963 0.963 0.964
0.961 0.962 0.962 0.962 0.962
0.962 0.962 0.962 0.963 0.963
0.960 0.961 0.961 0.961 0.961
0.959 0.959 0.959 0.960 0.960
0.957 0.958 0.958 0.958 0.959
0.958 0.958 0.959 0.959 0.959
0.957 0.957 0.957 0.958 0.958
0.955 0.955 0.956 0.956 0.956
0.953 0.954 0.954 0.954 0.955
0.955 0.955 0.955 0.956 0.956
0.953 0.953 0.954 0.954 0.954
0.951 0.952 0.952 0.952 0.953
0.950 0.950 0.950 0.951 0.951
0.951 0.951 0.952 0.952 0.952
0.949 0.950 0.950 0.950 0.951
0.947 0.948 0.948 0.949 0.949
0.946 0.946 0.946 0.947 0.947
0.947 0.948 0.948 0.949 0.949
0.946 0.946 0.946 0.947 0.947
0.944 0.944 0.944 0.945 0.945
0.942 0.942 0.942 0.943 0.943
0.944 0.944 0.945 0.945 0.945
0.942 0.942 0.943 0.943 0.943
0.940 0.940 0.941 0.941 0.941
0.938 0.938 0.938 0.939 0.939
Conductivity, in microsiemens per centimeter at 25°C 4000 5000 6000 7000 8000 9000 10000 11000 12000 13000 14000 15000 16000
392
0.996 0.996 0.996 0.996 0.996
0.996 0.996 0.996 0.996 0.996
1000
Salinity Correction Factors for Dissolved Oxygen in Water (based on conductivity) (Temp °C, temperature in degrees Celsius; salinity correction factors at 30 to 35°C are shown at the end of this table)
Temp ˚C
Table B3
L1282/Appendix B/frame Page 392 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
0.934 0.934 0.935 0.935 0.935
0.936 0.936 0.937 0.937 0.937
0.938 0.939 0.939 0.939 0.940
0.0 1.0 2.0 3.0 4.0
5.0 6.0 7.0 8.0 9.0
10.0 11.0 12.0 13.0 14.0
0.934 0.935 0.935 0.936 0.936
0.932 0.933 0.933 0.933 0.934
0.930 0.930 0.931 0.931 0.932
0.931 0.931 0.932 0.932 0.933
0.928 0.929 0.929 0.930 0.930
0.926 0.926 0.927 0.927 0.928
0.927 0.927 0.928 0.928 0.929
0.924 0.925 0.925 0.926 0.926
0.922 0.922 0.923 0.923 0.924
0.990 0.990 0.991 0.991 0.991
0.923 0.924 0.924 0.925 0.925
0.920 0.921 0.922 0.922 0.922
0.918 0.918 0.919 0.919 0.920
0.987 0.987 0.987 0.987 0.988
0.919 0.920 0.920 0.921 0.922
0.917 0.917 0.918 0.918 0.919
0.914 0.914 0.915 0.915 0.916
0.984 0.984 0.984 0.984 0.984
0.983 0.984 0.984 0.984 0.984
0.916 0.916 0.917 0.917 0.918
0.913 0.913 0.914 0.914 0.915
0.910 0.910 0.911 0.911 0.912
0.981 0.981 0.981 0.981 0.981
0.980 0.980 0.980 0.980 0.981
0.912 0.912 0.913 0.914 0.914
0.909 0.909 0.910 0.911 0.911
0.905 0.906 0.907 0.907 0.908
0.977 0.978 0.978 0.978 0.978
0.977 0.977 0.977 0.977 0.977
0.908 0.909 0.909 0.910 0.911
0.905 0.905 0.906 0.907 0.907
0.901 0.902 0.903 0.903 0.904
0.974 0.974 0.975 0.975 0.975
0.973 0.973 0.974 0.974 0.974
0.904 0.905 0.906 0.906 0.907
0.901 0.902 0.902 0.903 0.904
0.897 0.898 0.899 0.899 0.900
0.971 0.971 0.971 0.972 0.972
0.970 0.970 0.970 0.971 0.971
0.900 0.901 0.903 0.902 0.903
0.897 0.898 0.898 0.899 0.900
0.893 0.894 0.895 0.895 0.896
0.968 0.968 0.968 0.968 0.969
0.966 0.967 0.967 0.967 0.967
0.897 0.897 0.898 0.899 0.899
0.893 0.894 0.894 0.895 0.896
0.889 0.890 0.891 0.891 0.892
0.964 0.965 0.965 0.965 0.965
0.963 0.963 0.964 0.964 0.964
0.893 0.894 0.894 0.895 0.896
0.889 0.890 0.891 0.891 0.892
0.885 0.886 0.887 0.887 0.888
0.961 0.961 0.962 0.962 0.962
0.960 0.960 0.960 0.960 0.961
0.889 0.890 0.890 0.891 0.892
0.885 0.886 0.887 0.887 0.888
0.881 0.882 0.883 0.883 0.884
0.958 0.958 0.958 0.959 0.959
0.956 0.957 0.957 0.957 0.957
0.885 0.886 0.887 0.887 0.888
0.881 0.882 0.883 0.884 0.884
0.877 0.878 0.879 0.879 0.880
0.954 0.955 0.955 0.955 0.956
0.953 0.953 0.953 0.954 0.954
0.881 0.882 0.883 0.884 0.884
0.877 0.878 0.879 0.880 0.880
0.873 0.874 0.875 0.875 0.876
0.951 0.951 0.952 0.952 0.952
0.949 0.950 0.950 0.950 0.951
0.877 0.878 0.879 0.880 0.881
0.873 0.874 0.875 0.876 0.877
0.869 0.870 0.871 0.871 0.872
0.948 0.948 0.948 0.949 0.949
0.946 0.946 0.947 0.947 0.947
Conductivity, in microsiemens per centimeter at 25°C 17000 18000 19000 20000 21000 22000 23000 24000 25000 26000 27000 28000 29000 30000 31000 32000 33000
0.994 0.994 0.994 0.994 0.994
0.987 0.987 0.987 0.987 0.987
Temp ˚C
0.997 0.997 0.997 0.997 0.997
0.990 0.990 0.990 0.990 0.990
1.000 1.000 1.000 1.000 1.000
0.993 0.993 0.993 0.994 0.994
25.0 26.0 27.0 28.0 29.0
0.997 0.997 0.997 0.997 0.997
1.000 1.000 1.000 1.000 1.000
20.0 21.0 22.0 23.0 24.0
L1282/Appendix B/frame Page 393 Monday, June 18, 2001 11:29 AM
USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 393
0.0 1.0 2.0 3.0 4.0
0.865 0.866 0.867 0.867 0.868
0.861 0.862 0.862 0.863 0.864
0.856 0.857 0.858 0.859 0.860
0.852 0.853 0.854 0.855 0.856
0.934 0.935 0.935 0.936 0.936
0.848 0.849 0.850 0.851 0.852
0.931 0.931 0.932 0.932 0.933
0.844 0.845 0.846 0.847 0.848
0.927 0.928 0.928 0.929 0.929
0.840 0.841 0.842 0.843 0.844
0.924 0.925 0.925 0.926 0.926
0.921 0.922 0.922 0.923 0.923
0.836 0.837 0.838 0.839 0.840
0.921 0.921 0.922 0.922 0.923
0.918 0.918 0.919 0.919 0.920
0.832 0.833 0.834 0.835 0.836
0.917 0.918 0.918 0.919 0.919
0.914 0.915 0.915 0.916 0.917
0.828 0.829 0.830 0.831 0.832
0.914 0.914 0.915 0.915 0.916
0.911 0.911 0.912 0.912 0.913
0.823 0.825 0.826 0.827 0.828
0.910 0.911 0.911 0.912 0.913
0.907 0.908 0.908 0.909 0.910
0.904 0.904 0.905 0.906 0.906
0.819 0.821 0.822 0.823 0.824
0.907 0.907 0.908 0.909 0.909
0.903 0.904 0.905 0.905 0.906
0.900 0.901 0.901 0.902 0.903
0.815 0.816 0.818 0.819 0.820
0.903 0.904 0.905 0.905 0.906
0.900 0.901 0.901 0.902 0.903
0.896 0.897 0.898 0.899 0.899
0.811 0.812 0.814 0.815 0.816
0.900 0.901 0.901 0.902 0.903
0.896 0.897 0.898 0.898 0.899
0.893 0.893 0.894 0.895 0.896
0.807 0.808 0.809 0.811 0.812
0.896 0.897 0.898 0.898 0.899
0.893 0.893 0.894 0.895 0.896
0.889 0.890 0.891 0.891 0.892
0.803 0.804 0.805 0.807 0.808
0.893 0.894 0.894 0.895 0.896
0.889 0.890 0.891 0.891 0.892
0.885 0.886 0.887 0.888 0.888
0.799 0.800 0.801 0.803 0.804
0.890 0.890 0.891 0.892 0.892
0.886 0.886 0.887 0.888 0.889
0.882 0.882 0.883 0.884 0.885
Conductivity, in microsiemens per centimeter at 25°C 34000 35000 36000 37000 38000 39000 40000 41000 42000 43000 44000 45000 46000 47000 48000 49000 50000
0.938 0.938 0.938 0.939 0.939
0.925 0.925 0.926 0.926 0.927
0.908 0.909 0.909 0.910 0.911
Temp ˚C
0.941 0.941 0.942 0.942 0.943
0.928 0.929 0.929 0.930 0.930
0.911 0.912 0.912 0.913 0.914
0.944 0.945 0.945 0.946 0.946
0.932 0.932 0.933 0.933 0.934
0.915 0.915 0.916 0.917 0.917
25.0 26.0 27.0 28.0 29.0
0.935 0.936 0.936 0.937 0.937
0.918 0.919 0.920 0.920 0.921
394
0.939 0.939 0.940 0.940 0.941
0.922 0.923 0.923 0.924 0.924
0.942 0.943 0.943 0.944 0.944
0.926 0.926 0.927 0.927 0.928
20.0 21.0 22.0 23.0 24.0
0.929 0.930 0.930 0.931 0.931
0.940 0.941 0.941 0.942 0.942
15.0 16.0 17.0 18.0 19.0
0.933 0.934 0.934 0.934 0.935
Conductivity, in microsiemens per centimeter at 25°C 17000 18000 19000 20000 21000 22000 23000 24000 25000 26000 27000 28000 29000 30000 31000 32000 33000
Temp ˚C 0.937 0.937 0.938 0.938 0.938
Salinity Correction Factors for Dissolved Oxygen in Water (based on conductivity) (Temp °C, temperature in degrees Celsius; salinity correction factors at 30 to 35°C are shown at the end of this table) (Continued)
Table B3
L1282/Appendix B/frame Page 394 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
0.869 0.870 0.871 0.872 0.873
0.874 0.874 0.875 0.876 0.877
0.878 0.879 0.879 0.880 0.881
0.882 0.883 0.884 0.884 0.885
0.886 0.887 0.887 0.888 0.889
5.0 6.0 7.0 8.0 9.0
10.0 11.0 12.0 13.0 14.0
15.0 16.0 17.0 18.0 19.0
20.0 21.0 22.0 23.0 24.0
25.0 26.0 27.0 28.0 29.0
0.882 0.883 0.884 0.885 0.886
0.878 0.879 0.880 0.881 0.882
0.874 0.875 0.876 0.877 0.877
0.870 0.871 0.871 0.872 0.873
0.865 0.866 0.867 0.868 0.869
0.879 0.880 0.880 0.881 0.882
0.875 0.876 0.876 0.877 0.878
0.870 0.871 0.872 0.873 0.874
0.866 0.867 0.868 0.869 0.869
0.861 0.862 0.863 0.864 0.865
0.875 0.876 0.877 0.878 0.879
0.871 0.872 0.873 0.874 0.874
0.867 0.867 0.868 0.869 0.870
0.862 0.863 0.864 0.865 0.866
0.857 0.858 0.859 0.860 0.861
0.872 0.873 0.874 0.874 0.875
0.867 0.868 0.869 0.870 0.871
0.863 0.864 0.865 0.866 0.867
0.858 0.859 0.860 0.861 0.862
0.853 0.854 0.855 0.856 0.857
0.868 0.869 0.870 0.871 0.872
0.864 0.865 0.866 0.866 0.867
0.859 0.860 0.861 0.862 0.863
0.854 0.855 0.856 0.857 0.858
0.849 0.850 0.851 0.852 0.853
0.865 0.866 0.867 0.867 0.868
0.860 0.861 0.862 0.863 0.864
0.855 0.856 0.857 0.858 0.859
0.850 0.851 0.852 0.853 0.854
0.845 0.846 0.847 0.848 0.849
0.861 0.862 0.863 0.864 0.865
0.856 0.857 0.858 0.859 0.860
0.852 0.853 0.854 0.855 0.855
0.846 0.848 0.849 0.850 0.851
0.841 0.842 0.843 0.844 0.845
0.858 0.859 0.860 0.860 0.861
0.853 0.854 0.855 0.856 0.857
0.848 0.849 0.850 0.851 0.852
0.843 0.844 0.845 0.846 0.847
0.837 0.838 0.839 0.840 0.842
0.854 0.855 0.856 0.857 0.858
0.849 0.850 0.851 0.852 0.853
0.844 0.845 0.846 0.847 0.848
0.839 0.840 0.841 0.842 0.843
0.833 0.834 0.835 0.837 0.838
0.851 0.852 0.853 0.853 0.854
0.845 0.846 0.848 0.849 0.850
0.840 0.841 0.842 0.843 0.844
0.835 0.836 0.837 0.838 0.839
0.829 0.830 0.831 0.833 0.834
0.847 0.848 0.849 0.850 0.851
0.842 0.843 0.844 0.845 0.846
0.836 0.838 0.839 0.840 0.841
0.831 0.832 0.833 0.834 0.835
0.825 0.826 0.828 0.829 0.830
0.843 0.844 0.845 0.846 0.848
0.838 0.839 0.840 0.841 0.842
0.833 0.834 0.835 0.836 0.837
0.827 0.828 0.829 0.830 0.832
0.821 0.822 0.824 0.825 0.826
0.840 0.841 0.842 0.843 0.844
0.834 0.836 0.837 0.838 0.839
0.829 0.830 0.831 0.832 0.833
0.823 0.824 0.825 0.827 0.828
0.817 0.818 0.820 0.821 0.822
0.836 0.837 0.838 0.839 0.841
0.831 0.832 0.833 0.834 0.835
0.825 0.826 0.827 0.829 0.830
0.819 0.820 0.822 0.823 0.824
0.813 0.814 0.816 0.817 0.818
0.833 0.834 0.835 0.836 0.837
0.827 0.828 0.829 0.830 0.832
0.821 0.822 0.824 0.825 0.826
0.815 0.817 0.818 0.819 0.820
0.809 0.810 0.812 0.813 0.814
0.829 0.830 0.831 0.832 0.834
0.823 0.825 0.826 0.827 0.828
0.817 0.819 0.820 0.821 0.822
0.811 0.813 0.814 0.815 0.816
0.805 0.806 0.808 0.809 0.810
L1282/Appendix B/frame Page 395 Monday, June 18, 2001 11:29 AM
USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 395
0.795 0.796 0.797 0.798 0.800
0.801 0.802 0.804 0.805 0.806
0.807 0.809 0.810 0.811 0.812
0.814 0.815 0.816 0.817 0.819
5.0 6.0 7.0 8.0 9.0
10.0 11.0 12.0 13.0 14.0
15.0 16.0 17.0 18.0 19.0
0.810 0.811 0.812 0.814 0.815
0.804 0.805 0.806 0.807 0.809 0.806 0.907 0.809 0.810 0.811
0.800 0.801 0.802 0.804 0.805
0.793 0.794 0.796 0.797 0.798
0.786 0.788 0.789 0.790 0.792
0.802 0.804 0.805 0.806 0.807
0.796 0.797 0.798 0.800 0.801
0.789 0.790 0.792 0.793 0.794
0.782 0.783 0.785 0.786 0.788
0.798 0.800 0.801 0.802 0.804
0.792 0.793 0.794 0.796 0.797
0.785 0.786 0.788 0.789 0.790
0.778 0.779 0.781 0.782 0.794
0.795 0.796 0.797 0.799 0.800
0.788 0.789 0.791 0.792 0.793
0.781 0.782 0.784 0.785 0.787
0.774 0.775 0.777 0.778 0.780
0.791 0.792 0.794 0.795 0.796
0.784 0.785 0.787 0.788 0.789
0.777 0.778 0.780 0.781 0.783
0.770 0.771 0.773 0.774 0.775
0.787 0.788 0.790 0.791 0.792
0.780 0.781 0.783 0.784 0.786
0.773 0.774 0.776 0.777 0.779
0.767 0.767 0.768 0.770 0.771
0.783 0.785 0.786 0.787 0.789
0.776 0.778 0.779 0.780 0.782
0.769 0.770 0.772 0.773 0.775
0.761 0.763 0.764 0.766 0.767
0.779 0.784 0.782 0.784 0.785
0.772 0.774 0.775 0.777 0.778
0.765 0.766 0.768 0.769 0.771
0.757 0.759 0.760 0.762 0.763
0.776 0.777 0.778 0.780 0.781
0.768 0.770 0.771 0.773 0.774
0.761 0.762 0.764 0.765 0.767
0.753 0.755 0.756 0.758 0.759
0.772 0.773 0.775 0.776 0.777
0.764 0.766 0.767 0.769 0.770
0.757 0.758 0.760 0.761 0.763
0.749 0.751 0.752 0.754 0.755
0.768 0.769 0.771 0.772 0.774
0.760 0.762 0.763 0.765 0.766
0.753 0.754 0.756 0.757 0.759
0.745 0.746 0.748 0.750 0.751
0.764 0.766 0.767 0.769 0.770
0.757 0.758 0.760 0.761 0.763
0.749 0.750 0.752 0.753 0.755
0.741 0.742 0.744 0.746 0.747
0.760 0.762 0.763 0.765 0.766
0.753 0.754 0.756 0.757 0.759
0.745 0.746 0.748 0.749 0.751
0.737 0.738 0.740 0.741 0.743
0.756 0.758 0.760 0.761 0.763
0.749 0.750 0.752 0.753 0.755
0.741 0.742 0.744 0.745 0.747
0.732 0.734 0.736 0.737 0.739
0.753 0.754 0.756 0.757 0.759
0.745 0.746 0.748 0.750 0.751
0.737 0.738 0.740 0.742 0.743
0.728 0.730 0.732 0.733 0.735
396
0.797 0.798 0.800 0.801 0.802
0.790 0.792 0.793 0.794 0.796
Conductivity, in microsiemens per centimeter at 25°C 51000 52000 53000 54000 55000 56000 57000 58000 59000 60000 61000 62000 63000 64000 65000 66000 67000
Temp ˚C
0.0 1.0 2.0 3.0 4.0
Salinity Correction Factors for Dissolved Oxygen in Water (based on conductivity) (Temp °C, temperature in degrees Celsius; salinity correction factors at 30 to 35°C are shown at the end of this table) (Continued)
Table B3
L1282/Appendix B/frame Page 396 Monday, June 18, 2001 11:29 AM
NATURAL AND ENHANCED REMEDIATION SYSTEMS
0.943 0.943 0.944 0.944 0.945 0.945
0.940 0.940 0.941 0.941 0.941 0.942
0.936 0.937 0.937 0.938 0.938 0.939
0.933 0.934 0.934 0.935 0.935 0.935
0.779 0.780 0.778 0.783 0.784
0.775 0.776 0.774 0.779 0.781
0.771 0.773 0.771 0.776 0.777
0.764 0.766 0.767 0.768 0.770
0.772 0.774
0.768 0.769
0.760 0.762 0.763 0.765 0.766
0.946 0.947 0.947 0.947 0.948 0.948
0.782 0.784 0.781 0.786 0.788
0.768 0.769 0.771 0.772 0.774
30.0 31.0 32.0 33.0 34.0 35.0
0.786 0.787 0.785 0.790 0.791
0.771 0.773 0.774 0.776 0.777
3300
0.789 0.791 0.789 0.794 0.795
0.775 0.777 0.778 0.779 0.781
Conductivity, in microsiemens per centimeter at 25°C 17000 18000 19000 20000 21000 22000 23000 24000 25000 26000 27000 28000 29000 30000 31000 32000
0.988 0.988 0.988 0.988 0.988 0.988
0.793 0.794 0.792 0.797 0.798
0.779 0.780 0.782 0.783 0.785
Temp ˚C
0.991 0.991 0.991 0.991 0.991 0.991
0.797 0.798 0.799 0.801 0.802
0.783 0.784 0.785 0.787 0.788
0.930 0.930 0.931 0.931 0.932 0.932
0.985 0.985 0.985 0.985 0.985 0.985
0.927 0.927 0.928 0.928 0.929 0.929
0.981 0.982 0.982 0.982 0.982 0.982
0.923 0.924 0.924 0.925 0.925 0.926
0.978 0.978 0.979 0.979 0.979 0.979
0.920 0.920 0.921 0.922 0.922 0.923
0.975 0.975 0.975 0.976 0.976 0.976
0.917 0.917 0.918 0.918 0.919 0.919
0.972 0.972 0.972 0.973 0.973 0.973
0.913 0.914 0.914 0.915 0.916 0.916
0.969 0.969 0.969 0.969 0.970 0.970
0.910 0.911 0.911 0.912 0.912 0.913
0.966 0.966 0.966 0.966 0.967 0.967
0.907 0.907 0.908 0.908 0.909 0.910
0.962 0.963 0.963 0.963 0.963 0.964
0.903 0.904 0.905 0.905 0.906 0.906
0.959 0.959 0.960 0.960 0.960 0.961
0.900 0.901 0.901 0.902 0.903 0.903
0.956 0.956 0.957 0.957 0.957 0.957
0.896 0.897 0.898 0.899 0.899 0.900
0.893 0.894 0.895 0.895 0.896 0.897
Conductivity, in microsiemens per centimeter at 25°C 5000 6000 7000 8000 9000 10000 11000 12000 13000 14000 15000 16000
0.800 0.802 0.803 0.804 0.805
0.786 0.788 0.789 0.790 0.792
0.950 0.950 0.950 0.951 0.951 0.951
0.994 0.994 0.994 0.994 0.994 0.994
4000
0.804 0.805 0.806 0.808 0.809
0.790 0.791 0.783 0.794 0.795
0.953 0.953 0.953 0.954 0.954 0.954
0.997 0.997 0.997 0.997 0.997 0.997
3000
0.808 0.809 0.810 0.811 0.812
0.794 0.795 0.796 0.798 0.799
1.000 1.000 1.000 1.000 1.000 1.000
2000
0.811 0.812 0.814 0.815 0.816
0.797 0.799 0.800 0.801 0.803
30.0 31.0 32.0 33.0 34.0 35.0
1000
0.815 0.816 0.817 0.818 0.820
0.901 0.802 0.804 0.805 0.806
0
0.818 0.820 0.821 0.822 0.823
0.805 0.806 0.807 0.809 0.810
Temp ˚C
0.822 0.823 0.824 0.825 0.827
0.809 0.810 0.811 0.812 0.814
0.826 0.827 0.828 0.829 0.830
0.812 0.814 0.815 0.816 0.817
25.0 26.0 27.0 28.0 29.0
0.816 0.817 0.818 0.820 0.821
0.820 0.821 0.822 0.823 0.824
20.0 21.0 22.0 23.0 24.0
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USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 397
0.821 0.822 0.823 0.824 0.825 0.826
0.817 0.818 0.820 0.821 0.822 0.823
0.814 0.815 0.816 0.817 0.818 0.820
0.810 0.811 0.813 0.814 0.815 0.816
Source: USGS Groundwater Sampling Field Manual (Chapter 6).
0.824 0.825 0.826 0.828 0.829 0.830
0.807 0.808 0.809 0.810 0.812 0.813
0.803 0.804 0.806 0.807 0.808 0.809
0.862 0.863 0.864 0.865 0.866 0.867
0.800 0.801 0.802 0.803 0.805 0.806
0.859 0.860 0.861 0.862 0.863 0.863
0.796 0.797 0.799 0.800 0.801 0.803
0.855 0.856 0.857 0.858 0.859 0.860
0.793 0.794 0.795 0.797 0.798 0.799
0.852 0.853 0.854 0.855 0.856 0.857
0.789 0.790 0.792 0.793 0.795 0.796
0.849 0.850 0.851 0.851 0.852 0.853
0.786 0.787 0.788 0.790 0.791 0.792
0.845 0.846 0.847 0.848 0.849 0.850
0.782 0.783 0.785 0.786 0.788 0.789
0.842 0.843 0.844 0.845 0.846 0.847
0.779 0.780 0.781 0.783 0.784 0.785
0.838 0.839 0.840 0.841 0.842 0.843
0.775 0.776 0.778 0.779 0.781 0.782
0.835 0.836 0.837 0.838 0.839 8.400
398
0.828 0.829 0.830 0.831 0.832 0.833
0.866 0.867 0.868 0.868 0.869 0.870
0.831 0.832 0.833 0.834 0.836 0.837
0.869 0.870 0.871 0.872 0.873 0.874
30.0 31.0 32.0 33.0 34.0 35.0
0.873 0.873 0.874 0.875 0.876 0.877
Conductivity, in microsiemens per centimeter at 25°C 51000 52000 53000 54000 55000 56000 57000 58000 59000 60000 61000 62000 63000 64000 65000 66000 67000
0.876 0.877 0.878 0.879 0.879 0.880
Temp ˚C
0.879 0.880 0.881 0.882 0.883 0.883
0.890 0.890 0.891 0.892 0.893 0.893
30.0 31.0 32.0 33.0 34.0 35.0
0.883 0.884 0.884 0.885 0.886 0.887
Conductivity, in microsiemens per centimeter at 25°C 34000 35000 36000 37000 38000 39000 40000 41000 42000 43000 44000 45000 46000 47000 48000 49000 50000
Temp ˚C 0.886 0.887 0.888 0.889 0.889 0.890
Salinity Correction Factors for Dissolved Oxygen in Water (based on conductivity) (Temp °C, temperature in degrees Celsius; salinity correction factors at 30 to 35°C are shown at the end of this table) (Continued)
Table B3
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NATURAL AND ENHANCED REMEDIATION SYSTEMS
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USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING
Table B4
399
Standard Half-Cell Potentials of Selected Reference Electrodes as a Function of Temperature and Potassium Chloride Reference-Solution Concentration, in Volts (Liquid-junction potential included — multiply volts by 1000 to convert to millivolts; KCl, potassium chloride; Temp °C, temperature in degrees Celsius; M, molar; —, value not provided in reference) Silver:silver chloride
Calomel
Temp °C
3M KCI
3.5M KCI
Saturated KCI
3M KCI
10 15 20 25 30 35 40
0.220 0.216 0.213 0.209 0.205 0.202 0.198
0.215 0.212 0.208 0.205 0.201 0.197 0.193
0.214 0.209 0.204 0.199 0.194 0.189 0.184
0.260 — 0.257 0.255 0.253 — 0.249
3.5M KCI
4M KCI
Saturated KCI
Orion™ 96–78 Combination Electrode
0.254 0.251 0.248 0.244 0.241 0.238 0.234
0.256 0.253 0.249 0.246 0.242 0.238 0.234
0.256 — — — 0.252 — 0.250 0.246 0.248 0.244 — — 0.244 0.239
Source: USGS Groundwater Sampling Field Manual (Chapter 6).
Table B5
Eh of ZoBell’s Solution as a Function of Temperature (°C, degrees Celsius; mV, millivolts)
Temperature °C
Eh (mV)
10 12 14 16 18 20 22 24 25 26 28 30 32 34 36 38 40
467 462 457 453 448 443 438 433 430 428 423 418 416 407 402 397 393
Source: USGS Groundwater Sampling Field Manual (Chapter 6).
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400
Table B6
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Troubleshooting Guide for Eh Measurement (±, plus or minus; mV, millivolts; emf, electromotive force) Symptom
Eh of ZoBell’s solution exceeds theoretical by ± 5 mV Excessive drift Erratic performance Poor response when using paired electrodes
Possible Corrective Action • • •
• •
•
• •
• •
Check meter operation: Use shorting lead to establish meter reading at zero mV. Check/replace batteries. Check against backup meter. Check electrode operation: Check that electrode reference solution level is to the fill hole. Plug questionable reference electrode into reference electrode jack and another reference electrode in good working order of the same type into the indicator electrode jack of the meter; immerse electrodes in a potassium chloride solution, record mV, rinse off and immerse electrodes in ZoBell’s solution. The two mV readings should be 0 ± 5 mV. If using different electrodes (Ag:AgCl and Hg:HgCl2), reading should be 44 ± 5 mV for a good reference electrode. Polish platinum tip with mild abrasive (crocus cloth, hard eraser, or a 400–600-grit wet/dry Carborundum™ paper), rinse thoroughly with deionized water. Use a Kimwipe™ if these abrasives are not available. Drain and refill reference electrolyte chamber. Disconnect reference electrode. Drain and refill electrolyte chamber with correct filling solution. Wipe off connectors on electrode and meter. Use backup electrode to check the emf. Read emf with fresh aliquot of ZoBell’s solution; prepare fresh ZoBell’s solution if possible. Recondition electrode by cleaning with aqua regia and renewing filling solution — this is a last resort.
Source: USGS Groundwater Sampling Field Manual (Chapter 6).
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USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING
401
Test and calibrate field measurements. Select downhole or flowthrough-chamber system.
DOWNHOLE
FLOWTHROUGH CHAMBER
Lower sensors and pump to selected depth.
Install pump in monitoring well or plumbing for use of existing pump in a supply well. Install sensors in flowthrough chamber.
Adjust flow to desired purge rate and record rate and time purging began.
Turn on pump and adjust flow to desired purge rate, record rate and time purging began. Allow sensors to equilibrate at purge rate.
• Divert initial water to waste. • Correct chamber in-line from pump. • Adjust flow to chamber and eliminate backpressure: allow sensors to equilibrate
Record and monitor sequential sets of field measurement readings during withdrawal of trial well volume. • After two or more well volumes are purged and before final five or more readings are made, adjust flow rate to be used for sampling flow; flow must be sufficient for dissolved oxygen measurements. • For pH: divert flow from chamber and record measurement when water is quiescent. Redirect flow to chamber for next set of readings.
NO
Extend purge time. Document difficulty in field notes.
Are stabilization criteria being met?
YES
Record at least 5 measurements at intervals of 3 to 5 minutes or more.
Report the median of the last 5 or more readings and the time of measurement.
Figure B1
Field-measurement procedures using downhole and flowthrough-chamber systems for groundwater. Source: USGS Groundwater Sampling Field Manual (Chapter 6).
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402
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Test and calibrate field instruments.
• • • • • •
Purge well (see text for exceptions and instructions).
(2)
Field rinse precleaned sampler. Use clean/dirty hands technique. Lower sampler smoothly, without splashing, to desired depth of screened or open interval. (If using bailer, double check-valve type is recommended.) Raise sampler smoothly at a constant rate, keeping lines clean and off the ground. Place sampler in holding stand.
(3)
Withdraw subsamples from sampler. If using bailer, insert clean bottom-emptying device with groved hands; device should fit snugly over collection bottles and (or) measurement vessels. If a filtered sample is needed, filter in-line from sampler to bottle/vessel. Drain subsample without turbulence into collection bottles or measurement vessel. Rinse collection bottle(s) three times with sample (filtrate for filtered samples), fill to brim, cap tightly, and maintain at temperature of water source until measurement. Rinse sensors, stir bar, and measurement vessel three times with sample. For alkalinity, rinse with deionized water.
• •
• • •
• • •
Step (1)
Insert sensor(s) in measurement vessel. Wait for sensors to equilibrate to sample temperature. Don’t let sensors touch bottom or sides of vessel.
Swirl or stir gently to mix sample. Minimize streaming potential or vortex; keep sensor out of vortex. For pH, do not stir samples with conductivity less than 100 S/cm. When using magnetic stirrer, use smallest stir bar.
(4)
(5)
(6)
Record field measurement and method used on field form. Record median value of stabilized readings. If readings do not stabilize, extend number of measurements and record median of at least 5 or more sequential readings. If there is a constant trend toward lower or higher values, record the first value, the range of values, and the time period observed.
(7)
Repeat process from steps (4) through (7) on two or more subsamples from the same sample volume to document precision. Rinse sensors and equipment thoroughly with deionized water. Discard sample to waste, in accordance with applicable regulations.
(8)
Figure B2
Subsample field-measurement procedures for conductivity, pH, and alkalinity of groundwater. Source: USGS Groundwater Sampling Field Manual (Chapter 6).
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USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING
403
8000 7500 7000 ELEVATION, NGVD OF 1929, IN FEET
6500 6000 5500 5000 4500 4000 3500 3000 2500 2000 1500 1000 500 0 -500 -1000 0
20
40
60 80
100 100
120 140 160 180 200
VALUE TO SUBTRACT FROM ATMOSPHERIC PRESSURE, IN MILLIMETERS OF MERCURY
Figure B3
Factors used to correct atmospheric pressures adjusted to sea level. Source: USGS Groundwater Sampling Field Manual (Chapter 6).
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L1282/Appendix C/frame Page 405 Tuesday, June 19, 2001 1:16 PM
APPENDIX
C
Common and Scientific Names of Various Plants Common Name
Scientific Name
Alfalfa Algae stonewort Alyssum Bean Bean, bush Bermuda grass Birch, river Black locust Black willow tree Bladder campion Bluestem, big (prairie grass) Bluestem, little Bluestem, little Boxwood Buffalo grass Canada wild rye (prairie grass) Canola Cattail Cherry bark oak Clover Colonial bentgrass Colonial bentgrass Cottonwood, eastern Cottonwood (poplar) Crab apple
Medicago sativa Nitella Alyssum wulfenianum Phaseolus coccineus L. Phaseolus vulgaris cv. Tender Green Cynodon dactylon Betula nigra Robinia pseudoacacia Salix nigra Silene vulgaris Andropogon gerardi Vit. Andropogon scoparius Schizachyrium scoparius Buxaceae Buchloe dactyloides Elymus canadensis Brassica napus Typha latifolia Quercus falcata Genus trifolium Agrostis tenuis cv. Goginan Agrostis tenuis cv. Parys Populus deltoides Populus Malus fusca Raf. Schneid
405
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406
Crested wheatgrass (Hycrest)
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Agropyron desertorum (Fisher ex Link) Schultes Cypress, bald Taxodium distichum Duckweed Lemna minor Eastern Cottonwood Populus deltoides European milfoil/yarrow Achillea millefolium Felt leaf willow Salix alaxensis Fescue, hard Festuca ovina var. duriuscula Fescue, red Festuca rubra cv. Merlin Fescue, tall Festuca arundinacea Schreb. Four-wing saltbrush Aicanescens Grama, side oats (prairie grass) Bouteloua curtipendula Grama, blue Bouteloua gracilis Grass, cool season (colonial bentgrass) Agrotis tenuis Grass, warm season (Japanese lawngrass) Zoysia japonica Horseradish (roots) Armoracia rusticana Hybrid poplar Populus deltoides x nigra DN-34, Imperial California; Populus charkowiieensis x incrassata; Populus tricocarpa x deltoides Hycrest, crested wheatgrass Agropyron desertorum Indiangrass (prairie grass) Sorghastrum nutans Indian mustard Brassica juncea Indian ricegrass Oryza sativa subsp. indica Japanese lawngrass Zoysia japonica Jimson weed (thornapple) Datura innoxia Kenaf Hibiscus cannabinus L. cv. Indian Koa haole Leucaena leuccephala Kudzu Pueraria lobata Lambsquarter Chenopodium Legume Lespedeza cuneata (Dumont) Little bluestem (prairie grass) Schizachyrium scoparius Loblolly pine Pinus taeda L. Mesquite Prosopis Millet, Proso Panicum miliaceum L. Mulberry, red Morus rubra L. Mustard, Indian Brassica juncea Mustard weed Arabidopsis thaliana Oak, cherry bark Quercus falcata Oak, live Quercus virginiana Osage, orange Maclura pomifera (Raf.) Schneid Parrot feather Myriophyllum aquaticum thlaspi rotundifolium Pennycress Thlaspi rotundifolium Pennycress, alpine Thlaspi caerulescens Pennyworth Hydrocotyle umbellata
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COMMON AND SCIENTIFIC NAMES OF VARIOUS PLANTS
Poplar, cottonwood Poplar, hybrid
Poplar, yellow Red fescue Reeds Rice Sacaton, alkali Sea pink; wild thrift Salt marsh plant Sand dropseed Soybean Spearmint Sugarcane Sudangrass Sunflower Switchgrass (prairie grass) Tall fescue Thornapple (jimson weed) Thrift (wild); sea pink Tobacco Water hyacinth Water milfoil Water velvet Wheat grass, slender Wheat grass, western (prairie grass) Willow tree, black Willow tree, felt leaf
407
Populus Populus deltoides x nigra DN-34, Imperial California; Populus charkowiieensis x incrassata; Populus tricocarpa x deltoides; Populus charkowiieensis x incrassata Liriodendron tulipifera Festuca rubra cv. Merlin Phragmites Oryza sativa L. Sporobolus wrightii Armeria maritima Spartina alterniflora Sporobolu cryptandrus Glycine max L. Merr, cv. Davis Mentha spicata Saccharum officinarum Sorghum vulgare L. Helianthus annuus Panicum virgatum Festuca arundinacea Schreb. Datura innoxia Armeria maritima Nicotiana tabacum Eichhornia crassipes Myriophyllum spicatum Azolla pinnata Agropyron trachycaulum Agropyron smithii Salix nigra Salix alaxensis
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L1282_frame_IDX Page 409 Monday, June 18, 2001 9:13 AM
Index A
fermentation, 151 natural attenuation of, 97 sources of, 95 Bioaugmentation, 136–138 Biodegradation acceptable, 26 arsenic, 107 benzene, toluene, ethylbenzene, and xylene, 95–97 chemical oxidation pretreatment effects, 223 chemical structure and, 36 chemical substrate concentration effects, 27 chlorate, 108 chlorinated aliphatics, 99 chlorinated aromatics, 100–103 chlorobenzoates, 100–101 chlorophenols, 100–101 chromium, 106–107 cometabolism effects, 32 conditions necessary for, 26 definition of, 26 electron-acceptor-limited model of, 94–95 environmental impacts of, 32–34 first-order decay model, 92–94 lindane, 106 mercury, 107 metals, 105–106 natural attenuation effects, 90–109 nitrate, 107–108 nitrate esters, 104 nitroaromatics, 103–104 organic contaminants, 95 oxyanions, 107–108 oxygenated hydrocarbons, 98–99 pathways of, 35 pentachlorophenol, 101–102 perchlorate, 108 phyto-cover effects, 326–327 polychlorinated biphenyls, 101 polycyclic aromatic hydrocarbons, 98, 308
Absorption, 39 Acceptable biodegradation, 26 Adenosine triphosphate, 90–91 Adsorption, 39 Advection dispersion equation, 84–85 Aerobic oxidation cometabolism, 202–204 description of, 200–201 indications for, 200 MTBE degradation, 204–205 rates of, 201 Aging, 33 Alkane hydroxylase, 30 Alkane monooxygenase, 30 Alkylbenzene sulfonate, 34 Alkylhalides, aerobic oxidation of, 201 American Society of Testing and Materials, monitored natural attenuation support by, 64 Ammonia monooxygenase, 30 Ammonification, 197–198 Anaerobic organisms, 5 Analogue enrichment, 25 Anoxic environments, oxidants in, 57 Aquifer matrix, 47 Aquifer solids, 51 Arsenic biodegradation of, 107 in situ precipitation, 194–195 Atmospheric pressures, 383, 403
B Benzene, toluene, ethylbenzene, and xylene biodegradation of, 95–97 description of, 6
409
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410
NATURAL AND ENHANCED REMEDIATION SYSTEMS
primary, 26 rates of, 35–38 selenium, 108 structural effects, 34–35 treatment wetlands for, 306–309 ultimate, 26 Bioemulsifiers, 165 Biogeochemical definition of, 89 reactions, 89–90 Bioremediation, 6 Biostimulation, 136–138 Biosurfactants, 165–166 Biphenyl dioxygenase, 30 Boiling point, 18 Butane-utilizing bacteria, 202–203
C Carbon tetrachloride, reductive dechlorination of, 159 Chelates, 247 Chemical oxidation advantages of, 206–207 applications of description of, 218–219 1,4-dioxane, 222–223 biodegradation effects, 223 chemistry of, 208–218 concerns regarding, 207 description of, 205–206 mechanism of action, 206 oxidants used description of, 208–209 hydrogen peroxide, 208–209, 211–213 potassium permanganate, 209, 213–216 predictions, 219–220 Chlorate biodegradation, 108 Chlorinated aliphatics aerobic oxidation of, 201 biodegradation of, 99, 138–139 mixture of, 156 natural attenuation of, 75, 99 Chlorinated alkanes aerobic oxidation of, 202 treatment wetland removal of, 306 Chlorinated aromatics biodegradation of, 100–103, 138–139 natural attenuation of, 75, 100–103 Chlorinated ethenes, reductive dechlorination using, 145, 170–177 Chlorobenzene biodegradation, 100 Chlorobenzoate biodegradation, 100–101
Chloroethene reductive dehalogenases, 147 Chloroform, reductive dechlorination of, 159 Chlorophenol biodegradation, 100–101 Chromium biodegradation of, 106–107 in situ precipitation, 192–194 Colloidal organic matter, 49 Colloidal particles, 196–197 Cometabolism aerobic, 202–204 cautions associated, 28–29 chemicals subject to, 28 chlorinated solvents, 139 definition of, 27, 140 environmental consequences of, 28 polycyclic aromatic hydrocarbons, 25 reactions co-oxidation, 140 mechanisms of, 29–32 reactive dechlorination, See Reductive dechlorination types of, 28 Comprehensive Environmental Response, Compensation, and Liability Act, 314 Constructed treatment wetlands algae in, 279–280 bacteria in, 278–279 benefits of, 272 components of, 270–271 contaminant removal mechanisms hydraulic retention time, 297–299 partitioning and storage, 295–296 volatilization, 294–295 description of, 270 dissolved oxygen concentration, 294 in Europe, 272–273 fish in, 310 free water surface, 276 fungi in, 278–279 groundwater remediation use of, 299–310 horizontal flow systems characteristics of, 276–277 emergent macrophyte-based systems with, 285 hydrology of, 309–310 hydroperiods, 292 inorganics removal using, 309 issues regarding, 273–274 metals removal, 300–305 morphology of, 309–310 popularity of, 271 regulation of, 272, 275–276 research of, 274
L1282_frame_IDX Page 411 Monday, June 18, 2001 9:13 AM
INDEX
soils biological influences, 292 cation exchange capacity, 289–290 characteristics of, 287–289 clays, 287 hydric, 287 microbial processes, 292–293 mineral, 287 organic, 287–288 oxidation reactions, 290–291 peats, 287–288 pH, 292 reduction reactions, 290–291 subsurface flow, 276–277 toxic organics removal using, 306–309 in United States, 272–273 vegetation in description of, 279, 281–282 duckweeds, 283–284 emergent macrophyte-based systems description of, 284 with horizontal subsurface flow, 285 with vertical subsurface flow, 285 free-floating macrophyte-based systems, 282–284 multistage macrophyte-based systems, 287 submerged macrophyte-based systems, 285–287 water hyacinths, 282–283 vertical flow systems characteristics of, 277–278 emergent macrophyte-based systems with, 285 wildlife considerations, 274–275, 310 Contaminants availability of, 14 biodegradation of acceptable, 26 chemical structure and, 36 chemical substrate concentration effects, 27 cometabolism effects, 32 conditions necessary for, 26 definition of, 26 environmental impacts of, 32–34 pathways of, 35 primary, 26 rates of, 35–38 structural effects, 34–35 ultimate, 26 biological characteristics of, 26–38 boiling point of, 18 characteristics of, 16–17 cometabolism of, 27–32
411
dense nonaqueous-phase liquid definition of, 73 identification methods, 88–89 permanganate oxidation of, 220–221 transport of, 88 Henry's law constant, 19–20 hydrolysis of, 22–24 immobilization of, 196 inorganic biodegradation of, 105 treatment wetland removal of, 309 light nonaqueous-phase liquid, 73 light nonaqueous-phase liquids, 73 natural attenuation removal of, 72–77 nonaqueous-phase liquid definition of, 73–74 transfer of, 88 nonaqueous-phase liquids, 73–74 octanol/water partition coefficients, 20 photolytic reactions, 24–25 physical properties of, 365–382 sequestration of, 33–34 in soil-plant system, 242–243 solubility, 20–22 sorption coefficient of competitive, 49 description of, 38–39 factors that affect, 39, 48–51 kinetic considerations, 50 pH effects, 49 soil, 43–48 temperature effects, 48 source areas of, 8 sources zones for, 72–73 subsurface movement methods advection, 81–82 dispersion, 83–87 treatment wetland removal of hydraulic retention time, 297–299 partitioning and storage, 295–296 volatilization, 294–295 vapor pressure of, 18 Cooxidation, 28 Cosolvents, 49
D Dechlorination, See Reductive dechlorination Dehalobacgter restrictus, 145, 147 Dehalococcoides ethenogenes, 146 Dehalospirillum multivorans, 145, 147 Dehydrohalogenation, 23 Denitrification, in situ, 197–198
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412
NATURAL AND ENHANCED REMEDIATION SYSTEMS
Dense nonaqueous-phase liquid contaminants definition of, 73 identification methods, 88–89 permanganate oxidation of, 220–221 transport of, 88 Desulfitobacterium sp., 145 Desulfuromonas chloroethenica, 146 3,5-Dichlorobenzoate, reductive dechlorination of, 156 Dilution, natural attenuation effects, 79–81 1,4-Dioxane, ozone oxidation of, 222–223 Dispersion advection dispersion equation, 84–85 definition of, 83 mechanical, 84 molecular, 83–84 Dissolved organic matter, 49 Dissolved oxygen measurements of description of, 113 electrodes, 113–115 in field, 116 field calibrations, 115–116 salinity correction factors, 392–398 Distribution coefficient definition of, 45–46 estimating of, 87 Duckweed-based wetland systems, 283–284
E Earth anaerobic organisms, 5 history of, 5–6 Electron-acceptor-limited model, 94–95 Enhanced reductive dechlorination anaerobic oxidation, 158 biostimulation vs. bioaugmentation, 136–138 cometabolic, 142–144 electron acceptors, 159–160 electron donor selection for, 147–158 field studies of, 160–164 groundwater solute transport of chlorinated ethenes, 170–177 halorespiring microorganisms, 144–147 high constituent concentration areas, 169–170 low constituent concentration areas, 169 mechanisms of, 138–142 microorganisms for electron donors for, 147–158 halorespiring, 144–147 hydrogen production, 149–151 nutrients, 158 mixture of compounds effect, 155–158
oxidation-reduction potential effects, 142 in situ reactive zones for biofilm, 168–169 development considerations, 160–163 fermentation, 167–168 natural surfactant effect, 165–167 performance data, 177–183 reagent delivery, 164–165 studies of, 135–136 temperature effects, 158 Enterobacter agglomerans, 146 Environment, biodegradation effects on, 32–34 Environmental Protection Agency, monitored natural attenuation support by, 64 Estuaries, natural organic material in, 42 Evapotranspiration description of, 321 effective, 340–343 potential, 337–340
F Fermentation benzene, toluene, ethylbenzene, and xylene, 151 definition of, 150 hydrogen produced during, 149–151, 153–154 primary, 151 reductive dechlorination, 167–168 secondary, 151 First-order decay model, 92–94 Flocculation, 196 Fortuitous metabolism, 27
G Groundwater field-measurement procedures, 401–402 mobile water phase of, 121 phytoremediation of, 259–260, 263 rock/soil phase of, 121 treatment wetland remediation of, 299–310 velocity of, 82 Groundwater capture zone, 267 Growth rate, 37
H Half-velocity coefficient, 37 Halorespiration, 144–145
L1282_frame_IDX Page 413 Monday, June 18, 2001 9:13 AM
INDEX
413
Halorespiring microorganisms hydrogen consumption, 154 reductive dechlorination using, 144–147 types of, 144 Hazard definition of, 2 predicting of, 4 subjective probability of, 2 transport of, 4–5 Hazardous waste, 3 Henry's law constant, 19–20 Hexachlorbutadiene, treatment wetland removal of, 308 Hexachlorobenzene, treatment wetland removal of, 308 High molecular weight natural organic matter, 58 Humic soil organic matter, 42 Hydrocarbons, natural attenuation of, 75 Hydrogen fermentation-induced production of, 149–151, 153–154 microorganism production of competition, 152–153 description of, 149–151 Hydrogenolysis, 55 Hydrogen peroxide, in situ chemical oxidation uses of, 208–209, 211–213 Hydrolysis definition of, 22 equation, 22 plume effects, 23–24 rates of, 23 reaction products produced, 22–23 Hyperaccumulators, 246–247
I Inorganic contaminants biodegradation of, 105 treatment wetland removal of, 309 In situ chemical oxidation advantages of, 206–207 applications of description of, 218–219 1,4-dioxane, 222–223 biodegradation effects, 223 chemistry of, 208–218 concerns regarding, 207 description of, 205–206 mechanism of action, 206 oxidants used description of, 208–209 hydrogen peroxide, 208–209, 211–213
potassium permanganate, 209, 213–216 predictions, 219–220 In situ denitrification, 197–198 In situ precipitation of metals aquifer parameters, 195–196 arsenic, 194–195 chromium, 192–194 contaminant removal, 196–197 description of, 183, 188 principles of, 187–195 In situ reactive zones advantages of, 133–134 definition of, 132 description of, 8 design of, 132 effectiveness of, 132–133 metals, 134–135 nano-scale particle injections iron description of, 223–228 history of use, 225 organic contaminants treatable using, 231–232 particle production, 228–230 particle size, 226, 228 reductive process, 224–225 permeable sediment applications, 231 reductive dechlorination biofilm, 168–169 development considerations, 160–163 fermentation, 167–168 natural surfactant effect, 165–167 performance data, 177–183 reagent delivery, 164–165 schematic representation of, 132–133 types of, 134–135 Instantaneous reaction model, 94–95 Intrinsic remediation, See Monitored natural attenuation Ionic strength, 49 Ionizability, 50–51 Iron, nano-scale injections in in situ reactive zones description of, 223–228 history of use, 225 organic contaminants treatable using, 231–232 particle production, 228–230 particle size, 226, 228 reductive process, 224–225 IRZ, See In situ reactive zones ISCO, See In situ chemical oxidation Isotherms definition of, 43 linear sorption, 43–44 nonlinear, 44
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414
NATURAL AND ENHANCED REMEDIATION SYSTEMS
L Lakes natural organic material in, 42 recharging, 79 Landfill history of, 314–315 leachates, 273, 317 microbial activity in, 317 moisture flow and content, 333 waste stabilization phases in, 317–320 Landfill cover barrier-type, 316 capillary barrier, 321 conventional description of, 316–317 phyto-cover vs., 326 design features of, 316 evapotranspiration, 321 function of, 315–316 phyto-cover, See Phyto-cover requirements for, 315 Light nonaqueous-phase liquid contaminants, 73 Lindane biodegradation, 106 Liquid-filled macropores, 14–15 Low molecular weight natural organic matter, 58
M Marine waters, contaminant levels in, 33 Mass removal efficiency, 7 Maximum growth rate, 37 Mechanical dispersion, 84 Mercury, biodegradation of, 107 Metals, See also specific metal biodegradation of, 105–106 field filtration of, 121–124 filtered vs. unfiltered samples, 120–124 microbial transformation of, 190 natural attenuation of, 75 precipitates, 187 in situ precipitation aquifer parameters, 195–196 arsenic, 194–195 chromium, 192–194 contaminant removal, 196–197 description of, 183, 188 principles of, 187–195 in situ reactive zones, 134–135 soil concentrations of, 186, 241–242 treatment wetlands for, 299–310 Methane monooxygenases description of, 30 types of, 203
Methanotrophs, 202–203 Methylene chloride, reductive dechlorination of, 159 Micelles, 289–290 Microbial ecology, 6 Microorganisms electron donors, 92 hydrogen production by, 149–151 inorganic contaminants transformed by, 105 reproductive mechanisms of, 90–91 in rhizodegradation, 252 Molecular dispersion, 83–84 Monitored natural attenuation capacity, 77–78 definition of, 64 description of, 8 documenting of, 66 evaluative approaches, 65–69 monitoring and sampling of case study, 124–125 considerations, 109–110 description of, 109–113 dissolved oxygen, 113–116 field portable meters, 111 filtered vs. unfiltered metal samples, 120–124 low-flow sampling, 124 on-site, 111 oxidation-reduction potential, 117–119 pH, 119–120 in situ, 112–113 techniques for, 110–113 organizations that support, 64–65 patterns of contaminant sources, 72–77 description of, 71 questions for assessing, 71–72 processes that affect advection, 81–82 biodegradation, 90–109 dilution, 79–81 dispersion, 83–87 stabilization, 88–89 volatilization, 89 protocols, 70–71 terminology associated with, 64–65 MTBE, aerobic oxidation of, 204–205
N Nano-scale particle injections, in in situ reactive zones iron description of, 223–228
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history of use, 225 organic contaminants treatable using, 231–232 particle production, 228–230 particle size, 226, 228 reductive process, 224–225 permeable sediment applications, 231 Naphthalene dioxygenase, 30 Natural attenuation capacity, 77–78 definition of, 64 description of, 8 documenting of, 66 evaluative approaches, 65–69 monitoring and sampling of case study, 124–125 considerations, 109–110 description of, 109–113 dissolved oxygen, 113–116 field portable meters, 111 filtered vs. unfiltered metal samples, 120–124 low-flow sampling, 124 on-site, 111 oxidation-reduction potential, 117–119 pH, 119–120 in situ, 112–113 techniques for, 110–113 organizations that support, 64–65 patterns of contaminant sources, 72–77 description of, 71 questions for assessing, 71–72 processes that affect advection, 81–82 biodegradation, 90–109 dilution, 79–81 dispersion, 83–87 stabilization, 88–89 volatilization, 89 protocols, 70–71 terminology associated with, 64–65 Natural attenuation capacity, 77–78 Natural degradation, See Monitored natural attenuation Natural organic matter high molecular weight, 58 low molecular weight, 58 Natural surfactant effect, 165–167 Nitrate biodegradation, 107–108 Nitrate esters, 104 Nitrification, 198 Nitroaromatics biodegradation of, 103–104 natural attenuation of, 75, 103–104
415
Nonaqueous-phase liquid contaminants definition of, 73–74 transfer of, 88
O Octanol/water partition coefficients, 20 Organic contaminants, 95 Organic matter colloidal, 49 dissolved, 49 soil composition of, 42 critical level of, 46 definition of, 41–42 humic, 42 nonhumic, 42 sediments vs., 42–43 sorption coefficient effects, 49–50 Oxidants in chemical oxidation systems description of, 208–209 hydrogen peroxide, 208–209, 211–213 potassium permanganate, 209, 213–216 in natural systems, 56 Oxidation-reduction, See REDOX Oxidation-reduction potential definition of, 117 measurement of, 117–119 reductive dechlorination effects, 142 wetland soils, 290–292 Oxyanions biodegradation of, 107–108 natural attenuation of, 76, 107–108 Oxygen dissolved description of, 113 electrodes, 113–115 in field, 116 field calibrations, 115–116 salinity correction factors, 392–398 solubility in water, 384–391 Oxygenase, 30 Oxygenated hydrocarbons biodegradation of, 98–99 natural attenuation of, 75, 98–99 sources of, 98 Oxyhydroxides, 123 Ozone chemical structure of, 217 1,4-dioxane oxidation, 222–223 in situ chemical oxidation uses of, 208, 216–218
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P Particle diffusion, description of, 15 Passive remediation, See Monitored natural attenuation pe, 51–52 Pentachlorophenol biodegradation, 101–102 Perchlorate biodegradation of, 108 chlorite, 199 reduction of, 199–200 in situ reactive zone for, 199–200 Pesticide biodegradation, 104 pH definition of, 51 measurement of, 119–120 sorption coefficient effects, 49 wetland soil, 292 Photolytic reactions definition of, 24 direct, 24–25 function of, 24 indirect, 25 Photoreactions, See Photolytic reactions Phreatophytes, 259–260 Phytoaccumulation, 245–248, 259, 299 Phyto-cover system agronomic chemistry sampling of, 358–359 benefits of aesthetic, 327–328 ecological, 327–328 economics, 329 gas permeability, 327 maintenance, 329 public safety, 329 in situ biodegradation, 326–327 case study example of, 344–347 characteristics of, 321–323 components of, 329–330 conventional covers vs., 326 description of, 321, 332–333 designing of, 333–334 examples of, 323–326 illustration of, 323 irrigation/irrigation system considerations for, 329–330 guidelines for, 352–356 requirements, 362 maintenance active, 360–361 passive, 361 preventive, 359 models for designing, 333–335 nonsoil amendments to, 331 operation and maintenance schedule, 359–362
performance of hydrologic water balance, 332–335 potential evapotranspiration, 337–343 precipitation, 335 runoff, 335–337 water balance model, 343 plants, 322, 331–332 poplars, 321–322 repairs, 359 safety considerations, 359 sample application of, 344–347 schematic of, 334 site inspections, 348–349 soil moisture monitoring, 349–352 trees evaluations of, 356–358 leaves of, 357–358 root systems of, 349 selection of, 321–322, 331–332 soil moisture monitoring, 349–350 stem evaluations, 356 understory planting, 331 vegetative, 330–331 water balance model, 343 summary overview of, 347–348 Phytodegradation, 248–250, 258 Phytoextraction, See Phytoaccumulation Phytoremediation advantages of, 240 applications of, 258–259 decision tree for, 262 definition of, 240 description of, 244 disadvantages of, 240 groundwater contaminants, 259–260, 263 mechanism of action, 240 phytoaccumulation, 245–248, 259, 299 phytodegradation, 248–250, 258 phytostabilization, 250–251, 258 phytovolatilization, 251–252, 258 rhizodegradation, 252–256, 258 rhizofiltration, 256–259, 258, 299 sediments, 260–261, 263 soil-plant system chemicals metals, 241–242 organic compounds, 242–243 system design agronomic inputs, 266 contaminant levels, 265 groundwater capture zone, 267 irrigation, 266 maintenance, 266 overview of, 261–265 plant species, 265–266
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417
transpiration rate, 267 treatability, 266 Phytostabilization, 250–251, 258 Phytotransformation, See Phytodegradation Phytovolatilization, 251–252, 258 Plants chemical uptake in metals, 241–242 organic compounds, 242–243 enzyme systems of, 248–249 phyto-cover system, 322, 331–332 rhizodegradation uses of, 255 rhizofiltration uses of, 257 scientific names of, 405–407 in treatment wetlands description of, 279, 281–282 duckweeds, 283–284 emergent macrophyte-based systems description of, 284 with horizontal subsurface flow, 285 with vertical subsurface flow, 285 free-floating macrophyte-based systems, 282–284 multistage macrophyte-based systems, 287 submerged macrophyte-based systems, 285–287 water hyacinths, 282–283 Pollutants, 5 Pollution evaluations, 2 Polychlorinated biphenyls biodegradation of, 101 plant uptake of, 243 sources of, 101 Polycyclic aromatic hydrocarbons biodegradation of, 98, 308 cometabolism of, 25 half-life of, 295 hydrolysis of, 25 natural attenuation of, 98 sources of, 98 Poplars, phyto-cover use of, 321–322 Potassium permanganate, in situ chemical oxidation uses of, 209, 213–216 Precipitate, 187 Primary biodegradation, 26 Proton, 51
R Reactive zones, See In situ reactive zones Recharge definition of, 79
dilution estimations, 80 natural attenuation effects, 79–81 REDOX contaminant transformation, 58 measurement of, 111, 119 poise, 52–53 reactions description of, 53 oxidants, 56–57 oxidations, 53–54 reductants, 57–58 reductions, 54–56 wetland soils, 291 zones, 51 Reductants, abiotic environmental, 57 Reductive dechlorination anaerobic oxidation, 158 cometabolic, 142–144 description of, 55 electron acceptors, 159–160 electron donor selection for, 147–158 field studies of, 160–164 groundwater solute transport of chlorinated ethenes, 170–177 halorespiring microorganisms, 144–147 high constituent concentration areas, 169–170 low constituent concentration areas, 169 microorganisms for electron donors for, 147–158 halorespiring, 144–147 hydrogen production, 149–151 nutrients, 158 mixture of compounds effect, 155–158 oxidation-reduction potential effects, 142 in situ reactive zones for biofilm, 168–169 development considerations, 160–163 fermentation, 167–168 natural surfactant effect, 165–167 performance data, 177–183 reagent delivery, 164–165 temperature effects, 158 tetrachloroethane, 141–144 Remediation evolution of, 7–11 extractive techniques, 7–8 goal of, 14 phytoremediation, See Phytoremediation Resource Conservation and Recovery Act, 314 Retardation, equations, 85–86 Rhizodegradation, 252–256, 258 Rhizofiltration, 256–259, 258, 299 Rhizosphere, 254 Risk assessment of, 2–3
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description of, 2
S Sediments phytoremediation of, 260–261, 263 soil organic matter vs., 42–43 treatment wetland, 293 Selenium biodegradation, 108 Sequestration, 33 Smectites, 40 Sodium permanganate, 214 Soil characteristics of, 40–41 distribution coefficient of, 45 metals concentration in, 186, 241–242 sorption coefficients, 43–48 in treatment wetlands biological influences, 292 cation exchange capacity, 289–290 characteristics of, 287–289 clays, 287 hydric, 287 microbial processes, 292–293 mineral, 287 organic, 287–288 oxidation reactions, 290–291 peats, 287–288 pH, 292 reduction reactions, 290–291 Soil organic matter composition of, 42 critical level of, 46 definition of, 41–42 humic, 42 nonhumic, 42 sediments vs., 42–43 Soil water composition of, 42 functions of, 42 Solid waste, 318 Solubility definition of, 20–21 single compound in an immiscible liquid, 21 single compound is an immiscible liquid, 22 SOM, See Soil organic matter Sorption coefficients competitive, 49 description of, 38–39 factors that affect, 39, 48–51 kinetic considerations, 50 pH effects, 49 soil, 43–48
temperature effects, 48 wetland soils, 289 Source areas, 8 Source zones definition of, 72 types of, 72, 76 Streams, recharging, 79
T Temperature reductive dechlorination effects, 158 sorption coefficient effects, 48 Tetrachloroethane hydrogenolysis of, 141 phytodegradation of, 250 reductive dechlorination of, 135, 141–144, 152–153 treatment wetland removal of, 306 1,1,2,2-Tetrachloroethane, treatment wetland removal of, 306 Toluene dioxygenase, 30–31 Toluene monooxygenase, 30 Toxicity characteristic leaching procedure, 241 Transport processes, description of, 14–15 Treatment wetlands, See Wetlands, constructed treatment Trees, in phyto-cover system evaluations of, 356–358 leaves of, 357–358 replacement of, 362 root systems of, 349 selection of, 321–322, 331–332 soil moisture monitoring, 349–350 stem evaluations, 356 1,1,1-Trichloroethane hydrolysis effects, 23–24 reductive dechlorination of, 159 Trichloroethylene, enhanced reductive dechlorination of, 135 Trinitrotoluene biodegradation, 104
U Ultimate biodegradation, 26
V Vapor pressure, 18 Vicinal dehalogenation, 55
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Volatile organic compounds, Henry's law constant, 19 Volatilization definition of, 89 natural attenuation effects, 89 treatment wetland use of, 294–295
W Water hyacinth-based wetland systems, 282–283 Wetlands, constructed treatment algae in, 279–280 bacteria in, 278–279 benefits of, 272 components of, 270–271 contaminant removal mechanisms hydraulic retention time, 297–299 partitioning and storage, 295–296 volatilization, 294–295 description of, 270 dissolved oxygen concentration, 294 in Europe, 272–273 fish in, 310 free water surface, 276 fungi in, 278–279 groundwater remediation use of, 299–310 horizontal flow systems characteristics of, 276–277 emergent macrophyte-based systems with, 285 hydrology of, 309–310 hydroperiods, 292 inorganics removal using, 309 issues regarding, 273–274 metals removal, 300–305 morphology of, 309–310 popularity of, 271 regulation of, 272, 275–276 research of, 274 soils
biological influences, 292 cation exchange capacity, 289–290 characteristics of, 287–289 clays, 287 hydric, 287 microbial processes, 292–293 mineral, 287 organic, 287–288 oxidation reactions, 290–291 peats, 287–288 pH, 292 reduction reactions, 290–291 subsurface flow, 276–277 toxic organics removal using, 306–309 in United States, 272–273 vegetation in description of, 279, 281–282 duckweeds, 283–284 emergent macrophyte-based systems description of, 284 with horizontal subsurface flow, 285 with vertical subsurface flow, 285 free-floating macrophyte-based systems, 282–284 multistage macrophyte-based systems, 287 submerged macrophyte-based systems, 285–287 water hyacinths, 282–283 vertical flow systems characteristics of, 277–278 emergent macrophyte-based systems with, 285 wildlife considerations, 274–275, 310
Z ZoBell's solution, 399–400 Zwitterions, 217
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