Natural and Enhanced Remediation Systems

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NATURAL and ENHANCED REMEDIATION SYSTEMS

L1282_fm_frame Page 2 Monday, June 18, 2001 10:20 AM

NATURAL and ENHANCED REMEDIATION SYSTEMS Suthan S. Suthersan

LEWIS PUBLISHERS Boca Raton London New York Washington, D.C.


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


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



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).


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


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.



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.



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


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


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


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


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


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


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


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


4


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



5

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.


6


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



7

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,


8


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.



9

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.


10


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.



11

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.



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


14


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


CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS

15

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.



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


1

O R C

C

NH 2

O

R C

C

OR'

O

General Formula

O

Functional Group

Table 2.1 (Cont.) L1282_C02_frame Page 17 Monday, June 18, 2001 10:17 AM

17


18


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.



19

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


20


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.



21

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.


22


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



23

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


24


(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),



25

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 ( [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,


Log [Cs]


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.


40


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



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


42


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



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


44


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

L1282_C02_frame Page 45 Tuesday, June 19, 2001 12:58 PM


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


46


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)



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


48


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



49

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.


50


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.



51

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.


52


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



53

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


54


• • • • •

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:



55

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


56


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)



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:


58


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.

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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.”


MONITORED NATURAL ATTENUATION

65

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


66


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.



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



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 ( 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.



77

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.



3.4

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)



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.


82


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.



83

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)


84


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)



3.4.2

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


86


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)



87

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



89

• 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


90


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),



91

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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93

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


94


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


96


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





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.



111

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.


112


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., 1,000 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


IN SITU REACTIVE ZONES

169

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.

L1282_C04-B_frame Page 170 Tuesday, June 19, 2001 1:06 PM

170


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:



171

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)


172


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.



173

 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)


174


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)



175

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       


176


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.


Aqueous-phase concentration


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





179

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,


180


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.


IN SITU REACTIVE ZONES 181


182

Figure 4.21


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).



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


Figure 4.23

Concentration

Groundwater Contamination Concentrations vs. Time


184

Ethylene



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.

L1282_C04-B_frame Page 186 Tuesday, June 19, 2001 1:07 PM

186


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



187

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.


188


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.



189

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


190


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



191

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


192


[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)



193

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.


194


-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



195

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.


196


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.



197

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:


198


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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4.2.4

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

L1282_C04-C_frame Page 200 Monday, June 18, 2001 9:44 AM

200


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


202


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



203

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


204


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.



205

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


206


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.



207

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


208

4.4.3


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.



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


210

Figure 4.36


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.



211

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)


212


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



213

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


214


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)



215

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


216


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.



217

• • • •

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,


218


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



219

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


220


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.



DNAPL

221

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


222


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



223

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)


224


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).

L1282_C04-C_frame Page 225 Thursday, March 6, 2003 11:45 AM


225

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


226


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).



227

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.


228


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



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:


230


• 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.



4.5.2

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).


232


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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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.


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

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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


PHYTOREMEDIATION

Figure 5.6c

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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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261

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:


262


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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263

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.


264


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


PHYTOREMEDIATION

•

•

•

•

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


266


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.


PHYTOREMEDIATION

5.4.5

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


270


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


CONSTRUCTED TREATMENT WETLANDS

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,


272


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


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


288


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



289

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)


290


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.



291

When ORP >100 mV, conditions are termed aerobic because dissolved oxygen is available. When ORP

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