EPA/540/R-95/532

September 1995

BIOREMEDIATION OF HAZARDOUS WASTES:

Research, Development, and Field Evaluations

Biosystems Technology Development Program

Office of Research and Development

U.S. Environmental Protection Agency

U.S. Environmental Protection Agency

Ada, OK; Athens, GA; Cincinnati, OH; Gulf Breeze, FL;

and Research Triangle Park, NC
Disclaimer

The information in this document has been funded wholly or in part by the U.S. Environmental

Protection Agency (EPA) and has been reviewed in accordance with EPA’s peer and administrative

review policies and approved for presentation and publication. Mention of trade names or commercial products does not constitute endorsement or recommendation for use.

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Contents

Page

Executive Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ix

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xi

Section 1: Bioremediation Field Initiative . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

Intrinsic Bioremediation of Trichloroethylene at the St. Joseph Aquifer/Lake Michigan Interface: 

A Role for Iron and Sulfate Reduction

Jack Lendvay, Mike McCormick, Peter Adriaens, University of Michigan, Ann Arbor, MI . . . . . . . . . . 3

Modeling Intrinsic Remediation as Ground-Water Discharges to a Lake: 

The Trichloroethylene Plume at St. Joseph, Michigan

Sean M. Dean, Nikolaos D. Katopodes, University of Michigan, Ann Arbor, MI . . . . . . . . . . . . . . . . . 6

The U.S. Environmental Protection Agency’s Development of Bioventing

Gregory Sayles, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

Section 2: Field Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

A Review of Intrinsic Bioremediation of Trichloroethylene in Ground Water at Picatinny Arsenal,

New Jersey, and St. Joseph, Michigan

John T. Wilson, Don Kampbell, James Weaver, Barbara Wilson, U.S. EPA, Ada, OK;

Tom Imbrigiotta, Ted Ehlke, U.S. Geological Survey, Trenton, NJ . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

Intrinsic Bioremediation of a Gasoline Plume: Comparison of Field and Laboratory Results

Morton A. Barlaz, Melody J. Hunt, Sreenivas Kota, Robert C. Borden, North Carolina State

University, Raleigh, NC . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Toxicity Effects on Methanogenic Degradation of Phenol in Ground Water

Barbara A. Bekins, E. Michael Godsy, Ean Warren, U.S. Geological Survey, Menlo Park, CA . . . . . 20

A Multiphase, Multicomponent Numerical Model of Bioventing With Nonequilibrium Mass Exchange

Linda M. Abriola, John R. Lang, Klaus M. Rathfelder, University of Michigan, Ann Arbor, MI . . . . . . 22

Aromatic Hydrocarbon Biotransformation Under Mixed Oxygen/Nitrate Electron Acceptor Conditions

Liza P. Wilson, Peter C. D’Adamo, Edward J. Bouwer, The Johns Hopkins University,

Baltimore, MD. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24

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Nutrient Transport in a Sandy Beach

Brian A. Wrenn, Makram T. Suidan, B. Loye Eberhardt, Gregory J. Wilson, University of

Cincinnati, Cincinnati, OH; Kevin L. Strohmeier, Environmental Technologies and Solutions,

Inc., Covington, KY; Albert D. Venosa, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

Bioremediation of Crude Oil Intentionally Released on the Shoreline of Fowler Beach, Delaware

Albert D. Venosa, John R. Haines, U.S. EPA, Cincinnati, OH; Makram T. Suidan, Brian A.

Wrenn, B. Loye Eberhardt, Miryam Kadkhodayan, Edith Holder, University of Cincinnati,

Cincinnati, OH; Kevin L. Strohmeier, Environmental Technologies and Solutions, Inc.,

Covington, KY; Dennis King, Kingstat Consulting, Fairfield, OH; Bennet Anderson, Delaware

Department of Natural Resources and Environmental Control, Dover, DE . . . . . . . . . . . . . . . . . . . . . 29

Dynamics of Oil Degradation in Coastal Environments: Effect of Bioremediation Products and Some

Environmental Parameters

Marirosa Molina, Rochelle Araujo, U.S. EPA, Athens, GA; Jennifer R. Bond, DYNCORP,

Athens, GA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

Progress Toward Verification of Intrinsic Cobioremediation of Chlorinated Aliphatics

Mark Henry, Michigan Department of Natural Resources, Oscoda, MI . . . . . . . . . . . . . . . . . . . . . . . . 35

Development and Capabilities of the National Center for Integrated Bioremediation Research and

Development (NCIBRD)

Mark Henry, Michigan Department of Natural Resources, Oscoda, MI . . . . . . . . . . . . . . . . . . . . . . . . 36

Phytoremediation of Petroleum-Contaminated Soil: Laboratory, Greenhouse, and Field Studies

M. Katherine Banks, A. Paul Schwab, Kansas State University, Manhattan, KS . . . . . . . . . . . . . . . . 37

Intrinsic Bioremediation of Fuel Contamination in Ground Water at a Field Site

Don H. Campbell, Robert S. Kerr Environmental Research Laboratory, Ada, OK; 

T.H. Wiedemeier, Parsons Engineering Science, Inc., Denver, CO; J.E. Hansen, U.S. Air Force

Center for Environmental Excellence, Brooks Air Force Base, TX. . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

Section 3: Performance Evaluation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

Detoxification of Model Compounds and Complex Waste Mixtures Using Indigenous and Enriched

Microbial Cultures

K.C. Donnelly, Jeannine L. Capizzi, Ling-Yu He, Henry J. Huebner, Texas A&M University,

College Station, TX . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

Assessing the Genotoxicity of Complex Waste Mixtures

Larry D. Claxton, Virginia S. Houk, Sarah H. Warren, Thomas J. Hughes, and Susan E.

George, U.S. EPA, Research Triangle Park, NC . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

Contents (continued)

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Section 4: Pilot-Scale Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

In Situ Bioremediation of Trichloroethylene With Burkholderia Cepacia PR1: Analysis of Parameters

for Establishing a Treatment Zone

Richard A. Snyder, University of West Florida, Pensacola, FL; M. James Hendry, John R.

Lawrence, Environment Canada, Saskatoon, Canada . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

Characterization of Trichloroethylene-Degrading Bacteria From an Aerobic Biofilter

Alec Breen, Todd Ward, Ginger Reinemeyer, John Loper, Rakesh Govind, University of

Cincinnati, Cincinnati, OH; John Haines, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . 51

Anaerobic/Aerobic Degradation of Aliphatic Chlorinated Hydrocarbons in an Encapsulated 

Biomass Biofilter

Rakesh Govind, P.S.R.V. Prasad, University of Cincinnati, Cincinnati, OH; Dolloff F. Bishop,

U.S. EPA, Cincinnati, OH. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

Operation and Optimization of Granular Air Biofilters

Francis Lee Smith, George A. Sorial, Makram T. Suidan, Amit Pandit, Pratim Biswas, University

of Cincinnati, Cincinnati, OH; Richard C. Brenner, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . . . 57

Abiotic Fate Mechanisms in Soil Slurry Bioreators

John A. Glaser, Paul T. McCauley, U.S. EPA, Cincinnati, OH; Majid A. Dosani, Jennifer S. Platt,

E. Radha Krishnan, IT Corporation, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

Design and Testing of an Experimental In-Vessel Composting System

Carl L. Potter, John A. Glaser, U.S. EPA, Cincinnati, OH; Majid A. Dosani, Srinivas Krishnan,

Timothy A. Deets, E. Radha Krishnan, IT Corporation, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . 64

Integrated Systems To Remediate Soil Contaminated With Wood Treating Wastes

Makram T. Suidan, Amid P. Khodadoust, Gregory J. Wilson, Karen M. Miller, 

University of Cincinnati, Cincinnati, OH; Carolyn M. Acheson, Richard C. Brenner, U.S. EPA,

Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66

Biological Treatment of Contaminated Soils Using Redox Control

Margaret J. Kupferle, Tiehong L. Huang, Yonggui Shan, Maoxiu Wang, Guanrong You,

University of Cincinnati, Cincinnati, OH; Gregory D. Sayles, Carolyn M. Acheson, U.S. EPA,

Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

Development of a Sulfate-Reducing Bioprocess To Remove Heavy Metals From Contaminated 

Water and Soil

Munish Gupta, Makram T. Suidan, University of Cincinnati, Cincinnati, OH; Gregory D. Sayles,

Carolyn M. Acheson, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71

Development of Techniques for the Bioremediation of Chromium-Contaminated Soil and Ground Water

Michael J. McInerney, Nydia Leon, Veronica E. Worrell, John D. Coates, University of

Oklahoma, Norman, OK. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

Contents (continued)

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Bioremediation of Chlorinated Pesticide-Contaminated Sites Using Compost

James C. Young, Jean-Marc Bollag, Raymond W. Regan, Pennsylvania State University,

University Park, PA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75

Reductive Electrolytic Dechlorination

John W. Norton, Jr., Makram T. Suidan, University of Cincinnati, Cincinnati, OH; Carolyn M.

Acheson, Albert D. Venosa, U.S. EPA, Cincinnati, OH. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76

Biological Ex Situ Treatment of Soil Contaminated With Polynuclear Aromatic Hydrocarbons

Carl L. Potter, U.S. EPA, Cincinnati, OH; Roy C. Haught, IT Corporation, Cincinnati, OH. . . . . . . . . 77

Effectiveness of Gas-Phase Bioremediation Stimulating Agents (BSAs) for Unsaturated 

Zone In Situ Bioremediation

James G. Uber, Ronghui Liang, R. Scott Smith, University of Cincinnati, Cincinnati, OH;

Paul T. McCauley, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78

Section 5: Process Research. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 81

Monitoring Crude Oil Mineralization in Salt Marshes: Use of Stable Carbon Isotope Ratios

Andrew W. Jackson, John H. Pardue, Louisiana State University, Baton Rouge, LA . . . . . . . . . . . . . 83

Mercury and Arsenic Biotransformation

Ronald S. Oremland, U.S. Geological Survey, Menlo Park, CA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 86

Monod Degradation Kinetics of Quinoline in Natural and Microbially Enriched Methanogenic Microcosms

E. Michael Godsy, Ean Warren, Barbara A. Bekins, U.S. Geological Survey, Menlo Park, CA . . . . . 87

Stimulating the Biotransformation of Polychlorinated Biphenyls

John F. Quensen, III, Stephen A. Boyd, James M. Tiedje, Michigan State University, East

Lansing, MI. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89

Bioaugmentation for In Situ Co-metabolic Biodegradation of Trichloroethylene in Ground Water

Junko Munakata Marr, Perry L. McCarty, Stanford University, Stanford, CA; V. Grace Matheson,

Larry J. Forney, James M. Tiedje, Michigan State University, East Lansing, MI; Stephen

Francesconi, Malcolm S. Shields, University of West Florida, Pensacola, FL; 

P.H. Pritchard, U.S. EPA, Gulf Breeze, FL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

Biodegradation of Chlorinated Solvents

Larry Wackett, Lisa Newman, Sergey Selifonov, University of Minnesota, St. Paul, MN; 

Peter Chapman, Michael Shelton, U.S. EPA, Gulf Breeze, FL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96

Biological and Nutritional Factors Affecting Reductive Dechlorination of Chlorinated Organic Chemicals

Dingyi Ye, National Research Council, Athens, GA; W. Jack Jones, U.S. EPA, Athens, GA . . . . . . . 98

Predicting Heavy Metal Inhibition of the In Situ Reductive Dechlorination of Organics at the Petro

Processor’s Superfund Site

John H. Pardue, Louisiana State University, Baton Rouge, LA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

Contents (continued)

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Effect of Primary Substrate on the Reduction of 2,4-Dinitrotoluene

Jiayang Cheng, Makram T. Suidan, University of Cincinnati, Cincinnati, OH; Albert D. Venosa,

U.S. EPA, Cincinnati, OH. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 104

Surfactants in Sediment Slurries: Partitioning Behavior and Effects on Apparent Polychlorinated

Biphenyl Solubilization

Jae-Woo Park, John F. Quensen, III, Stephen A. Boyd, Michigan State University, East

Lansing, MI. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 108

Partial Characterization of an Anaerobic, Aryl, and Alkyl Dehalogenating Microorganism

Xiaoming Zhang, National Research Council, Athens, GA; W. Jack Jones, John E. Rogers,

U.S. EPA, Athens, GA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109

Microbial Degradation of Petroleum Hydrocarbons in Unsaturated Soils: The Mechanistic Importance of

Water Potential and the Exopolymer Matrix

Patricia A. Holden, James R. Hunt, Mary K. Firestone, University of California, Berkeley, CA . . . . . 111

Metabolic Indicators of Anaerobic In Situ Bioremediation of Gasoline-Contaminated Aquifers

Harry R. Beller, Martin Reinhard, Alfred M. Spormann, Stanford University, Stanford, CA. . . . . . . . . 113

Contaminant Dissolution and Biodegradation in Soils Containing Nonaqueous-Phase Organics

Larry E. Erickson, L.T. Fan, J. Patrick McDonald, George X. Yang, Satish K. Santharam,

Kansas State University, Manhattan, KS. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114

Protein Expression of Mycobacteria That Metabolize Polycyclic Aromatic Hydrocarbons

David E. Wennerstrom, University of Arkansas for Medical Sciences, Little Rock, AR;

Carl E. Cerniglia, National Center for Toxicological Research, Jefferson, AR . . . . . . . . . . . . . . . . . . . 115

Section 6: Hazardous Substance Research Centers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117

Co-metabolic Biodegradation Kinetics of Trichloroethylene in Unsaturated Soils

Karen L. Skubal, Peter Adriaens, University of Michigan, Ann Arbor, MI. . . . . . . . . . . . . . . . . . . . . . . 119

The Effect of Water Potential on Biodegradation Kinetics and Population Dynamics

Astrid Hillers, Peter Adriaens, University of Michigan, Ann Arbor, MI. . . . . . . . . . . . . . . . . . . . . . . . . . 120

Anaerobic-Aerobic Bioventing Development

Gregory Sayles, U.S. EPA, Cincinnati, OH; Munish Gupta, Makram T. Suidan, University of

Cincinnati, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

Development of Co-metabolic Bioventing: Laboratory Tests

Gregory Sayles, U.S. EPA, Cincinnati, OH; Alan Zaffiro, Jennifer Platt, IT Corporation,

Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123

Contents (continued)

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Evaluating the Environmental Safety of Using Commercial Oil Spill Bioremediation Agents

Jeffrey L. Kavanaugh, University of West Florida, Pensacola, FL; C. Richard Cripe, Carol B.

Daniels, U.S. EPA, Gulf Breeze, FL; Rochelle Araujo, U.S. EPA, Athens, GA; Joe E. Lepo,

University of West Florida, Pensacola, FL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125

UNIFAC Phase Equilibrium Modeling To Assess the Bioavailability of Multicomponent

Nonaqueous-Phase Liquids Containing Polycyclic Aromatic Hydrocarbons

Catherine A. Peters, Princeton University, Princeton, NJ . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 126

Field Evaluation of Pneumatic Fracturing Enhanced Bioremediation

Sankar N. Venkatraman, David S. Kosson, Rutgers University, Piscataway, NJ;

Thomas M. Boland, John R. Schuring, New Jersey Institute of Technology, Newark, NJ. . . . . . . . . . 127

Solids Suspension Characteristics Related to Slurry Biotreatment Performances

J.-W. Jim Tzeng, Paul T. McCauley, John A. Glaser, U.S. EPA, Cincinnati, OH . . . . . . . . . . . . . . . . . 128

Contents (continued)

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Executive Summary

The U.S. Environmental Protection Agency’s (EPA’s)

Office of Research and Development (ORD) hosted the

eighth annual Symposium on Bioremediation of Hazardous Wastes: Research, Development, and Field Evaluations in Rye Brook, New York, August 8-10, 1995. More

than 250 people attended, including leading bioremediation researchers, field personnel from federal, state,

and local agencies, and representatives from industry

and academia. Three speakers opened the symposium

with introductions and background information on bioremediation research.

Fran Kremer, Coordinator of the Bioremediation Field

Initiative and Symposium Chairperson, began by introducing several members of the scientific steering committee of the Biosystems Technology Development

Program (BTDP). The BTDP draws on ORD scientists

who possess unique skills and expertise in biodegradation, toxicology, engineering, modeling, biological and

analytical chemistry, and molecular biology.

George Pavlou, Deputy Director of EPA’s Region 2

Emergency and Remedial Response Division in New

York City, presented a summary of recent advances in

bioremediation technology. Mr. Pavlou noted that new

techniques have increased the availability of in situ contaminants to biological degradation, and they have also

increased the range of contaminants that can be treated

biologically. EPA scientists have learned that microorganisms are not only able to transform simple hydrocarbons but also such toxic and resistant contaminants as

chlorinated aromatics and heavy metal salts. Mr. Pavlou

provided a regional perspective on these developments,

describing how bioremediation has been put to use in

Region 2.

Timothy Oppelt, Director of the National Risk Management Research Laboratory in Cincinnati, Ohio, discussed the future of bioremediation on the national

level. He explained that bioremediation has been essential to the development of cost-effective cleanup technology and that it has already been used at more than

450 sites. This technology could potentially save hundreds of millions of dollars in future cleanup costs. While

bioremediation is now predominantly used for petroleum

decontamination, it will be applied to a wider range of

sites in the future. Mr. Oppelt concluded by describing

the reorganization of research under ORD and by warning that funding shortages may hinder the development

of bioremediation technology.

The 33 papers and 22 posters presented at the conference highlighted recent program achievements and research projects aimed at bringing bioremediation into

more widespread use. Taken as a whole, these topic areas

represent a comprehensive approach to bioremediation

of hazardous waste sites. The presentations were organized into five key research and program areas:

• Bioremediation Field Initiative: This initiative was instituted in 1990 to collect and disseminate performance data on bioremediation techniques from field

application experiences. The Agency assists regions

and states in conducting field tests and in carrying

out independent evaluations of site cleanups using

bioremediation. Through this initiative, tests are under way at Superfund sites, Resource Conservation

and Recovery Act corrective action facilities, and underground storage tank sites. Three papers presented at the symposium were devoted to this key

program area.

• Field Research: Once a bioremediation approach has

proven effective in a laboratory or pilot-scale treatability study, it must be monitored and evaluated at a

field site. The objective of this level of research is

to demonstrate that the particular bioremediation

process performs as expected in the field. For most

bioremediation technologies, certain key factors concerning applicability, such as cost effectiveness, cannot be thoroughly evaluated until the approach is

scaled up and field tested. Ten papers and two posters provided information on recent field research.

• Performance Evaluation: Performance evaluation involves assessing the extent and rate of cleanup for

particular bioremediation processes as well as monitoring the environmental fate and effects of contaminants and their biological byproducts. Two papers

and one poster addressed this area.

• Pilot-Scale Research: Pilot-scale research provides

information on the operation and control of bioremediation technologies and the management of process-related residuals and emissions to enable the

full-scale application of a technology. Given the expanding base of experience with various bioremediation methods, the need for pilot-scale research is

ix
increasing. Ten papers and three posters were presented concerning research based on microcosms of

field sites.

• Process Research: Process research involves isolating and identifying microorganisms that carry out

biodegradation processes as well as developing techniques for modeling and monitoring such processes.

This research is fundamental to the development of

new biosystems for treatment of environmental pollutants in surface waters, sediments, soils, and subsurface materials. Nine papers addressed this critical

area, focusing on the role of metals and chlorinated

organics in bioremediation. In addition, six poster

presentations discussed process research.

• Hazardous Substances Research Centers: In addition

to presentations on research being carried out under

the BTDP, the symposium included nine poster presentations from the EPA Hazardous Substance Research Centers (HSRC). The scientists and engineers

involved in HSRC conduct EPA research sponsored

by the following centers: the Northeast Hazardous

Substance Research Center (Regions 1 and 2), the

Great Lakes and Mid-Atlantic Hazardous Substance

Research Center (Regions 3 and 5), the South/

Southwest Hazardous Substance Research Center

(Regions 4 and 6), the Great Plains and Rocky Mountain Hazardous Substance Research Center (Regions 7 and 8), and the Western Region Hazardous

Substance Research Center (Regions 9 and 10).

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Introduction

Bioremediation is one of the most promising technological approaches to the problem of hazardous waste. This

process relies on microorganisms such as bacteria or

fungi to transform hazardous chemicals into less toxic

or nontoxic substances. There are several reasons why

such biological transformation is often more attractive

than direct chemical or physical treatment. Microorganisms typically:

• Directly degrade contaminants rather than merely

transferring them from one medium to another.

• Employ metabolic degradation pathways that can terminate with benign waste products (e.g., carbon dioxide and water).

• Derive the food energy necessary to degrade contaminants from the contaminants themselves.

• Can be used in situ to minimize disturbance of the

cleanup site.

For these reasons, microorganisms can be effective,

economical, and nondisruptive tools for eliminating hazardous chemicals. Until recent years, however, the use

of bioremediation was limited by the lack of a thorough

understanding of biodegradation processes, their appropriate applications, their control and enhancement in

environmental matrices, and the engineering techniques required for broad application of the technology.

Because the U.S. Environmental Protection Agency (EPA)

believes that bioremediation offers an attractive alternative to conventional methods of hazardous waste cleanup,

it has developed a strategic plan for its acceptance

and use by the technical and regulatory communities.

The Agency’s strategic plan is centered on site-directed

bioremediation research to expedite the development

and use of relevant technology. EPA’s Office of Research

and Development (ORD) developed an integrated Bioremediation Research Program to advance the understanding, development, and application of bioremediation solutions to hazardous waste problems threatening

human health and the environment.

Related bioremediation studies are being carried out at

five EPA Hazardous Substance Research Centers

(HSRCs) under the direction of ORD’s National Center

for Extramural Research and Quality Assurance

(NCERQA). EPA was authorized to establish these centers by provisions in the 1986 amendments to the Superfund law calling for research into all aspects of the

“manufacture, use, transportation, disposal, and management of hazardous substances.”

EPA’s bioremediation research efforts have produced

significant results in the laboratory, at the pilot scale, and

in the field. The many accomplishments include aquifer

restoration, soil cleanup, process characterization, and

technology transfer. Research also focuses on extending the range of substances that can be treated with

biological agents. This symposium was held to present

and discuss recent developments in bioremediation research undertaken during 1994 and 1995 under the

Biosystems Technology Development Program.

In this document, abstracts of paper and poster presentations from the symposium are organized within five

key research and program areas:

• Bioremediation Field Initiative

• Field research

• Performance evaluation

• Pilot-scale research

• Process research

The last section of this document includes abstracts of

presentations on bioremediation research performed as

part of the HSRC program.

xi
Section 1

Bioremediation Field Initiative

The Bioremediation Field Initiative is one of the major components of EPA’s Bioremediation

Research Program. The Initiative was undertaken in 1990 to expand the nation’s field experience

in bioremediation techniques. The Initiative’s goals are to more fully assess and document the

performance of full-scale bioremediation applications, to create a database of current field data on

the treatability of contaminants, and to assist regional and state site managers using or considering

bioremediation. The Initiative is currently tracking bioremediation activities at more than 400 sites

under government and private-sector jurisdiction, in both the United States and Canada. Performance evaluations are currently being conducted at nine sites, three of which were reported on at

this symposium.

Data were presented from work at the St. Joseph, Michigan, Superfund site on the use of iron as

an electron acceptor and the potential for natural attenuation of chlorinated solvents. A presentation

was given on the modeling of the natural attenuation of solvents at the ground-water and lake

interface. Studies also were carried out on the design and field applications of bioventing in the

bioremediation of jet fuel spills.

1
Intrinsic Bioremediation of Trichloroethylene at the St. Joseph Aquifer/

Lake Michigan Interface: A Role for Iron and Sulfate Reduction

Jack Lendvay, Mike McCormick, and Peter Adriaens

University of Michigan, Ann Arbor, Michigan

Introduction

The anaerobic aquifer at the St. Joseph, Michigan, National Priorities List (NPL) site was contaminated with

trichloroethylene (TCE), which has been shown to have

dechlorinated to cis- and trans-dichloroethylene (DCE),

vinyl chloride (VC), ethylene, and ethane. These products occur as a result of natural attenuation processes,

presumably under methanogenic conditions (1).

The flux of all alkyl halides into Lake Michigan is of major

public concern because of the suspected carcinogenicity of VC. As the plume moves toward the aerobic surface water, the dominant redox conditions can be

expected to change because of wave action and vertical

seepage, which promote the interchange of oxygen-rich

lake water and anaerobic ground water. This presentation provides preliminary results of laboratory investigations geared toward determining the prevailing redox

processes and the potential for natural attenuation of

chlorinated solvents at the interface.

Background

Three of the most important redox processes in the

natural anaerobic environment are the coupling of the

oxidation of organic matter to iron (Fe) (III) reduction, to

sulfate reduction, and to methanogenesis. These three

processes are considered mutually exclusive which, in

the anaerobic subsurface environment, results in the

development of spatially or temporally distinct redox

zones. It has been demonstrated in aquatic sediment

and aquifer samples that Fe(III)-reducing bacteria can

outcompete sulfate reducers, as well as methanogens,

for organic matter (2, 3).

Mineral-bound Fe(III) has been shown to contribute significantly to the total oxidation capacity of both pristine

and contaminated aquifers, as it often represents the

most abundant anaerobic terminal electron acceptor (4,

5). The speciation of iron in aquifer solids is greatly

influenced by microbial processes, particularly under

redox conditions favoring sulfate- and Fe(III)-reduction.

Depending on the temporal or spatial succession in the

development of subsurface redox conditions, Fe(II) produced by iron reducers precipitates as iron sulfides once

sulfate-reducing conditions develop, or as iron oxides

such as magnetite (Fe3O4) in the absence of sulfide.

Alternatively, biogenically produced sulfide may precipitate as FeS(1-x) (mackinawite) after reductive dissolution

of Fe(III) minerals. The iron sulfide and iron oxide minerals thus formed may then contribute significantly to the

reduction capacity of aquifer solids and, in turn, play a

major role in the fate of organic contaminants. The

presence of these precipitated minerals is direct evidence for past or present iron- or sulfate-reducing conditions.

Materials and Methods

A three-stage iron analysis was performed on sediments

collected from the same depths in sampling wells 55AB

(upstream) and 55AD (near shore) to evaluate the occurrence of oxidized and reduced iron minerals, and to

calculate inorganic reducing equivalents present in aquifer materials.

• Bioavailable iron (Fe(III)): Microbial Fe(III)-reduction

has been shown to predominantly use amorphous

oxyhydroxides and goethite as terminal electron acceptors. Quantitation of these minerals can be approximated by extracting sediments according to

Lovley and Phillips (6).

• Ferrous monosulfides and amorphous iron oxides

(Fe(II)): Reduced Fe(II) minerals resulting from microbial iron and sulfate reduction, predominantly ferrous sulfides and iron oxides, can be quantified using

a 24-hr extraction with 0.5 M hydrochloric acid (HCl)

(7). As this extraction removes bioavailable iron as

well, FeS can be approximated by substration.

• Siderite, crystalline iron oxide, and magnetite (Fe(II)):

This extraction represents the precipitated ferrous

3
iron fraction in the absence of sulfate reduction. Wet

sediment samples were extracted with 5 M HCl for a

21-day period, and analyzed according to Heron and

Christensen (5).

This information was then used in conjunction with available data on ground-water chemistry (pH, redox, carbonate, sulfate, sulfide) and compiled in a chemical

equilibrium model (MINEQL+) to predict speciation of

iron. Precipitation of ferrous iron solids was based on

stability constants from the literature.

Results

The analysis of samples from wells 55AB and 55AD

indicated the presence of similar concentrations of

bioavailable Fe(III) (5 mmol equiv./kg) (Table 1).

Whereas precipitated iron oxides and ferrous monosulfides increased dramatically toward the shore, however,

solids representative of the absence of sulfate reduction

significantly decreased (12 versus 5 mmol equiv./kg

aquifer material, after subtraction of bioavailable iron).

This observation, based on iron extraction data, was

confirmed by chemical modeling, which predicted a predominant occurrence of siderite in the sampling point

farthest from the lakeshore and increasing occurrence

of iron sulfide and mackinawite closer to shore and into

the lake. Iron- and sulfate-reducing activity was relatively easily stimulated in these sediments.

Ongoing dechlorination experiments under sulfate- and

iron-reducing conditions, using TCE and VC, have resulted in the production of t-DCE from TCE under ironreducing conditions. Whether the production of the trans

rather than the cis isomer is indicative of an abiotic

dechlorination mechanism (by Fe(II)) remains to be elucidated. VC did not dechlorinate under sulfate-reducing

conditions during the 2 months monitored to date.

Information available from previous field studies in collaboration with the Robert S. Kerr Laboratories (Dr. John

Wilson) suggested that neither methanotrophic nor

methanogenic activity predominates in the contaminant

plume under the shoreline and Lake Michigan, based on

redox potential and oxygen and methane measurements. Redox potentials and an increase in soluble iron

concentrations between both shore samples and below

the Lake Michigan bottom, however, suggest that iron

reduction may be the dominant process in regions near

plume emergence (8).

References

1. Wilson, J.T., J.W. Weaver, and D.H. Kampbell. 1994. Intrinsic

bioremediation of TCE in ground water at an NPL Site in St.

Joseph, Michigan. In: U.S. EPA Symposium on Intrinsic Bioremediation of Ground Water. EPA/540/R-94/515. Washington, DC.

2. Barcelona, M.J., and T.R. Holm. 1991. Oxidation-reduction capacities of aquifer solids. Environ. Sci. Technol. 25:1,565-1,572.

3. Chapelle, F.H., and D.R. Lovley. 1992. Competitive exclusion of

sulfate reduction by Fe(III)-bacteria: A mechanism for producing

discrete zones of high-iron ground water. Ground Water 30:29-36.

4. Heron, G., T.H. Christensen, and J.C. Tjell. 1994. Oxidation capacity of aquifer sediments. Environ. Sci. Technol. 28:153-158.

5. Heron, G., and T.H. Christensen. 1995. Impact of sediment-bound

iron on redox buffering in a landfill leachate-polluted aquifer (Vejen,

Denmark). Environ. Sci. Technol. 29:187-192.

Table 1. Determination of Iron and Sulfate Equivalents in St. Joseph Sediments

Sample Location and

Analysis

Chemical Species

Analyzed

Chemical

Concentration Redox Equivalents

Well 55AB Iron Extractions (mg/kg soil) (mmol equiv./kg soil)

1-hr/0.5 M HCl Limited Fe(II) 219.2 ± 15.8 3.9 ± 0.3

1-day/0.5 M HCl Limited Fe(II) 210.6 ± 30.6 3.8 ± 0.6

1-hr/0.25 M HCl and 0.25 M Limited Fe(II) and 269.6 ± 29.7 4.7 ± 0.5

NH2OH ⋅ HCl Bioavailable Fe(III)

21-day/5.0 M HCl Limited Fe(II) 881.8 ± 66.2 15.8 ± 1.2

Sulfate Extractions (mg/L) (mmol equiv./L)

Ion chromatography Soluble sulfate 25.8 ± 3.5 2.2 ± 0.3

Well 55AD Iron Extractions (mg/kg soil) (mmol equiv./kg soil)

1-hr/0.5 M HCl Limited Fe(II) 333.4 ± 11.5 6.0 ± 0.2

1-day/0.5 M HCl Limited Fe(II) 445.5 ± 16.9 8.0 ± 0.3

1-hr/0.25 M HCl and 0.25 M Limited Fe(II) and 270.8 ± 8.2 4.9 ± 0.2

NH2OH ⋅ HCl Bioavailable Fe(III)

21-day/5.0 M HCl Limited Fe(II) 735.9 ± 48.7 13.2 ± 0.9

Sulfate Extractions (mg/L) (mmol equiv./L)

Ion chromatography Soluble sulfate 29.3 ± 4.2 2.4 ± 0.4

4
6. Lovley, D.R., and E.J.P. Phillips. 1987. Rapid assay for microbially

reducible ferric iron in aquatic sediments. Appl. Environ. Microbiol.

53:1,536-1,540.

7. Heron, G., C. Crouzet, A.C.M. Bourg, and T.H. Christensen. 1994.

Speciation of Fe(II) and Fe(III) in contaminated aquifer sediments

using chemical extraction techniques. Environ. Sci. Technol.

28:1,698-1,705.

8. Adriaens, P., J.V. Lendvay, N. Katopodes, S. Dean, J.T. Wilson,

and D. Kampbell. 1995. Intrinsic bioremediation of chlorinated solvents at the St. Joseph, Michigan Aquifer/Lake Michigan Interface.

EPA/600/R-95/012. In: Proceedings of the 21st Annual RREL Research Symposium, Abstract.

5
Modeling Intrinsic Remediation as Ground-Water Discharges to a Lake: 

The Trichloroethylene Plume at St. Joseph, Michigan

Sean M. Dean and Nikolaos D. Katopodes

Department of Civil and Environmental Engineering, University of Michigan, 

Ann Arbor, Michigan

Introduction

Contamination of ground water by chlorinated solvents

is widespread and has been in the forefront of public and

regulatory concern for the last decade. As a result,

considerable research efforts, both in the laboratory and

in the field, have addressed the potential for using biological processes to degrade these pollutants via either

in situ or onsite bioremediation technologies. Natural

attenuation has been observed to be responsible for

removal or partial transformation of both chlorinated and

nonchlorinated organic contaminants. At several sites,

naturally occurring reductive dechlorination has been

found to be responsible for the anaerobic transformation

of trichloroethylene (TCE) to lesser chlorinated intermediates, such as c- and t-dichloroethylene (DCE) and

vinyl chloride (VC), and to ethylene (1, 2). Because

aquifers become oxygenated near groundwater and

lake interfaces, concerns have been raised with respect

to surface water contamination by VC. In the current

study, a modeling approach addresses the fate of TCE

and its lesser chlorinated transformation products at this

anaerobic/aerobic interface. The modeling effort describes and predicts the fate of the chlorinated solvents

at and near the interface, taking into account groundwater flow rates and microbial degradation rates, as well

as the oxygenating effects of wave action and lake water

intrusion near the shore line.

Numerical Model

The numerical model for the simulation of the hydrodynamic, physicochemical, and biological processes that

take place at the lake-aquifer interface is validated

based on specific data from an application site at St.

Joseph, Michigan. Although the hydrodynamic processes are truly three-dimensional, most of the phenomena of interest, such as migration of TCE, DCE, and VC

into the lake and transfer of dissolved oxygen into the

aquifer from water infiltrating through the surf region,

can be modeled by a two-dimensional model on the

vertical plane.

Model Components

Due to the significant time-scale difference between the

near-shore circulation and wave runup and breaking in

the lake compared with the flow in the porous media,

two separate models are constructed for the corresponding hydrodynamic phenomena. The resulting flow

fields are then integrated in a single mass transport and

contaminant fate model. All three components of the

model are two-dimensional, covering a vertical plane

extending from a location inland where uniform flow and

mass flux are observed in the aquifer to a distance

inside the lake where most near-shore current activity

has diminished. The various modules are verified by

analytical solutions and intermodel comparisons.

Ground-Water Module

In the porous media, a finite-element module for variably

saturated flow has been constructed on a vertical plane.

This module uses pressure heads calculated by the lake

module as a boundary condition at the lake-aquifer interface. For the St. Joseph site, this module has been

used in conjunction with two separate grids to take

advantage of naturally defined boundary conditions.

A coarse grid has been developed extending from Lake

Michigan to a ground-water divide approximately 900 m

inland. Zero flux boundaries are prescribed at the

ground-water divide and an underlying clay layer. A

seasonally varying flux boundary condition at the ground

surface reflects recharge from rainfall. To focus on the

lake-aquifer interface region, a refined grid has been

developed extending from Lake Michigan to a point

approximately 100 m inland. One year’s output from

running this module on the coarse grid defines a sea6
sonally varying inland boundary condition for the refined

grid.

Lake Module

The near-shore/free-surface flow simulation is based on

the numerical solution of the Navier-Stokes equations

by means of the finite-element method. For turbulent

flow, a widely accepted two-equation closure model is

employed together with certain approximations near the

bed and free-surface boundaries. At high Reynolds

numbers, an upwind formulation known as the PetrovGalerkin method of weighted residuals is introduced for

the suppression of nonlinear instabilities. The model can

predict the vertical structure of the flow from the seepage face between the aquifer and the lake to the free

surface. Wave action is incorporated, and special attention is focused on wave runup and breaking. The beach

is assumed to be a porous bed so that water from the

surf and break region is allowed to infiltrate and reach

the aquifer.

For the complete formulation of the problem, bed permeability resulting in seepage through the surf region

would be computed by simultaneous solution of the

free-surface flow problem in the lake with the associated

unsaturated flow problem in the subsurface domain. In

this model, bed seepage is introduced as a boundary

condition. This eliminates the difficulty of having to deal

simultaneously with two time scales without affecting the

robustness of the model. The bed seepage is averaged

over time to provide an interface boundary condition to

the ground-water flow module.

Contaminant Transport Module

The contaminants are assumed to be well mixed laterally. The contaminant fate and transport module uses a

finite-element model to solve the two-dimensional transport equation based on the flow fields computed by the

ground-water and lake modules. This accounts for contaminant transport due to advection and dispersion in

the aqueous phase and for interphase mass transfer

due to sorption and volatilization.

Microbial transformations are incorporated using modified Monod kinetics to describe a source/sink term in

transport equation. Rate constants have been estimated

by a parallel experimental effort focusing on microbial

interactions. The microbial biomass is assumed to be

immobile below a limiting concentration at which sloughing occurs.

Conclusion

A two-dimensional finite element has been developed to

simulate the transport and biodegradation of chlorinated

solvents at and near ground-water and lake interfaces.

Example simulations consider the effects that factors

such as heterogeneities in the porous media, uncertainties in parameter estimation, and varying recharge

through the beach have on the location and concentration of a plume of chlorinated solvents.

References

1. Freedman, D.L., and J.M. Gossett. 1989. Biological reductive

dechlorination of tetrachloroethylene and trichloroethylene to ethylene under methanogenic conditions. Appl. Environ. Microbiol.

55(9):2,144-2,151.

2. Vogel, T.M., and P.L. McCarty. 1985. Biotransformation of

tetrachloroethylene to trichloroethylene, dichloroethylene, vinyl

chloride, and carbon dioxide under methanogenic conditions. Appl.

Environ. Microbiol. 49(5):1,080-1,083.

7
The U.S. Environmental Protection Agency’s

Development of Bioventing

Gregory D. Sayles

National Risk Management Research Laboratory,

U.S. Environmental Protection Agency, Cincinnati, Ohio

Research conducted in the mid to late 1980s by the U.S.

Air Force (1, 2), researchers in the Netherlands (3-6),

Texas Research Institute (7, 8), Battelle Memorial Institute (2, 9, 10, 11), Utah State University (11), and the

U.S. Environmental Protection Agency (EPA) (12),

among others, suggested that delivering air to the

vadose zone to promote biodegradation could be a lowcost means of cleaning fuel-contaminated vadose zone

soils. This approach was motivated by the attempt to

solve two different remediation development problems:

• Soil vacuum extraction for treatment of contaminated

vadose zones involved costly off-gas treatment and

only removed the volatile fraction of the contamination.

• Oxygen delivery to the vadose zone to promote aerobic biodegradation using the approaches attempted

in promoting biodegradation in ground water (namely,

delivering oxygen saturated water or aqueous solutions of hydrogen peroxide or nitrate to the contaminated area) was neither efficient nor cost-effective.

A process was needed that could deliver oxygen by

introducing air into the vadose at a rate that minimized

volatilization of the contamination. Several groups simultaneously developed what is now known as bioventing.

EPA recognized the potential cost savings of such a

technology over traditional remediation approaches and

began an aggressive bioventing development program

in 1990. The mission of the program, in essence, was

to develop bioventing so that it could be applied at as

many contaminated sites as possible. To date, EPA’s

program has demonstrated or is currently developing

the use of bioventing for the following situations:

• For operation with air injection.

• In cold climates.

• With soil warming.

• For jet fuel/aviation fuel.

• For nonfuel contaminants such as acetone, toluene,

polycyclic aromatic hydrocarbons (PAHs), and trichloroethylene (TCE).

Table 1 provides a list of EPA’s involvement in bioventing

research and development. The cumulative knowledge

of EPA, the Air Force, and Battelle Memorial Institute

regarding bioventing of fuel-contaminated sites was distilled in Principles and Practices Manual for Bioventing,

to be released in late 1995.

The next frontier for aerobic bioventing is the application

of the process to sites contaminated with chlorinated

solvents. EPA is currently involved in two laboratory and

field projects to develop “co-metabolic bioventing.”

Co-metabolic bioventing is the promotion of the aerobic

biodegradation of chlorinated solvents, such as TCE, in

the vadose zone by delivering oxygen and, if necessary,

a volatile co-metabolite to the contaminated site. EPA

projects will consider two scenarios: 1) the co-metabolite

is a co-contaminant of the chlorinated solvent, and thus

only air must be delivered; and 2) the co-metabolite

must be delivered with the air stream and must therefore

be volatile.

In summary, EPA and its collaborators, primarily the Air

Force, the U.S. Coast Guard, and Battelle, has been

successful in developing bioventing into an inexpensive,

robust process, applicable to the cleanup of many contaminated sites.

References

 1. Miller, R.N. 1990. A field-scale investigation of enhanced petroleum hydrocarbon biodegradation in the vadose zone combining

soil venting as an oxygen source with moisture and nutrient additions. Ph.D. dissertation. Utah State University, Logan, UT.

8
 2. Miller, R.N., C.C. Vogel, and R.E. Hinchee. 1991. A field-scale

investigation of petroleum hydrocarbon biodegradation in the

vadose zone enhanced by soil venting at Tyndall AFB, Florida.

In: Hinchee, R.E., and R.F. Olfenbuttel, eds. In situ bioreclamation. Stoneham, MA: Butterworth-Heinemann. pp. 283-302.

 3. Staatsuitgeverij. 1986. Proceedings of a Workshop (March 21-

21). Bodembeschermingsreeeks No. 9. Biotechnologische

Bodemsanering. Rapportnr. 851105002. ISBN 90-12-054133, Ordernr. 250-154-59; Staatsuitgeverij Den Haag, The Netherlands.

pp. 31-33.

 4. van Eyk, J., and C. Vreeken. 1988. Venting-mediated removal of

petrol from subsurface soil strata as a result of stimulated evaporation and enhanced biodegradation. Med. Fac. Landbouww.

Riiksuniv. Gent 53(4b):1,873-1,884.

 5. van Eyk, J., and C. Vreeken. 1989. Model of petroleum mineralization response to soil aeration to aid in site-specific, in situ

biological remediation. In: Jousma et al., eds. Groundwater contamination: Use of models in decision-making. Proceedings of

the International Conference on Ground-water Contamination.

Boston, MA/London, England: Kluwer. pp. 365-371.

 6. van Eyk, J., and C. Vreeken. 1989. Venting-mediated removal of

diesel oil from subsurface soil strata as a result of stimulated

evaporation and enhanced biodegradation. In: Hazardous waste

and contaminated sites, Envirotech, Vienna. Vol. 2, Session 3.

ISBN 389432-009-5. Essen, Germany: Westarp Wiss. pp. 475-485.

 7. Texas Research Institute. 1980. Laboratory-scale gasoline spill

and venting experiment. Interim Report No. 7743-5:JST. American

Petroleum Institute.

 8. Texas Research Institute. 1984. Forced venting to remove gasoline vapor from a large-scale model aquifer. Final Report No.

8210l-F:TAV. American Petroleum Institute.

 9. Hinchee, R.E., and M. Arthur. 1991. Bench-scale studies of the

soil aeration process for bioremediation of petroleum hydrocarbons. J. Appl. Biochem. Biotech. 28/29:901-906.

10. Hinchee, R.E., and S.K. Ong. 1992. A rapid in situ respiration test

for measuring aerobic biodegradation rates of hydrocarbons in

soil. J. Air Waste Mgmt. Assoc. 42(10):1,305-1,312.

11. Dupont, R.R., W.J. Doucette, and R.E. Hinchee. 1991. Assessment of in situ bioremediation potential and the application of

bioventing at a fuel-contaminated site. In: Hinchee, R.E., and R.F.

Olfenbuttel, eds. In situ bioreclamation: Applications and investigations for hydrocarbon and contaminated site remediation.

Stoneham, MA: Butterworth-Heinemann. pp. 262-282.

12. Wilson, J.T., and C.H. Ward. 1986. Opportunities for bioremediation of aquifers contaminated with petroleum hydrocarbons. J.

Ind. Microbiol. 27:109-116.

13. Ostendorf, D.W., and D.H. Kampbell. 1990. Boioremediated soil

venting of light hydrocarbons. Haz. Waste Haz. Mat. 1(4):319-334.

14. Kampbell, D.H., and J.T. Wilson. 1991. Bioventing to treat fuel

spills from underground storage tanks. J. Haz. Mat. 28:75-80.

15. Kampbell, D.H., J.T. Wilson, and C.J. Griffin. 1992. Performance

of bioventing at Traverse City, Michigan. In: U.S. EPA. Symposium on Bioremediation of Hazardous Wastes. EPA/600/R-92/126.

pp. 61-64.

Table 1. Summary of EPA Bioventing Research and Development

Project/

Location Dates Contaminants Scale Approach/Results References

Coast Guard

Station, Michigan

1990-1991 Aviation fuel Lab, pilot First test of air-injection bioventing. Showed

that air injection near the water table could

induce biodegradation in the vadose zone and

in ground water.

13-16

Hill AFB, Utah 1990-1994 Jet fuel (JP-4) Full First full-scale air-injection bioventing. Showed

that low-rate air injection could supply

biodegradation oxygen demand and produce

no measurable surface emissions of volatile

organic compounds.

17

Eielson AFB,

Alaska

1991-1994 Jet fuel (JP-4) Lab, pilot Showed that bioventing in cold climates is

feasible and that simple soil warming

techniques can increase the rate of

biodegradation when bioventing, thereby

decreasing the time required for remediation.

18

Reilly Tar SF

Site, MN

1992- PAHs Pilot Attempting to show that PAHs at wood treating

sites can be remediated with bioventing.

20

Greenwood

Chemical SF

Site, Virginia

1993-1995 Acetone,

toluene, others

Pilot Attempting to show that bioventing of sites

contaminated with non-fuel, aerobically

biodegradable organics and in low permeability

soils is feasible.

N/A

Dover AFB,

Delaware

1995- TCE Lab, pilot Attempting to show that TCE can be treated

with bioventing if the necessary co-metabolite

is present, either as a co-contaminant or

delivered as a volatile organic in the injected

air stream.

N/A

Principles and

Practices Manual

1995

release

Fuels Two

volumes

Volume 1: Principles of Bioventing

Volume 2: The Practice of Bioventing

N/A

9
16. Kampbell, D.H., C.J. Griffin, and F.A. Blaha. 1993. Comparison

of bioventing and air sparging for in situ bioremediation of fuels.

In: U.S. EPA. Symposium on Bioremediation of Hazardous

Wastes: Research, Development, and Field Evaluations.

EPA/600/R-93/054. pp. 61-67.

17. Sayles, G.D., R.C. Brenner, R.E. Hinchee, and R. Elliott. 1994.

Bioventing of jet fuel spills II: Bioventing in a deep vadose zone

at Hill AFB, Utah. In: U.S. EPA. Symposium on Bioremediation

of Hazardous Wastes: Research, Development and Field Applications. EPA/600/R-94/075. pp. 22-28.

18. Sayles, G.D., R.C. Brenner, R.E. Hinchee, A. Leeson, C.M. Vogel,

and R.N. Miller. 1994. Bioventing of jet fuel spills I: Bioventing in

a cold climate with soil warming at Eielson AFB, Alaska. In: U.S.

EPA. Symposium on Bioremediation of Hazardous Wastes: Research, Development and Field Applications. EPA/600/R-94/075.

pp. 15-21.

19. Leeson, A., R.E. Hinchee, J. Kittel, G. Sayles, C. Vogel, and R.

Miller. 1993. Optimizing bioventing in shallow vadose zones in

cold climates. Hydrological Sci. J. 38(4).

20. McCauley, P.T., R.C. Brenner, F.V. Kremer, B.C. Alleman, and

D.C. Beckwith. 1994. Bioventing soils contaminated with wood

preservatives. In: U.S. EPA. Symposium on Bioremediation of

Hazardous Wastes: Research, Development and Field Applications. EPA/600/R-94/075. pp. 40-45.

10
Section 2

Field Research

Field research is essential for evaluating the performance of full-scale bioremediation precesses

and for conducting accelerated testing on technologies that are appropriate for scaled-up application. For example, problems associated with the use of bacteria used in the laboratory include

optimizing the activity of the organism under site conditions and defining the risks associated with

introducing a non-native microorganism to the site. The objective of this level of research is to

demonstrate that the particular bioremediation process performs as expected in the field. Researchers at the symposium provided information on several ongoing field experiments.

Field studies conducted at St. Joseph, Michigan, and Rocky Point, North Carolina, sought to

determine the extent of intrinsic bioremediation in subsurface contaminant plumes. EPA found

extensive dechlorination of trichloroethylene contamination by endogenous microorganisms at St.

Joseph. At Rocky Point, researchers found considerable agreement between laboratory models of

hydrocarbon biodegradation and field observations.

At Pensacola, Florida, researchers obtained data from a site contaminated with creosote to

construct a model of toxic inhibition of bioremediation. Another group of scientists constructed a set

of models to better understand the processes of soil vapor extraction and bioventing.

Research also was performed on the use of a mixed oxygen/nitrate electron acceptor condition to

degrade aromatic hydrocarbons.

Finally, a group of presentations dealt with the use of bioremediation to clean petroleum contamination in beaches and wetlands. Researchers performed studies to determine how often nutrients

should be applied to beaches to support bacterial growth, as well as directly studied the process of

bacterial petroleum degradation on beaches. Another study examined what environmental influences might affect the biodegradation of petroleum in wetland areas.

There were two field research poster presentations. The first of these described the National Center

for Integrated Bioremediation Research and Development (NCIBRD) and its work at Wurtsmith Air

Force Base in Michigan. The second presentation described use of plant growth to enhance

bioremediation in the field.

11
A Review of Intrinsic Bioremediation of Trichloroethylene in Ground Water at

Picatinny Arsenal, New Jersey, and St. Joseph, Michigan

John T. Wilson, Don Kampbell, James Weaver, and Barbara Wilson

U.S. Environmental Protection Agency, Ada, Oklahoma

Tom Imbrigiotta and Ted Ehlke

U.S. Geological Survey, Trenton, New Jersey

Reductive dechlorination occurs frequently in large

trichloroethylene (TCE) plumes. TCE is transformed

largely to cis-dichloroethylene (cis-DCE), then to vinyl

chloride, and finally to compounds that do not contain

organic chlorine. This abstract evaluates the rate and

extent of natural reductive dechlorination of TCE in two

large plumes with similar properties.

Description of the Plumes

Both plumes originated in a release of liquid TCE. The

plume at St Joseph, Michigan, originates from an industrial park, while the plume at the Picatinny Arsenal, New

Jersey, originates in a release from a degreasing vat at

a plating shop. Cross sections of the plumes are depicted in Figures 1 and 2. Both plumes have high concentrations of TCE in the core of the plume (over 25,000

μg/L), are devoid of oxygen or nitrate, contain low concentrations of iron (II) and methane (generally less than

10 mg/L), and have relatively low concentrations of

sulfate (generally less than 15 mg/L). Both plumes have

concentrations of dissolved organic carbon that are elevated over background. The ground water in both

plumes is cold (near 10°C). The water is hard, with pH

near neutrality.

Both plumes discharge to surface water. The interstitial

seepage velocities of the plumes are very similar. The

seepage velocity of the TCE plume at St. Joseph (corrected for retardation) is near 0.1 m/day, while the velocity of the plume on the Picatinny Arsenal varies from

0.3 to 1.0 m/day. For purposes of calculation, 0.3 m/day

is used in this abstract.

Monitoring

The plume at St. Joseph was characterized by four

transects that extended across the plume, perpendicular

to ground-water flow. At each point in each transect,

water was sampled in 1.5-m vertical intervals extending

from the water table to a clay layer at the bottom of the

aquifer. Each transect contains at least 20 sampling

points. Table 1 compares the average concentration of

TCE, cis-DCE, and vinyl chloride in each transect, as

well as the highest concentration encountered. The

most distant transect was sampled from the sediments

of Lake Michigan. The plume was encountered approximately 1.5 m below the sediment surface, 100 m from

the shore line.

The plume of TCE at the Picatinny Arsenal is monitored

by a series of well clusters installed along the centerline

of the plume. Table 1 presents data from the monitoring

well in a cluster that had the highest concentration of

TCE. The data were collected in 1989.

Extent of Attenuation

Dechlorination in the plume at St. Joseph is extensive.

Vinyl chloride and cis-DCE accumulated near the spill,

then were degraded as the plume moved downgradient

(Table 1). Dechlorination in the plume at Picatinny Arsenal was also extensive. Comparing the location of the

highest concentration with the point of discharge,

dechlorination destroyed approximately 90 percent of

the TCE. Vinyl chloride and cis-DCE did not accumulate

to an appreciable extent. Because the plume at Picatinny Arsenal discharged to surface water before

dechlorination was complete, the U.S. Army installed

and continues to operate a pump-and-treat system on

the plume.

Comparison of Attenuation due to

Dilution and Dechlorination

The plume at St. Joseph has high concentrations of TCE

at its core, while the concentration of chloride in the

aquifer is low. This makes it possible to estimate the

13
contribution of dilution by comparing the accumulation

of chloride from reductive dechlorination to attenuation

of chloroethenes. Table 2 portrays the accumulation of

chloride and reduction of total organic chlorine along the

flow path. Table 2 compares water from the most concentrated sample in each transect. Based on Koc relationships and the fraction of organic carbon in the aquifer, approximately 60 percent of the TCE in the aquifer

should be in solution. TCE was largely depleted, and

sorption of cis-DCE and vinyl chloride in the aquifer

Figure 1. Cross section of the plume at St. Joseph, Michigan, as it leaves the industrial park and enters the sediments under Lake

Michigan. Concentrations are in μg/L total chloroethenes.

Figure 2. Cross section of the plume at Picatinny Arsenal, New Jersey, as it moves from its source near Building 24 and discharges

at Green Pond Brook.

14
should be minimal. We will assume that the organic

chlorine in ground water represents the pool of chlorine

available for dechlorination to chloride.

Near the source, the concentration of chloride plus potential biogenic chloride minus background chloride was

145 mg/L. Only 38 percent of this quantity was actually

chloride. Total organic and inorganic chlorine attenuated

with distance downgradient. By the time the plume

reached the lake, the concentration of total chlorine

(minus background) was 43 mg/L, which is significantly

higher than background. Apparently the plume was attenuated three- to four-fold due to dilution. Total attenuation of chloroethenes was at least 100,000-fold.

Kinetics of Reductive Dechlorination in

Ground Water

Table 3 compares first-order rate constants calculated

between transects in the plume at St. Joseph and between monitoring wells in the plume at Picatinny Arsenal. Field-scale estimates of rates are also compared

with attenuation in microcosms constructed from material collected along the flow path at Picatinny Arsenal.

There is surprising agreement in the rates of dechlorination of TCE within the same plume, between plumes,

and between microcosm studies and field-scale estimates. Nine separate estimates vary less than an order

of magnitude. The rates of degradation of vinyl chloride

and cis-DCE were comparable to the rates of degradation of TCE (Table 3).

The rates of attenuation in the two plumes are as slow

as humans experience time. In particular, they are slow

compared with the time usually devoted to site characterization. In plumes with a long residence time, on the

order of decades, however, they have significance for

protection of waters that receive the plumes.

Table 1. Attenuation of TCE in Ground Water With Distance From the Source and Residence Time in the Aquifer (1-3)

Average conc. (μg/L)

Highest conc. (μg/L)

Location

Distance From Source

(m)

Time in Aquifer

(y) TCE cis-DCE

Vinyl

Chloride

St. Joseph 130 3.2 6,500

68,000

 8,100 

128,000 

 930

4,400

390 9.7 520

8,700

 830

9,800

 450

1,660

550 12.5 15

 56

 18

 870

 106

 205

855 17.9 < 1

 1.4

 < 1

 0.8

 < 1

 0.5

Picatinny 240 2.2 25,000 220 4 

320 2.9 10,000 35 1 

460 4.2 1,400 310 6 

Table 2. Comparison of the Relative Attenuation of TCE, cis-DCE, and Vinyl Chloride With the Attenuation of Chloride in the

Plume at St. Joseph, Michigan (1, 2)

Highest concentration

Distance From

Source (m)

Chloride Ion

(mg/L)

Organic Chlorine

(mg/L)

TCE

(μg/L)

c-DCE

(μg/L)

Vinyl Chloride

(μg/L)

Background 14

130 55 104 4,000 128,000 4,400 

390 109 15 8,700 9,800 1,660 

550 71 0.8 11 828 205 

855 57 < 0.1 1.4 0.8 0.5

15
References

1. Semprini, L., P.K. Kitanidis, D.H. Kampbell, and J.T. Wilson. Anaerobic transformation of chlorinated aliphatic hydrocarbons in a

sand aquifer based on spatial chemical distributions. Water Resources Res. In press.

2. Wilson, J.T., J.W. Weaver, and D.H. Kampbell. 1994. Intrinsic

bioremediation of TCE in ground water at an NPL site in St.

Joseph, Michigan. In: Symposium on Intrinsic Bioremediation of

Ground Water. EPA/540/R-95/515. pp. 154-160.

3. Martin, M., and T.E. Imbrigiotta. 1994. Contamination of ground

water with trichloroethylene at the Building 24 site at Picatinny

Arsenal, New Jersey. In: Symposium on Intrinsic Bioremediation

of Ground Water. EPA/540/R-95/515. pp. 143-153.

4. Ehlke, T.A., T.E. Imbrigiotta, B.H. Wilson, and J.T. Wilson. 1991.

Biotransformation of cis-1,2-dichloroethylene in aquifer material

from Picatinny Arsenal, Morris County, New Jersey. In: U.S. Geological Survey Toxic Substances Hydrology Program—Proceedings of the Technical Meeting. Water-Resources Investigations

Report 91-4034. pp. 689-697.

5. Wilson, B.H., T.A. Ehlke, T.E. Imbrigiotta, and J.T. Wilson. 1991.

Reductive dechlorination of trichloroethylene in anoxic aquifer material from Picatinny Arsenal, New Jersey. In: U.S. Geological Survey Toxic Substances Hydrology Program—Proceedings of the

Technical Meeting. Water-Resources Investigations Report 91-

4034. pp. 704-707.

Table 3. Rates of Reductive Dechlorination of TCE, cis-DCE, and Vinyl Chloride in Ground Water (residence time refers to time in

the segment of the plume being described, or incubation time of microcosms) (1-5)

Apparent Loss Coefficient (1/yr)

Location

Distance From

Source (m)

Time From

Source (yr)

Residence

Time (yr) TCE cis-DCE

Vinyl

Chloride

Field Scale Estimates

St. Joseph 130 to 390 3.2 to 9.7 6.5 0.38 0.50 0.18

390 to 550 9.7 to 12.5 2.8 1.3 0.83 0.88

550 to 855 12.5 to 17.9 5.4 0.93 3.1 2.2 

Picatinny 240 to 460 2.2 to 4.2 2.0 1.4 Produced Produced

320 to 460 2.9 to 4.2 1.3 1.2 Produced Produced

 0 to 460 0.0 to 4.2 4.2 1.0 

240 to 320 2.2 to 2.9 0.7 1.6 

 0 to 250 0.0 to 2.3 0.5 

Laboratory Microcosm Studies

Picatinny 240 2.2 0.5 0.64 0.52

320 2.9 0.5 0.42 9.4 

460 4.2 0.5 0.21 3.1 

16
Intrinsic Bioremediation of a Gasoline Plume: Comparison of Field and

Laboratory Results

Morton A. Barlaz, Melody J. Hunt, Sreenivas Kota, and Robert C. Borden

North Carolina State University, Raleigh, North Carolina

Introduction

Assessing the potential for natural bioremediation in the

subsurface is complicated by site-specific conditions

and the methods used to estimate biodegradation rates.

Controlled laboratory experiments often are necessary

to verify biological loss of a compound and to assess

factors that influence biodegradation. The effect of removing samples from such a stable environment and

placing them in laboratory microcosms, however, is not

understood. In situ columns have been used to measure

biodegradation on a limited basis, and little is known

about their reliability. In this paper, we use laboratory

microcosms and in situ column experiments to estimate

intrinsic biodegradation rates of benzene, toluene, ethylbenzene, and xylene (BTEX) isomers in the subsurface.

Site Background

This research was conducted at a petroleum-contaminated aquifer in the southeastern coastal plain near

Rocky Point, North Carolina. The plume is characterized

by negligible dissolved oxygen and redox potentials of

-100 to -200 mV due to intrinsic biodegradation of BTEX

(1). The dominant electron acceptors within the plume

are sulfate and iron (1). The midpoint of the plume is

characterized by high dissolved iron (Fe)(II) (greater

than 40 mg/L) and low SO4-2 concentrations (less than

4 mg/L). Toluene and o-xylene are nearly depleted (less

than 20 μg/L), whereas high quantities of benzene,

ethylbenzene, and m-,p-xylene remain (greater than

500 μg/L).

Experimental Methods

Laboratory Microcosms

Multiple replicate microcosms with no headspace were

constructed in an anaerobic chamber under aseptic conditions using blended aquifer sediment and ground

water recovered under anaerobic conditions. Microcosm

preparation was designed to simulate ambient conditions to the maximum extent possible. Microcosms

were spiked with approximately 10,000 μg/L BTEX

(2,000 μg/L of each compound) and incubated in anaerobic containers stored at the ambient ground-water

temperature, 16°C. Because it cannot be distinguished

from m-xylene by the analytical procedure used, pxylene was not added. BTEX loss was monitored by

destructively sampling three live and three abiotic microcosms at monthly intervals for 300 days. A final time

point was taken 100 days later (after 400 days).

In Situ Columns

The in situ columns were similar to a system used

previously (2). Each column consisted of a 1-m long

chamber where sediment and ground water were isolated from the surrounding aquifer. Two sets of columns,

Group A and Group B, were installed at the midpoint

area of the plume. Each set contained three individual

columns: two live and one abiotic control. After installation, the columns were filled with anaerobic ground

water containing BTEX, which had been recovered from

nearby wells. The contaminant concentrations in

water added to Group A were 1,300, 40, 1,800, 700,

and 15 μg/L for benzene, toluene, ethylbenzene,

m-,p-xylene, and o-xylene, respectively. The approximate concentrations of compounds added to Group B

columns were 200, 200, 600, 1,600, and 800 μg/L for

benzene, toluene, ethylbenzene, m-,p-xylene, and

o-xylene, respectively. The abiotic control columns were

prepared by adjusting the pH to less than 2 with hydrochloric acid (HCl). Tracer tests were conducted on all

columns before each experiment began to ensure that

they were properly installed.

Results

A distinct order of compound disappearance was measured in the laboratory incubations: m-xylene degradation began with no lag period, followed by toluene,

17
o-xylene, benzene, and ethylbenzene (Figure 1). The

rate of m-xylene loss slowed once toluene loss began; once toluene and o-xylene were below 20 μg/L

(120 days), the rate of m-xylene loss increased. The

aquifer material was obtained in an area of the plume

where toluene and o-xylene concentrations were very

low (less than 50 μg/L) but significant quantities of m-,pxylene remained (greater than 1,000 μg/L). Thus, the

microbial population appeared to have an initial preference for m-xylene, but switched to toluene and o-xylene

after a 22-day acclimation period. Benzene began to

biodegrade once m-xylene was depleted and was at or

below 10 μg/L in all microcosms at the final sampling

(403 days). First-order decay rates (K) were determined

during the time of loss for each compound (Table 1).

All live and abiotic in situ columns exhibited an initial

concentration decrease of several hundred micrograms/per liter between the injection water and the first

sample taken from the chamber. This initial loss is attributed to sorption. After the sorption loss, the initial compound concentrations were less than 500 μg/L in most

columns. The concentrations of hydrocarbons in the

abiotic columns remained fairly constant or declined

slowly after the initial decrease, indicating biological

activity or short circuiting did not occur in the control

columns.

In Group A, benzene and m-,p-xylene exhibited significantly higher losses in the live columns relative to the

abiotic columns. The concentration of m-,p-xylene decreased in the live columns after an initial lag of 85 to

121 days (Figure 2). Benzene concentrations remained

constant in both live columns for 155 days, after which

time decreases attributed to biological activity were

measured (data not shown). Initial toluene and o-xylene

concentrations were too low (less than 50 μg/L) to accurately measure concentration changes. Figure 2

shows the measured m-,p-xylene loss in the Group A

columns and illustrates the timeframe used to calculate

the decay rates. Results from the two live columns were

pooled to estimate the live decay rate. In Group B,

significant biological loss of toluene occurred with no

apparent lag time. The short sampling period of 75

days was not adequate to measure losses of the other

compounds.

Comparison of In Situ Columns and

Laboratory Microcosms

Biological loss for three of the five BTEX compounds

occurred over similar periods in the laboratory and in situ

experiments. In both cases, toluene degradation was

followed by m-,p-xylene and benzene. This order is

consistent with previous field investigations (1). Ethylbenzene loss was minimal in the laboratory microcosms

during the 400 days of incubation, and no ethylbenzene

degradation was measured during the 7 months of in

situ monitoring. Loss of o-xylene was not observed in

the Group B columns, but fairly rapid depletion concurrent with toluene loss was measured in the laboratory.

The initial concentration of o-xylene (less than 500 μg/L)

Table 1. Comparison of Microcosm and In Situ Column

Biodegradation Ratesa

Compound

Laboratory Rate

(percent -day-1)

(time interval in

days)

In Situ Column Rate

(percent -day-1)

(time interval in

days)

Benzene 2.37 (184 to 403) 0.41 (121 to 251)

Toluene 4.46 (22 to 120) 1.15 (13 to 75)

m,p-xylene 2.04 (0 to 184) 1.43 (121 to 251)

o-xylene 5.59 (37 to 120) NS

Ethylbenzene 0.19 (0 to 403) NS 

a

 Rate calculated is the difference between the live and abiotic loss

rates assuming first-order model.

NS = The difference between the live and abiotic loss rate is not

significant at the 95-percent confidence level.

Figure 1. BTEX biodegradation in laboratory microcosms.

Figure 2. m-,p-Xylene biodegradation in Group A in situ

columns.

18
was possibly too low to stimulate in situ degradation, or

the 75-day monitoring period could have been too short.

Although the monthly sampling frequency was consistent for both types of measurements, the length of monitoring was shorter in the in situ columns due to the

limited sample volume available. Thus, direct comparison of decay rates between the two types of measurements is difficult. Given these limitations, the measured

rates are comparable in both columns and microcosms.

The slightly lower decay rates measured in the in situ

columns may be due to the lower initial concentrations

used in these experiments. Biological decay was demonstrated in the controlled column and microcosm experiments. Use of in situ columns could provide a practical link between laboratory evaluations and full-scale

field studies.

References

1. Borden, R.C., C.A. Gomez, and M.T. Becker. 1995. Geochemical

indicators of natural bioremediation. Ground Water 82(2):180-89.

2. Gillham, R.W., R.C. Starr, and D.J. Miller. 1990. A device for in situ

determination of geochemical transport parameters; two biochemical reactions. Ground Water 28(6):858-862.

19
Toxicity Effects on Methanogenic Degradation of Phenol in Ground Water

Barbara A. Bekins, E. Michael Godsy, and Ean Warren

U.S. Geological Survey, Menlo Park, California

Introduction

At an abandoned creosote works located near Pensacola, Florida, the shallow ground water is contaminated

with phenolic compounds, heterocyclic compounds, and

polyaromatic hydrocarbons. Based on the use of unlined

disposal ponds during the 80 years of operations at the

plant, the contaminants have probably been present in

the ground water for several decades. A methanogenic

consortium in the aquifer is degrading some of the compounds, and the concentrations of the degradable fraction drop to less than 1 percent of the source values by

100 to 150 m downgradient from the nonaqueous phase

(1). In spite of the long exposure time to a continuous

source, the results of acridine orange direct counts and

most probable number determinations indicate a low,

uniform microbial density (1). The continued existence

of low microbial numbers suggests that some factor is

limiting growth. Several possibilities, such as microbial

transport, nutrient limitation, predation, and toxicity,

have been examined. Of these, toxicity appears to be

the most promising explanation. The toxicity of creosote

compounds to various organisms has been studied for

a long time (2). Very little work has been done, however,

on the effects of creosote on methanogenic consortia

known to be especially sensitive to toxic compounds (3).

The results of our work indicate that the dimethylphenols

and methylphenols present in the ground water at this

site inhibit the degradation of phenol. Furthermore, incorporating these inhibition effects into a one-dimensional model of the aquifer predicts a steady-state

degradation profile and zero net growth of the active

methanogenic consortium.

Toxicity Assay Results

A serum bottle assay similar to that described by Owens

et al. (3) was performed to determine which of the

compounds present in the ground water might be toxic

to the methanogenic consortium. These compounds

were grouped by class and added at three concentration

levels equivalent to 1.5, 1.0, and 0.5 times the highest

measured field value. The classes of compounds tested

were 1) indene, benzothiophene, 2-methylnaphthalene,

biphenyl, flourene, and 2-naphthol; 2) 2-, 3-, and 4-

methylphenol; 3) 2,4- and 3,5-dimethylphenols; 4) quinoline and isoquinoline; 5) 2(1H)-quinolinone and

1(2H)-isoquinolinone; and 6) all of the preceding compounds combined. Duplicate 100-mL serum bottles containing the target compounds, enriched methanogenic

culture derived from the aquifer, mineral salts, and phenol as the growth substrate (at the highest concentration

observed in the field) were prepared in an anaerobic

glove box and capped. The volume of gas produced in

each bottle was monitored by allowing a wetted glass

syringe inserted through the septum to equilibrate with

atmospheric pressure. Figure 1 shows the gas production in the bottles containing Mixtures 2, 3, 4, and 6 at

the 1.5 times concentration as well as a control containing only phenol. Methylphenols and dimethylphenols

showed a substantial toxicity effect, whereas nitrogen

heterocycles had a smaller effect and polyaromatic hydrocarbons had no measurable effect (data not shown).

Figure 1. Gas production by the aquifer microbes with phenol

as the growth substrate and various inhibitor concentrations equal to 1.5 times the maximum value observed in the aquifer. The results are not shown for

quinolinone, which was similar to quinoline, or for

polyaromatic hydrocarbons, which were similar to

the positive control.

20
In the bottles with concentrations equal to those in the

field, dimethylphenols had a substantial effect, and

methylphenols had a smaller effect. In the 0.5 times

concentration bottles, methylphenols and dimethylphenols had a slight effect. Preliminary results showing the

buildup of fatty acids suggest that the intermediate steps

in the degradation process are being inhibited.

Model Results

The monitoring of ground-water concentrations of the

degradable compounds for more than 12 years shows

that the concentration profiles are constant in time. The

existence of a steady-state degradation profile of each

substrate together with a low, uniform microbial density

indicates that the microbial numbers do not change with

time. In theory, the functional form of the Monod growth

expression cannot be balanced by a constant decay

rate. To address this problem, toxicity effects are incorporated into the following equations for one-dimensional

substrate transport with degradation and microbial

growth:

R ∂S

∂t

= D ∂2

S

∂x2 − v

∂S

∂x − μm

Y

B

θ











S

Ks







1 +

Sc

Kc







+ S +

S2

Ki











(Eq. 1)

dB

dt = 









μm











S

Ks







1 +

Sc

Kc







+ S +

S2

Ki











− kd











B

(Eq. 2)

where R is the retardation factor, S is phenol concentration, v is the flow velocity, D is dispersion, μm is the

maximum growth rate, Y is the yield, B is biomass, θ is

porosity, KS is the half saturation constant, Ki

 and KC are

haldane and competitive inhibition constants, Sc is the

concentration of the inhibiting compound, and kd is the

biomass decay or maintenance rate. When only the

toxicity of phenol is incorporated using the Haldane

inhibition model, the predicted growth is about 50 percent lower but still much higher than the only published

decay rate (4). In addition, the equations do not produce

a steady-state solution. Incorporating the effect of the

dimethylphenol toxicity produces a steady-state solution

for phenol and microbial concentrations that matches

the character of the data (Figure 2). The values of the

parameters used in the solution are given in the figure

caption. The model results show that the aquifer concentrations take about 6 years to evolve to a steady

state, while the microbial population takes about 25 years.

The population of aquifer microorganisms oscillates

as they adjust their distribution to account for two

competing effects: 1) the maximum concentration of phenol near the source should lead to maximum growth and

substrate utilization there, and 2) the maximum concentrations of dimethylphenols near the source lead to

maximum inhibition of growth and substrate utilization.

The result is a tradeoff between a location where the

growth substrate concentrations are higher versus one

farther from the source where the inhibitor concentrations are lower. The final microbial concentrations stabilize at about an order of magnitude higher 50 m from the

source than immediately adjacent to it.

References

1. Godsy E.M., D.F. Goerlitz, and D. Grbic-Galic. 1992. Methanogenic biodegradation of creosote contaminants in natural and

simulated ground-water ecosystems. Ground Water 30:232-242.

2. Mayfield, P.B. The toxic elements of high-temperature coal tar

creosote. 1951. Proc. Am. Wood-Preserver’s Assoc. 47:62-85.

3. Owen, W.F., D.C. Stuckey, J.B. Healy, L.Y. Young, and P.L.

McCarty. 1979. Bioassay for monitoring biochemical methane potential and anaerobic toxicity. Water Research Res. 13:485-492.

4. Bekins, B.A., E.M. Godsy, and D.F. Goerlitz. 1993. Modeling

steady-state methanogenic degradation of phenols in groundwater. J. Contam. Hydrol. 14:279-294.

Figure 2. Solution to Equation 1 for phenol after 2,000, 6,000,

and 10,000 days (left), and coupled solutions to Equation2 for microbe concentrations at the same times.

The values of the parameters used were R = 1.01,

S(0) = 26 mg/L, v = 1.0 m/d, D = 1.0 m2

/d,μm = 0.111,

Y = 0.013, Binit = 0.005 mg/L(1.6 x 106

 per 100g),θ =

0.38, Ks = 1.33, Ki = 250 mg/L, Kc = 0.52 mg/L, Sc = 23

exp(x2

/(2 (47)2)) mg/L (an empirical fit to the observed

dimethylphenol concentrations), and kd =0.0326 d-1

21
A Multiphase, Multicomponent Numerical Model of Bioventing With

Nonequilibrium Mass Exchange

Linda M. Abriola, John R. Lang, and Klaus M. Rathfelder

Department of Civil and Environmental Engineering, University of Michigan, 

Ann Arbor, Michigan

Introduction

Soil vapor extraction (SVE) and bioventing (BV) are

common remediation practices for unsaturated soils

contaminated with volatile organic compounds (VOCs).

These methods have been demonstrated to be effective at comparatively low costs. The efficiency of these

techniques is known to be restricted by soil characteristics; by mass transfer limitations between phases,

including liquid/solid, liquid/gas, and liquid/microorganism

phases; by the availability of oxygen; and by system

design and operation parameters (1). Assessment of

SVE/BV systems is often hindered by the complex interplay of physical, chemical, and biological processes.

Consequently, design and operation of these systems

are typically based on engineering experience and/or

simple design equations. Numerical models of SVE/BV

systems can be valuable tools for the investigation of

the effects of various processes on system performance

and for optimal system design. In this work, a numerical

model is presented that has been specifically developed

to incorporate the complete range of processes occurring at the field scale and to include interphase mass

transfer rate limitations.

Model Formulation

Three fluid phases are modeled: gas, aqueous, and a

nonaqueous phase liquid (NAPL). The gas and aqueous

phases may flow simultaneously in response to applied

pumping/injection or density gradients. The movement

of these phases is described by standard macroscopically averaged flow equations (2). The NAPL phase is

assumed to be at an immobile residual saturation.

Changes in NAPL saturation, therefore, result solely

from interphase mass transfer.

The NAPL may be a mixture of an unrestricted number

of organic components. The gas phase is assumed to

be composed of nitrogen and oxygen (the two major

constituents of air), water vapor, volatile components of

the NAPL, and a single limiting nutrient. The aqueous

phase is composed of water, oxygen, soluble components of the NAPL, and the limiting nutrient. Sorption to

the soil particles is restricted to components of the

NAPL. The migration of each component in each phase

is described by standard macroscopically averaged

transport equations (2).

Quantification of the biotransformation processes follows the conceptual approach of Chen et al. (3). Biodegradation is assumed to occur only within the aqueous

phase by an indigenous, spatially heterogeneous, mixed

microbial population that is present as attached microcolonies. There is no biomass transport or detachment

or sloughing of the microcolonies, and biomass growth

does not affect permeability. Monod-type kinetic expressions are employed to describe biophase utilization of

substrates, oxygen, and a limiting nutrient, as well as

growth of the microbial population. Additionally, a minimum biophase concentration reflecting the indigenous

population is maintained when growth is restricted due

to oxygen, substrate, or nutrient limitations.

A linear driving force expression is used to model nonequilibrium interphase exchange. Interphase partitioning processes included in the model are: volatilization

and dissolution of components from the NAPL;

gas/aqueous exchange of oxygen, water vapor, and the

components of the NAPL; sorption of the NAPL components to the soil particles through the aqueous phase;

and rate-limited uptake by the biophase of oxygen,

substrate (components of the NAPL), and the limiting

nutrient.

Numerical Solution

The flow and transport equations are solved in two

space dimensions (vertical cross section or radial geometry) using a standard Galerkin finite element method

22
with linear triangular elements. A set-iterative scheme is

used for computational efficiency. The sets of coupled

flow, transport, and biodegradation equations, as well as

multiple equations within sets, are decoupled and

solved sequentially. Decoupling is accomplished by lagging, either by one iteration or one time step, the coupling terms which are phase density and interphase

mass exchange. Iteration within and between equation

sets is performed to account for nonlinearities and ensure solution accuracy. Numerical solutions of the flow,

transport, and biodegradation equations have been independently verified with analytical solutions and intermodel comparisons. A detailed description of the model

and example simulations are presented in Lang et al. (4).

Demonstration of SVE and BV Simulations

Hypothetical field-scale SVE and BV systems are presented to demonstrate model capabilities. The modeled

scenario involves the remediation of a residual NAPL

distributed within a layered soil system. Here the

nonuniform initial NAPL distribution was generated with

a multiphase flow model. NAPL contamination is present

in both the unsaturated zone and capillary fringe.

SVE operations are examined by simulating relatively

large pumping rates to an extraction well positioned in

the center of the residual NAPL zone. Removal efficiency in the test simulations is shown to be sensitive to

mass transfer rates, permeability contrasts, the initial

NAPL distribution, and, to a lesser extent, pumping rate

and well screen position.

A BV operation is also modeled by simulating small gas

injection rates at a well located in the center of the

contamination zone. Removal efficiency is shown to be

sensitive to flow rate, interphase mass transfer, biodegradation rates, and NAPL distribution.

An example of the complex interplay between chemical

and biological processes in BV systems is demonstrated

in simulation results shown in Figure 1. Here contaminant removal is compared for BV systems run at a

comparatively low flow (0.1 pore volumes/day) and high

flow (1 pore volume/day). Total mass removed (Figure

1a) is greater at high flow. Due to nonequilibrium interphase partitioning occurring at the high flow, however,

the difference in mass removed is less than the proportionate difference in flow rate. The rate of contaminant

interphase partitioning also affects the quantity of mass

removed by biodegradation (Figure 1b), which is far

greater at high flow. At low flow, interphase partitioning

of the contaminant is approximately at equilibrium, resulting in downgradient aqueous concentrations that are

greater than an inhibitory threshold. Consequently,

biodegradation in the low-flow scenario is restricted to

the region upgradient of the NAPL contamination

zone. Nonequilibrium partitioning at high flow rates

produces downgradient aqueous concentrations below the inhibitory threshold, resulting in enhanced

biodegradation. This greater degradation in the

high-flow system produces a reduction in the contaminant mass in the gas phase arriving at a downgradient extraction point (Figure 1c).

Conclusions

A numerical model of SVE/BV systems has been developed that incorporates the complete compositional and

biological processes representative of field conditions.

Example simulations demonstrate the model capabilities and illustrate the complex interplay of chemical,

physical, and biological processes occurring in SVE/BV

systems.

References

1. Rathfelder, K., J.L. Lang, and L.M. Abriola. 1995. Soil vapor extraction and bioventing: Applications, limitations, and future research directions. Reviews of Geophysics IUGG Quadrennial

Report. In press.

2. Abriola, L.M. 1989. Modeling multiphase migration of organic

chemicals in groundwater systems—A review and assessment.

Environmental Health Perspectives 83:117-143.

3. Chen, Y.-M., L.M. Abriola, P.J.J. Alverez, P.J. Anid, and T.M. Vogel.

1992. Modeling transport and biodegradation of benzene and toluene in sandy aquifer material: Comparisons with experimental

measurements. Water Resources Res. 28(7):1,833-1,847.

4. Lang, J.L., K.M. Rathfelder, and L.M. Abriola. 1995. A multiphase,

multicomponent numerical model of bioventing with non-equilibrium mass exchange. In: Proceedings of the Bioremediation Symposium, San Diego, CA. April.

Figure 1. Contaminant removal versus time for hypothetical BV scenarios at comparatively low and high flow rates. 

23
Aromatic Hydrocarbon Biotransformation Under Mixed Oxygen/Nitrate Electron

Acceptor Conditions

Liza P. Wilson, Peter C. D’Adamo, and Edward J. Bouwer

Department of Geography and Environmental Engineering, The Johns Hopkins University,

Baltimore, Maryland

Introduction

Biodegradation of contaminants associated with sediments and ground water under mixed oxygen/nitrate

electron acceptor conditions may prove to be more successful and feasible than remediation under strict aerobic or anaerobic conditions. In particular, the low level of

oxygen may allow subsurface microorganisms to attack

the aromatic rings of many organic compounds (using

oxygenases), with nitrate serving as the electron acceptor to complete the degradation. Providing nitrate to

the subsurface is less expensive than maintaining aerobic conditions, and, as nitrate is highly soluble, it is

easier to maintain a residual concentration in ground

water.

A laboratory investigation is being conducted to provide

a better understanding of the effect of dual oxygen/nitrate electron acceptor conditions on the biodegradation

of monocyclic and polycyclic aromatic hydrocarbon mixtures in aqueous solution. The specific objectives of the

research are 1) to quantify the stoichiometry and kinetics

of biodegradation of a mixture of aromatic hydrocarbons

under microaerophilic conditions (defined as less than

or equal to 2 mg/L O2), and 2) to assess the relative

efficacy of bioremediation under microaerophilic conditions compared with strict aerobic or denitrification conditions in the laboratory, using batch microcosms and

aquifer sediment columns.

Microaerophilic Biodegradation by an

Enrichment of Aquifer Bacteria

The microaerophilic biodegradation of a mixture of aromatic compounds was investigated by varying combined concentrations of oxygen and nitrate. Batch

microcosms were prepared using a liquid enrichment of

aquifer bacteria as inocula; a mixture of benzene, toluene, ethylbenzene, m-xylene, naphthalene, and phenanthrene as substrate; and oxygen and nitrate as electron acceptors.

The results of this study indicated that the level of oxygen had a significant affect on the extent of biodegradation of most of the aromatic hydrocarbons. Analysis of

the consumption of electron acceptors indicated that

both nitrate and oxygen acted as electron acceptors

during biodegradation of the mixture of aromatic hydrocarbons. Denitrification may be inhibited by oxygen levels above 1 mg/L (1). In this study, toluene and

naphthalene biodegradation was favored at microaerophilic oxygen levels between 1.5 and 2 mg/L. The data

(measurements of oxygen and nitrate not shown here)

suggested that oxygen and nitrate were used sequentially to biodegrade naphthalene and toluene, respectively (i.e., denitrification was inhibited until oxygen was

depleted)(Table 1).

At lower levels of oxygen (0.5 to 1 mg/L), toluene and

ethylbenzene biodegradation was favored. The mechanism for biodegradation of toluene and ethylbenzene at

very low oxygen levels (less than or equal to 1 mg/L)

Table 1. Aromatic Hydrocarbons Degraded Under the

Various Combinations of Oxygen and Nitrate

Investigated (degradation is removal greater than or

equal to 10 percent relative to killed controls)

Oxygen (mg/L)

0 0.5 1 1.5 2 8

Nitrate 

(mg/L)

10 T T, E T, N B, T, E, m-X, N, P

50 T T, E T, E T, N T, N B, T, E, m-X, N, P

150 T T, E T, E T, N T, N B, T, E, m-X, N, P

400 T T, E T, E T, N T, N B, T, E, m-X, N, P

B = benzene, T = toluene, E = ethylbenzene, m-X = m-xylene,

N = naphthalene, and P = phenanthrene

24
may be quite different than at higher levels (greater than

or equal to 1.5 mg/L). The data indicated that at low

levels (less than or equal to 1 mg/L) of oxygen, nitrate

played a role in biodegradation of both toluene and

ethylbenzene. Evidence for simultaneous utilization of

nitrate and oxygen has been documented (2). Oxygen

levels below 1 mg/L may not inhibit denitrification and

may actually be beneficial by increasing cell numbers

(3, 4).

Benzene was recalcitrant under denitrifying and microaerophilic conditions. Extensive benzene mineralization, however, was observed under aerobic conditions

(Figure 1). Although the majority of benzene biodegradation occurred in the presence of oxygen, partial transformation of the parent compound to intermediates and

carbon dioxide was observed in the absence of oxygen.

Microaerophilic Biodegradation in the

Presence of Sediments

A laboratory study of the stoichiometry of microaerophilic biodegradation in the presence of sediments is

being conducted. Aquifer sediments may affect the

stoichiometry of microaerophilic biodegradation by exerting an additional oxygen demand from natural organic matter or reduced metals, and by providing surfaces for microbial attachment that are not present in

sediment-free microcosms. In this study, aquifer sediments were used as inocula in microcosms instead of a

liquid enrichment of aquifer bacteria. Initial results indicate that toluene biodegradation under denitrifying conditions occurs after a significant lag time in microcosms

containing sediment as inocula when compared with

denitrification in liquid enrichment microcosms. This delay in denitrifying activity is likely due to the development

of a sufficient denitrifying population. No biodegradation

of aromatic compounds was observed under microaerophilic conditions (microaerophilic oxygen consumed by

sediment demands). Biodegradation under aerobic conditions (7 mg/L), however, exceeded what was observed

in the liquid enrichment microcosms. Studies of sediment microcosms and columns are ongoing.

Acknowledgment

This research was made possible by the generous support of the U.S. Environmental Protection Agency

Robert S. Kerr Environmental Research Laboratory

(Project CR-821907).

References

1. Kuhn, E.P., J. Zeyer, P. Eicher, and R.P. Schwarzenbach. 1988.

Anaerobic degradation of alkylated benzenes in denitrifying laboratory columns. Appl. Environ. Microbiol. 54(2):490-496.

2. Ottow, J.C.G., and W. Fabig. 1985. Influence of oxygen aeration

on denitrification and redox level in different batch cultures. In:

Caldwell, D.E., J.A. Brierly, and C.L. Brierly, eds. Planetary ecology. New York: Van Nostrad Reinhold. pp. 427-440.

3. Hutchins, S.R. 1991. Optimizing BTEX biodegradation conditions.

Environ. Toxicol. Chem. 10:1,437-1,448.

4. Su, J., and D. Kafkewitz. 1994. Utilization of toluene and xylenes

by a nitrate reducing strain of Pseudomonas maltophilia under low

oxygen and anoxic conditions. FEMS Microbiol. Ecol. 15:249-258.

Figure 1. Conversion of benzene to intermediates, cells, and

carbon dioxide.

25
Nutrient Transport in a Sandy Beach

Brian A. Wrenn, Makram T. Suidan, B. Loye Eberhardt, and Gregory J. Wilson

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Kevin L. Strohmeier

Environmental Technologies & Solutions, Inc., Covington, Kentucky

Albert D. Venosa

Risk Reduction Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio

Introduction

Bioremediation of beaches that are contaminated with

crude oil is expected to be limited by the availability of

nutrients. Addition of nutrients, such as nitrogen and

phosphorous, has been shown to stimulate the rate of

crude oil biodegradation in the beach environment (1,

2). For bioremediation to be effective, the nutrient concentration in contact with the oil-contaminated beach

material must be high enough to allow hydrocarbon-degrading bacteria to grow at their maximum rates.

Oil often strands in the intertidal zone (3). When this

occurs, washout will probably dominate the nutrient residence time in the bioremediation zone. The periodic

flooding that occurs when the tide rises is one of several

important nutrient transport mechanisms. Wave action

can also affect nutrient transport in the intertidal region

of beaches (4), as can the flow of fresh ground water

from inland (5, 6). The objective of this research was to

characterize the nutrient transport rates for a low-energy

sandy beach. The results were used to determine the

nutrient application rate that is required for effective

bioremediation in this type of environment.

Methods

The tracer study was conducted on a long, uniform

stretch of sandy beach south of Slaughter Beach, Delaware, near the southern end of Delaware Bay. The

beach is composed largely of coarse sand with some

gravel, which lies on top of an impermeable peat layer.

The beach has a slope of approximately 11 percent in

the upper intertidal zone.

Eight replicate plots (5 m x 10 m) were established in

the upper intertidal zone. The tops of the plots were

placed approximately at the spring high-tide line. The

plots were divided into two blocks of four plots each. A

water soluble tracer, either lithium or sodium nitrate, was

applied to one plot in each block at each of four different

stages in the tidal cycle (Figure 1): spring tide (full

moon), falling midtide (i.e., as the tide proceeded from

spring to neap), neap tide (last quarter moon), and rising

midtide (i.e., as the tide proceeded from neap to spring).

(Spring tide is the point in the lunar tidal cycle at which

the difference between high and low tide is greatest.

Neap tide is the point at which the difference is smallest.) Lithium nitrate was used as the tracer for the

Figure 1. The actual and expected tidal elevations are shown,

with the start times for the four tracer transport experiments. The elevations of the tops and bottoms of

the plots are also shown. All elevations are relative

to a benchmark that was placed behind the beach in

a dune area well above the maximum high tide line.

26
spring- and neap-tide experiments, and sodium nitrate

was used for both midtidal experiments. The tracer was

dissolved in fresh water (the initial concentration was

20 g/L) and applied to the plots with a sprinkler system. Tracer was always applied at low tide, and the

first samples were collected shortly afterward.

The spring- and neap-tide plots contained multiport

wells that were used to collect water samples from

discrete depths within the beach. These wells were

placed along a transect through the middle of the plots

at 2.5-m intervals, beginning 2.5 m above the top of the

plots and continuing to 5 m below the bottom of the

plots. The wells had sample ports every 6 in. The top

port of each well was installed 3 in. below the beach

surface. Sand samples were collected from randomly

selected positions within five subsections that were

marked off in each plot. The locations of these subsections corresponded to the positions of the wells in the

spring and neap-tide plots.

Sand samples were collected as 10-in. cores using a

5-in. auger. Samples that were analyzed for lithium were

kept as two separate 5-in. cores: an upper sample (0 to

5 in.) and a lower sample (5 to 10 in.). The cores taken

for nitrate analysis, on the other hand, were composited

into a single 10-in. (0 to 10 in.) sample. Lithium was

extracted into 1 M ammonium acetate and analyzed by

atomic absorption spectrophotometry (7). Nitrate, which

was extracted and analyzed in the field, was analyzed

by the cadmium reduction autoanalyzer method following extraction into 2 M potassium chloride (KCl) (8).

Results and Discussion

The average lithium concentrations in the two 5-in. core

samples for the spring- and neap-tide tracer experiments are shown in Figure 2. The concentrations measured in samples collected from the five subsections of

each plot were pooled to obtain the plot averages that

are shown in this figure. The plot averages for each of

the replicate plots are shown independently in this figure. The average lithium concentration was reduced to

zero very rapidly following the spring-tide application

(Figure 2A), but it was washed out more slowly after the

neap-tide application (Figure 2B). Washout of nitrate

following the two midcycle applications behaved similarly; nitrate was washed out quickly following the falling

midtidal application, and it persisted for a relatively long

time following the rising midtidal application (data not

shown). Although the initial lithium concentrations

were higher in the samples from the upper 5-in., it

disappeared more rapidly in this region than it did from the

5- to 10-in. region.

It is clear from Figure 1 that one of the major differences

among the four tracer experiments is the extent to which

the plots were covered with water at high tide. Wave

action also contributed to plot coverage, and the total

coverage is the sum of both effects. The cumulative

effects of plot coverage on nutrient retention are shown

in Figure 3, in which the ratio of the remaining sand

nitrate concentration to its initial concentration is plotted

as a function of the maximum extent of plot coverage

that had occurred before collecting each sample (e.g., if

Figure 2. Lithium concentrations in sand samples collected in

the bioremediation zone following tracer application

at spring tide (A) and neap tide (B). The concentrations for the upper (0 to 5 in.) and lower (5 to 10 in.)

samples are plotted separately for each experiment.

Figure 3. Relationship between the fraction of nitrate remaining

and the maximum extent to which water had covered

the plots before the samples were collected. Data

from all four tracer experiments are included in this

plot.

27
the first high tide following tracer application covered 75

percent of the plot with water and subsequent high tides

covered only 50 percent, the maximum coverage for all

of the data collected after the first high tide is reported

as 75 percent). Data from all four tracer experiments,

calculated as described for the lithium data plotted in

Figure 2, were used to construct this plot. Figure 3

shows a strong correlation between the maximum extent of plot coverage and the remaining nitrate concentration, suggesting that nutrient retention in the

bioremediation zone of a sandy beach can be predicted

based solely on the extent of water coverage.

The simplest explanation for the results shown in Figure

3 is that the tracers become diluted by mixing with bulk

seawater when the plots are covered by the rising tide,

and they are washed away when the tide recedes. This

explanation is consistent with a model for nutrient transport in a beach in Prince William Sound, Alaska (9).

Pore-water data, however, show that the tracer movement is predominantly downward into the beach (Figure

4). These data are plot averages, calculated as described above, for all pore-water samples collected from

each depth within the plot. Samples collected from wells

outside of the plots were not used to compute these

averages.

Figure 4 shows that nutrient is probably removed from

the bioremediation zone by advective flow through the

porous matrix of the beach, not by mixing with bulk

seawater.

Conclusions

Effective bioremediation requires a sufficient supply of

the growth-limiting substrate to be available to the bacteria responsible for biodegradation. For bioremediation

of oil-contaminated beaches, nutrients such as nitrogen

and phosphorus are expected to be most important (1,

2), and washout is expected to be the dominant nutrient

removal mechanism. In many cases, it will be desirable

to optimize nutrient application rates to minimize costs

and to reduce the opportunity for eutrophication of

adjacent bodies of water. Our data suggest a simple

method for determining the frequency required for

nutrient application.

Nutrient retention in the bioremediation zone of the intertidal region of a sandy beach varies with the lunar

tidal cycle. Our tracers were washed out of the bioremediation zone very quickly when they were applied during

and shortly after the spring tide (when the high tide

reached its maximum elevation), but they persisted

through several tidal cycles when applied around neap

tide, when the lowest high tides occurred. Our data

suggest that the differences in nutrient retention time are

related to the maximum extent to which our experimental plots were covered by water at high tide. Total coverage appears to be more important than coverage due

to the tide alone. Therefore, wave activity and tidal

elevation must both be considered to determine the

appropriate fertilization frequency. When water-soluble

fertilizers are used, visual inspection of the extent to

which the contaminated area is covered by water

during high tide is a reliable alternative to expensive

and time-consuming chemical analyses for nutrient

concentration.

References

1. Lee, K., G.H. Tremblay, and E.M. Levy. 1993. Bioremediation:

Application of slow release fertilizers on low-energy shorelines. In:

Proceedings of the 1993 Oil Spill Conference, American Petroleum

Institute, Washington, DC. pp. 449-454.

2. Pritchard, P.H., and C.F. Costa. 1991. EPA’s Alaska oil spill bioremediation project. Environ. Sci. Technol. 25:372-379.

3. Payne, J.R., and C.R. Phillips. 1985. Petroleum spills in the marine

environment: The chemistry and formation of water-in-oil emulsions and tar balls. Chelsea, MI: Lewis Publishers.

4. Brown, A.C., and A. McLachlan. 1990. Ecology of sandy shores.

New York, NY: Elsevier.

5. Glover, R.E. 1959. The pattern of fresh-water flow in a coastal

aquifer. J. Geophys. Res. 64:457-459.

6. Nielsen, P. 1990. Tidal dynamics of the water table in beaches.

Water Resources Res. 26:2,127-2,134.

7. American Public Health Association. 1989. Direct air-acetylene

flame method 3111B. In: Standard methods for the examination of

water and wastewater, 17th ed., Washington, DC.

8. Keeney, D.R., and D.W. Nelson. 1982. Nitrogen—inorganic forms.

In: Page, A.L., R.H. Miller, and D.R. Keeney, eds. Methods of soil

analysis, Part 2. Chemical and microbiological properties. Madison, WI: American Society of Agronomy, Inc. pp. 643-698.

9. Wise, W.R., O. Guven, F.J. Molz, and S.C. McCutcheon. Nutrient

retention time in a high permeability, oil-fouled beach. J. Environ.

Engin. In press.

Figure 4. Average lithium concentrations in pore water collected at discrete depths below the beach surface for

three time points during the spring tide tracer experiment. All samples were collected at high tide. 

28
Bioremediation of Crude Oil Intentionally Released on the 

Shoreline of Fowler Beach, Delaware

Albert D. Venosa and John R. Haines 

U.S. Environmental Protection Agency, Cincinnati, Ohio

Makram T. Suidan, Brian A. Wrenn, B. Loye Eberhart, Miryam Kadkhodayan, and Edith Holder

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Kevin L. Strohmeier

Environmental Technologies and Solutions, Inc., Covington, Kentucky

Dennis King

Kingstat Consulting, Fairfield, Ohio

Bennet Anderson

Delaware Department of Natural Resources and Environmental Control, Dover, Delaware

Introduction

A major factor contributing to the equivocal findings of

past field studies (1-5) was that conclusions were usually based on comparisons between one large treatment

plot and one large control plot. The problem with this

type of experiment is that no replicate plots are established to provide a basis for estimating experimental

error. The collection of numerous subsamples from one

treatment plot and one control plot, termed pseudoreplication (6), is statistically invalid for drawing inferences

on treatment effects because no experimental error can

be computed. An experiment lacking replication is an

uncontrolled experiment because it does not control for

among-replicate variability inherent in the experimental

material, introduced by the experimenter, or arising from

chance occurrences. To eliminate uncontrollable and

unknown environmental factors that could skew results

in one direction, several replicate plots must be set up

in random fashion on the beach surface. The experimental approach described herein was carefully designed to allow for valid and statistically authentic

comparisons between treatments.

The goals of the study were 1) to obtain sufficient statistical and scientifically credible evidence to determine

whether bioremediation with inorganic mineral nutrients

or microbial inoculation enhances the removal of crude

oil contaminating mixed sand and gravel beaches, and

2) to compute the rate at which such enhancement takes

place.

Materials and Methods

The plan was to maximize the effectiveness of bioremediation by maintaining a certain level of nitrogen, in the

form of nitrate (agricultural grade sodium nitrate), and

phosphorus, in the form of tripolyphosphate (sodium

tripolyphosphate), in contact with the degrading populations so that they would be able to grow at their maximal

rates at all times. In a previous study (7), we had shown

that the minimum nitrate-N concentration needed by oil

degraders to grow on hydrocarbons at an accelerated

rate under semicontinuous flow conditions was approximately 1.5 mg/L. It was also known that, if the incoming

tide completely submerged the plot, the levels of nitrate

in the interstitial pore-water diluted to undetectable limits

(8). Thus, to maximize bioremediation during spring

tides, we reasoned that nutrients would have to be

applied every day. To achieve the target 1.5 mg/L interstitial pore-water concentrations, we assumed a 100-

fold safety factor to account for dilution. The amount of

nitrate-N needed under these circumstances was thus

calculated to be about 55 g/m2

, applied once daily to

each plot.

The approach used to assess treatment effects in the

field study was a randomized complete block (RCB)

29
design with repeated measures. Five areas of beach

were selected based on the homogeneity of geomorphology within each area. Each area (“block”) was large

enough to accommodate four experimental units or test

plots. The blocks were situated in a row on the beach

parallel to the shoreline. Three treatments were tested

on oiled plots: no-nutrient control, water-soluble nutrients, and water-soluble nutrients supplemented with a

natural microbial inoculum from the site. The inoculum

was grown by isolating a mixed culture from the site and

adding it to a 55-gal drum containing Delaware Bay

seawater, the same Bonny Light crude oil, and the same

nutrient mix used on the beach. A fourth treatment, an

unoiled and untreated plot, served as a background

control for microbiological characterization and baseline

bioassays. The four treatments were randomized in

each of the five blocks so that whatever inferences could

be ascertained from the data would be applicable to the

entire beach, not just the test plots.

Results and Discussion

The mean interstitial nitrate-N concentrations measured

over the course of the 14-week investigation were:

0.8 ± 0.3 mg/L in the unamended control plots, 6.3 ±

2.7 mg/L in the nutrient-treated plots, and 3.5 ± 1.7 mg/L in

the inoculum-treated plots. These results indicate that

background nutrient levels on Fowler Beach were high

enough to sustain nearly maximum oil degrader growth.

Figure 1 is a summary of the hopane-normalized alkane

(Figure 1) and aromatic (Figure 2) oil components remaining after the first 8 weeks of the study. In regards

to the alkanes, statistical analysis of variance (ANOVA)

showed that the difference between both treated plots

and the control plots were highly significant at Weeks 2,

4, and 8 (p < 0.01) but not at Week 6. Differences

between the nutrient-treated and inoculum-treated plots

were not significant at any time. Clearly, substantial

natural biodegradation was taking place on the control

plots without addition of nutrients. This observation is

consistent with observations made on the background

nutrient levels existent on Fowler Beach. Addition of

nutrients significantly enhanced the natural biodegradation rates of the alkane fraction, but not to the extent

expected on a less eutrophic beach.

With respect to the aromatic components, results of the

ANOVA revealed no significant differences among any

of the treatments at Weeks 0, 2, 4, and 6, although

substantial biodegradation of the polycyclic aromatic

hydrocarbons (PAHs) occurred on all plots. At Week 8,

statistically significant differences between the treated

and untreated plots were evident. Most of the disappearance occurred among the two- and three-ring PAHs and

the lower alkyl-substituted homologues (data not

shown). The four-ring PAHs began to show evidence of

biodegradation during the eighth week of the study.

These results suggest that biostimulation may not always be necessary to promote bioremediation if sufficient nutrients are naturally present at a spill site in high

enough concentrations to effect natural cleanup. The

evidence suggests that nutrient application to maintain

a residual nitrate concentration in the interstitial waters

at high enough levels to sustain maximum biodegradative metabolism resulted in a significant enhancement of

alkane and, to a lesser extent, aromatic biodegradation

over natural attenuation. Bioaugmentation (i.e., supplementation with a bacterial inoculum indigenous to the

area), however, did not appear to result in further enhancement.

References

1. Pritchard, P.H., and C.F. Costa. 1991. EPA’s Alaska oil spill bioremediation project. Environ. Sci. Technol. 25:372-379.

2. Sveum, P., and A. Ladousse. 1989. Biodegradation of oil in the

Arctic: Enhancement by oil-soluble fertilizer application. In: Proceedings of the 1989 International Oil Spill Conference. Washington, DC: American Petroleum Institute.

Figure 1. Decline in hopane-normalized total alkanes during

the first 8 weeks.

Figure 2. Decline in hopane-normalized total PAHs during the

first 8 weeks.

30
3. Sveum, P. and Ladousse, A. 1989. “Biodegration of Oil in the Artic:

Enhancement by Oil-Soluble Fertilizer Application.” In: Proceedings of the 1989 International Oil Spill Conference. Washington,

DC: American Petroleum Institute. 

4. Rosenburg, E., R. Legmann, A. Kushmaro, R. Taube, R. Adler, and

E.Z. Ron. 1992. Petroleum bioremediation—a multiphase problem. Biodegradation 3:337-350.

5. Lee, K., G.H. Tremblay, and E.M. Levy. 1993. Bioremediation:

Application of slow-release fertilizers on low-energy shorelines. In:

Proceedings of the 1993 International Oil Spill Conference. Washington, DC: American Petroleum Institute.

6. Hurlbert, S.H. 1984. Pseudoreplication and the design of ecological field experiments. Ecol. Monographs 54(2):187-211.

7. Venosa, A.D., J.R. Haines, M.T. Suidan, B.A. Wrenn, K.L.

Strohmeier, B.L. Eberhart, E.L. Holder, and X. Wang. 1994. Research leading to the bioremediation of oil-contaminated beaches.

In: Symposium on Bioremediation of Hazardous Wastes: Research, Development, and Field Evaluations, June 28-30, 1994,

San Francisco, CA. EPA/600/R-94/075. pp. 103-108.

8. Venosa, A.D., M.T. Suidan, B.A. Wrenn, J.R. Haines, K.L.

Strohmeier, E.L. Holder, and B.L. Eberhart. 1994. Nutrient application strategies for oil spill bioremediation in the field. In: Twentieth Annual RREL Research Symposium, March 15-17, 1994,

Cincinnati, OH. EPA/600/R-94/011. pp. 139-143.

31
Dynamics of Oil Degradation in Coastal Environments: Effect of Bioremediation

Products and Some Environmental Parameters

Marirosa Molina and Rochelle Araujo

U.S. Environmental Protection Agency, Athens, Georgia

Jennifer R. Bond

DYNCORP, Athens, Georgia

Introduction

Oil extraction, refining, and transshipment are often located in coastal regions, putting wetland systems at risk

of exposure to spilled oil and oil products. The inaccessibility of wetlands and the fragile nature of those ecosystems preclude mechanical cleanup of oil, making

bioremediation a preferred option. Moreover, the high

level of indigenous microbial activity suggests a potential for biodegradation, especially if fertilizer additions

can relieve environmental nutrient limitations.

Bioremediation strategies that have been proposed for

oil spills in wetlands include fertilization, solubilization of

oil, and bioaugmentation with oil-degrading bacteria.

Although bioaugmentation has been demonstrated to be

effective in engineered systems, the ability of introduced

organisms to establish themselves in the complex web

of microbial relationships that characterize wetlands is

questionable. Moreover, supplies of oxygen and nutrients may be insufficient for concurrent degradation of oil

and natural substrates. In this research, we propose a

system for assessing the efficacy of bioremediation in

wetlands and for testing the effectiveness of several

bioremediation products. We also present data on the

dynamics of the bacterial populations and the relative

rates of degradation of natural substrates and oil in

wetlands.

Methods and Materials

Sediment microcosms were constructed using glass columns (10-cm internal diameter, 20-cm length) fitted with

fritted glass supports and filled with homogenized marsh

sediments from Sapelo Island, Georgia. Seawater or

artificial seawater was adjusted to a brackish salinity of

20‰ and exchanged on a tidal basis. An artificially

weathered Alaska North Slope (ANS) crude oil was

added at low tide to cover the sediment surface to a

depth of 0.5 mm.

Bioremediation products consisted of bacterial preparations enriched in oil degraders, nutrients, surfactants, or

combinations thereof. Products were applied in a manner consistent with the manufacturer’s recommendations. In addition to commercial products, inorganic

nutrients were added to microcosms to test the potential

response of indigenous bacteria. Ground Spartina alterniflora, a salt marsh grass common to the coast of

Georgia, was used as an alternate natural organic substrate. The grass (4.5 g) was added to the sediment

surface in an amount equivalent to a 1-yr standing stock

of aboveground biomass.

After 3 months, the residual hydrocarbons were extracted and analyzed by gas chromatography/mass

spectrometry (GC/MS). Efficacy was assessed on the

basis of reduced concentrations of specific components

of oil, including straight-chained and branched alkanes

and aromatics, expressed as ratios to conserved internal markers. Microbial populations were estimated using

modified heterotrophic and oil most probable number

(MPN) techniques (1). Heterotrophic bacteria were also

quantified using standard plate counts.

Deoxyribonucleic acid (DNA) samples were extracted

from each microcosm using a modification of the method

of Tsa and Olson (2). Target DNA was filtered onto

maximum-strength Nytran membranes. Sediment DNA

was loaded onto the membranes in triplicate. The membranes were hybridized with a Pseudomonas 23S rRNA

oligoprobe (Group I). Detection was carried out using

Rad-Free Lumi-Phos 530 substrate sheet and exposure

to x-ray film for 3 hr at room temperature. The signal was

quantified using densitometry.

32
Results and Discussion

The alkane fraction of oil was degraded in the presence

or absence of bioremediation products, as indicated by

the absence of C13 and C14 compounds and the greatly

diminished peaks for C13 through C33, including the

branched alkanes, pristane, and phytane. Addition of

inorganic nutrients to sediments containing only indigenous bacteria resulted in greater depletion of the alkane

fraction than did additions of products composed of

surfactants or bacterial enrichments. Moreover, in the

presence of surfactants, the extent and range of degradation was less than that in the control treatment, suggesting inhibition of microbial activity.

The aromatic fractions of the oil (Figure 1) were degraded to a lesser extent in all treatments than were the

alkane fractions (data not shown), although the rank

order of the treatments was the same. Both the range

of compounds degraded and the extent to which they

were degraded were significantly greater in the treatment containing nutrients than in the presence of the

commercial products.

The abundance of oil-degrading bacteria (Figure 2) was

not consistent with the extent of oil degradation observed. In the absence of bioremediation products, the

number of oil degraders (1.51 x 103

 bacteria/g sediment)

after 3 months of experimentation was not appreciably

different from that observed in the initial Sapelo Island

sediment (1.38 x 103

). Even at such low numbers, however, the microbial community was capable of degrading

the alkane fraction of oil, especially when fertilized with

inorganic nutrients. No enhanced degradation resulted

from the addition of a product purported to be enriched in oil-degrading bacteria, regardless of the relatively high numbers of total heterotrophs and potential

oil-degrading bacteria that the treatment yielded.

The inhibition of oil degradation by the addition of surfactant was consistent with the microbial numbers,

which show that, in the presence of the surfactant, the

abundance of oil degraders was less than in the treatments containing nutrients or added oil degraders. The

heterotrophic bacteria, however, were not affected by

the addition of surfactants, as indicated by the numbers

in this treatment being comparable with the numbers

obtained in the presence of added nutrients (data not

shown). The surfactant product could have been used

as a carbon or nutrient source by the heterotrophic

bacteria, thus increasing their numbers. Furthermore,

the surfactant-based product may be inhibitory to the

indigenous oil degraders, which were only enhanced in

the presence of inorganic nutrients.

Enumeration of oil-degrading bacteria gives no information about the source of the active bacteria because the

oil-degrading bacteria may have been of sediment or

product origin. DNA hybridizations with 16S rRNA

oligoprobes indicated that Pseudomonas Group I was

significantly reduced in the presence of the microbial

product. This suggests that addition of the bacterial

preparation suppressed some indigenous populations,

including perhaps the indigenous oil-degrading bacteria.

None of the other treatments produced numbers of

Figure 1. Residual aromatics in surface layer of microcosms

after 3 months in the presence of bioremediation

agents.

Figure 2. Distribution of oil degraders in wetlands microcosms

after 3 months of exposure to bioremediation agents.

33
Pseudomonads that were significantly different from

those observed in the initial sediment; data on other

sediment microbial groups will be forthcoming.

Under nutrient-amended conditions, both alkane and

aromatic fractions were degraded relative to the 521 oil;

however, additions of Spartina alterniflora detritus decreased the extent of degradation for both fractions. The

results for aromatic constituents are shown in Figure 3.

Given that microbial activities in wetlands ecosystems

are usually limited by oxygen and nutrients, oil degradation may compete with the degradation of natural organic substrates for these substances. In addition, the

hemicellulose fraction of lignocellulose constitutes a

readily degradable carbon source that may compete

with the utilization of petroleum hydrocarbons by the

indigenous microbial community.

In the treatments containing no oil, the number of oil

degraders was constant over the 3-month period (approximately 6.31 x 105

 bacteria/g of sediment), which

indicates that addition of nutrients only does not increase the population of degraders. The two treatments

containing oil produced an increase of oil degraders

during the second month of experiment (Figure 4), which

may correspond to the initial, rapid degradation of the

alkane fraction. By the third month, the oil degraders had

decreased, possibly reflecting the slower rate of degradation of aromatic compounds.

The numbers of heterotrophic bacteria were higher in

the presence of Spartina and oil together than in any of

the other treatments. Although the numbers of oil degraders in both treatments containing oil were very similar, the extent of oil degradation was significantly

different. The diminished oil degradation in the presence

of oil and Spartina suggests that the combination of

substrates may antagonize the activity of indigenous oil

degraders, either directly or by competition for nutrients

and/or oxygen. Given the abundance of Spartina alterniflora in coastal regions of the United States, the interactions between oil-degrading and other heterotrophic

bacteria and the impact of natural substrates should not

be overlooked in the design of remediation strategies.

References

1. Brown, E.J., and J.F. Braddock. 1990. Sheen screen, a miniaturized most-probable-number method for enumeration of oil-degrading microorganisms. Appl. Environ. Microbiol. 56:3,895-3,896.

2. Tsai, Y., and B.H. Olson. 1991. Rapid method for direct extraction

of DNA from soil and sediment. Appl. Environ. Microbiol. 57:1,070-

1,074.

Figure 3. Residual aromatics in surface layer of microcosms

after 3 months in the presence of a natural organic

substrate.

Figure 4. Heterotrophic bacteria and oil degraders in surface

layer of microcosms after 3 months of exposure to

oil.

34
Progress Toward Verification of Intrinsic Cobioremediation of 

Chlorinated Aliphatics

Mark Henry

Michigan Department of Natural Resources, Oscoda, Michigan

A plume consisting of chlorinated aliphatic and aromatic

hydrocarbons mixed with JP-4 (benzene, toluene, ethylbenzene, and the xylenes; BTEX) from a former Air

Force fire-fighting training area shows evidence of continual natural bioremediation. This site is being characterized and monitored by the National Center for

Integrated Bioremediation Research and Development

(NCIBRD) at Wurtsmith Air Force Base, a decommissioned installation located in lower northeast Michigan.

Wurtsmith is bounded by the Au Sable River to the south

and west and by Van Etten Lake to the north and east.

The base sits on a 20-m bed of homogeneous glacial

alluvial sand and gravel aquifer underlain by a thick clay

aquitard. The average ground-water depth in the study

area is 6 m. The hydraulic conductivity has been reported to be 3e-3 cm/sec (v = 2.6 m/day) at the site,

which has resulted in a narrow 50 m x 300+ m plume.

The plume is monitored on a quarterly basis through the

use of dedicated bladder pumps installed in 37 monitoring wells at the site. Local ground-water elevations are

continually recorded by a datalogging network of recorder wells.

The site has been characterized through 2 years of

quarterly sampling of the well and pieziometer network,

as well as direct analysis of continuous cores (gathered

by resident Geoprobe sampling equipment) across the

plume. This information is supplemented by a weekly

monitoring of the vertical temperature profile of the site

and periodic soil gas profiles.

In general, the site has a large amount of residual

fuel/solvent residing near the interface of the water

column and capillary fringe, extending at least 125 m

from the source. Soil gas measurements in the vadose

zone near the source indicate that the interstices contain approximately 65 percent methane, 30 percent

carbon dioxide, ppb level hydrogen sulfide and nitrogen,

and virtually no oxygen. The ground water beneath the

free product has almost no dissolved oxygen and has

depressed redox potential, increased electrolytic conductivity, depressed pH levels, and increased concentrations of reduced iron. BTEX levels steadily decrease

over the length of the plume. While perchloroethylene

(PCE) and trichloroethylene (TCE) levels are significant

(greater than 1,500 mg/kg) in the solids, only trace levels

are found in the ground water. As the dissolved plume

moves downgradient, the predominant chlorinated species are cis-1,2-dichloroethylene and vinyl chloride.

On the fringe of the contamination, BTEX metabolites

such as m,p-toluic acid and salicylic acid have been

identified.

The disappearance of TCE, PCE, and BTEX and the

appearance of bacterial metabolites of these compounds over the length of the plume suggest that these

contaminants are being bioattenuated within the same

plume. Changes in redox potential, temperature, and pH

support this assumption. It remains to be seen whether

or not these processes are interrelated, and are perhaps

influenced by bacterially mediated iron or manganese

reduction.

35
Phytoremediation of Petroleum-Contaminated Soil:

Laboratory, Greenhouse, and Field Studies

M. Katherine Banks

Department of Civil Engineering, Kansas State University, Manhattan, Kansas

A. Paul Schwab

Department of Agronomy, Kansas State University, Manhattan, Kansas

Common environmental problems associated with the

pumping and refining of crude oil are the disposal of

petroleum sludge and pipeline leaks. Contaminants are

often treated by incorporation into the soil. If the soil is

frequently tilled and fertilized, soil microorganisms will

be stimulated and organic contaminants biodegraded.

Unfortunately, the biodegradation rate of more recalcitrant and potentially toxic contaminants, such as the

polycyclic aromatic hydrocarbons (PAHs), is rapid at first

but declines quickly. Biodegradation of these compounds is limited by their strong adsorption potential and

low solubility.

Recent research suggests that vegetation may play an

important role in the biodegradation of toxic organic

chemicals in soil. The establishment of vegetation on

hazardous waste sites may be an economical, effective,

low-maintenance approach to waste remediation and

stabilization. The use of plants for remediation may be

especially appropriate for soils contaminated by organic

chemicals to depths of less than 2 m. The beneficial

effects of vegetation on the biodegradation of hazardous

organics are two-fold: organic contaminants may be

taken up by the plant and accumulated, metabolized, or

volatilized; and the rhizosphere microflora may accelerate biodegradation of the contaminants.

Completed greenhouse studies indicate that vegetative

remediation is a feasible method for cleanup of surface

soil contaminated with petroleum products. A field demonstration is necessary, however, to exhibit this new

technology to the industrial community. In this project,

several petroleum-contaminated field sites have been

chosen in collaboration with three industrial partners.

These sites have been thoroughly characterized for

chemical properties, physical properties, and initial contaminant concentrations. A variety of plant species have

been established on the sites, including warm and cool

season grasses and legumes. Soil analyses for the

target compounds over time indicate that the interaction

between plants and rhizosphere microflora significantly

enhances remediation of the contaminated soils. Continued monitoring will allow us to assess the efficiency

and applicability of this remediation approach.

37
Section 3

Performance Evaluation

In an effort to evaluate the performance of various bioremediation technologies, researchers assess

the extent and rate of cleanup for particular bioremediation methods. They also study the environmental fate and effects of compounds and their byproducts because remediation efforts at a

contaminated site can produce intermediate compounds that can themselves be hazardous. Thus,

another important aspect of performance evaluation projects involves assessing the risk of potential

health effects and identifying bioremediation approaches that best protect public health.

To this end, EPA’s National Health and Environmental Effects Research Laboratory (NHEERL)

developed an integrated program to address 1) the toxicity of known hazardous waste site contaminants, their natural breakdown products, and their bioremediation products; 2) the development of

methods to screen microorganisms for potential adverse health effects; 3) the potential for adverse

effects when chemical/chemical chemical/microorganism interactions occur; and 4) the development of methods to better extrapolate toxicological bioassay results to the understanding of potential

human toxicity.

Two performance evaluation papers were presented at the symposium studying the toxicity of

hazardous waste mixtures before and after bioremediation. One group of microorganisms (Pseudomonas aeruginosa and Phanerochaete chrysosporium) were found to be unable to significantly

degrade or eliminate the mutagenic activity of a mixture of several aromatic chemicals.

39
Detoxification of Model Compounds and Complex Waste Mixtures Using

Indigenous and Enriched Microbial Cultures

K.C. Donnelly, Jeannine L. Capizzi, Ling-Yu He, and Henry J. Huebner

Texas A&M University, College Station, Texas

Introduction

Biological degradation of organic compounds may be

considered an economical tool for remediating hazardous waste contaminated environments. While some environments may be too severely contaminated for initial

in situ treatment to be effective, most contaminated

media will use some form of biological degradation in

the final treatment phases. Traditional studies monitor

the success of bioremediation by the loss of a parent

compound or class of chemicals. Prior studies, however,

have observed both an increase and decrease in toxicity

as a result of degradation (1-3). This study uses both

genotoxicity bioassays and chemical analysis to monitor

the efficacy of biological degradation of model compounds and a complex chemical mixture in soils.

Methods

A Weswood sandy loam soil was allotted in 7-kg portions

into 40 stainless steel boxes. Four soil treatment categories included: 1) no amendment; 2) amendment with

three model chemicals, i.e., benzo(a)pyrene (BAP),

2,4,6-trinitrotoluene (TNT), and pentachlorophenol

(PCP); 3) amendment with a wood preserving bottom

sediment waste (WPW); and 4) amendment with both

the model chemicals and WPW. Additionally, each of the

four soil treatment categories were divided into threebox subunits and inoculated with either an indigenous

culture, a Pseudomonas aeruginosa culture, or a

Phanerochaete chrysosporium culture. Thus, a total of

40 boxes were prepared, including the soil with no

chemical amendment and the soil with indigenous, bacteria, or fungi inocula (NI, NB, NF); a wood preserving

waste amended soil with each of the three inocula (WI,

WB, WF); soil amended with the model chemicals and

each of the microbial inocula (MI, MB, MF); soil

amended with both the model chemicals and wood preserving waste and each inolula (CI, CB, CF); and sterilized soil receiving each of the chemical amendments

(St-N, St-W, St-M, St-C). The soil boxes were stored in

a greenhouse under constant temperature and humidity,

and monitored to maintain moisture content. Samples

were collected on Days 0, 90, 180, and 360 posttreatment. The solvent was extracted with methylene chloride and methanol, then evaluated for genotoxicity in the

Salmonella/microsome assay. Benzo(a)pyrene and

pentachlorophenol were quantified using gas chromatography (GC), whereas TNT was quantified using highperformance liquid chromatography (HPLC).

Results

The data presented in Table 1 describe the quantity of

solvent-extractable organics recovered from the soil

treatments using methylene chloride and methanol. The

data indicate that smaller quantities of organics were

recovered from the treatments with indigenous organisms on 0 and 30-day sampling periods. The soil treatment with the wood preserving waste, model chemicals

(CI), and indigenous organisms was not appreciably

changed during the 360-day incubation. The waste and

bacteria treatment (WB) yielded an average of 32 mg/g

solvent-extractable organics on Day 0 and 20 mg/g on

Day 360. Overall, none of the treatments amended with

wood-preserving waste displayed an appreciable reduction over the initial 360 days of treatment.

The extract of the unamended Weswood soil induced an

average negative mutagenic response at Days 0 and 90.

In general, treatments with the model chemicals, wood

preserving waste, or combined treatments induced an

average positive mutagenic response at all time points.

With metabolic activation, the methylene chloride extract of the MB treatment induced an average of 167 net

TA98 revertants on Day 0 and 122 net revertants on Day

90. A slight increase in the mutagenic response was

observed in the majority of extracts of samples collected

on Day 180. The methylene chloride extracts of the

waste amended soils (WI, WB, and WF) collected on

Day 180 induced responses with activation that ranged

from 46 net revertants (WI) to 63 net revertants per mg

41
of residue (WB). The methylene chloride extracts of

the combined waste and model chemical soils (CI,

CB, and CF) collected on Day 180 induced responses

that ranged from 71 net revertants (CB) to 76 net

revertants per mg of residue (CF and CI). Most of the

methanol extracts induced a stronger response than

was observed in the methylene chloride extract. At

least one solvent-extractable fraction from each of the

soils collected on Day 180 induced a positive mutagenic response.

Discussion

Approximately 1 year after the application of both wood

preserving waste and model chemicals to a Weswood

soil, the level of solvent-extractable organics was not

appreciably reduced. Although chemical analysis suggested reductions in the model chemicals in some of the

treatments, contaminant concentrations were all detectable after 360 days. The data do suggest that the presence of oils and hydrophobic chemicals in the waste

amended soil may have limited the availability of chemicals for microbial degradation. Changes in mutagenicity

over the initial year of the study indicate that the bioavailability of chemicals was increased in some treatments.

These results indicate that 360 days of treatment was

insufficient to eliminate the mutagenic activity of soils

amended with either the model chemicals or the wood

preserving waste. Data collection will continue for an

additional 720 days to monitor the chemical and toxicological changes associated with each treatment.

References

1. Aprill, W., R.C. Sims, J.L. Sims, and J.E. Matthews. 1990. Assessing detoxification and degradation of wood preserving and petroleum wastes in contaminated soil. Waste Manag. and Res.

8:45-65.

2. George, S.E., D.A. Whitehouse, and L.D. Claxton. 1992. Gentoxicity of 2,3,4-trichlorophenoxyacetic acid biodegradation products in

the Salmonella reversion and lamba prophage-induction bioassays. Environ. Toxicol. Chem. 116:733-740.

3. Donnelly, K.C., P. Davol, K.W. Brown, M. Estiri, and J.C. Thomas.

1987. Mutagenic activity of two soils amended with a woodpreserving waste. Environ. Sci. Technol. 21:57-64.

Table 1. Solvent Extractable Organics Recovered From Soil Treatments (mg residue/g soil)

Time (days after treatment)

Treatmenta 0 30 60 90 180 360

WI 19.80 ± 3.2b 23.18 ± 1.6 29.45 ± 2.2 24.18 ± 0.8 20.28 ± 1.8 18.25 ± 0.8

WB 32.25 ± 1.1 27.25 ± 2.5 26.10 ± 2.8 23.81 ± 1.2 25.24 ± 3.8 19.78 ± 1.6

WF 21.04 ± 0.9 28.35 ± 2.0 30.84 ± 4.6 23.79 ± 2.0 21.19 ± 1.7 18.14 ± 1.0

CI 19.17 ± 3.5 29.26 ± 1.7 28.16 ± 3.2 26.67 ± 2.7 24.06 ± 0.7 20.21 ± 0.4

CB 26.45 ± 1.6 25.42 ± 0.9 26.35 ± 1.0 23.54 ± 0.4 23.97 ± 1.8 19.62 ± 1.1

CF 23.36 ± 2.0 31.51 ± 0.6 29.90 ± 1.9 26.17 ± 0.2 21.71 ± 2.6 19.57 ± 0.9

StC 27.05 30.31 32.88 36.60 26.95 28.78

a

 WI = wood preserving waste with indigenous bacteria.

 WB = wood preserving waste with P. aeruginosa.

 WF = wood preserving waste with P. chrysosporium.

 CI = combined model chemicals and wood preserving waste with indigenous bacteria.

 CB = combined model chemicals and wood preserving waste with P. aeruginosa.

 CF = combined model chemicals and wood preserving waste with P. chrysosporium.

 StC = sterilized soil with combined model chemicals and wood preserving waste.

b

 Values are the mean of three replicates ± standard deviation.

42
Assessing the Genotoxicity of Complex Waste Mixtures

Larry D. Claxton, Virginia S. Houk, Sarah H. Warren, Thomas J. Hughes, and Susan E. George

Environmental Carcinogenesis Division, National Health and Environmental Effects Research

Laboratory, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina

Introduction

Some chemicals associated with environmental spills

and hazardous waste sites can induce permanent genetic alterations in all organisms. These changes, called

mutations, can have deleterious effects on individuals

and their descendants. In human populations, mutations

are known to increase the incidence of cancer and

genetic diseases and may play a role in numerous other

diseases. In nonhuman biota, mutations may alter the

balance of an ecosystem or change the virulence of

pathogens.

In spite of 20 years of research concerning environmental mutagenesis, current knowledge concerning the

identity and effects of most environmental mutagens

(including genotoxic carcinogens) is limited, and data on

their effects on other living organisms and the ecosystem are practically nonexistent. Because their mode of

action theoretically does not depend on a threshold limit,

genotoxicants are a class of toxic substances that

should be examined when found in even low concentrations in the environment.

The major objective of this paper is to review the U.S.

Environmental Protection Agency’s (EPA’s) research in

the area of genetic toxicology as it relates to bioremediation. Readers can refer to a completed manuscript (1)

for more in-depth information, examples of studies and

data, and a discussion of the regulatory context.

Research Targeted by the Risk

Assessment Paradigm

The complexity of environmental situations complicates

the monitoring, evaluation, and risk assessment of hazardous waste sites and spills. The composition of pollutant mixtures may consist of a few or thousands of

individual compounds, and remediation efforts produce

additional components that add to the complexity of

evaluating risk. EPA research targets efforts that

enhance the risk assessment process. Research, therefore, can be placed into the four categories associated

with the risk assessment process: hazard identification,

exposure assessment, dose-response assessment, and

risk assessment modeling and calculation methods.

Hazard Identification Research

Hazard identification research strengthens the risk assessment process by developing methods that are more

reliable and more cost effective, and, when possible,

give some sense of relative toxicity when identifying

toxic agents in the environment. This research involves

the development of new assay methods (e.g., the spiral

mutagenicity assay), the enhancement of existing assays (e.g., the modification of the prophage assay), the

integration of bioassay and analytical chemistry methods, and the modification of methods for use with complex environmental mixtures rather than with single

environmental chemicals.

Exposure Assessment Research

Generally, a hazardous substance must come into contact with a specific biological component (e.g., deoxyribonucleic acid) before a toxic response can occur.

Measuring or estimating the potential level of contact

between a toxicant and its reactive target (or a surrogate

for that target) is exposure assessment. Exposure assessment is done on a population and not an individual

basis. Quantitative toxicological assays and analytical

chemistry methods have been developed to determine

relative amounts of toxicity potential existing in different

environmental (e.g., treated versus untreated) sites, to

determine the relative bioavailability of environmental

toxicants and their metabolites, and to follow the change

in environmental concentrations of toxicants over time.

In collaboration with other EPA laboratories, the National

Health and Environmental Effects Research Laboratory (NHEERL) has demonstrated the usefulness of

bioassays to enhance and calibrate exposure assessment methods.

43
Dose-Response Assessment

Dose-response assessments quantitate the potency of

an environmental agent(s) for a specific outcome. Although the NHEERL bioremediation program does not

support whole animal carcinogenicity and mutagenicity

studies, the Environmental Carcinogenesis Division

(ECD) of NHEERL does develop, in support of other

programs, information relevant for bioremediation hazard and risk assessments. In addition, ECD has developed a statistical software system for analyzing, in a

quantitative manner, short-term mutagenicity tests.

Risk Assessment Modeling Research

When information is limited or only a general characterization is required, a qualitative assessment can be

made; however, there is often a need to be as quantitative as possible in risk assessments. ECD, therefore, not

only develops quantitative data but also examines

mechanisms and developing models to help scientists

understand to what extent the toxic effect is relevant to

human populations. Mechanistic research, for example,

may demonstrate that rodents respond to a toxin in a

manner not relevant to humans, alleviating the need for

a quantitative human risk assessment. Likewise, mechanistic research may demonstrate whether or not synergism should be considered when evaluating a

multichemical site.

The purpose of the presentation given at this conference

is to catalog NHEERL bioremediation research, to show

its relevance, and to provide examples of how the information obtained will be useful in future assessments.

Reference

1. Claxton, L.D., V.S. Houk, and S.E. George. 1995. Integration of

complex mixture toxicity and microbiological analyses for environmental remediation research. In: de Serres, F.J., and A.

Bloom, eds. Ecotoxicology and human health. Boca Raton, FL:

Lewis Publishers. In press.

44
Section 4

Pilot-Scale Research

By studying bioremediation processes under actual site conditions on a small scale, researchers

can gather critical information on issues such as operation, control, and management of residuals

and emissions before moving to full-scale research. This is a critical intermediate step in which the

success of laboratory experiments are further tested in an expanded but controlled setting.

Pilot-scale evaluations covered many different tools for bioremediation, including biofilters, compost

piles, and slurry bioreactors. One paper studied the optimization of biofilters for use in removing

volatile organic contaminants from the air, while another sought to establish the optimal operating

conditions for compost media. Finally, a third paper found that many organic contaminants become

concentrated in the foam produced in slurry bioreactors and suggested that this effect could be

utilized to isolate contaminants from the rest of the slurry.

Two papers focused on the use of combinations of aerobic and anaerobic conditions to degrade

recalcitrant chlorinated wastes. Work was presented on biofiltration with gel beads that provide an

oxic environment at their surface and an anoxic environment at their centers. The authors found

that such filtration was highly effective at degrading trichloroethylene (TCE). A system was evaluated

for land treatment that involved switching between anaerobic and aerobic conditions.

Research continued on the integration of physical and biological processes to clean up organic

wastes. One paper showed how wood preserving wastes can be first removed from their soil matrix

by washing and distillation processes and then degraded in a bioreactor.

Two presentations covered research into optimizing the biodegradation of TCE. The first was the

result of the study of a variety of parameters that might affect the success of field-scale TCE

treatment. The next characterized several potentially useful bacteria isolated from a biofiltration

device.

Two papers investigated the use of microorganisms to clean up inorganic wastes. One technique

mentioned involves the biological reduction of sulfates into sulfides. This technique helps to remove

metal cations from solution. The other technique involves the biological reduction of metal cations

into less harmful oxidation states. The authors of the final paper tested the ability of microorganisms

to reduce chromium (VI).

Poster presentations described several pilot-scale research projects. The poster projects within this

category involved the testing of a compost bioreactor, a biofilm-electrode reactor, and a laminar-type

flow reactor. 

45
In Situ Bioremediation of Trichloroethylene With Burkholderia Cepacia PR1:

Analysis of Parameters for Establishing a Treatment Zone

Richard A. Snyder

Center for Environmental Diagnostics and Bioremediation, University of West Florida,

Pensacola, Florida

M. James Hendry and John R. Lawrence

National Hydrology Research Institute, Environment Canada, Saskatoon, Canada

Introduction

The use of chloroethenes, including trichloroethylene

(TCE), has led to an extensive contamination of groundwater resources in the United States. In situ bioremediation of this contaminant is being pursued with the

aerobic microorganism Burkholderia cepacia (formerly

Pseudomonas cepacia) G4 (1). Mutant strains G4

PR123 and PR131 constitutively produce a toluene orthomonooxygenase (TOM) that mineralizes TCE. The gene

for this enzyme is located on the degradative plasmid

TOM of B. cepacia (2). The constitutive expression of

this gene is the result of secondary transposition following Tn5 insertion mutagenesis, which also confers

kanamycin (km) resistance to these bacteria. PR123

contains a single entire Tn5 in the chromosome and an

IS50R of Tn5 in the plasmid. PR131 bears an entire Tn5

and an IS50R in the TOM31c plasmid. The IS50 elements

are at nearly the same locations and are thought to be

responsible for the constitutive expression of tom.

This project involves ground-water flow control with

sealed sheet piling to funnel ground water through a

narrow gate area in which treatment technologies can

be applied. Parameters for the establishment of a biological treatment zone with PR1 have been investigated

in the laboratory. Transport and survival characteristics

of the bacterium have been examined in ground water

and sediment from a targeted release site (the aquifer

under the Canadian Armed Forces Base, Borden, Ontario). Monitoring techniques have been developed for

tracking PR1 populations in the treatment zone and

determining the extent of dispersal and survival beyond

the treatment zone.

The functional integration of non-native bacteria in natural microbial communities and maintenance of bacterial

populations above normal environmental background

concentrations provides a challenge for both microbial

ecologists and applied microbiologists. Monitoring of

population dynamics and trophic interactions is critical

for successful bioaugmentation applications. Risk assessment associated with uncontained biotechnological

introductions also requires careful monitoring of the

survival and dispersal of released microorganisms and

altered genes. Although releases of non-native or recombinant bacteria have not been reported to result in

adverse environmental effects to date, there is a responsibility to ensure that released microorganisms will be

constrained by the selective pressures of the target

environment.

Tracking

Selective media provide a first step in tracking the organism. Phenol, o-cresol, and phthalic acid combined

with kanamycin (km) were tested for growth of PR1 and

isolates from the Borden aquifer. Growth of PR1 was

optimal on 20 mM phthalate medium. In aquifer sediment slurries, numbers of native bacteria isolated on this

medium range from 0 to less than 5 x 105

 colony forming

units (CFU) and direct epifluorescence microscopy

counts of bacterivorous protists are less than 8 x 102

mL-1 in unamended incubations.

A monoclonal antibody (mab) specific to G4 lipopolysaccharide (LPS) (3) has been used both in direct immunofluorescent counts and to confirm CFU of PR1 on

phthalate plates. Details of the production and testing of

this mab can be found in Winkler et al. (3).

Polymerase chain reaction (PCR) primers targeting the

Tn5 insertion junctions have been developed and

tested. The primers were designed based on the

47
assumption that the insertion points would be unique for

PR1. The PCR products resulting from this primer set

are of slightly different size for PR123 and PR131, reflecting the slightly shifted location of the IS50R in the plasmid in these two strains. Attempts to extract indigenous

bacterial deoxyribonucleic acid (DNA) from Borden

aquifer sediments and to amplify product from these

extractions with the PR1 primers have failed. We have

demonstrated that PCR amplification with these primers

will work with whole, intact cells added directly to the

PCR reaction. The current limit of detection for positive

amplification from cell suspensions (equivalent to pore

water samples) is 1 x 103

 cells/mL-1, but from sediment

slurries the detection limit drops an order of magnitude.

These results indicate that this assay will be most useful

for large volumes of pore water samples collected on

Sterivex filters (Millipore Corp.) for extraction of DNA (4).

Potential host range for the TOM plasmid from PR131 in

24 random isolates from Borden aquifer sediments (R2A

medium; 5) was assessed by direct filter matings. Overnight cultures of PR131 grown in lactate medium and

aquifer isolates in R2A medium were pelleted and resuspended in R2A medium to an optical density of 1.0 @

600 nm. One milliliter of donor and recipient were mixed

and filtered on a 0.2 μm pore-size polycarbonate filter.

Filters were placed on R2A plates and incubated overnight at 30°C. Filters were then removed and vortexed

in R2A broth and suspensions plated on selective media

predetermined by antibiotic screening. Out of 24 isolates, 42 percent were positive for TOM31c transfer by

PCR detection, and 80 percent of these were positive

for tom activity as indicated by a positive transformation

of trifluoromethyl phenol (TFMP) (1) and mineralization

of TCE in standing cell assays. These data indicate a

wide potential host range within the target environment

for TOM, highlighting the need to ensure tracking capability for both the organism to be introduced and its

associated genetic elements. These native bacterial

strains bearing functional TOM plasmids may also be

better adapted for bioremediation application in the target environment than PR1.

Analysis of PR1 Transport in Borden

Aquifer Material

Transport characteristics of PR1 in Borden sand are

important in determining the retardation and dispersal of

cells within a treatment zone and in the downstream

aquifer. Cell retardation is important to allow contaminated water to flow past the inoculated populations of

PR1. A series of experiments was carried out at three

scales: 3.8-cm, 10-cm, and 40-cm length columns

packed with sterile and nonsterile Borden sand and

commercially available silica sand. Artificial ground

water (AGW) was used as the solute. All columns were

packed under standing water to minimize air entrapment

and tamped with a glass rod to attain a uniform bulk

density of 1.8 g/cm-3 and a porosity of 0.4. Ground-water

velocity was set to the approximate velocity of water in

the target site. Bacterial suspensions were pulsed

through for 1 void volume, and the effluents were collected in sterile vials with a chromatography fraction

collector. Numbers of PR1 were determined by plating

on selective media. Chloride (Cl) ion as a conservative

tracer was measured with Ag/AgCl electrodes calibrated

against Cl ion analysis by ion chromatography.

Both the Cl tracer and bacterial breakthrough curves

displayed similar patterns between columns, indicting

good replication. All bacterial breakthrough curves exhibited a notable pulse of bacteria corresponding to the

breakthrough of the Cl tracer. Reduced peak concentrations of bacteria relative to breakthrough concentrations

of Cl (greater than 99 percent) indicate irreversible sorption of PR1 onto the geologic media. Well-defined tailing

was also observed over the duration of the experiments,

indicating reversible sorption of PR1 to the geologic

media. Peak heights and tailing were three orders of

magnitude lower in Borden material than silica sand,

possibly due to clays and iron coatings on Borden

sands. Breakthrough of PR1 was not affected by sterilization of the sediment or the addition of a cotransported

bacterium. These results were integrated with predation

loss rates of PR1 added to Borden sediment in slurry

microcosms to develop a predictive model with both

physical and biological parameters for the transport and

fate of the organism within the Borden aquifer.

Survival of PR1 in Aquifer Microcosms

Much of the target environment consists of anaerobic

saturated sediments. Survival of the organism beyond

an oxygenated target zone will likely depend on its ability

to withstand conditions of little or no oxygen. Plate

counts of a suspension of cells with Nitrogen (N2) gas

flushed through the head space indicate little effect on

PR131 viability through 6 days. After this point, culturable

numbers drop precipitously but maintained a low population level through 25 days of anaerobic conditions.

No PR1-specific viruses could be isolated from the target environment. Samples of aquifer sediment were

incubated with growing PR1 cells as an enrichment, and

supernatants were tested on PR1 overlay plates to look

for cleared viral plaques.

Survival of PR1 above 1 x 107

 cells/mL-1 is of interest in

establishing needed inoculation densities for effective

bioremediation (6). Survival below this concentration

and long-term integration of the organism into the native

microbial community was of interest for risk assessment. A series of experiments has been conducted in

ground water, sediment slurries, and flow columns to

examine population dynamics of PR1 and the bacterivorous protists from the Borden aquifer that are the

primary vector for loss of PR1. Use of a monoclonal

48
antibody (mab) to PR1 LPS (3) allows enumeration of

PR1 after PR1 numbers have been reduced to the

background level of phthalate utilizers in the system.

The loss of PR1 cells and corresponding tracking data

by mab indicates that with increasing inoculation density, PR1 populations are sustained for longer periods

prior to rapid loss. This delay in the loss of cell numbers

can be attributed to exceeding the maximum response

of the bacterivorous protists in the system.

Data on the time of sustained PR1 populations and the

loss rate of PR1 cells after this delay were compiled for

all sediment slurry incubations. The delay in rapid loss

of PR1 was incorporated into a regression analysis of

time of sustained PR1 populations above 1 x 107

 as a

function of inoculation density. A half-life estimation by

regression analysis incorporates loss rate data after the

period of sustained population numbers as a function of

inoculation density. This former analysis provides an

estimated lifetime for a pulse of PR1 into the treatment

zone. The latter analysis provides an extinction coefficient for cells in the downstream aquifer.

An aquifer sediment column has been established to

test the response of PR1 within the flow regime of the

target system. A commercial spring water (GMW) was

used as the diluent. A chromatography column was fitted

with cut gas chromatography (GC) vials closed with

Teflon septa to provide sampling ports. Teflon tubing and

fittings were used throughout the setup. Tygon tubing

connected to a constant temperature recirculating bath

was used to jacket the column and maintain 15°C. Flow

was controlled at the column outflow to the flow rate in

the target environment (2 cm/day-1). Cells and TCE were

added by syringe pump. An overflow ensured constant

supply of diluent, cell suspension, or TCE solution to the

top of the sediment. Pore-water samples were taken by

syringe (2:300 μL each), and used for plate counts,

direct counts of bacteria and protists, and TCE analysis

(alternate sampling periods).

With an inoculum of 1 x 108

 cells/mL-1 for 1 void volume,

a population of greater than 107

 CFU/mL-1 pore water

was maintained for 5 days at in the top portion of the

column. Column data expressed as depth profiles showing a roughly linear decrease in PR131 numbers with

distance traveled through the column from Days 1

through 5, and the combined effect of predation and

elution decreased numbers in the upper portion of the

column at Days 8 and 10. By Day 15, the pulse has been

eliminated at the upper and lower portions of the column, leaving residual cells in the central portion. Bacterivorous protists increased in proportion to the

numbers of PR1 added. These organisms form resistant

cysts on sediment surfaces when food is not available.

Decreases observed over time for protist numbers are

likely due mainly to encystment of these organisms after

depletion of PR131. This results in a reservoir of

bacterivores capable of responding to subsequent additions of bacteria.

Integration of PR1 Into Stable Microbial

Consortia

Persistence of a non-native bacterium introduced into

an environment is dependent on the ability of that organism to find refuges from predation and compete with

native bacteria. One such refuge may be in biofilms. To

examine the ability of PR1 to integrate into biofilms, PR1

was introduced into existing biofilms developed from

Borden aquifer material and into a defined microbial

community (including protists) grown with input of TCE

and dibutyl phthalate. PR1 was found not to stick to

pre-existing biofilms of Borden aquifer origin, but successfully integrated into the degradative community

growing in the presence of TCE and dibutyl phthalate.

The addition of these substances apparently provided a

competitive advantage to PR1 in this system. One year

post-inoculation, PR1 was located in biofilms from this

system by fluorescent monoclonal antibody and scanning confocal laser microscopy. PR1 was found throughout the biofilms as scattered cells and microcolonies.

This information suggests that in the presence of TCE it

may be possible to maintain and grow PR1 within a

treatment zone in the target aquifer.

Modeling of Transport and Fate of PR1 in

Borden Aquifer

Using the transport data and loss rates due to predation,

preliminary modeling exercises have been conducted.

These initial approximations assume grazing losses will

be the same for pore water and attached bacteria, and

do not address excystment/encystment processes of

the protists, but do incorporate protist grazing as a

dynamic response to variable bacterial density. In all

simulations, predation reduces the bacterial concentrations in the effluent from a modeled transport column.

With an increased predation constant, peak breakthrough numbers are reduced, but tailing is affected

more significantly. Tailing, which can be attributed to the

process of reversible sorption, is of concern for transport

of bacteria to greater distances in geologic media.

Thus, reductions in tailing by the protist response reduces the concern of offsite migration of the introduced

microorganism.

Conclusions

Data collected from these laboratory studies indicate

that biological interactions, particularly with bacterivorous protists, will limit the survival of PR1 introduced into

the system and may provide for a natural containment

of the bacterium. Maintenance of PR1 numbers high

enough for mineralization activity will likely require repeated additions of sufficient cells to exceed protist

49
maximum response. Preliminary data on TCE additions

suggest that TCE may impede protist activity until it is

mineralized, providing a gradient of protist response in

proportion to TCE removal. Ultimately, the utility of laboratory analyses will be gauged against data from the

field release.

Acknowledgments

The authors thank M.S. Shields and S.C. Francisconi for

advise and assistance in the research. This work was

supported by the U.S. Environmental Protection

Agency’s Gulf Breeze Environmental Research Laboratory through cooperative research agreement 822568 to

RAS.

References

1. Shields, M.S., S.O. Montgomery, S.M. Cuskey, P.J. Chapman, and

P.H. Pritchard. 1991. Mutants of Pseudomonas cepacia G4 defective in catabolism of aromatic compounds and trichloroethylene.

Appl. Environ. Microbiol. 57:1,935-1,941.

2. Shields, M.S., Reagin, M.J., Gerger, R.R., Cambell, R., Somerville,

C.C. 1995. TOM: A new aeromatic degradative plasmid from Burkholderia (Pseudomonas) cepacia G4. Appl. Environ. Microbiol.

61:1,352-1,356. 

3. Winkler, J., K.N. Timmis, and R.A. Snyder. 1995. Tracking the

response of Burkholderia cepacia G4 5223 PR1 in aquifer microcosms. Appl. Environ. Microbiol. 61:448-455.

4. Somerville, C.C., I.T. Knight, W.L. Straube, and R.R. Colwell. 1989.

Simple, rapid method for direct isolation of nucleic acids from

aquatic environments. Appl. Environ. Microbiol. 55:548-554.

5. Reasoner, D.J., and E.E. Geldreich. 1985. A new medium for the

enumeration and subculture of bacteria from potable water. Appl.

Environ. Microbiol 49:1-7.

6. Krumme, M.L., K.N. Timmis, and D.F. Dwyer. 1993. Degradation

of trichloroethylene by Pseudomonas cepacia G4 and the constitutive mutant strain G4 5223 PR1 in aquifer microcosms. Appl.

Environ. Microbiol. 59:2,746-2,749.

50
Characterization of Trichloroethylene-Degrading Bacteria 

From an Aerobic Biofilter

Alec Breen, Todd Ward, Ginger Reinemeyer, John Loper, and Rakesh Govind

University of Cincinnati, Cincinnati, Ohio

John Haines

U.S. Environmental Protection Agency, Cincinnati, Ohio

Introduction

The microbial community colonizing a vapor-phase

biofilter was examined to determine the population(s)

capable of trichloroethylene (TCE) degradation. The

community had been exposed to low levels of TCE

continuously for 24 months and maintained degradation

in the absence of a canonical co-metabolite. Although

low levels of autotrophic ammonia-oxidizing bacteria

were present, nitrapyrin inhibitor studies suggested that

alternative bacteria were responsible for TCE oxidation.

In addition, replacement of ammonia with nitrate did not

affect TCE degradation. Incubation of biofilter biomass

in a toluene or benzene atmosphere resulted in a turbid

culture within 2 to 3 days. In light of these observations,

aromatic hydrocarbon-oxidizing bacteria were pursued

as putative candidates mediating TCE degradation.

A significant fraction of the culturable heterotrophs

(greater than 80 percent) were capable of growth on

toluene or benzene. This study describes the naturally

occurring TCE-degradative populations that became established over time in the biofilter. Individual isolates

were tested for TCE-degradative capacity under several

growth conditions. The pure cultures tested were all

capable of co-metabolic TCE degradation. These organisms had persisted in the biofilter regardless of conditions that would exert a negative selective pressure due

to generation of the TCE-derived epoxide during aerobic

TCE degradation. Several isolates were selected for

further study. Initial sampling of the biofilter yielded three

isolates: Rhodococcus sp. TA1, Pseudomonas putida

TA2, and Nocardia sp. AR1. Both the Rhodococcus and

the P. putida could be repeatedly isolated from the biofilter. Two other organisms, P. putida DC1 and Burkholderia cepacia GR3, were isolated more recently. This

consortium appears relatively resistant to the toxic effects of TCE oxidation at the concentrations used in the

biofilter.

Background

Halogenated aliphatic compounds are a major class of

industrially important chemicals that have become significant environmental contaminants with mutagenic and

carcinogenic potential. A widespread ground-water contaminant, TCE can undergo co-metabolic oxidation by

a variety of physiologically diverse bacteria (1). Cometabolic TCE degradation by aromatic hydrocarbon

utilizing bacteria was originally reported by Nelson et al.

(2). Since that time, efforts to employ these organisms

to ameliorate TCE contamination problems in situ and

in reactors have been conducted. This study examines

the toluene-degrading bacteria surviving in a vaporphase biofilter. Although only a mineral salts medium

was supplied to the biofilter, toluene oxidizers survived

and TCE degradation was maintained at a level of 20

percent of a 21-ppmv gas stream.

Experimental Methods

Biofilter and Sampling

The biofilter consisted of ceramic plates in a stainless

steel casing. The initial inoculum was a municipal sludge

sample that was acclimated to a volatile organic compound (VOC) mixture (benzene, toluene, ethylbenzene,

and TCE) for a period of 3 months. At this point, all VOCs

except TCE were removed. The biofilter was operated

at a gas-flow rate of 520 mL/min, and had an empty-bed

residence time of 1.9 min. TCE inlet concentration was

21 ppmv. A mineral salts solution was applied to the

biofilter at a flow rate of 357 mL/day. Biofilter sampling

was conducted by opening the biofilter and scraping

biomass off the ceramic matrix. VOC-degrading bacteria

51
were isolated by incubation of biofilter material in a

mineral salts medium, with the appropriate VOC supplied in the vapor phase.

TCE Mineralization Assays

Degradation experiments using 14C-TCE were conducted in 20.0 mL vials, with teflon-lined silicone septa

closures allowing injection into the vial. An inner vial

containing 0.4 N NaOH served as a CO2 trap. Sterile

control vials were subtracted from experimental values

when determining conversion of TCE to CO2 or soluble

products. All data represent a mean value of triplicate

vials.

Biochemical and Genetic Characterization of

Biofilter Bacteria

Aromatic hydrocarbon utilizing bacteria were isolated as

follows. Biofilter biomass was inoculated into mineral

salts medium, and the medium was exposed to toluene

or benzene vapor as a sole carbon source for 48 hr.

Cultures were then plated onto mineral salts medium

and grown in a toluene or benzene atmosphere for 10

days. Organisms that appeared were picked and

streaked onto mineral salts plates and grown again with

the carbon source in the vapor phase. The isolates were

then checked for purity and TCE mineralization capability. Isolates were sent to Microbial ID, Inc. (Newark,

Delaware), for fatty acid methyl ester (FAME) analysis.

16S rDNA sequencing was carried out by amplifying the

27 to 321 base pair region of the 16S rDNA gene by

polymerase chain reaction (PCR). The primers used are

forward ’5AGAGTTTGATCCTGGCTCAG-3’ (positions

27-46) and reverse ’5AGTCTGGACCGTGTCTCAGT-3’

(positions 321-301). Both forward and reverse DNA

strands were sequenced. Gene probes for characterized toluene/TCE co-metabolic oxygenases were obtained from other researchers. The todABC probe was

obtained from Dr. D.T. Gibson, and the tbu probe was

obtained from Dr. A. Byrne and Dr. R.H. Olsen (3, 4).

Results and Discussion

The predominant bacteria in the biofilter were shown to

be degrading TCE by toluene/benzene oxygenase cometabolic route. The biofilter community had not been

exposed to these compounds for over 24 months, yet

these organisms persisted and were shown to be the

key population mineralizing TCE. These isolates are of

interest because they arose spontaneously from a naturally occurring population and were maintained in the

continual presence of TCE. It might be expected that

TCE-oxidizing organisms would be selected against in

such a system (5). Mixed cultures mineralized 14C-TCE

(data not shown) and exhibited very little diversity when

plated out onto a mineral salts medium and grown with

toluene or benzene vapor as a sole carbon source.

These plates generally had only one or two colony

morphotypes shown to be a P. putida (designated TA2)

and a Rhodococcus (designated TA1). A Nocardia sp.

(designated AR1) was isolated during an early sampling

time but was not reisolated. Two additional organisms,

B. cepacia GR3 and P. putida DC1, were more recently

isolated.

The predominant toluene degrader isolated from the

biofilter was TA2. Growth on glucose, succinate, or tryptophan completely inhibited TCE mineralization by TA2.

A gene probe for the alpha subunit of the tbu toluene

monooxygenase strongly hybridizes with TA2 (Figure 1).

This probe will also hybridize to two other toluene

monooxygenases, tmo and tom (data from our laboratory and unpublished information from M.S. Shields,

University of West Florida, and R.H. Olsen, University

of Michigan; therefore, the mode of toluene oxidation is

not definitively established. Slot blot analysis of biofilter

isolates probed with tbu is shown in Figure 1.

Rhodococcus sp.TA1 did not hybridize strongly to either

the tbu or tod toluene oxygenase probes, indicating the

uniqueness of its toluene oxygenase. In addition to TA1

and TA2, three other TCE-co-metabolizing organisms

from the reactor were investigated, and their properties

are listed in Table 1. All organisms are being evaluated

to determine basal levels of TCE catabolic activity as

toluene is depleted. The effect of acclimation to alternative substrates on TCE degradation is also being

examined.

Figure 1. Slot blot analysis of biofilter DNA extracts probed

with the tbu monooxygenase alpha component:

A) P. putida DC1, B) B. cepacia GR3, C) P. putida

TA2, D) Nocardia sp. AR1, and E) Rhodococcus

sp. TA1.

52
References

1. Ensley, B.D. 1991. Biochemical diversity of trichloroethylene metabolism. Arm. Rev. Microbiol. 45:283-299.

2. Nelson, M.K.J., S.O. Montgomery, E.J. O’Neil, and P.H. Pritchard.

1986. Aerobic metabolism of trichloroethylene by a bacterial isolate. Appl. Environ. Microbiol. 52:383-384.

3. Byrne, A.M., J. Kukor, and R.H. Olsen. 1995. Sequence of the

gene cluster encoding toluene-3-monooxygenase from Pseudomonas pickettii PKO1. Gene. In press.

4. Wackett, L.P., and S.R. Householder. 1989. Toxicity of trichloroethylene to Pseudomonas putida F1 is mediated by toluene

dioxygenase. Appl. Environ. Microbiol. 55:2,723-2,725.

5. Alvarez-Cohen, L., and P.L. McCarty. 1991. A co-metabolic

biotransformation model for halogenated aliphatic compounds exhibiting product toxicity. Environ. Sci. Tech. 25:1,381-1,387.

Table 1. Properties and Growth Substrates of TCE-Co-metabolizing Biofilter Isolates

Organism Oxidase Indolea Benzene Ethylbenzene Phenol o-Cresol m-Cresol p-Cresol

TA1 − −+ + + + + + + + + + + + + + + +

TA2 + ++ + + + + + −+ + −

AR1 − ++ + + + + + + + + + + + + + + −

DC1 + −+ + + + + + _ −− −

GR3 + ++ + + + + + −+ − −

a

 Conversion of indole to indigo.

53
Anaerobic/Aerobic Degradation of Aliphatic Chlorinated Hydrocarbons in an

Encapsulated Biomass Biofilter

Rakesh Govind and P.S.R.V. Prasad

Department of Chemical Engineering, University of Cincinnati, Cincinnati, Ohio

Dolloff F. Bishop

National Risk Management Research Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Introduction

During the past decade, public awareness and concern

about the quantity and diversity of persistent (recalcitrant to degradation) synthetic chemicals produced by

industry have increased. Release of these chemicals

into the environment is inevitable, and hence there is a

strong need to control, direct, and improve the processes for degradation of these chemicals. Large differences in the rates and mechanisms of biodegradation

of various compounds under oxic and anoxic conditions

exist. Consequently, sequential anoxic and oxic conditions, enabling cooperation of anaerobic and aerobic

bacteria, are desirable for rapid complete mineralization

of many polyhalogenated compounds (1-6).

Encapsulation of Biomass in Hydrogels

The goal of this research is to use hydrogel-encapsulated bacteria for simultaneous creation of oxic and

anoxic zones inside the hydrogel bead. Further, the oxic

and anoxic zones created inside the hydrogel bead can

be successfully used to mineralize chlorinated compounds such as trichloroethylene (TCE) and perchloroethylene (PCE). Hydrogel beads with encapsulated

bacteria can be used to mineralize chlorinated compounds present in air, ground water, and soil, and can

also be used to promote ecological competitiveness of

laboratory-grown cultures that are specially adapted for

biodegradation of specific environmental pollutants.

Preparation of New Gel Material (7)

The gelation procedure used is as follows: Silica sol,

obtained commercially as LUDOX colloidal silica SM

grade, is mixed with 1 to 3 percent sodium alginate in

the following proportion: silica sol (90 wt percent to

99 wt percent) : 1 to 3 percent sodium alginate solution

(1 wt percent to 10 wt percent). The mixture, after adjustment of pH between 7 and 8 by 5N HCl, is mixed

with active aerobic and anaerobic biomass cells, with

the initial cell wt percent varying from 2.0 to 10.0. The

mixture is stirred, then poured into a petri dish to a depth

of 5 mm. Calcium chloride solution (0.1 molar) is poured

on top of the mixture in the petri dish. The silica sol and

sodium alginate mixture immediately gel due to the

diffusion of calcium forming calcium alginate on the

outer surface, and the gel is allowed to cure from 10 min

to 24 hr. During the curing process, the pH of the silica

sol decreases, thereby forming silica gel with pockets of

calcium alginate inside the silica gel. The survival of

active cells is maximized by using a combination of silica

sol and sodium alginate. Once the silica sol has formed

silica gel, stainless-steel wire mesh cylinders 2.5 mm in

diameter and 5.0 mm long (open at both ends) are

pushed into the gel layer, thereby enclosing the gel

inside the wire mesh. The use of the stainless steel

mesh gives the silica gel/calcium alginate bead structural strength so that the beads can be packed in a bed

without compaction.

Experimental Studies Conducted

A 40-mL bioreactor (1.9 cm inner diameter) consisting

of a jacketed cylinder was constructed from borosilicate

glass. The reactor was packed randomly with the gel

beads. Air was passed at a controlled rate through the

bioreactor, and nutrient solution was trickled down from

the top of the bioreactor counter current to the air flow.

The air was contaminated with chlorinated pollutants,

such as TCE or PCE, using a syringe pump that injected

the liquid contaminant into the air line through a septum.

The concentration of the contaminant in the air stream

54
was varied within the following range: toluene 0 to

100 ppmv; TCE 0 to 25 ppmv; and PCE 0 to 25 ppmv.

The reactor temperature was maintained at 25o

C by circulating water from a constant temperature bath through the

jacket of the bioreactor. Nutrient solution was trickled

down the bioreactor at a flowrate of 1 liter per day, and

the nutrient composition was as follows: KH2PO4 (85 mg/L),

K2HPO4 (217.5 mg/L), Na2HPO4⋅2H2O (334 mg/L), NH4Cl

(25 mg/L), MgSO4⋅7H2O (22.5 mg/L), CaCl2 (27.5 mg/L),

FeCl3⋅6H2O (0.25 mg/L), MnSO4⋅H2O (0.0399 mg/L),

H3BO3 (0.0572 mg/L), ZnSO4⋅7H2O (0.0428 mg/L),

(NH4)6Mo7O24 (0.0347 mg/L), FeCl3⋅EDTA (0.1 mg/L), yeast

extract (0.15 mg/L), and formate (50 mg/L). Results obtained

are shown in Table 1.

Carbon and chlorine balances were made by monitoring

the increase in carbon dioxide in the exit air, and increase in chloride ion concentration in the exit nutrients

was analyzed by an ion chromatograph. The chlorine

balance was developed at steady-state conditions within

an error band of 15 percent of the calculated increase

in chloride ion concentration.

The proposed degradation pathway was shown to be a

partial dehalogenation in the anoxic zone followed by

oxic biodegradation of the anoxic degradation products

in the outer aerobic zone of the gel bead. The anoxic

zone was created due to oxygen consumption in the

aerobic zone by the oxic degradation of the partially

dehalogenated products as they diffused out from the

anoxic zone.

A mathematical model was developed to describe the

diffusion of TCE and oxygen, and consumption of oxygen due to aerobic degradation of the dehalogenated

products. At the outer surface of the gel bead (denoted

by dimensionless position of 1.0) the oxygen concentration is about 8 mg/L due to presence of air outside the

bead. As oxygen diffuses inside the gel bead, it is consumed due to aerobic degradation of the dehalogenated

products diffusing outwards. At some point in the interior

of the gel bead, oxygen is completely consumed producing an anoxic zone in the interior portion of the gel bead.

It is in this anoxic zone that dehalogenation of TCE

occurs. The formate in the nutrient medium is rapidly

absorbed by the gel bead and provides the organic

carbon source needed for partial dehalogenation of TCE

in the anoxic zone by anaerobic microbiota. Other potential carbon sources for anaerobic microorganisms

are acetate and other carboxylic acids.

Experiments also were conducted with perchloroethylene (PCE) at an inlet concentration of 25 ppmv.

Results obtained are shown in Table 2. Chloride ion

balances were obtained at steady-state to prove that

complete mineralization of PCE had occurred. Each

experiment had to be conducted for over 5 days to

achieve a stable exit concentration of chloride ion in the

exit nutrients. No other by products were observed in the

exit gas phase at the above operating conditions.

Conclusions

The hydrogel-encapsulated biomass reactor is capable

of biodegrading trichloroethylene (TCE) and perchloroethylene (PCE) through an anaerobic/aerobic degradation mechanism. Experimental results indicate that

the degradation of TCE and PCE is complete, and the

empty-bed gas phase residence time for complete

removal is less than a few minutes. Further studies

are ongoing to quantitate the transport parameters

and apply the process for treatment of TCE or PCE

in ground water.

References

1. Abramowitz, D.A. 1990. Aerobic and anaerobic biodegradation of

PCBs: A review. Crit. Rev. Biotechnol. 10:241-251.

2. Beunink, G., and H.J. Rehm. 1988. Synchronous anaerobic and

aerobic degradation of DDT by an immobilized mixed culture system. Appl. Environ. Microbiol. 151:95-100.

3. Beunink, J., and H.J. Rehm. 1990. Coupled reductive and oxidative degradation of 4-chloro-2-nitrophenol by a co-immobilized

mixed culture system. Appl. Microbiol. Biotechnol. 29:72-80.

4. Fathepure, B.Z., and T.M. Vogel. 1991. Complete degradation of

polychlorinated hydrocarbons by a two-stage biofilm reactor. Appl.

Environ. Microbiol. 57:3,418-3,422.

5. Fogel, S., R.L. Lancione, and A.E. Sewall. 1982. Enhanced biodegradation of Methoxychlor in soil under sequential environmental

conditions. Appl. Environ. Microbiol. 44:113-120.

Table 1. Percent degradation of TCE in the bioreactor at

various air flowrates. Inlet TCE concentration was

25 ppmv, and nutrient flowrate was 1 liter/day.

Air Flowrate

(mL/min)

Percent Degradation

of TCE

35 100.0

40 67.2

50 40.7

60 22.1

65 10.8

Table 2. Percent degradation of PCE in the bioreactor at

various air flowrates. Inlet PCE concentration was

25 ppmv, and nutrient flowrate was 1 liter/day.

Air Flowrate

(mL/min)

Percent Degradation

of PCE

10 100.0

15 86.7

20 72.4

30 41.8

50 12.8

55
6. Kastner, M. 1991. Reductive dechlorination of tri- and tetrachloroethylenes depends on transition from aerobic to anaerobic conditions. Appl. Environ. Microbiol. 57:2,039-2,046.

7. Bishop, D.F., and R. Govind. 1995. New hydrogel material for

degradation of persistent pollutants in immobilized film bioreactors.

U.S. Patent Application.

56
Operation and Optimization of Granular Air Biofilters

Francis Lee Smith, George A. Sorial, Makram T. Suidan, Amit Pandit, and Pratim Biswas

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Richard C. Brenner

Risk Reduction Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio

Introduction

Since enactment of the 1990 amendments to the Clean

Air Act, the control and removal of volatile organic compounds (VOCs) from contaminated air streams have

become major public concerns (1). Consequently, considerable interest has evolved in developing more economical technologies for cleaning contaminated air

streams, especially large, dilute air streams. Biofiltration

has emerged as a practical air pollution control (APC)

technology for VOC removal (2-4). In fact, biofiltration

can be a cost-effective alternative to the more traditional

technologies, such as carbon absorption and incineration, for removal of low levels of VOCs in large air

streams (5, 6). Such cost-effectiveness stems from a

combination of low energy requirements and microbial

oxidation of the VOCs at ambient conditions.

Our biofiltration research has focused on expanding the

range of application of biofiltration technology to the

treatment of high VOC loadings at consistently high

removal efficiencies. The preliminary period of our research was dedicated to the pilot-scale comparison of

three different types of biological attachment media: a

patented peat mixture and two synthetic inorganic media, one channelized and the other pelletized. The biofilters containing the latter two media were operated as

trickle bed air biofilters (TBABs), called such because

the media received a steady application of water. After

18 months of testing, the pelletized medium (Celite

6-mm R-635 Bio-Catalyst Carrier) was demonstrated to

be significantly better than the other two for handling

high VOC loadings (7-9). Subsequent work to evaluate

the performance and behavior of biofilters using the

R-635 pelletized medium produced two significant findings: first, that an increase in the biofilter operating

temperature permits a significantly higher practical VOC

loading (i.e., a significantly smaller required media volume), and second, that biofilter performance decreases

substantially with the buildup of back pressure due

to the accumulation of biomass within the media bed

(10, 11).

Working exclusively with this pelletized medium, our

continued research focused on the development of

strategies for long-term operation with high VOC loadings at sustained high-removal efficiencies. This research effort demonstrated that this objective could be

achieved using a biomass removal and control strategy

employing periodic backwashing of the media with

water. Backwashing (the upflow washing of the fluidized

media with water) gently removes excess biomass from

the media, circumventing the problems noted earlier for

this medium. A second finding of this research was that

NO3-N as the sole nitrogen source was superior to

NH3-N. The use of NO3-N resulted in lower volatile

suspended solids (VSS): chemical oxygen demand

(COD) and VSS:N ratios. In other words, for a given set

of operating conditions, less biomass is produced and

less nitrogen is consumed. Finally, it was also observed

that both the recovery of the VOC removal efficiency

with time after backwashing (unsteady state) and the

VOC removal efficiency with depth (at near steady state)

were superior when using NO3-N.

This paper discusses the continuing research being performed for the development of biofiltration as an efficient, reliable, and cost-effective VOC APC technology.

The objectives of this effort were to investigate the removal efficiencies of TBABs under high toluene loadings

and low residence times, and to evaluate the associated

development and control of excess biomass with time.

The biofilter operational period between backwashings

was evaluated to determine its effect on the stability of

biofilter performance. Backwashing variables, including

backwashing frequency and backwashing duration,

were evaluated.

57
Methodology

The biofilter apparatus used in this study consisted of

four independent, parallel biofilter trains, each containing 3.75 ft of pelletized Celite 6-mm R-635 biological

attachment medium. A detailed schematic, equipment

description, and typical system operation are given elsewhere (8). Each biofilter had a circular cross section with

a 5.75-in. internal diameter (ID). The air feed was massflow controlled, and the VOC (liquid toluene) was metered by syringe pumps into the air feed stream. Each

biofilter was fed a buffered nutrient feed solution containing all necessary macro- and micronutrients with a

sodium bicarbonate buffer, described elsewhere (8). For

each biofilter, the sole nitrogen nutrient source was NO3-

N. The flow directions of the air and nutrients were

downward. All biofilters were insulated and independently temperature controlled at 32.2°C.

Results

Each biofilter was loaded with clean, sterilized pellets

and seeded with backwashing water from a similar,

previous run. For each biofilter, a significant dip in performance occurred from Days 17 to 26 due to an error

in preparing the nutrient solution that resulted in feeding

insufficient NO3-N. After the target VOC loading was

achieved, at a COD:NO3-N ratio of 50:1, it was maintained for the duration of the run.

Three different backwashing strategies were tested on

all biofilters sequentially. After restart following backwashing, effluent samples were collected to determine

the recovery of the VOC removal efficiency with time.

On days when no backwashing occurred, samples were

collected along the length of the bed to determine the

VOC removal efficiency with respect to depth.

Biofilter A

This biofilter was started up at 50 ppmv toluene influent

concentration, 1.33 min empty bed residence time

(EBRT), and 21 mmol NO3-N per day. On Day 17, the

biofilter was backwashed for the first time. Detailed

schematic and backwashing descriptions are given

elsewhere (12). The procedure used was to recycle 70

L of 32.2°C tap water through the bed, bottom to top, at

a rate of 57 L/min to induce full media fluidization at a

bed expansion of about 40 percent. At the end of the

backwashing period, the media was flushed at the same

rate with another 50 L of clean, 32.2°C tap water. For

this first backwashing strategy, the period and frequency

were 1 hr twice per week. The target influent VOC

concentration (500 ppmv toluene) and loading (6.2 kg

COD/m3

 day) were reached on Day 53. On Day 129, the

second backwashing strategy was started using a period and frequency of 2 hr twice per week. On Day 171,

the third and final backwashing strategy was started

using a period and frequency of 1 hr every 2 days. The

performance of Biofilter A is shown in Figure 1.

Biofilter B

This biofilter was started up at 50 ppmv toluene influent

concentration, 0.67 min EBRT, and 21 mmol NO3-N per

day. On Day 17, the biofilter was backwashed for the

first time using the first backwashing strategy of 1 hr

twice per week. The target influent VOC concentration

(250 ppmv toluene) and loading (6.2 kg COD/m3

 day)

were reached on Day 35. On Day 129, the second

backwashing strategy of 2 hr twice per week was begun,

and on Day 171, the third strategy of 1 hr every 2 days

was begun. The performance of Biofilter B is shown in

Figure 2.

Biofilter C

This biofilter was started up at 50 ppmv toluene influent

concentration, 1.0 min EBRT, and 21 mmol NO3-N per

day. On Day 17, the biofilter was backwashed for the

first time using the first backwashing strategy of 1 hr

twice per week. The target influent VOC concentration

(250 ppmv toluene) and loading (4.1 kg COD/m3

 day)

Figure 1. Performance of Biofilter A with backwashing.

58
were reached on Day 53. On Day 129, the second

backwashing strategy of 2 hr twice per week was begun,

and on Day 171, the third strategy of 1 hr every 2 days

was begun. The performance of Biofilter C is shown in

Figure 3.

Biofilter D

This biofilter was started up at 50 ppmv toluene influent

concentration, 2.0 min EBRT, and 21 mmol NO3-N per

day. On Day 17, the biofilter was backwashed for the

first time using the first backwashing strategy of 1 hr

twice per week. The target influent VOC concentration

(500 ppmv toluene) and loading (4.1 kg COD/m3

 day)

were reached on Day 53. On Day 129, the second

backwashing strategy of 2 hr twice per week was begun,

and on Day 171, the third strategy of 1 hr every 2 days

was begun. The performance of Biofilter D is shown in

Figure 4.

Conclusions

The performances of the four biofilters with respect to

the three backwashing strategies were similar, although

clearly affected by both the loading and the residence

time. The effectiveness of the three strategies increased

from the first through the third strategy. This shows that

both backwashing duration and frequency are very important parameters for control of the biofilters’ VOC

removal efficiency. The third and best strategy, however,

actually had less total backwashing time per week. At

the higher loading, the greater than 90 percent removal

efficiencies of both biofilters were unexpectedly high for

the third backwashing strategy but below the sustained

99.9 percent achieved by the lower loaded biofilters. It

can also be seen that for a given loading, the performance at the lower EBRTs is more sensitive to the backwashing strategy employed. Both of these effects were

anticipated; what was not anticipated was that this

pelletized medium would perform so well for any backwashing strategy at a loading of 6.2 kg COD/m3

 day.

These findings, as well as biofilter recovery of performance after backwashing, will be presented.

Acknowledgment

This research was supported by Cooperative Agreement CR-821029 with the U.S. Environmental Protection Agency.

Figure 2. Performance of Biofilter B with backwashing.

Figure 3. Performance of Biofilter C with backwashing.

59
References

 1. Lee, B. 1991. Highlights of the Clean Air Act Amendments of

1990. J. Air Waste Manag. Assoc. 41(1):16-31.

 2. Leson, G., and A.M. Winer. 1991. Biofiltration: An innovative air

pollution control technology for VOC Emissions. J. Air Waste

Manag. Assoc. 41(8):1,045-1,054.

 3. Leson, G., F. Tabatabal, and A.M. Winer. 1992. Control of hazardous and toxic air emissions by biofiltration. Paper presented

at the Annual Meeting and Exhibition of the Air & Waste Management Association, Kansas City, MO, June 21-26.

 4. Ottengraf, S.P.P. 1986. Exhaust gas purification. Rehn, H.J., and

G. Reed, eds. In Biotechnology, Vol. 8. Weinham, Germany: VCH

Verlagsgesellschaft.

 5. Ottengraf, S.P.P. 1986. Biological elimination of volatile xenobiotic

compounds in biofilters. Bioprocess Eng. 1:61-69.

 6. Severin, B.F., J. Shi, and T. Hayes. 1993. Destruction of gas

industry VOCs in a biofilter. Paper presented at the IGT sixth

International Symposium on Gas, Oil, and Environmental Technology. Colorado Springs, CO, November 29 - December 1.

 7. Smith, F.L., G.A. Sorial, P.J. Smith, M.T. Suidan, P. Biswas, and

R.C. Brenner. 1993. Preliminary evaluation of attachment media

for gas phase biofilters. Paper presented at the U.S. EPA Symposium on Bioremediation of Hazardous Wastes: Research, Development, and Field Evaluation. Dallas, TX, May 4-6.

 8. Sorial, G.A., F.L. Smith, P.J. Smith, M.T. Suidan, P. Biswas, and

R.C. Brenner. 1993. Development of aerobic biofilter design criteria for treating VOCs. Paper no. 93-TP-52A.04. Presented at

the 86th Annual Meeting and Exhibition of Air & Waste Management Association. Denver, CO, June 13-18.

 9. Sorial, G.A., F.L. Smith, P.J. Smith, M.T. Suidan, P. Biswas, and

R.C. Brenner. 1993. Evaluation of biofilter media for treatment of

air streams containing VOCs. In: Proceedings of the Water Environment Federation 66th Annual Conference and Exposition,

Facility Operations Symposia, Volume X. pp. 429-439.

10. Sorial, G.A., F.L. Smith, M.T. Suidan, P. Biswas, and R.C. Brenner. 1994. Evaluation of the performance of trickle bed biofilters—

Impact of periodic removal of accumulated biomass. Paper no.

94-RA115A.05. Presented at the 87th Annual Meeting and Exhibition of Air & Waste Management Association. Cincinnati, OH,

June 19-24.

11. Smith, F.L., G.A. Sorial, M.T. Suidan, P. Biswas, and R.C. Brenner.

1994. Pilot-scale evaluation of alternative biofilter attachment media for the treatment of VOCs. Paper presented at the U.S. EPA

Symposium on Bioremediation of Hazardous Wastes: Research,

Development, and Field Evaluations. San Francisco, CA, June

28-30.

12. Smith, F.L., M.T. Suidan, G.A. Sorial, P. Biswas, and R.C. Brenner. 1994. Trickle bed biofilter performance: Biomass control and

N-nutrient effects. Paper no. AC946004. Presented at the Water

Environment Federation 67th Annual Conference and Exposition,

Facility Operations Symposia. Chicago, IL, October 15-19.

Figure 4. Performance of Biofilter D with backwashing.

60
Abiotic Fate Mechanisms in Soil Slurry Bioreactors

John A. Glaser and Paul T. McCauley

National Risk Management Research Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Majid A. Dosani, Jennifer S. Platt, and E. Radha Krishnan

IT Corporation, Cincinnati, Ohio

Introduction

Soil slurry treatment of contaminated soil has been

shown to offer a viable technology for soil bioremediation. This technology, however, has not sufficiently progressed to be a durable, reliable, and cost-effective

treatment option (1).

The use of aggressive mixing energy to provide conditions for improved contact between soil contamination

and microorganisms capable of degrading the contamination is the hallmark of slurry treatment technology. A

more complete description of pollutant mass transfer

during the treatment phase is required that includes

treatment fate mechanisms attributable to biotic and

abiotic processes. Losses attributable to abiotic means

can be overlooked in field application of the technology,

because limited questions can be successfully addressed at field scale. Discussions with U.S. Environmental Protection Agency (EPA) regional personnel and

inspection of active field-scale soil slurry bioreactor operations have identified operational problems, such as

foaming, that could result in possible abiotic loss (2).

Field bioslurry operations have adopted various approaches to reduce foaming: 1) addition of defoaming

agents, 2) reduction of the rotational speed of the agitator, and 3) reduction of gas flow through the bioreactor

system. The foaming phenomenon is generally considered a nuisance, rather than a potential beneficial removal mechanism. Where pollutants have a specific

gravity less than water once desorbed from the slurried

soil, the pollutants would rise to the surface, as in a

flotation process. One of our working hypotheses was

that foam formation could be related to this pollutant

release process. If this analysis has merit, it is possible

that the operational strategy used in the field is counterproductive, because a separated contaminant phase is

re-entrained with partially cleaned soil material.

We have conducted two bench-scale slurry bioreactor

treatability studies at the EPA Testing & Evaluation Facility in Cincinnati, Ohio, which were designed to assess

operating factors leading to foam formation, and to identify the most advantageous means to deal with foaming.

The initial study was previously presented as a general

treatability study for treatment of creosote contamination

in a soil (3). During this previous study, foaming became

a major problem for operation. Use of a defoamer controlled foaming conditions, as did reduction of the mixer

rotational speed and gas flow in the more extreme

cases. Subsequent studies devoted specifically to investigating the causes and conditions of foaming and

using a different batch of soil from the same site as the

earlier study showed little foaming at the beginning of

the study.

Methodology and Experimental Designs

Foam Study (First Study)

A soil from St. Louis Park, Minnesota, contaminated with

polynuclear aromatic hydrocarbons (PAHs) was used to

assess the importance of foaming conditions to the

performance of bench-scale slurry reactors. The design

of the bench-scale experimental bioslurry reactor has

been reported (3). Operational slurry volume was 6 L,

representing 75 percent of the total reactor volume.

To evaluate the conditions and causes of foam formation, a subsequent study was designed. This investigation used the monitoring conditions specified

for the treatability study, and was conducted using six bench-scale slurry reactors. Each reactor

was loaded at 30 percent solids, with an initial volume

61
of 6 L. The study design employed two reactors that

were permitted to develop foam, two reactors in which

foam formation was suppressed through the use of a

defoamer, and two reactors in which formaldehyde was

used to suppress biological activity. The study duration

was 1 week, with foam sampling based on the cumulative production throughout the study period.

Foam and Scale Study

Experimental variables selected for the foam and scale

study were 30 percent soil solids loading, two treatment

conditions of condensed foam (removed), and foam

retention within the reactors. Each condition was replicated with a single replicated control based on the foam

retention condition. The foam-retained conditions were

maintained through the use of a dispersant (Westvaco,

Reax 100M). The soil was classified to a minus 3/16 in.

dimension; no provisions were developed for preliminary dispersion of the soil solids or sand exclusion.

Scale was collected from the reactor walls after the

mobile soil solids were removed. The scale was strongly

fixed to the reactor wall, and some effort was required

to chip the scale away from the reactor.

General Reactor Operation Condition

The control reactors were operated under abiotic conditions to serve as bioinactive control reactors. Formaldehyde was used as a biocide in these reactors and

maintained at 2 percent residual concentration.

The following monitoring and operating conditions were

held constant for the reactors:

• Dissolved oxygen: greater than 2 mg/L

• pH: range of 6 to 9

• Ambient temperature: recorded daily

• Treatment duration: 10 weeks

• Nutrients: C:N:P ratio = 100:10:1

• Antifoam: as needed to control foam

Results

In a previous treatability study, a high solids control

reactor showed the greatest amount of foaming (3). The

amount of foaming was surprising because foam formation was expected to be related to the formation of

biosurfactants by the microbiota. Addition of formaldehyde to the control reactor was the only other explanation for the foam formation observed. Other reactors had

foaming problems, but this reactor was very noticeable

by contrast. Higher solids loading was also observed to

contribute to foam formation.

In contrast to the earlier studies in which foaming was

observed, the foaming study showed little foam formation. A small amount of foaming occurred on the first day

of operation. Figure 1 shows the increased concentration of the t-PAH analytes in the foam, which is five times

greater than the concentration found in the initial suspended soil slurry. A second attempt to evaluate foam

formation is shown in Figure 2. These data show a

decrease in the foam concentration factor, and are probably more realistic than the first foam study results. The

second study was also designed to evaluate the deposition of t-PAH analytes as part of a scale formed in the

reactor. Figures 3 and 4 show the results of t-PAH

deposition in the scale under conditions in which the

foam was condensed (removed from the reactor) and

retained within the reactor. From inspection of the results, it is clear that the higher molecular weight components of the t-PAHs were deposited in the scale in

quantities 10 to 90 percent above the initial suspended

slurry concentration. A mass balance analysis will be

presented that puts the importance of these abiotic fate

mechanisms into perspective.

Differences in physical characteristics of the soil and

operation of the bioslurry reactor between the two studies may have contributed to decreased foam formation

in the foam study. Although soil for the same site was

used for both the treatability study and foam study, the

batch of soil used for the foam study was coarser (less

than 1/4 in.) which may have resulted in lower PAH

Figure 1. Foam composition and concentration factor (Study 1).

Figure 2. Foam composition and concentration factor (Study 2).

62
concentrations in the foaming study and to decreased

foam formation. Furthermore, the air flow rate for reactors in the foam study (1 ft3

/min) was approximately

20 percent of that used in the treatability study (5

ft3

/min), which also may have contributed to decreased

foam formation.

Conclusions

Foam formation continues to be an unpredictable and

poorly understood event associated with slurry treatment. The results of our studies are based on the benchscale reactor and may exaggerate the abiotic fate

mechanisms due to high surface-to-volume ratio considerations. The concentration effects associated with foam

formation indicate that foam removal may be desirable

to optimize slurry reactor performance. Future studies

will endeavor to evaluate foam separation as part of the

slurry process. 

References

1. U.S. EPA. 1990. Engineering bulletin: Slurry biodegradation.

EPA/540/2-90/016. Cincinnati, OH.

2. Jerger, D.E., S.A. Erickson, and R.D. Rigger. 1994. Full-scale

slurry phase biological treatment of wood-preserving wastes at a

Superfund site. Draft manuscript.

3. Glaser, J.A., M.A. Dosani, P.T. McCauley, J.S. Platt, E.J. Opatken,

and E.R. Krishnan. 1994. Soil slurry bioreactors: Bench scale studies. In: U.S. EPA Twentieth Annual RREL Research Symposium:

Abstract Proceedings. EPA 600/R-94-011. Cincinnati, OH. p. 127.

Figure 4. Scale concentrations for foam retained reactors

(Study 2).

Figure 3. Scale concentrations for foam condensed reactors

(Study 2).

63
Design and Testing of an Experimental In-Vessel Composting System

Carl L. Potter and John A. Glaser

U.S. Environmental Protection Agency, Cincinnati, Ohio

Majid A. Dosani, Srinivas Krishnan, Timothy A. Deets, and E. Radha Krishnan

IT Corporation, Cincinnati, Ohio

Introduction

The goal of this compost research is to evaluate the

potential use of compost systems in remediation of soils

contaminated with hazardous chemicals. We have developed bench-scale composters to evaluate factors

controlling compost treatment at large scale. We are

currently studying the ability of compost microorganisms

to biodegrade polynuclear aromatic hydrocarbons

(PAHs) in in-vessel reactors located at the U.S. Environmental Protection Agency’s Test & Evaluation (T&E)

Facility in Cincinnati, Ohio.

Composting differs from other ex situ soil treatment

systems in that bulking agents are added to the compost

mixture to increase porosity and serve as sources of

easily assimilated carbon for biomass growth. Aerobic

metabolism generates heat, resulting in significant temperature increases that bring about changes in the microbial ecology of the compost mixture.

Optimal conditions for composting may vary depending

on many factors, but generally aerobic conditions with

45° to 55°C (mesophilic temperature range), 40 to 60

percent moisture, and a carbon-to-nitrogen ratio of 20:1

to 30:1 have been considered best. Mesophilic composting in the range of 35°C to 50°C might prove to be

the most effective at destroying certain wastes. Maintaining temperature below 50°C, however, may not always be cost effective if cooling requires too much

energy.

In an active compost pile, temperature can easily exceed 55°C, and temperatures above 70°C have been

reported. When the temperature exceeds 55°C, called

the thermophilic stage, most bacteria are killed. Organisms capable of sporulation, such as some bacteria (2)

and fungi (3, 4), will sporulate and remain dormant until

aerobic activity slows; the temperature falls back into the

mesophilic range when they re-emerge.

Reactor Design

Ten 55-gal, insulated stainless steel compost reactors

have been fabricated to provide the closely monitored

and controlled conditions required for treatability studies. These fully enclosed, computer-monitored, benchscale reactors hold about 1/4 yd3

 total compost mixture.

The reactor units stand upright with air flowing vertically

up through the compost mixture for 23 hours per day.

Enclosed units permit on-line analysis of oxygen, carbon

dioxide, and methane at inlet and exit locations. A datalogging system accumulates data and transmits them to

the PC-based central data system that monitors and

controls each reactor. XAD traps in the exit line of each

composter permit trapping of volatile organic compounds (VOCs) for analysis.

The bottom of each reactor contains a conical collection

system for periodic sampling of any leachate leaving the

reaction mixture. The space above the leachate collection system holds 2 in. of gravel. Mass balance studies

on soil contaminants are possible by direct sampling of

the reaction mixture at different depths through bung

holes in the lid, together with capture of VOCs and

leachate leaving the reactor.

Periodic determination of compost moisture content in

each reactor unit permits adjustment of total moisture

content in the compost matrix to 40 to 50 percent.

Moisture condensers inside compost units promote retention of moisture within the reactor. Otherwise, with

typical air flows, each unit could lose significant amounts

of water daily. If moisture falls below 40 percent, a water

distribution system inside the reactor may be used to add

water to the reaction mixture without opening the reactor.

The cylindrical reactor design permits mixing of reactor

contents by rolling each unit on a drum roller at desired

intervals. Mixing can be used to break up anaerobic

pockets and to avoid packing of the compost mixture. All

64
reactors are mixed simultaneously by placing them on

rollers over a modified conveyor belt that forces the

reactors to turn in unison. Baffles inside the reactors

promote mixing during rolling.

Insulation between the reactor core and outer shell reduces heat loss from the reactor during aerobic activity.

Heating coils provide the option of warming the reactor

to accelerate composting during startup. Each composter houses five thermocouples connected to a central

computer for on-line temperature measurements. Thermocouples reside at four equally spaced locations within

the compost mixture, and a fifth thermocouple tracks

ambient temperature outside the reactor. If the mean

temperature of the middle two reactor thermocouples

exceeds a predetermined high value, the computer

switches that unit to high air flow (60 L/min) to cool the

reaction mixture. After the high-temperature unit cools

to a specified low temperature, the computer switches

the unit back to low air flow (5 L/min) to reduce further

heat loss from the reaction mixture.

Methods

Current studies focus on defining acceptable operating

conditions and process characteristics to establish suitable parameters for treatment effectiveness. Parameters of interest include aeration, moisture dynamics,

heat production, and physical and chemical properties

of the compost mixture. Growth of microorganisms and

disappearance of parent compounds serve as indicators

of parameter suitability.

A 24-day treatability study, using field soil from the Reilly

Tar Pit Superfund site near Minneapolis, Minnesota, was

conducted to evaluate performance of the compost reactor system. The soil was contaminated with creosote and

contained 22 PAHs that were measured during the study.

The study design included five replicated treatment conditions involving different ratios of corn cobs to soil and

different air flow rates in 10 reactors. Soil/bulking agent

compositions evaluated in this study were 50:50 (four

reactors) and 30:70 (two reactors) ratios of corn cobs to

soil (50 percent soil and 70 percent soil, respectively).

Selected air flow rates were 5 and 10 L/min.

Results and Discussion

Temperatures in reactors with 50 percent soil and

moisture content of about 50 percent or less climbed to

the upper mesophilic and lower thermophilic ranges.

Temperatures in reactors with moisture content above

53 percent failed to increase much above 30°C. This

might indicate that higher moisture content restricted air

flow through the compost mixture, resulting in insufficient aerobic activity to attain high temperatures. Reactors with air flows of 5 and 10 L/min exhibited similar

temperatures within the compost mixture.

Reactors with 70 percent soil in the compost remained

relatively cool throughout the entire run, never reaching

the mid to upper mesophilic temperature range. These

reactors tended to maintain higher moisture content

throughout the study. Fewer corn cobs to absorb excess

moisture in the mixture may have resulted in flooding of

the pore space, blocking of air flow through the mixture,

and reduced drying.

Total heterotrophic populations increased from a range of

107

-108 to 109-7.6 x 1010 (60- to 300-fold increases) in

reactors during the first 24 hr of composting. Heterotroph

counts ranging from 1.6 x 109

 to 1.4 x 1010 remained after

24 days in reactors with 50 percent soil, but had returned

to around 2 x 108

 in reactors with 70 percent soil.

Small PAHs (two to three rings) were reduced by averages of 50 and 30 percent in compost mixtures of 50

and 70 percent soil, respectively, after 24 days. Large

PAHs (four to six rings) were not decreased under any

treatment condition after 24 days. Continued evaluation

of the compost mixture will provide more information on

the long-term ability of composting to destroy large PAHs.

Future investigations will include application to pentachlorophenol and other soil contaminants yet to be

specified. Evaluation of pollutant mass balance and

biotransformation products is an important aspect of

future research.

To judge the abilities of microorganisms to degrade

hazardous wastes in soil under various composting conditions, emphasis will be placed on diagnosing population changes throughout treatment and identifying

microbial species responsible for biodegradation of contaminants. Early microbiological studies have focused

on enumerating total microorganisms and determining

the presence of PAH degraders. Future studies will focus on characterizing changes in biological activity during the four stages of composting, and on identifying the

microbial species responsible for significant biodegradation of PAHs during each composting stage. Reappearance of fungi and other mesophiles (e.g., Actinomycetes)

during the cooling stage is also of interest.

References

1. Nakasaki, K., M. Sasaki, M. Shoda, and H. Kubota. 1985. Change

in microbial numbers during thermophilic composting of sewage

sludge with reference to CO2 evolution rate. Appl. Environ. Microbiol. 49(1):37-41.

2. Strom, P.F. 1985. Identification of thermophilic bacteria in solidwaste composting. Appl. Environ. Microbiol. 50(4):906-913.

3. Fogarty, A.M., and O.H. Tuovinen. 1991. Microbiological degradation of pesticides in yard waste composting. Microbiol. Rev.

June:225-233.

65
Integrated Systems To Remediate Soil Contaminated With Wood Treating Wastes

Makram T. Suidan, Amid P. Khodadoust, Gregory J. Wilson, and Karen M. Miller

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Carolyn M. Acheson and Richard C. Brenner

National Risk Management Research Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Introduction

Approximately 15 percent of Superfund Records of Decision (RODs) are directed towards sites contaminated

with wood treating wastes (1). Several types of pollutants characterize these sites, including pentachlorophenol (PCP), creosote, polycyclic aromatic hydrocarbons

(PAHs), other hydrocarbons, and heavy metals such as

copper, chromium, arsenic, and zinc (2). A process (Figure 1) that integrates soil washing with sequential anaerobic and aerobic biotreatment is being developed to

cost-effectively remediate soil contaminated with these

wood treating wastes. Soil washing facilitates degradation by mobilizing the target compounds and expanding

the range of feasible remediation technologies (3). To

reduce costs and the volume of PCP-bearing liquid, the

soil wash liquid is concentrated via distillation, and the

recovered ethanol and water is recycled to the first soil

washing unit. The remainder of the wash solution is

initially bioremediated in an anaerobic environment.

Mineralization of the target compounds is completed

aerobically (4). Process development began by independently evaluating soil washing and target compound

bioremediation. PCP-contaminated soils were the initial

focus, but this work is currently being extended to include soils contaminated with both PCP and PAHs.

Based on preliminary results, the integrated process will

meet the target cleanup level in 73 to 55 percent of the

RODs directed towards PCP remediation, resulting in

soil with a residual PCP level of 8 to 13 mg/kg, respectively (2).

Soil Washing/Solvent Extraction Studies

An equimass (50 percent) mixture of ethanol and water

(5) was found to be the optimal solution to remove PCP

from a variety of spiked soils in a bench-scale soil

washing process. This soil washing method removes

PCP at levels comparable with those achieved through

the analytical techniques of sonication and soxhlet extraction. Starting with initial spike levels of 85 to 100

mg/kg, 70 to 100 percent of the PCP added to the soil

was removed by washing, depending on soil particle

size, contamination age, and soil washing format. PCP

is extracted from soil in a 30-min contact time. The

availability of residual PCP on soils of 20 x 40, 100 x

140, and greater than 200 U.S. meshsize has been

Figure 1. Integrated soil treatment process. evaluated through a serial procedure: soils were

66
washed with 50 percent ethanol solution, rinsed with

water, and finally treated by soxhlet or sonication extraction using methanol/methylene chloride. Less than 4

percent of the residual PCP (less than 0.6 mg/kg) was

removed from the soil by the final sonication or soxhlet

extraction, demonstrating the limited availability of the

residual PCP. The solvent washing of soil with mixtures

of water and ethanol is also being investigated for PAHcontaminated soils, using four compounds on the U.S.

Environmental Protection Agency’s list of priority pollutants as model compounds: naphthalene, acenaphthene,

pyrene, and benzo(b)fluoranthene (6). A more ethanolrich mixture may be required to effectively mobilize PAHs

from soil.

A sequential soil washing train is being optimized for

PCP-contaminated soils in which ex situ soil washing is

performed with the 50 percent ethanol solution in three

batch-wash stages. After washing the soil for 30 min in

each stage, the washed soil is recovered from the soilsolvent slurry via vacuum filtration of the slurry, and a

fresh batch of solvent is added to the soil in the next

stage. Preliminary design data indicate that a series

of three 20-mL solvent washes will clean 5 g of soil

(1:12 soil:solvent ratio) as effectively as a single extraction of 100 mL cleans 1 g of soil (1:100 soil:solvent

ratio). Additional optimization will further decrease the

soil:solvent ratio.

Biological Treatment Studies

In the integrated process, the distillate bottoms will be

fed to an anaerobic fluidized-bed granular activated

carbon (GAC) reactor. Two of these reactors were constructed and operated for over 40 months, evaluating

variables such as PCP loading and reactor empty bed

contact time (EBCT) (7). The reactor volume is 10 L with

a 1-L recycle loop. Based on this evaluation, the following optimal operating variables were identified: EBCT,

2.3 hr; ethanol loading, 33.3 g/day (loading rate 6.3 g

chemical oxygen demand/L day); and PCP loading,

4.8 g/day (loading rate 0.55 g/L day). When the GAC

reactor operated at an EBCT of 2.3 hr, on a molar basis,

greater than 99.97 percent of the influent PCP was

dechlorinated to monochlorophenol (MCP). In addition,

data from the extraction of the reactor GAC during the

operating period indicated negligible accumulation of

PCP on the surface of the GAC. An aerobic fluidized-bed

GAC reactor will polish the effluent from the anaerobic

GAC reactor to attain complete mineralization of PCP.

Operation of the aerobic reactor has recently been

initiated.

An additional anaerobic fluidized-bed GAC reactor has

been constructed to evaluate the biotreatment of chemically synthesized solutions of the four PAHs and PCP in

ethanol. Greater than 99 percent transformation of the

influent PCP concentration of 100 mg/L has been

achieved in the reactor, while operating the reactor with

an EBCT of 9.3 hr. The reactor effluent data represent

greater than 99 percent removal for naphthalene, acenaphthene, and pyrene, and greater than 90 percent

removal for benzo(b)fluoranthene in the reactor.

References

1. U.S. EPA. 1994. Innovative treatment technologies: Annual status

report, 6th ed. EPA/542/R-94/005. Cincinnati, OH.

2. U.S. EPA. 1992. Contaminants and remedial options at wood preserving sites. EPA/600/R-92/182. Cincinnati, OH.

3. U.S. EPA. 1990. Soil washing treatment. Engineering bulletin.

EPA/540/2-90/017. Cincinnati, OH.

4. Khodadoust, A.P., J.A. Wagner, M.T. Suidan, and S.I. Safferman.

1994. Solvent washing of PCP contaminated soils with anaerobic

treatment of wash fluids. Water Environ. Res. 66:692.

5. U.S. EPA. 1986. Microbial decomposition of chlorinated aromatic

compounds. EPA/600/2-86/090. Cincinnati, OH.

6. Keith, L.H., and W.A. Telliard. 1979. Priority Pollutants I: A perspective view. Environ. Sci. Technol. 13:416.

7. Wilson, G.J., J.A. Wagner, A.P. Khodadoust, M.T. Suidan, and R.C.

Brenner. 1994. The evaluation of empty bed contact time on the

biodegradation of pentachlorophenol using an anaerobic GAC

fluidized-bed. In: Proceedings of the National Conference on Environmental Engineering: Critical Issues in Water and Wastewater

Treatment. American Society of Civil Engineers. p. 624.

67
Biological Treatment of Contaminated Soils Using Redox Control

Margaret J. Kupferle

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Gregory D. Sayles

National Risk Management Research Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Tiehong L. Huang, Yonggui Shan, Maoxiu Wang, and Guanrong You

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Carolyn M. Acheson

National Risk Management Research Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Introduction

Land treatment is a well-understood, cost-effective

means of conducting aerobic biological treatment of

soils contaminated with aerobically biodegradable compounds, such as petroleum. Common contaminants in

soil also include highly chlorinated organics that are not

readily biodegraded aerobically, such as pentachlorophenol (PCP), polychlorinated biphenyls (PCBs), and

1,1,1-trichloro-2, 2-bis(p-chlorophenyl)ethane (DDT).

These compounds may, however, be efficiently degraded using a sequential anaerobic/aerobic treatment

strategy. A cost-effective process to treat soils contaminated with these highly chlorinated contaminants

is needed. A modified type of prepared-bed land treatment that incorporates variable redox states (i.e., anaerobic and aerobic phases) is being evaluated in this

project. The first pilot-scale study, using PCP-contaminated soil from the American Wood Products site

in Lake City, Florida, is in progress at the U.S. Environmental Protection Agency’s Test and Evaluation

(T&E) Facility in Cincinnati, Ohio.

Methodology

The pilot-scale study is being conducted in soil pans that

simulate a prepared-bed land treatment unit with permeate collection. Each Plexiglas enclosure contains four

soil pans suspended in a controlled-temperature water

bath maintained at 20 ± 2°C. Each soil pan (13 in. x

13 in.) is loaded with contaminated soil to a depth of

8 in. over a graded gravel underdrain system, which is

separated from the soil layer by a coarse mesh stainless

steel screen. During the anaerobic phase, permeate

recycle is optional from the underdrain to the top of the

pan at flowrates of 5 to 20 mL/min. Anaerobic conditions

are maintained in the soil by flooding the pans with clean

creek water from the site. Aerobic conditions are produced and maintained in the soil by draining the water

and tilling the soil in a manner consistent with landfarming techniques.

Initial Anaerobic Phase

Based on a review of the literature, which is dominated

by studies evaluating spiked soils, three variables were

selected for study in the initial anaerobic phase: 1) permeate recycle vs. no recycle; 2) addition of a supplemental organic source (ethanol or anaerobic sewage

sludge); and 3) soil PCP contamination level. The experimental design was a three-factor analysis of variance with replication. The amounts of soil, ethanol,

anaerobic digestor sludge (32.8 g dry solids/L, 60 percent volatile solids), and site water initially added to each

of the 24 pans are summarized in Table 1. In situ oxidation-reduction potential (ORP) probes were placed in

one pan representative of each of the 12 treatments.

Four probes were buried in each pan, two in opposite

corners approximately 1 in. from the soil/gravel interface

and two in the remaining corners approximately 1 in.

from the soil/water interface near the top of the reactor.

68
Recycle flow rates were maintained at 8 to 10 mL/min.

The pH of the water flooding the pans was measured

in situ each week. The soil in the pans was sampled and

analyzed for PCP and its less chlorinated phenolic metabolites, hydrocarbons, and percent moisture on a

monthly basis. Hydrocarbons were present in the site

soil in significant quantities (1,000 to 2,500 mg/kg dry

soil) because diesel fuel serves as a carrier for PCP in

wood treating operations. The water flooding the pans

was analyzed for PCP and less chlorinated phenolic

metabolites each time a soil sample was collected.

Aerobic Phase

After 6 months, all of the pans except four of the sludgeamended pans (2D, 3A, 4B, and 5C) were converted to

aerobic conditions. After the free water was drained from

the soil, the soil was tilled three times a week for 4

weeks, until it dried to less than 10 percent total moisture. When the conversion phase was completed, some

of the pans were continuously supplied with air at a low

flow rate (in addition to weekly tilling of all pans) and/or

amended with poultry manure (see Table 1). Each

month, the soil in the pans is sampled and analyzed for

PCP and its less chlorinated phenolic metabolites, as

well as hydrocarbons. Moisture content and water addition volumes are measured weekly to maintain moisture

content in a constant range.

Results

Anaerobic conditions were established in the soil pans

after the first week. The measured ORPs ranged between -150 to -500 mV (versus Ag/AgCl reference electrode). No apparent correlation was found with respect

to probe depth, soil type, or treatments. The soil sample

PCP concentration data showed no significant amount

of PCP removal in any of the treatments after 6 months.

Changes in PCP concentration in the flood water of

several of the pans were noted, however. After 2 months,

the PCP concentration in the flood water of the soil pans

containing sludge dropped from 15 to 55 mg/L to less

than 0.5 mg/L. After 4 months, the PCP concentration

also dropped to less than 0.5 mg/L in the two replicate

pans with low-PCP soil treated without recycle or supplemental organic source. The less chlorinated phenolic

metabolites were not detected as intermediates in the

flood water from any of these pans. No degradation of

hydrocarbons was noted in any of the pans in the initial

anaerobic phase, but degradation has occurred in the

aerobic phase. The presence of hydrocarbons may have

interfered with the bioavailability of PCP in the initial

anaerobic phase. To test this hypothesis, aerobically

treated soil will be reconverted to anaerobic conditions

once the hydrocarbon concentration has been reduced.

Another possibility is that appropriate anaerobic PCP

degraders are not present in sufficient quantities in the

soil pans. A bench-scale study using soil from the

same source as the pilot-scale study has been initiated to investigate the effect of amendment with PCPacclimated culture.

Conclusions

Adaptation of the pilot-scale land-treatment units to anaerobic operation has been evaluated. Flooding the soil

with water successfully creates a low redox (anaerobic)

state. The in situ ORP probes constructed for the project

work well. Monthly sampling intervals and the analytical

techniques used adequately characterize system behavior. The tilling strategy used in the conversion from

anaerobic to aerobic operation was successful. The

presence of significant amounts of hydrocarbon cocontamination may have affected PCP degradation,

suggesting that a more appropriate treatment sequence

may be aerobic-anaerobic-aerobic. This observation reinforces the importance of technology evaluation with

soils characteristic of those found at actual sites.

69
Table 1. Soil Pan Operation Summary

Initial Anaerobic Phase Treatmentsb Aerobic Phase

Treatmentsc

Soil (kg as is)

Pana Lowd Highe Ethanol

(mL)

Sludge

(L)

Water

(L)

Recycle

(mL/min)

Air

(mL/min)

Manure

(g/pan)

6B 36 - - - 16 - - -

1C 36 - - - 16 - 10 750

2B 36 - - - 16 8-10 - 750

6Cf 36 - - - 16 8-10 10 -

1Af 36 - 17.2 - 16 - - -

3B 36 - 17.2 - 16 - 10 750

3C 36 - 17.2 - 16 8-10 - 750

2Cf 36 - 17.2 - 16 8-10 10 -

4C 36 - - 5.16 11 8-10g - -

5Cf 36 - - 5.16 11 - Not converted

4Bf 36 - - 5.16 11 8-10g Not converted

6D 36 - - 5.16 11 - 10 -

2A - 35 - - 16 - - -

1Df - 35 - - 16 - 10 750

5A - 35 - - 16 8-10 - 750

4Df - 35 - - 16 8-10 10 -

3Df - 35 17.2 - 16 - - -

5B - 35 17.2 - 18h - 10 750

6A - 35 17.2 - 16 8-10 - 750

5Df - 35 17.2 - 16 8-10 10 -

1B - 35 - 4.88 12 - - -

2Df - 35 - 4.88 12 - Not converted

3Af - 35 - 4.88 10.5h 8-10g Not converted

4A 35 - 4.88 12 8-10g 10

a

 Pan location for treatments (Pans A-D in Boxes 1-6) randomly assigned for statistical purposes.

b

 Initial anaerobic phase from August 11, 1994, to March 2, 1995.

c

 Aerobic phase from March 2, 1995, to July 7, 1995.

d

 Soil from 5 ft depth at site containing approximately 250 mg PCP per kg dry soil.

e

 Soil from 12 ft depth at site containing approximately 650 mg PCP per kg dry soil.

f

 In situ ORP probes added to pan during initial anaerobic phase.

g

 Recycle was set at 9 mL/min initially but was discontinued after the first week due to extremely low flow (less than 1 mL/min) through

sludge-amended soils. h

 Amount of water required to maintain a constant depth of 2 in. above soil surface varied somewhat.

70
Development of a Sulfate-Reducing Bioprocess To Remove Heavy Metals From

Contaminated Water and Soil

Munish Gupta and Makram T. Suidan

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Gregory D. Sayles and Carolyn M. Acheson

U.S. Environmental Protection Agency, Cincinnati, Ohio

Introduction

Acid mine drainage is characterized by low pH (1.5 to

3.5) and high concentrations of sulfate and dissolved

heavy metals. Bacterial sulfate reduction has been identified as a potentially cost-effective process for removing

metals from mine drainage (1, 2). Sulfate-reducing bacteria convert sulfate to sulfide using an organic carbon

source as the electron donor. The sulfide precipitates

the various metals present in the wastewater, yielding a

very low concentration of dissolved metals in the effluent. In this study, acetic acid was used as the carbon

source for two reasons: it is relatively inexpensive, and

being an acid, it can effectively leach out metals from

contaminated soils such as mine tailings.

Reactor Selection

To effectively treat metal-contaminated wastewater, a

reactor must establish an anaerobic environment to support sulfate reduction, resulting in metal precipitation as

metal sulfides, and must provide an efficient clarifier to

remove metal precipitates from the effluent. Because

sludge (metal precipitates and biomass solids) would

accumulate and eventually clog the reactor, ease of

sludge removal or cleaning is an important consideration in selecting a reactor. Two reactors were evaluated:

an upflow anaerobic filter packed with plastic Pall rings

and an anaerobic sludge blanket reactor. To clean the

reactor, sludge can be removed from the bottom of the

sludge blanket reactor. The same technique can be

used for the filter; however, it may be more difficult due

to packing material. The cleaning of the filter, therefore,

was an additional aspect of research.

Reactor Operation and Performance

Two filters (A and B) and one sludge blanket reactor

were operated at a temperature of 30°C and at a pH of

7.2, optimal pH for sulfate-reducing microorganisms.

The feed concentration of metals (shown in Table 1)

used in this study were among the highest concentrations observed at mines in Montana and Colorado.

After an initial acclimation, Filter A was fed the metals

listed in Table 1, while Filter B and the sludge blanket

reactor were fed iron at a concentration equal to the

sum of the molar concentrations of all the metals fed

to Reactor A. Table 1 shows characteristic effluent

Table 1. Influent and Effluent Concentrations (mg/L) for Filter A

Effluent Concentration

Influent

Constituent

Influent

Concentration Filtered Total

Iron 840 0.09 0.497

Zinc 650 0.14 0.310

Manganese 280 5.4 5.60 

Copper 130 0.02 0.022

Cadmium 2.3 0.02 0.019

Lead 2.1 0.005 0.005

Arsenic 1.5 0.01 0.01 

Acetate 3000 11.0 —

Sulfate 5000 800 —

Total sulfide 0 31 —

71
concentrations for Filter A, operating at a hydraulic

detention time of 5 days.

During this period, the level of sludge rose above the

packed bed in Filter A. To investigate whether sludge

withdrawal from the bottom would control the sludge

height in the filter, 1 L of sludge was removed. Sludge

withdrawal lowered the sludge level and the filter continued to operate efficiently, with less than 1 percent

change in effluent conditions. A similar situation in Filter

B was also corrected in the same fashion. Sludge withdrawal from the bottom can, therefore, control the accumulation of sludge and prevent clogging of the filters.

The sludge blanket reactor did not perform very well as

a clarifier. Although the soluble iron concentration in the

effluent was less than 0.25 mg/L, the total concentration

was as high as 25 mg/L and varied between 18 and

22 mg/L. This reactor had a high concentration of total

suspended solids in the effluent compared with Reactors A and B. It was concluded that this type of reactor

was not effective in clarification and was unable to meet

the requirements. Therefore, the operation of the sludge

blanket reactor was discontinued.

Conclusions

Compared with an anaerobic sludge blanket reactor, an

upflow anaerobic filter packed with Pall rings was found

to be a very efficient reactor for the treatment of water

contaminated with heavy metals. The filter, unlike the

sludge blanket reactor, worked very well as a clarifier,

and all metals except manganese were reduced to a

concentration close to drinking-level standards. Sludge

withdrawal from the bottom of the filter can be used to

remove accumulating sludge, and, therefore, the filter

can be operated continuously. Ongoing work will evaluate the performance of the filters as a function of hydraulic retention time, lower temperatures, and pH. Sludge

removal frequency will also be optimized.

References

1. Dvorak, D.H., R.S. Hedin, H.M. Edenborn, and P.E. McIntire. 1992.

Treatment of metal-contaminated water using sulfate reduction:

Results from pilot-scale reactors. Biotech. & Bioeng. 40:609-616.

2. Kuyucak, N., D. Lyew, P. St. Germain, and K.G. Wheeland. 1991.

In situ bacterial treatment of AMD in open pits. Presented at the

Second International Conference on the Abatement of Acidic

Drainage, Montreal, Canada, September 16-18.

72
Development of Techniques for the Bioremediation of Chromium-Contaminated

Soil and Ground Water

Michael J. McInerney, Nydia Leon, Veronica E. Worrell, and John D. Coates

University of Oklahoma, Norman, Oklahoma

The potential for biotic Cr(VI) reduction in samples from

a Cr(VI)-contaminated aquifer (Elizabeth City, North

Carolina) was evaluated by inoculating aquifer material

into anaerobically prepared mineral salts medium that

did not contain chemical reductant. In inoculated microcosms, the Cr(VI) concentration decreased after 5 days

incubation at 25°C, and almost all of the Cr(VI) was gone

after 25 days (Figure 1). Little or no change in Cr(VI)

concentration was observed in uninoculated controls, or

in sterile controls prepared by autoclaving, boiling, or the

addition of HgCl2 or chloramphenicol. Hydrogen reduced the lag time before Cr(VI) reduction occurred but

did not markedly affect the rate of Cr(VI) reduction (Figure 2). The addition of other exogenous electron donors

such as glucose, acetate, formate, or benzoate did not

affect the rate or lag time associated with Cr(VI) reduction in microcosms compared with controls that lacked

an exogenous electron donor. The addition of phenol,

lactate, and ethanol to microcosms inhibited Cr(VI) reduction. Subsequent addition of Cr(VI) to microcosms

with benzoate as the electron donor decreased the lag

time and increased the rate of Cr(VI) reduced compared

with that observed initially.

The effect of sodium sulfate, sodium nitrate, amorphous

ferric hydroxide (each at 10 mM) on Cr(VI) reduction

was tested with benzoate-amended microcosms. The

presence of sulfate and nitrate inhibited the reduction of

Cr(VI) compared with microcosms that did not receive

any of the three additional electron acceptors. Sulfide

levels remained unchanged during the course of the

experiments. In bottles with nitrate, nitrite accumulated after 8 days and decreased after 16 days. Ferrichydroxide-supplemented microcosms reduced Cr(VI)

to a much greater extent than unsupplemented control

cultures; ferrous iron production coincided with Cr(VI)

reduction.

Two facultative bacteria that can reduce Cr(VI) were

isolated from Elizabeth City aquifer material, and one

bacteria was isolated from an aquifer underlying a landfill in Norman, Oklahoma, using a mineral salts medium

with 5 mM benzoate, 500 μM Cr(VI). All three isolates

are gram-negative, motile rods that grow singly, in pairs,

and in branched chains. On agar medium, the isolates

formed shiny, smooth, pink colonies and produced a

diffusible green pigment. Upon initial isolation, Cr(VI)

was rapidly reduced (Figure 3); however, the rate and

Figure 1. Biological reduction of Cr(VI) with aquifer material.

Figure 2. Effect of electron donors on Cr(VI) reduction with

aquifer material. 

73
extent of Cr(VI) reduction decreased with repeated

transfer of the culture in benzoate-Cr(VI) medium. The

use of Cr(VI) was dependent on the presence of an

electron donor and an active inoculum. In addition to

benzoate, other substrates supported Cr(VI) reduction

(Table 1). Increases in cell numbers were observed

when the electron donor and Cr(VI) were both present. In the absence of Cr(VI) or electron donor, little

or no increase in cell number was observed: less than

6 x 105

 cells/mL.

Four observations supported the conclusion that the

decrease in Cr(VI) concentration was a biologically mediated reduction process: 1) Cr(VI) concentrations decreased faster and to a greater extent in nonsterile

versus sterile microcosms; 2) phenol, ethanol, and lactate inhibited Cr(VI) reduction in microcosms; 3) repeated additions of Cr(VI) to microcosms decreased the

lag time and stimulated the rate of Cr(VI) reduction; 4)

bacteria were isolated and capable of using Cr(VI) as an

electron acceptor. Iron hydroxide stimulated Cr(VI) reduction in microcosms, most likely by an indirect mechanism involving the production of ferrous iron. The extent

and rate of Cr(VI) reduction by aquifer microcosms was

not affected when exogenous electron donors, with the

exception of hydrogen, were added. This indicates that

the aquifer material had sufficient levels of endogenous

electron donors to support Cr(VI) reduction.

Figure 3. Cr(VI) reduction by strain NLB.

Table 1. Electron Donors That Support Cr(VI) Reduction by

Strain NLB

Additions

Cr(VI)

Reduced

(μM)

Increase in Cell

Numbers

([cells/mL] x 106)

Fumarate 414 0 

Purine/Pyrimidine mixa 337 31.8

p-Toluic acid 184 30.1

Lactate 169 62.8

Phenoxyacetate 166 56.3

Malate 80 3.8

Benzoate 68 142.9 

Phenol 59 19.3

Ethanol 32 7.8

No addition 0 0 

a

 Mix contains 0.5 mM each: adenine, guanine, thymine, uracil.

74
Bioremediation of Chlorinated Pesticide-Contaminated Sites Using Compost

James C. Young, Jean-Marc Bollag, and Raymond W. Regan

Pennsylvania State University, University Park, Pennsylvania

Many sites throughout the United States are contaminated with chlorinated pesticides. Of particular interest

to this project are those sites contaminated with chlordane and toxaphene. One objective is to determine the

feasibility of using compost as a culture medium for

mediating the biodegradation of these pesticides. A second objective is to determine major pathways of chlordane and toxaphene biodegradation that lead to

mineralization. These objectives are particularly challenging because chlordane and toxaphene each consist

of several chlorinated cyclic hydrocarbons that individually may follow different biodegradation pathways, or

may be only partially dechlorinated.

Biodegradation of chlordane and toxaphene and other

chlorinated pesticides is expected to require an organic

co-substrate as a carbon source for the growth of acclimated microorganisms that enzymatically are capable

of dechlorinating the pesticides through reductive or

oxidative reactions. Co-substrates considered for use in

field applications include milk solids, sugar, blood meal,

sewage solids, methane, or, in the current project,

compost.

The test program includes the development and operation of a pilot-scale compost reactor that contains

a mixture of 10 percent municipal-sludge compost,

10 percent spent-mushroom compost, 40 percent

grass, and 40 percent alfalfa hay to provide an environment suitable for the culture of chlordane- and toxaphenedegrading microorganisms. This compost is used to amend

various contaminated-soil matrices followed by analysis of

the fate of the pesticide. Residual pesticides are monitored using gas chromatography, thin layer chromatography, and mass spectroscopy. Test parameters include

soil type, compost-soil ratio, moisture level, oxidationreduction potential, pH, presence of sulfates and nitrates, and the effect of supplemental soluble organic

co-substrates.

75
Reductive Electrolytic Dechlorination

John W. Norton, Jr., and Makram T. Suidan

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Carolyn M. Acheson and Albert D. Venosa

National Risk Management Engineering Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Reductive dehalogenation is the only known mechanism

to biologically degrade some highly chlorinated organic

compounds, including pentachlorophenol (PCP) (1),

and occurs primarily in anaerobic environments (2). A

biofilm-electrode reactor (BER) was constructed to

evaluate PCP dechlorination as a function of ethanol

concentration and the presence of an electrical current.

The BER was operated as follows: current, 20 mA;

hydraulic retention time, 0.38 days; PCP feed, 5 mg/L;

ethanol feed range, 0 to 100 mg/L. The best observed

dechlorination occurred when 5 mg/L PCP and 25 mg/L

ethanol were fed to the reactor. The effluent under these

conditions contained 0.013 mg/L PCP, 0.26 percent of

the feed concentration. At lower ethanol levels, PCP was

not as effectively dechlorinated. The trichlorophenols (TCP)

and dichlorophenols (DCP) displayed a two- to threefold increase in effluent concentration as the substrate

ethanol was decreased, particularly at concentrations less

than 10 mg/L. The monochlorophenols (MCP), however,

reached a maximum of 0.014 mM at ethanol concentrations of 10 to 25 mg/L. The total dechlorination decreased significantly when the ethanol was removed

from the feed, indicating that the ethanol stabilized

dechlorination.

After characterizing the ethanol requirements in the system, the role of the current in the dehalogenation of the

PCP was evaluated by turning off the current. Electrical

current was shown to play a necessary role in dechlorination, although it is unknown whether this role was the

result of the hydrogen generation or the low reducing

potential surface formed on the cathode of the anodecathode cell. Two trials of reactor operation without current were conducted. Each current removal trial caused

a reduction in PCP dehalogenation, demonstrated by

the successive appearance of the higher chlorinated

phenols in the effluent. The two trials displayed very

different temporal behavior, however. The first removal

resulted in a quick rise in effluent PCP concentrations,

increasing one order of magnitude in a few hours. Following the second current removal, a much slower appearance of chlorinated phenols was observed; the

effluent PCP concentration increased one order of magnitude in approximately 6 days. During each trial, after

the current was reapplied, the system recovered. Each

trial showed a recovery pattern similar to the failure

preceding it, the first trial showing a quick recovery and

the second trial showing a much slower recovery. The

causes of the different behaviors have not been characterized. We are presently evaluating the role of electrical

current by varying the current while keeping the ethanol

concentration constant at 10 mg/L.

References

1. Mohn, W., and J. Tiedje. 1992. Microbial reductive dehalogenation.

Microb. Rev. 56:482-507.

2. Suflita, J., A. Horowitz, D. Shelton, and J. Tiedje. 1982. Dehalogenation: A novel pathway for the anaerobic biodegradation of

haloaromatic compounds. Science 218:1,115-1,117.

76
Biological Ex Situ Treatment of PAH-Contaminated Soil

Carl L. Potter

U.S. Environmental Protection Agency, Cincinnati, Ohio

Roy C. Haught

IT Corporation, Cincinnati, Ohio

The goal of this project is to evaluate the potential of

biological ex situ soil treatment systems (biopiles) to

remediate soils contaminated with hazardous chemicals. A laminar-type flow pilot-scale reactor with a volume of 3 yd3

 has been constructed at the U.S.

Environmental Protection Agency’s Test & Evaluation

(T&E) Facility in Cincinnati, Ohio. Laminar-type flow

from one side of the reactor to the other may provide

even aeration to all areas of the reactor while avoiding

the use of pipes inside the reactor. This design greatly

facilitates loading and unloading of the reactor and is

readily scalable to larger systems.

Passing smoke through the reactor for visual observation of flow indicated uniform, laminar-type flow through

the empty reactor. Further testing involved filling the

reactor with vermiculite or a synthetic soil, flushing with

argon, and then passing air through the reactor to evaluate air flow through this uniform solid matrix. Oxygen

probes, located at 27 positions within the reactor, indicated rapid and uniform air saturation of the system.

Analysis of gas flow through an empty reactor and

through uniform matrices allowed evaluation of reactor

performance without confounding effects of soil inhomogeneities that may lead to nonuniform aeration of the

reactor space.

The reactor uses pulsed air flow through the pile to

permit maximum distribution of air within the soil. Air is

driven into the soil during pulse action, then allowed to

diffuse in all directions during the rest interval.

Soil contaminated with polynuclear aromatic hydrocarbons (PAHs) from the Reilly Tar Pit Superfund site in St.

Louis Park, Minnesota, has been brought to the T&E

Facility for research on soil aeration and effectiveness

of this ex situ reactor design for biological treatment of

contaminated soils. Micronutrients were adjusted to

100:20:1 phosphorus:carbon:nitrogen, and 0.5 percent

by weight cow manure was added to the soil. A 10-week

treatability study is under way to evaluate disappearance of parent PAHs and microorganism population

changes in this reactor system.

77
Effectiveness of Gas-Phase Bioremediation Stimulating Agents (BSAs) for

Unsaturated Zone In Situ Bioremediation

James G. Uber, Ronghui Liang, and R. Scott Smith

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Paul T. McCauley

Water and Hazardous Waste Treatment Research Division, National Risk Management Research

Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio

Background

Successful in situ bioremediation in the unsaturated

zone requires that water, oxygen, nutrients, primary substrate, and perhaps co-metabolites be available to the

microorganism via physical transport mechanisms. Any

of these substances may be called a bioremediation

stimulating agent (BSA), given that a shortage of any

one may adversely affect the performance of an in situ

bioremediation system. Other potential BSAs include

substances (e.g., surfactants) that are not ordinarily required for microbial growth but that may enhance substrate or nutrient bioavailability.

Much work has focused on engineering approaches to

deliver BSAs at field scale. Little research has been

conducted, however, to evaluate which in situ delivery

approaches are best for transporting BSAs to microorganisms. Given the complexity of two-phase (gas/water)

or three-phase (gas/water/nonaqueous-phase liquid)

fluid and contaminant transport in the unsaturated zone,

considerable uncertainty exists about the ultimate distribution of BSAs in contaminated soils. Further, microbial

growth processes affect fluid and contaminant transport

not only through biochemical reactions but also through

a spatial-temporal influence on fluid permeability. (Plugging of pore spaces by microorganisms can reduce

wetting fluid permeability by greater than 99 percent.)

The present study will identify in situ BSA delivery strategies that are most likely to achieve a uniform BSA spatial

distribution and, therefore, most likely to improve bioremediation field performance. As a byproduct of this

work, the project aims to identify and measure the fundamental physical and microbial processes that affect

bioremediation performance enhancement through BSA

delivery methods.

Because of strong capillary forces that affect the distribution and movement of wetting fluids in unsaturated

soils, gas-phase BSAs are more likely to achieve uniform in situ spatial distribution. It is, in fact, well known

that movement of water in the unsaturated zone often

occurs in discrete fingers that occupy a small fraction of

the total pore space. Relatively little is known, however,

about the characteristics of in situ gas-phase BSA transport, including physical factors that may lead to complex

and undesirable flow patterns (e.g., interactions of water

saturation and air permeability), chemical transport factors that may limit gas-phase BSA spatial distribution

(e.g., BSA solubility), and dynamic microbial factors that

may affect BSA transport in the field (e.g., microbial BSA

utilization rates and plugging that leads to heterogeneous effects on air/water permeability). Because of the

promise of gas-phase BSAs and the significant unknowns regarding their effectiveness, this project will

focus on effective gas-phase addition of nutrients, cometabolites, oxygen, and moisture.

Objectives

The objectives of this work are to:

• Evaluate the effectiveness of field systems for gasphase delivery of BSAs to the unsaturated zone for

enhancing in situ bioremediation performance. These

BSAs include nutrients (organophosphates), cometabolites, surfactants or solvents, and water vapor.

• Identify and measure the physical and microbial factors affecting the bioavailability of gas-phase BSAs

in the unsaturated zone, including uneven spatial distribution of BSAs at the pore- and core-scales and

complex changes in unsaturated zone air permeability caused by microbial-growth dynamics.

78
• Develop visible light tomography (VLT) systems that

allow visualization of in situ unsaturated zone physical and microbial processes for controlled evaluation

of alternative BSAs and delivery systems.

• Use the VLT systems to evaluate alternative BSAs

for remediation of aged contaminated soils in controlled but realistic environments.

Accomplishments

We have designed and constructed two different laboratory systems for observing dynamic fluid distribution

in the unsaturated zone under simulated BSA delivery.

A three-dimensional column system has been designed

to collect data on fluid migration through discrete fingers

in disturbed and undisturbed soil cores, and will be used

to measure the limitations on BSA delivery caused

by fingering under a variety of soil conditions and

fluid application rates. This work will be completed

by December 1995. The columns are 30 cm in diameter

and comprised of stackable 10-cm sections separated

by 1-mm spacers. The columns rest on a base that

allows manipulation of the bottom pressure boundary

condition, and the side boundary condition is manipulated through 1-mm gaps between rings. Water is applied uniformly to the top surface of the soil columns via

a carefully designed air-atomizing nozzle. After the fluid

flow is developed, a dye mixture marks the locations of

any preferential flow pathways. The pathways will be

exposed at the surfaces of each 10-cm ring, and the

complete three-dimensional character of each pathway

will be recorded. Different color dyes will be used to

investigate the persistence of individual fingers when

fluid application is cycled, allowing the soil to drain to

varying water contents between application.

A two-dimensional, vertical, thin-slab visible light tomography (VLT) system has been designed to visualize

and measure the interactions between gas-phase BSA

and liquid-phase flow in a controlled environment. The

system will also serve as a bioremediation simulator to

measure the effectiveness of various gas-phase BSAs,

and to visualize dynamic microbial-growth processes

under simulated in situ bioremediation conditions (with

and without BSA addition). The vertically oriented chamber dimensions are 1 m x 2 m x 1 cm. The top boundary

will be either open or closed to the atmosphere, and will

be capable of having controlled amounts of liquid added

uniformly over the slab length. Side boundaries will be

either closed or open to the atmosphere, thereby providing the ability to control gas-phase BSA injection or

extraction (simulating the operation of BSA injection or

extraction wells). The bottom boundary will be either

open to the atmosphere or, via a manifold, will simulate

water table conditions.

The advantage of the thin-slab system is the ability to

visualize the complex flow and microbial processes occurring in the unsaturated zone under simulated in situ

conditions. A bank of high frequency fluorescent lamps

will illuminate the system from the back. Because light

transmission is related to water saturation, the water

distribution can be easily visualized without the use of

dyes. Fluorescent gases will be investigated for visualization of the gas movement, as will color-marked gasand liquid-phase pH indicator solutions. Data will be

recorded via a CCD camera and a data acquisition

system so that actual fluid flow and microbial processes

can be recorded and visualized. The CCD camera system will collect data at a spatial resolution on the order

of the pore scale (approximately 0.5 to 1.0 mm).

79
Section 5

Process Research

Process research involves isolating and identifying microorganisms that carry out biodegradation

processes and the environmental factors affecting these processes. It also deals with the development of techniques for modeling and monitoring biodegradation. Through this research, scientists

establish the building blocks of new biosystems for treatment of pollutants in surface waters,

sediments, soils, and subsurface materials. Thorough evaluation is critical at this level of research

because a firm scientific foundation can facilitate the scaling up of a promising technology. Process

research is being conducted on many environmental pollutants.

Two research projects quantified the extent of biodegradation of organic compounds using carbon

isotopes. The methods employed in this research can be used both in the laboratory, where target

compounds are isotopically labeled, and in the field, where contaminants contain different concentrations of isotopes than the surrounding environment. The methods were applied in the measurement of petroleum degradation and the demethylation of organometallic compounds.

Another project determined the kinetic rate constants of anaerobic degradation of quinoline by

methanogenic bacteria. This system is potentially of great practical significance because quinoline

is found in wastes at wood preserving plants.

Several papers focused on the use of microorganisms to degrade alkyl halides and polychlorinated

aromatics. Research included the characterization of a bacterial enzyme (toluene 2-monooxygenase) that can degrade trichloroethylene and of a strain of Pseudomonas cepacia G4 that

expresses the enzyme constitutively. Further research covered the effects of varying environmental

conditions on the bioremediation of chlorinated compounds.

Other process research projects studied heavy metal inhibition of the bioremediation of polychlorinated aromatics and the effects of different primary substrates on the reduction of 2,4-dinitrotoluene.

Primary substrates are used to sustain the growth of microorganisms when the target contaminant

cannot act as a food source.

Several poster presentations at the symposium involved process research. The presentations

covered methods for monitoring bioremediation, the use of surfactants in sediments, and characterization of an anaerobic dehalogenating microorganism. Another presentation dealt with the

potential of Mycobacterium to mineralize polycyclic aromatic hydrocarbons.

81
Monitoring Crude Oil Mineralization in Salt Marshes: 

Use of Stable Carbon Isotope Ratios

Andrew W. Jackson

Department of Civil and Environmental Engineering, Louisiana State University, 

Baton Rouge, Louisiana

John H. Pardue

Wetland Biogeochemistry Institute, Louisiana State University, Baton Rouge, Louisiana

Introduction

The ability to monitor mineralization of hydrocarbons is

of prime importance in a successful remediation strategy. Hydrocarbon mineralization must be ensured, because hydrocarbons can be sorbed, transformed, or

buried, or otherwise be undetected but still pose threats

to the existing system ecology. One successful technique has been monitoring changes in oil composition

relative to a stable, nondegradable compound (1, 2).

Two disadvantages to this method exist, however: 1) its

inability to demonstrate mineralization instead of transformation, and 2) its inability to measure absolute oil

degradation, because only “resolved” compounds are

quantified.

A promising new technique for the detection and quantification of hydrocarbon mineralization is the use of

stable carbon isotope ratios (3). Carbon dioxide gas

ratios vary from of 12C to 13C depending on the source

of the gas. Crude oils are more depleted in 13C, and thus

the mineralization of oil produces CO2 with lower δ13C

values. Oil has a δ13C value of -29 to -32 (0/00) depending on the source of the oil. Salt marshes are predominantly colonized by C3 plants, and CO2 evolved from

these soils has δ13C ratios of -14.4 to -17.7 (0/00) (4).

If biodegradation is occurring in a contaminated salt

marsh, the δ13C value of the produced CO2 should

decrease due to the presence of 13C-depleted CO2 from

the crude oil. If this occurs, it would be possible to qualify

and quantify hydrocarbon degradation by measuring

total CO2 production and changes in the 13C signature

of CO2 produced from the marsh.

Theoretical

The rate of CO2 produced from each carbon source can

be easily computed using three equations describing

CO2 production and the δ13C signature:

Ro + Ri = Rt (Eq. 1)

Ro

Rt

+

Ri

Rt

= 1 (Eq. 2)

Ro

Rt

(So) +

Ri

Rt

(Si) = St

(Eq. 3)

where Ro and Ri

 are the rates of CO2 production from

the crude oil and indigenous carbon sources respectively, and Rt is the total rate of CO2 production. So and

Si

 are the δ13C signatures of the crude oil and indigenous carbon, and St is the measured δ13C signature of

the produced CO2. So,Si

, St, and Rt are experimentally

determined. Ro and Rm can then be determined from

Equations 2 and 3. This assumes that CO2 is generated

from only these two carbon pools and that there is no

addition of atmospheric CO2.

Results and Discussion

Kinetic Experiments

Microcosm studies showed rapid and nearly complete

(greater than 90 percent reduction in the hopane ratio)

degradation of parent alkanes in the fertilized treatments

under completely mixed, aerated conditions. In the unfertilized treatment, less than a 10 percent reduction was

observed in the hopane ratio of the alkanes (Figure 1A).

Polycyclic aromatic hydrocarbon (PAH) degradation

83
was evident in both unfertilized and fertilized treatments

(Figure 1B).

A definite decrease in the δ13C signature of evolved

CO2 was observed in both treatments contaminated with

oil (Figure 2). The measured δ13C signature of CO2

evolved from noncontaminated marsh soil is 17.10/00±

0.2. The fertilized treatment approaches the δ13C ratio

of pure oil (29.10/00) and varies between -27 and -290/00.

The nonfertilized treatment varies between -25 and

-270/00.

CO2 production was correspondingly greater in the fertilized treatment. The average rate of total CO2 production for the nonoiled-fertilized, oiled-unfertilized, and

oiled-fertilized microcosms was 0.174, 0.396, 1.86 mg

CO2-C/day-gram soil, respectively. The large increase in

CO2 production between the oiled and the nonoiled

treatments suggests that the increase in CO2 production

is from the mineralization of the crude oil and not from

indigenous carbon sources.

The amount of CO2-C produced from crude oil can be

computed, using the δ13C signatures, CO2 production

data, and the isotope dilution equations. The predictive

ability of these equations is supported by the similarity

between the pseudo first-order rate constants in the

reduction of hopane ratios and the calculated CO2 production from the crude oil carbon pool, 0.082 and 0.087

day-1, respectively.

Field Experiment

A small-scale field experiment was conducted to verify

the ability to detect δ13C signature changes in CO2

evolved in situ from oil-contaminated soils. Significant

decreases in the δ13C signature of evolved CO2 was

detected 5 weeks after oiling in fertilized and unfertilized

treatments. The alkane- and PAH-hopane ratios decreased for all treatments. The C1 and C2-phenanFigure 1. (A) Total (C15-C44) alkane-hopane ratio versus time

in aerated microcosms. (B) Total PAH (phenanthrene,

C1, C2; naphthalene, C1, C2) hopane ration versus

time.

Figure 2. Delta-13C ratios versus time in aerated microcosms.

Figure 3. Rates of CO2-C mineralized from crude oil in fertilized

and unfertilized salt marsh soils (calculated from isotope dilution equations).

84
threne, however, appear to be stable over the period of

this experiment.

The CO2 production rates and the δ13C ratios measured

were used to calculate the CO2 produced from crude oil

(Figure 3). No mineralization of crude oil was detected

until Week 2, and the majority of mineralization appears

to begin at Week 5. The fertilized treatments appeared

to show higher mineralization rates before the unfertilized and to mineralize at a more even prolonged rate.

The unfertilized treatments have a more intense rate

of mineralization but for only one sampling date.

Amendments of fertilizer inconclusively increased degradation, as evidenced by hopane ratios of specific oil

components.

The importance of the δ13C data is the data’s ability to

calculate mineralization rates directly. They measure the

final product, while monitoring hopane ratios only measures the disappearance of the parent compound, not

mineralization. These experiments support the ability to

use δ13C ratios in conjunction with CO2 production to

qualitatively and quantitatively monitor crude oil degradation.

References

1. Bragg, J.R., R.C. Prince, E.J. Harner, and R.M. Atlas. 1994. Nature

368:413-418.

2. Bragg, J.R., R.C. Prince, E.J. Harner, and R.M. Atlas. 1993. In:

Proceedings of the 1993 International Oil Spill Conference. pp.

435-447.

3. Aggarwal, P.K., and R.E. Hinchee. 1991. Environ. Sci. Technol.

25:1,178-1,180.

4. Chmura, G.L., R.A. Socki, and R. Abernethy. 1987. Oecologia

74:264-271.

85
Mercury and Arsenic Biotransformation

Ronald S. Oremland

U.S. Geological Survey, Menlo Park, California

This presentation will cover our recent findings with

regard to bacterial processes affecting 1) methylmercury demethylation and 2) the dissimilatory reduction of

arsenic (As) (V).

Methylmercury Oxidative Degradation

Potentials in Contaminated and Pristine

Sediments of the Carson River, Nevada

Sediments from mercury-contaminated and uncontaminated reaches of the Carson River, Nevada, were

assayed for sulfate-reduction, methanogenesis, denitrification, and monomethylmercury (MeHg) degradation.

Demethylation of 14C-MeHg was detected at all sites, as

indicated by the formation of 14CO2 and 14CH4. Oxidative

demethylation was indicated by the formation of 14CO2

and was present at significant levels in all samples.

Oxidized/Reduced demethylation product (ORDP) ratios (e.g., 14CO2/

14CH4) generally ranged from 4.0 in

surface layers to as low as 0.5 at depth. Production of

14CO2 was most pronounced at sediments surfaces that

were zones of active denitrification and sulfate-reduction, but was also significant within zones of methanogenesis. In a core taken from an uncontaminated site

having more oxidized, coarse-grained sediments, sulfate-reduction and methanogenic activities were very

low, and 14CO2 accounted for 98 percent of the product

formed from 14C-MeHg. No relationship was apparent

between the degree of mercury contamination of the

sediments and the occurrence of oxidative demethylation. Sediments from Fort Churchill, the most contaminated site, however, were most active in terms of

demethylation potentials. Inhibition of sulfate reduction

with molybdate resulted in significantly depressed

ORDP ratios, but overall demethylation rates were comparable between inhibited and uninhibited samples.

Addition of sulfate to sediment slurries stimulated

production of 14CO2 from 14C-MeHg, while 2-bromoethane- sulfonic acid blocked production of 14CH4.

These results reveal the importance of sulfate-reducing

and methanogenic bacteria in oxidative demethylation

of MeHg in anoxic environments.

The Dissimilatory Reduction of As(V) to

As(III) in Anoxic Sediments and as an

Electron Acceptor for Growth of Strain

SES-3

Anoxic sediment slurries amended with millimolar levels

of As(V) achieved a complete reduction of this oxyanion

to As(III) upon incubation. As reduction was enhanced

when slurries were provided with the electron donors H2,

lactate, or glucose, although no effect was achieved with

acetate or succinate. Aerobically incubated slurries did

not reduce As(V), nor did formalin-killed or autoclaved

controls. Even though acetate did not stimulate As reduction, the oxidation of 2-14C-acetate to 14CO2 in anoxic slurries could be coupled with the abundance of

As(V). The selenium (Se) (VI) respiring anaerobe strain

SES-3 was found to be capable of achieving growth by

carrying out the dissimilatory reduction of As(V) to

As(III). Although growth parameters were meager (e.g.,

Ym = 0.53 g cells/mole lactate; maximal cell density =

9.2 x 107

 cells/mL), the ability to reduce As(V) to As(III)

was constitutive and occurred rapidly in either selenateor nitrate-grown cells. These results suggest that the

reduction of As(V) to As(III) in nature may be achieved

by bacteria-like strain SES-3 carrying out dissimilatory

As(V) reduction. 

86
Monod Degradation Kinetics of Quinoline in Natural and Microbially Enriched

Methanogenic Microcosms

E. Michael Godsy, Ean Warren, and Barbara A. Bekins

U.S. Geological Survey, Menlo Park, California

Introduction

The methanogenic biodegradation of nitrogen-containing heterocyclic compounds found in wastes from petroleum refineries, coke operations, coal tar production,

and wood preservation has not been studied in detail.

Quinoline, the largest single component in creosote (1),

is first oxidized to 2(1H)-quinolinone, which is then degraded to CH4 and CO2. In this study, the Monod nogrowth kinetic constants for the oxidation of quinoline

and the Monod kinetic constants for the methanogenesis of 2(1H)-quinolinone were determined under natural

and microbially enriched methanogenic conditions using

nonlinear regression analysis (2). In microcosms simulating natural aquifer conditions, it was necessary to

model the oxidation and subsequent methanogenesis

independently (1). In microbially enriched microcosms,

the two must be coupled to include the biomass increase from the methanogenesis of 2(1H)-quinolinone

as below:

dQ

dt = − knXa Q

Kn + Q

dQn

dt = knXa Q

Kn + Q − μmaxXa Qn

Y(Ks + Qn)

dX

dt = μmaxXa Qn

Ks + Qn

where

Q = quinoline, mg⋅L-1

kn = no-growth oxidation constant, day-1

Xa = active biomass, mg⋅L-1

Kn = one-half saturation no-growth coefficient, 

mg⋅L-1

μmax = maximum specific growth rate, mg⋅L-1⋅day-1

Qn = 2(1H)-quinolinone, mg⋅L-1

Ks = one-half saturation coefficient, mg⋅L-1

Y = growth yield, mg biomass⋅mg substrate-1

Materials and Methods

The study site is located adjacent to an abandoned

wood preserving plant within the city limits of Pensacola,

Florida (1). The wood preserving process consisted of

steam pressure treatment of pine poles with creosote

and/or pentachlorophenol. For more than 80 years,

large but unknown quantities of waste waters (consisting of extracted moisture from the poles, cellular debris,

creosote, pentachlorophenol, and diesel fuel from the

treatment processes) were discharged to nearby surface impoundments. These impoundments were unlined

and in direct hydraulic contact with the underlying sandand-gravel aquifer. Contamination of the ground water

resulted from the accretion of wastes from these impoundments. Methanogenesis in the aquifer was simulated using microcosms containing approximately 3 kg

of freshly collected anaerobic aquifer material in a 4-L

glass serum bottle with 2.5 L of prereduced anaerobically sterilized mineral salts solution. Approximately

40 mg/L of quinoline was added, simulating a concentration similar to that found in the aquifer (1). The microcosms were prepared, incubated, and sampled in an

anaerobic glove box containing an O2-free atmosphere

maintained at 22°C to 24°C. Microbially enriched microcosms were prepared as above but were batch fed for

three cycles by removing 50 percent of the liquid volume

and replacing that volume with fresh mineral salts containing enough quinoline to bring the final concentration

back to 40 mg/L. After the last feeding cycle, the liquid

culture was removed from the sediment, resulting in

liquid-only culture which was batch fed for six more

cycles.

Substrate concentrations were determined at approximately 4-day intervals by high-performance liquid

chromatography. Total biomass concentrations were

determined at approximately 20-day intervals by total

protein using the Coomassie brilliant blue staining procedure of Gälli (3).

87
Results and Conclusions

The oxidation of quinoline, a reaction that is endergonic

(4), is modeled using derived Monod no-growth kinetic

constants. This oxidation is uncoupled with the degradation of 2(1H)-quinolinone in microcosms simulating

natural conditions, as shown by the complete and

stoichiometric oxidation to 2(1H)-quinolinone before the

onset of methanogenesis. In microbially enriched microcosms, however, the oxidation of quinoline is linked to

the degradation of 2(1H)-quinolinone; the increase in

biomass from methanogenesis must be included in the

equations describing the oxidation of quinoline (Figure

1). The kinetic values derived from the microbially enriched, all-liquid microcosm experiments were not significantly different from those values from sand-filled

natural microcosms (Table 1). The Monod kinetic constants for both the oxidation and subsequent methanogenesis are representative of values describing

substrate utilization in an oligotrophic and somewhat

hostile environment (4). 

It is still unclear, however, what number of microbial

populations are involved and to what extent each of the

populations influences the steps in the biodegradation

of quinoline. This uncertainty can be seen by the high

concentration of biomass capable of the oxidation of

quinoline in natural microcosms, suggesting that the

ability to oxidize quinoline is not unique to just this

consortium but may be common to many of the individual members of the creosote-degrading consortia. The

enrichment procedure has altered the microbial population of the natural microcosms by potentially removing

all of the microorganisms that can oxidize quinoline but

are not directly involved in the methanogenesis of 2(1H)-

quinolinone. These results suggest that as long as the

culture is derived from the contaminated aquifer, enrichment does not alter the kinetics of quinoline oxidation

and subsequent methanogenesis of 2(1H)-quinolinone.

The size of the various active microbial populations,

however, must be known before fate-and-transport modeling can be attempted.

References

1 . G odsy, E .M., D.F. G oe rlitz, and D. G rb iŒ-G aliŒ. 1 992. Anae rob ic

b iode gradation of cre osote contam inants in natural and sim ulate d

grou nd w ate r e cosyste m s. G round W ate r 30: 232-242.

2. Monod, J. 1 949. T he grow th of bacte rial cultu re s. Ann. R ev. Microb iol. 3: 371 -394.

3. G älli, R . 1 987. Biode gradation of dichlorom e thane in w aste w ate r

u sing fluidize d be d b iore actor. Appl. Microbiol. Biote chnol. 27: 206-

21 3.

4. G odsy, E .M. 1 993. Me thanoge nic biode gradation of cre osote -derive d contam inants in natural and sim ulate d grou nd w ate r e cosyste m s. P h.D. disse rtation. Stanford U nive rsity, Stanford, C A. p. 1 55.

Figure 1. Quinoline oxidation and 2(1H)-quinolinone methanogenesis in microbially enriched laboratory microcosms. 

Table 1. Monod Kinetic Constants %95 Percent Confidence

Interval for Parameters Determined by Nonlinear

Regression for Both Natural and Microbially

Enriched Microcosms

Kinetic Constant

Natural

Microcosms

Microbially

Enriched

Microcosms

kn, day-1 0.31 ± 0.06 0.29 ± 0.02

Kn, mg⋅L-1 2.0 ± 1.4 7.58 ± 4.0 

μmax, day-1 0.09 ± 0.06 0.14 ± 0.07

Ks, mg⋅L-1 11.4 ± 0.6 33.1 ± 11.5

Y, mg⋅mg-1 0.03 0.07

Starting biomass, mg⋅L-1

Oxidation

17.3 1.39

Starting biomass, mg⋅L-1

Methanogenesis

0.003 1.39

88
Stimulating the Biotransformation of Polychlorinated Biphenyls

John F. Quensen, III, Stephen A. Boyd, and James M. Tiedje

Michigan State University, East Lansing, Michigan

Introduction

The discovery that polychlorinated biphenyls (PCBs)

can be reductively dechlorinated by microorganisms under anaerobic conditions has stimulated interest in the

development of a sequential anaerobic/aerobic biotreatment process for their destruction. While the aerobic

degradation of PCBs is generally limited to congeners

with four or fewer chlorines, the anaerobic process can

dechlorinate more highly substituted congeners, producing products that are aerobically degradable. Indeed, all products from the anaerobic dechlorination of

Aroclor 1254 (1) have been shown to be aerobically

degradable by one or more strains of aerobic bacteria

(2). Also, the high proportion of monochlorinated

biphenyls that can accumulate as a result of anaerobic

PCB dechlorination may serve to induce PCB-degrading enzymes in aerobic microorganisms (3). More

highly chlorinated congeners can be aerobically cometabolized but are not inducing substrates (4).

A greater understanding of the factors controlling the

anaerobic dechlorination of PCBs is necessary before a

successful sequential anaerobic/aerobic biotreatment

process can be developed for PCBs. In particular, it is

important to determine how to stimulate more rapid and

complete dechlorination in areas where the natural rate

and/or extent of dechlorination is limited. The general

approach we have taken is to identify the most probable

site-specific factors limiting in situ PCB dechlorination,

then to apply treatments to alleviate the limitation(s).

During the past year of this project, we have focused on

enhancing the dechlorination of PCBs present in River

Raisin (Michigan) and Silver Lake (Massachusetts)

sediments.

River Raisin Sediment Experiment

In a previous project, we found that little in situ dechlorination of the PCBs present in River Raisin sediment

collected near Monroe, Michigan, had occurred. PCBdechlorinating microorganisms were found to exist in the

sediment, however. The sediment supported dechlorination in laboratory assays when spiked with additional

PCBs and inoculated with PCB-dechlorinating microorganisms (meaning inhibitory compounds were not present), and the PCBs already present in the sediments

were bioavailable because they were dechlorinated under the conditions of our treatability assay. In fact, individual congeners in the contaminated sediment

decreased 30 to 70 percent in 24 weeks at rates nearly

identical to rates for the same congeners freshly spiked

into noncontaminated sediments.

The treatability assays were conducted using air-dried

River Raisin sediments. The sediments were slurried

with an equal weight of non-PCB-contaminated sediments and reduced anaerobic mineral medium (RAMM).

The slurry was then inoculated with microorganisms

eluted from Hudson River sediment to ensure that PCBdechlorinating microorganisms were present, and with

2’,3,4-trichlorobiphenyl (34-2-CB) in a small volume of

acetone. The 34-2-CB was added because the addition

of a single PCB congener (or other halogenated aromatic compound) can sometimes “prime” the dechlorination of PCBs already present in a contaminated

sediment (5). Non-PCB-contaminated Red Cedar River

sediments were added to provide a source of unidentified nutrients. The RAMM included essential mineral

salts and a chemical reductant (Na2S) to lower the initial

redox potential. We have conducted separate experiments with slurries made from both air-dried and always

wet River Raisin sediments to help determine what aspect of our treatability assay fosters the dechlorination

of the PCBs present in these sediments.

Materials and Methods

Air-Dried River Raisin Sediments

With slurries made from the air-dried sediments, the

factors considered were 1) addition of 34-2-CB, 2) addition of the mineral salts in RAMM, 3) addition of Na2S,

and 4) addition of the non-PCB-contaminated sediment.

89
The River Raisin sediment was added to Balch tubes

(1 g per tube). An additional 1 g of non-PCB-contaminated Red Cedar River sediment was added to the

appropriate treatments. Inocula for each treatment were

prepared by eluting PCB-dechlorinating microorganisms from Hudson River sediments with a medium appropriate to the treatment (i.e., with or without mineral

salts, with or without reductant), and 7 mL of an inoculum was added to each tube using an anaerobic technique. 34-2-CB in a small volume of acetone was added

to one treatment, while the rest received the same volume of acetone. The tubes were sealed with Teflon-lined

rubber stoppers and aluminum crimps. Autoclaved treatments served as controls. A tube was sacrificed for each

sample. Triplicate samples were taken at 8-week intervals, extracted, and analyzed for PCBs using capillary

gas chromatography with electron capture detection.

Wet River Raisin Sediments

The same four factors described above plus the necessity of inoculating with Hudson River microorganisms

were considered in an experiment with River Raisin

sediments that had been kept wet since the time of

collection. Portions of the sediment were mixed with the

appropriate medium (i.e., with or without salts, reductant, or inoculum) in a small Erlenmeyer flask on a

magnetic stirrer in an anaerobic chamber. Red Cedar

River sediments, acetone with or without 34-2-CB, and

PCB-dechlorinating microorganisms eluted from Hudson River sediments were also added as appropriate to

each treatment. Portions (7 mL) of the slurries were then

dispensed to Balch tubes, and the tubes were sealed

with Teflon-lined rubber stoppers and aluminum crimps.

The sampling and analytical procedure was the same

as the experiment with slurries made from air-dried

sediment.

Results and Discussion

Decreases in the concentrations of certain PCB congeners in the live samples relative to the autoclaved controls were used to compare the effectiveness of the

various treatments. These congeners (245-25-CB and

235-24-CB in chromatographic peak 42, and 34-34-CB

and 236-34-CB in peak 49) were chosen because each

peak represents more than 2 mole percent of the congeners initially present, and because they could not

have been formed in significant quantities from the

dechlorination of other PCBs present.

In the experiment with slurries made from the air-dried

sediment, approximately 50 percent of the congeners

present in each indicator peak were dechlorinated in the

treatment receiving 34-2-CB (Figure 1). No dechlorination was apparent in any of the other treatments.

In the experiment with slurries made from wet sediments, no inoculation was required for PCB dechlorination, and some dechlorination occurred in all treatments.

Thus, merely making a slurry from the River Raisin

sediments appears to have stimulated some dechlorination of the PCBs present in them. The most extensive

dechlorination, however, occurred in the treatment receiving 34-2-CB, showing that dechlorination could be

enhanced by “priming.” Somewhat surprisingly, the addition of Red Cedar River sediments inhibited dechlorination (Figure 2); perhaps the additional organic matter

served as a sorptive sink for some of the PCBs.

Silver Lake Sediment Experiments

Although there is evidence that the PCBs present in

Silver Lake sediments have undergone in situ dechlorination, these sediments do not support PCB dechlorination in laboratory experiments. These sediments have

high concentrations of several metals, especially zinc,

copper, lead, and chromium. We suspect that the metals

are present mainly in a reduced state in situ and become

partially oxidized and therefore more toxic to dechlorinating microorganisms during the sediment handling

Figure 1. Decrease of indicator PCB congeners due to dechlorination in slurries prepared from air-dried River Raisin sediments.

90
required to set up dechlorination experiments. We also

have previously shown that high concentrations of zinc

can inhibit PCB dechlorination even after highly reduced

conditions are restored. The experiments reported here

were designed to stimulate dechlorination by reducing

the bioavailability of toxic metals through chelation or

precipitation in both a model system and in Silver Lake

sediments.

Materials and Methods

General Procedure

Anaerobic sediment slurries containing PCBs were inoculated with a PCB-dechlorinating microbial consortium eluted from PCB-contaminated Hudson River

sediments. Treatments consisted of the addition of metal salts and/or amendments to precipitate or chelate

metals. Autoclaved slurries served as negative or sterile

controls, while untreated slurries served as positive controls. Triplicate samples were sacrificed at 4-week intervals, solvent extracted, and analyzed for PCBs using

capillary gas chromatography with electron capture detection. The course of PCB dechlorination was followed

by calculating the average meta plus para chlorines for

each treatment versus incubation time. No dechlorination from the ortho positions was evident. Dechlorination

patterns were evaluated by assessing changes in specific congener concentrations over time.

Model System

Anaerobic slurries of non-PCB-contaminated Hudson

River sediment were spiked with Aroclor 1242 (500 μg/g

sediment) and inoculated with PCB-dechlorinating microorganisms eluted from PCB-contaminated Hudson

River sediments. We consistently observed dechlorination of the Aroclor 1242 in such preparations. Zinc (Zn)

or Lead (Pb) (as chloride salts) was added at solution

concentrations of 500 μg/mL to induce metal toxicity.

Amendments of FeSO4, ethylene diamine triacetic acid

(EDTA), and citrate were added individually to samples

before incubation to test their effectiveness in alleviating

the toxicity of Zn and Pb.

Silver Lake Sediment Slurries

Anaerobic slurries of Silver Lake sediments were spiked

with 34-2-CB and inoculated with Hudson River microorganisms. The 34-2-CB was added so that we could

monitor the dechlorination of a freshly added PCB congener in addition to the PCBs already present in the

sediment. Experimental treatments consisted of the addition of FeSO4, EDTA, and citrate, as in the model

system described above.

Results and Discussion

In the model system, ZnCl prevented Aroclor 1242

dechlorination while PbCl decreased the extent of

dechlorination. EDTA, citrate, and FeSO4 amendments

all reversed the inhibitory effect of PbCl while EDTA and

FeSO4 eliminated the inhibition by ZnCl. In fact, FeSO4-

amended treatments exhibited more extensive dechlorination than the unamended positive controls (i.e., those

without PbCl or ZnCl additions). Apparently, the FeSO4

greatly stimulated dechlorination from para positions. In

all non-FeSO4-amended slurries exhibiting dechlorination, dechlorination occurred primarily from the meta

positions to yield ortho and para substituted products

(pattern M). But in FeSO4-amended treatments, the major products were 2-CB, 2-2-CB, and 26-CB, indicating

that dechlorination occurred from both meta and para

positions (pattern C). We have often noted that the para

dechlorination activity present in Hudson River sediments is lost during storage of the sediments. It appears

that addition of FeSO4 somehow “rescues” this dechlorination activity.

Figure 2. Decrease of indicator PCB congeners due to dechlorination in slurries prepared from wet River Raisin

sediments.

91
In the Silver Lake sediment slurries, the added 34-2-CB

was dechlorinated in citrate- and FeSO4-amended slurries, but not in EDTA-amended slurries. There was no

indication of further dechlorination of the PCBs already

present in the sediments.

References

1. Quensen, J.F., III, S.A. Boyd, and J.M. Tiedje. 1990. Dechlorination

of four commercial polychlorinated biphenyl mixtures (Aroclors) by

anaerobic microorganisms from sediments. Appl. Environ. Microbiol. 56:2,360-2,369.

2. Bedard, D.L., R.E. Wagner, M.J. Brennan, M.L. Haberl, and J.F.

Brown, Jr. 1987. Extensive degradation of Aroclors and environmentally transformed polychlorinated biphenyls by Alcalignes eutrophus H850. Appl. Environ. Microbiol. 53:1,094-1,102.

3. Masse, R., F. Messier, L. Peloquin, C. Ayotte, and M. Sylvestre.

1984. Microbial biodegradation of 4-chlorobiphenyl, a model compound of chlorinated biphenyls. Appl. Environ. Microbiol. 41:947-

951.

4. Furukawa, K., F. Matusumura, and K. Tonomura. 1978. Alcaligenes and Acinetobacter strains capable of degrading polychlorinated biphenyls. Agric. Biol. Chem. 42:543-548.

5. Bedard, D.L., H.M. VanDort, R.J. May, K.A. DeWeerd, J.M. Principe, and L.A. Smullen. 1992. Stimulation of dechlorination of Aroclor 1260 in Woods Pond sediment. In: General Electric Company

research and development program for the destruction of PCBs,

11th progress report. Schenectady, NY: General Electric Corporation Research and Development, pp. 269-280.

92
Bioaugmentation for In Situ Co-metabolic Biodegradation of 

Trichloroethylene in Ground Water

Junko Munakata Marr and Perry L. McCarty

Stanford University, Stanford, California

V. Grace Matheson, Larry J. Forney, and James M. Tiedje

Michigan State University, East Lansing, Michigan

 Stephen Francesconi and Malcolm S. Shields

University of West Florida, Pensacola, Florida

 P.H. Pritchard

U.S. Environmental Protection Agency, Gulf Breeze, Florida

Introduction

Trichloroethylene (TCE), a common ground-water contaminant, has been found to be fortuitously degraded

(co-metabolized) by organisms grown on a variety of

substrates (1, 2). Addition of such substrates can lead

to two significant problems in ground-water aquifers.

First, organisms stimulated by substrate addition may

be unable to degrade TCE. Second, the most promising

compounds for inducing TCE degradation, phenol and

toluene, are themselves hazardous substances and

therefore cause regulatory concern. To address the first

issue, aquifers may be bioaugmented with wild-type

strains known to be effective at TCE degradation. Perhaps ideally, both problems can be overcome through

the use of mutant strains known both to degrade TCE

efficiently and to grow on a nontoxic substrate. Laboratory studies were conducted to investigate these two

alternatives.

Laboratory Studies

Bacterial Cultures

The wild-type strain evaluated was Pseudomonas

cepacia G4 (G4), a strain isolated from a holding pond

at an industrial waste treatment facility in Pensacola,

Florida (2). This organism co-metabolizes TCE using

toluene ortho-monooxygenase (TOM), which is induced

by phenol or toluene (3). The mutant used was P.

cepacia G4 PR1301 (PR1), a chemically induced mutant

of G4 that constitutively expresses TOM while grown on

substrates such as lactate. A more complete description

of PR1 is presented at this meeting.

Soil Microcosms

Small-column microcosms (17 cm3) were constructed

using aquifer material from a test area at Moffett Federal

Air Station. Column fluids were exchanged every 2 to 3

days by pumping 10 mL of solutions held in gas-tight

glass barrel syringes through the column influent port

with a syringe pump. At the start of each fluid exchange

period, 1 mL of bacterial culture was added to the microcosms followed by 9 mL of oxygenated ground water

containing about 200 μg/L TCE and/or primary substrates. Microcosm effluent samples were collected during each exchange for analysis.

Detection of Bacteria

A deoxyribonucleic acid (DNA) probe specific for both

strains of G4 was constructed using polymerase chain

reaction (PCR) to amplify segments of G4 DNA between

repetitive extragenic palindromic (REP) sequences. The

REP-PCR reaction can be performed directly on environmental samples and therefore does not require extraction of DNA before amplification. The method was

tested using ground-water and sediment samples containing indigenous bacterial populations with and without added G4, with parallel plate counts on R2A agar.

93
Results

Soil Microcosms

The 6.5-mg/L, phenol-fed, nonbioaugmented column

followed a pattern similar to that observed in the field (4)

and consumed approximately 60 μg/L TCE relative to a

nonfed control. Columns augmented with induced G4

without a primary substrate achieved similar levels of

TCE degradation. With the addition of 15 mg/L lactate,

degradation increased to 100 μg/L TCE in G4-amended

columns, but no such increase occurred in a PR1-

amended column. In these lactate-fed columns, the G4

was pregrown on phenol while the PR1 was pregrown

on lactate. When columns amended with either G4 or

PR1 were fed 6.5 mg/L phenol, 130 μg/L TCE was

degraded. The results are summarized in Figure 1.

Detection of Bacteria

The REP-PCR probe was able to detect 10 colony

forming units (CFUs) of G4 against a background of 105

nontarget CFUs contained in 1 μL of template (Figure

2). The probe’s sensitivity compares favorably to other

PCR-based detection methods (5).

Application of REP-PCR to aqueous effluent samples

from the soil microcosms produced mixed results (Figure 3). G4 was not detected in the control or phenol-only

microcosms and gave a strong signal in phenol- and

lactate-fed microcosms augmented with G4, as anticiFigure 1. Average TCE levels during column incubation for the period between Days 37 and 56 for various column treatments. (Bars

indicate 95-percent confidence intervals.)

Figure 2. Sensitivity of G4 REP-PCR products using the strain

G4 GF13 probe; agarose gel electrophoresis of REPPCR reactions and hybridization of the Southern blot

to GF13.

94
pated. In microcosms containing unfed G4 or phenol-fed

PR1, however, responses were weak or absent. The

reason for the latter results has not yet been determined.

Conclusion

Bioaugmentation with G4 or PR1 and phenol feed provides a means for enhancing native activity toward TCE.

Addition of phenol to aquifers could be avoided by simply adding G4 previously induced for the TCE-degrading

enzyme. Lactate enhanced activity of preinduced G4

toward TCE. In PR1-augmented systems, however, lactate did not support the same level of activity toward

TCE as did phenol. G4 and PR1 were identified in

constructed samples with a sensitivity of 1 in 104

. Detection in aqueous column samples gave some unexpected results, the causes for which remain to be

elucidated.

References

1. Wilson, J.T., and B.H. Wilson. 1985. Biotransformation of trichloroethylene in soil. Appl. Environ. Microbiol. 49:242-243.

2. Nelson, M.J.K., S.O. Montgomery, E.J. O’Neill, and P.H. Pritchard.

1986. Aerobic metabolism of trichloroethylene by a bacterial isolate. Appl. Environ. Microbiol. 55:383-384.

3. Shields, M.S., S.O. Montgomery, S.M. Cuskey, P.J. Chapman, and

P.H. Pritchard. 1991. Mutants of Pseudomonas cepacia G4 defective in catabolism of aromatic compounds and trichloroethylene.

Appl. Environ. Microbiol. 57:1,935-1,941.

4. Hopkins, G.D., J. Munakata, L. Semprini, and P.L. McCarty. 1993.

Trichloroethylene concentration effects on pilot field-scale in situ

groundwater bioremediation by phenol-oxidizing microorganisms.

Environ. Sci. Technol. 27:2,542-2,547.

5. Thiem, S.M., M.L. Krumme, R.L. Smith, and J.M. Tiedje. 1994.

Use of molecular techniques to evaluate the survival of a microorganism injected into an aquifer. Appl. Environ. Microbiol.

60:1,059-1,067.

Figure 3. Detection of G4 in column effluents, ninth column

exchange (Day 20); agarose gel electrophoresis of

REP-PCR reactions of column effluents and probe

GF13 hybridization of the Southern blot.

95
Biodegradation of Chlorinated Solvents

Larry Wackett, Lisa Newman, and Sergey Selifonov

University of Minnesota, St. Paul, Minnesota

Peter Chapman and Michael Shelton

U.S. Environmental Protection Agency, Gulf Breeze, Florida

General Scope of Research

Research is being conducted on the bacterial metabolism of chlorinated aliphatic compounds, with a focus

on oxidative mechanisms of biodegradation. Pseudomonas cepacia G4 oxidizes trichloroethylene (TCE) and

related chlorinated alkenes with relatively little loss of

activity over time (1), and the molecular basis of this

observation is being elucidated. In vivo experiments are

delineating the substrate range and concentration limits

of P. cepacia G4 for chlorinated solvents. In vitro experiments are defining the properties of toluene 2-monooxygenase, the enzyme catalyzing the oxidation of TCE.

Purification and Properties of Toluene

2-Monooxygenase

Toluene 2-monooxygenase activity was monitored in

vitro via a sensitive radiometric assay using [14C]-toluene (2). Chromatography of cell-free extracts revealed

that this was a three-component oxygenase system. All

three components have now been purified to homogeneity. In vitro reconstitution of the three proteins and

reduced nicotinamide adenine dinucleotide (NADH)

yielded an active enzyme system that oxidizes toluene

to ortho-cresol and this, subsequently, to 3-methylcatechol. One component is a flavoprotein containing

a 2Fe2S cluster that accepts electrons from NADH

(Table 1). A second component is a low molecular

weight protein that stimulates activity but has no obvious

redox-active functional group (Table 1). The largest

component has an α2β2γ2 subunit structure (Table 1).

This component is implicated as the hydroxylase component as it alone will oxidize toluene in the presence of

dithionite + methyl viologen + O2 or hydrogen peroxide.

The hydroxylase component contains four to six iron

atoms per holoenzyme. Spectroscopically, this component resembles the soluble methane monooxygenase

hydroxylase component from Methylosinus trichosporium OB3b (3).

Table 1. Molecular Properties of Purified Components

Property Hydroxylase

Small

Component Reductase

Subunit structure (αβγ)2 Monomer Monomer

Subunit molecular

 masses (kDa)

5.4, 37.7, 13.5 10.4 40.0

Molecular mass (kDa)

 Gel filtration 216 19.3 45.8

 Native PAGE 190 — —

 SDS-PAGE 211 10.5 41.8

 Calculated (aa

 quantitation)

210 10.4 40.0

Metal content

 Iron content (mol/mol) 5.3 ND 2.3

Inorganic S-

 content

 (mol/mol)

ND ND 2.9

FAD (mol/mol) ND ND 1.2

pI 4.5 4.3 5.8

Absorption maxima 282 nm 277 nm 270, 341,

423, 457

nm

Specific activity

 (units/mg)

1.7a 79.4a 512.0b

Percent recovery 40 27 30

a

 One unit is defined as 1 nmol [14C]-toluene/min at 23°C. b

 One unit is defined as 1 μmol cytochrome c reduced/min at 23°C.

ND = not detected.

— = not determined.

PAGE = polyacrylamide gel electrophoresis

SDS = sodium dodecylsulfate

FAD = flavin adenine dinucleotide

pI = isoelectric point

96
In Vivo Studies With P. Cepacia G4

P. cepacia G4 was shown to grow on aromatic ring

compounds other than toluene and phenol. P. cepacia

also oxidized non-growth-supporting aromatic and aliphatic substrates. Examples of the aromatic substrates

that were investigated in some detail include naphthalene and indene. The oxidation of TCE by P. cepacia G4

has been studied in detail. The major oxidation product

is glyoxylic acid. The effects of TCE on P. cepacia G4

also were studied to determine how resistant the organism is to variable concentrations of TCE. Unlike Pseudomonas putida F1, P. cepacia G4 was not detectably

toxified by low concentration of TCE or by metabolites

generated by oxidative mechanisms. High TCE concentrations, however, exerted a solvent effect that could

markedly depress cell division rates and even cause

cell death.

References

1. Folsom, R.R., P.J. Chapman, and P.H. Pritchard. 1990. Phenol and

trichloroethylene degradation by Pseudomonas cepacia G4: Kinetics and interactions between substrates. Appl. Environ. Microbiol.

56:1,279-1,285.

2. Yeh, W.-K., D.T. Gibson, and T.N. Liu. 1977. Toluene dioxygenase:

A multicomponent enzyme system. Biochem. Biophys. Res.

Comm. 78:401-410.

3. Fox, B.G., W.A. Froland, J.E. Dege, and J.D. Lipscomb. 1989.

Methane monooxygenase from Methylosinus trichosporium OB3b:

Purification and properties of a three-component system with high

specific activity from a type II methanotroph. J. Biol. Chem.

264:10,023-10,033.

97
Biological and Nutritional Factors Affecting Reductive Dechlorination of

Chlorinated Organic Chemicals

Dingyi Ye

National Research Council, Athens, Georgia

W. Jack Jones

U.S. Environmental Protection Agency, Athens, Georgia

Introduction

Halogenated organic chemicals are of major public concern because these compounds are usually toxic and

persistent in the environment, and they tend to accumulate in soils, sediments, and biota. Polychlorinated

biphenyls (PCBs) and organochlorine pesticides are environmental pollutants of great concern due to a history

of heavy use, toxicities, persistence in the environment,

and wide distribution in environmental media.

PCBs are a mixture of chlorinated biphenyls consisting

of 209 possible congeners. These compounds were

widely used for almost 50 years, with several hundred

million pounds having been released into the environment. An organochlorine pesticide of concern is the

insecticide toxaphene, a complex mixture of chlorinated

camphenes. Toxaphene consists of more than 177 derivatives and was heavily used in the United States

before 1982. Estimates indicate that about 233,688 metric tons of toxaphene was manufactured in the United

States from 1964 to 1982. Toxaphene, like PCBs and

other organochlorines contaminants, is a global pollutant. Many studies have shown toxaphene to be relatively persistent and bioaccumulated by biota (1).

The objectives of the present research were to study

factors affecting anaerobic transformation of PCBs and

organochlorine pesticides (e.g., toxaphene) and to develop techniques to enhance their in situ bioremediation.

The preliminary goals were 1) to characterize the anaerobic microbial dechlorination of PCBs in Sheboygan

River, Wisconsin, sediment, 2) to examine the effects

of nutrients, Fe°, and electron carriers on dechlorination

of PCBs, and 3) to examine the anaerobic biotransformation of toxaphene using indigenous and PCBdechlorinating microorganisms.

Toxaphene Biotransformation

Materials and Methods

A Hudson River (HR) pasteurized enrichment culture

capable of reductive dechlorination of PCBs was used

for initial toxaphene experiments. The enrichment culture was originally pasteurized at 85°C for 15 min and

subsequently transferred at monthly intervals (1 percent

v/v transfer, repasteurized at 90°C for 10 min at each

transfer). The inoculum was serially diluted, and the

highest dilution (10-6) retaining PCB-dechlorination activity was inoculated to the medium without PCBs; this

culture was used as the inoculum for PCB and

toxaphene biotransformation studies. Two milliliters of

the PCB-dechlorinating inoculum was anaerobically

transferred to 28-mL culture tubes containing 2 mL revised anaerobic mineral medium (RAMM) and 1 g of

sterile, uncontaminated (toxaphene-free) pond sediment. Toxaphene was subsequently added at a final

concentration of 500 μg/g dry sediment. Control sediment samples were autoclaved three times (121°C for

1 hr each time) on consecutive days, with addition of

toxaphene occurring on Day 4. All cultures were incubated at 25°C in the dark.

Results and Discussion

Anaerobic transformation of toxaphene by the pasteurized HR enrichment was evident as indicated by

changes in the gas chromatography (GC) isomerdistribution patterns. GC chromatograms of the autoclaved control and a sample inoculated with the HR

pasteurized inoculum after 7 months of incubation are

presented in Figure 1. Several peaks representative of

toxaphene isomers are numbered to facilitate comparison of the control and experimental chromatograms.

Only a very minor change was observed for a late

98
eluting peak (#39) in the control and live experimental

samples; thus, this peak was chosen as an internal

reference peak to which other peaks were normalized

(based on peak height). In comparison with the sterile

control, appreciable changes in many peaks were noted

in live samples after 7 months of incubation. The percentage changes in specific peaks for the live sample

compared with the sterile control are presented in Figure

2. Increases in some peaks were observed, indicating

accumulation of some dechlorination products. Some

early eluting peaks also decreased, however, suggesting that in addition to reductive dechlorination, anaerobic degradation may have occurred. The possible

anaerobic degradation of toxaphene will be further investigated by identifying polar transformation products.

Changes in toxaphene isomer-distribution patterns were

also observed following GC analyses of autoclaved controls (data not shown). As mentioned previously, sterile

controls were prepared by autoclaving sediments at

121°C for 1 hr on 3 consecutive days. Thus, it is highly

unlikely that the observed transformations in the sterile

controls were biologically mediated. The observed

changes are likely due to either abiotic (chemical)

transformation or enhanced sorption of toxaphene to the

autoclaved soil matrix. It has been previously reported

that specific toxaphene isomers were transformed in

sterile sediments and in a sand-Fe(II)/Fe(III) system (2).

Phototransformation of toxaphene has also been

reported (2). On the other hand, the high Koc (soil organic carbon partition coefficient) value reported for

toxaphene suggests that the chemical mixture should be

strongly sorbed to soil particulates (1); any differences

in Koc among the isomers may influence the toxaphene

isomer distribution patterns in long-term sediment incubations. Thus, the changes noted in the isomer distribution pattern (after correcting for abiotic transformations

observed in sterile controls) are likely isomers of

toxaphene that were subject to transformation by the

inoculated microorganisms and that were relatively resistant to abiotic transformation. In these inoculated experimental cultures, CH4 production was not observed,

and a meta-directed dechlorination of amended PCB

congeners (Aroclor 1242) was confirmed in separate

experiments. These results suggest that the HR pasteurized enrichment culture was capable of anaerobically transforming toxaphene. The HR pasteurized

enrichment is easily maintained and cultivated and,

therefore, may be of potential use in the remediation of

toxaphene-contaminated soils. Additional studies are

under way to evaluate the effectiveness of this enrichment culture for remediation of historically contaminated

soils.

PCB Biotransformation

Materials and Methods

PCB biotransformation experiments were performed

with PCB-contaminated (approximately 500 ppm) Sheboygan River (SR) sediment. For abiotic transformation

experiments, PCB-contaminated sediment was slurried

with anoxic site water inside an anaerobic glove box,

Figure 1. GC profile of toxaphene in A) autoclaved and B) live

experimental microcosms (inoculated with enrichment from Hudson River) after 28 weeks of anaerobic

incubation. Numbered peaks are for reference only.

Figure 2. Percentage change (relative to control sample) of select toxaphene peaks in live experimental microcosm

after 28 weeks of anaerobic incubation.

99
homogenized, then amended with Fe°. Of the slurry

containing 1 g sediment (dry weight), 1.8 mL was transferred to replicate 28-mL serum tubes. Half of the tubes

were spiked with 300 μg Aroclor 1242 as an available

PCB source. Fe°, pyrite, and degassed, sterile distilled

water were then added. Control samples consisted of

autoclaved sediment slurries as described above. All

cultures were incubated at 25°C in the dark.

An experiment to assess the effect of pasteurization on

microorganisms eluted from SR sediment was conducted as described by Ye et al. (3). Finally, eluted

microorganisms from historically contaminated (PCB)

SR sediment were subjected to pasteurization (85°C for

20 min) and used as inocula to assess their potential for

reductive dechlorination of amended Aroclor 1242,

1248, and 1254. Aroclors were added individually at a

final concentration of 500 μg/g dry sediment.

Results and Discussion

Fe°-Amended Experiments

Several studies document the Fe°-mediated reductive

dechlorination of trichloroethylene (TCE) and other chlorinated compounds (4). Our preliminary results of Fe°-

amended SR sediment slurries, however, indicated that

no dechlorination of PCBs occurred after anaerobic incubation for 2 weeks at 20°C in either live or sterile

samples. Further, no evidence of reductive dechlorination of PCBs was observed in the SR sediment slurries

spiked with Aroclor 1242, indicating that bioavailability

of PCB congeners was not a limiting factor for the

Fe°-mediated dechlorination.

A prolonged incubation time is usually necessary to

achieve biologically mediated reductive dechlorination

of PCBs. Thus, it was not surprising to find no evidence

of dechlorination in the live experimental samples, especially because the SR sediment had been stored at

2°C to 4°C for approximately 1 yr before use. It is likely

that a significant amount of time is necessary for the

dechlorinating population to recover to a level to affect

significant dechlorination. Additional experimental results with SR sediment (without nutrient amendment)

indicated that approximately 4 weeks of incubation was

required before detectable PCB dechlorination was

observed.

Investigators at the U.S. Environmental Protection

Agency Athens Research Laboratory have recently

demonstrated Fe°-mediated reductive dechlorination of

other halogenated compounds. We have, however, no

evidence of PCB dechlorination in Fe°-amended samples under similar experimental conditions. These results suggest that PCBs are more resistant to chemical

(abiotic) dechlorination than other chlorinated compounds examined to date.

Pasteurization of SR Sediment

CH4 production was not observed in experimental microcosms inoculated with the pasteurized microorganisms from SR sediment; nonpasteurized cultures,

however, were actively methanogenic. These observations are consistent with previous pasteurization experiments using HR sediments as inocula (3). The

pasteurized cultures preferentially removed meta chlorines, while the untreated cultures removed both meta

and para chlorines from selective PCB congeners. In the

present study, dechlorination of Aroclor 1254 was observed by the untreated inocula after 6 weeks of incubation; however, no significant dechlorination of Aroclor

1254 was evident in experiments inoculated with the

pasteurized inocula after 12 weeks of incubation (Figure

3). These results suggest that anaerobic, spore-forming

microorganisms in the SR sediment exhibit similar

dechlorinating pathways as the microorganisms in the

Figure 3. GC profiles of Aroclor 1254 in A) pasteurized, B) autoclaved, and C) live experimental microcosms (inoculated with SR-eluted microorganisms) after 12 weeks

of incubation.

100
HR sediment. Stimulation of in situ PCB dechlorination

may be possible through the addition of a suitable spore

germinant or growth substrate.

References

1. Saleh, M.A. 1991. Toxaphene: Chemistry, biochemistry, toxicity

and environmental fate. Rev. Environ. Contam. Toxicol. 118:1-85.

2. Williams, R.R., and T.F. Bidleman. 1978. Toxaphene degradation

in estuarine sediments. J. Agric. Food Chem. 26:280-282.

3. Ye, D., J.F. Quensen, III, J.M. Tiedje, and S.A. Boyd. 1992. Anaerobic dechlorination of polychlorolbiphenyls (Aroclor 1242) by

pasteurized and ethanol-treated microorganisms from sediments.

Appl. Environ. Microbiol. 58:1,110-1,114.

4. Matheson, L.J., and P.G. Tratnyek. 1994. Reductive dechlorination

of chlorinated methanes by iron metal. Environ. Sci. Technol.

28:2,045-2,053.

101
Predicting Heavy Metal Inhibition of the In Situ Reductive Dechlorination of

Organics at the Petro Processor’s Superfund Site

John H. Pardue

Department of Civil and Environmental Engineering, Louisiana State University, 

Baton Rouge, Louisiana

Introduction

Transition metals and synthetic organic compounds are

common co-contaminants at waste sites that are candidates for biological treatment. The inhibition of microbial

decomposition of natural organic matter by certain transition metals has been widely documented (1); however,

the inhibition of anaerobic degradation processes (e.g.,

reductive dechlorination) is poorly understood.

Inhibition Characteristics

Inhibition characteristics of a model heavy metal (cadmium, Cd) on a model chlorinated aromatic (2,3,4-

trichloroaniline, 2,3,4-TCA) was determined in the

laboratory. Laboratory microcosm experiments were

conducted in three anaerobic flooded soils with varying

properties. Dechlorination of 2,3,4-TCA to monochloroanilines occurred when total pore-water Cd concentrations were below a critical threshold level. Inhibition

occurred across a continuum of Cd concentrations in

several soils, but a completely inhibited threshold concentration was readily identified (Figure 1). Dechlorination kinetics and metabolites differed with soluble metal

concentration. Speciation of soluble Cd was necessary

to predict whether inhibition would occur, particularly in

the presence of high concentrations of organic ligands

such as humic acids (Table 1). Estimation of metal pools

using selective extractions and measurement of acidvolatile sulfide (AVS) provided additional information but

did not adequately predict whether inhibition of dechlorination would occur. These results demonstrated

the importance of quantification and speciation of

pore-water metals in predicting potential inhibition of

anaerobic biodegradation reactions such as reductive

dechlorination.

Figure 1. Normalized 2,3,4-TCA dechlorination rates (k/kcontrol)

versus soluble Cd concentrations in three flooded

soils: bottomland hardwood (BLH), rice paddy, and

freshwater marsh.

Table 1. MINTEQA2 Results From Pore Water of

Representative Rice and Marsh Soil Suspensions

(estimated pore-water humic acid concentrations

were 1 mg/L in the rice soil and 55 mg/L in the

marsh soil)

Soil

Total

Soluble

Cd (mg/L) Cd Species

Equilibrated

Mass

Distribution,

%

2,3,4-TCA

Dechlorination

Inhibited

RS 0.195 Cd+2 43.7 +

CdCl+ 5.9

CdSO4 (aq) 1.6

Cd-Humate 48.5

MS 0.350 Cd+2 1.0 −

Cd-Humate 98.7

102
Site History

The Petro Processor’s, Inc., site is a high-priority Superfund site near Baton Rouge, Louisiana. The site served

as a chemical waste pit from the early 1960s to the late

1970s. An estimated 60,000 tons of chlorinated organic

waste, primarily hexachlorobutadiene and hexachlorobenzene (HCB), was deposited in several unlined,

diked pits. A spill event resulted in contamination of

stream sediments in an adjacent bottomland hardwood

wetland. Heavy metal contamination is contiguous with

chlorinated organic contamination (primarily HCB) in

these sediments. These sediments are the site of a

bioremediation field trial directed at enhancing reductive

dechlorination of HCB (2).

Predicting Heavy Metal Inhibition

Sampling is being conducted to determine if metal inhibition of reductive dechlorination can be predicted in the

field at the Petro Processor’s site. Characteristics of

inhibition of model compounds (described above) are

being used to develop a strategy for predicting inhibition.

Parallel laboratory studies are being used to confirm that

these same inhibition characteristics would be observed

for HCB in the sediments. Laboratory studies using

2,3,4-TCA indicated that noninhibited soils could be

adequately predicted using the AVS/SEM (simultaneously extracted metal) concept. This concept has been

used to predict the toxicity of metals to benthic organisms (3). In studies with 2,3,4-TCA, soils in which molar

metal concentrations exceeded molar AVS were not

always inhibited, requiring further metal speciation and

prediction of “free,” uncomplexed metal concentrations

using MINTEQA2. Spatial and seasonal information of

AVS and SEM and observations of lower chlorinated

benzene samples are being collected in the field. Sections of the bayou where molar SEM exceeds molar AVS

are undergoing further metal speciation studies.

References

1. Duxbury, T. 1985. Ecological aspects of heavy metal responses in

microorganisms. Adv. Microbiol. Ecol. 8:185-235.

2. Constant, W.D., J.H. Pardue, R.D. DeLaune, K. Blanchard, and

G.A. Breitenbeck. 1995. Enhancement of in situ microbial degradation of chlorinated organic waste at the Petro Processor’s Superfund site. Env. Progress 14:51-60.

3. Di Toro, D.M., J.D. Mahony, D.J. Hansen, K.J. Scott, M.B. Hicks,

S.M. Mays, and M.S. Redmond. 1990. Toxicity of cadmium in

sediments: The role of acid volatile sulfides. Environ. Toxicol.

Chem. 9:1,489-1,504.

103
Effect of Primary Substrate on the Reduction of 2,4-Dinitrotoluene

Jiayang Cheng and Makram T. Suidan

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Albert D. Venosa

Risk Reduction Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio

Introduction

2,4-Dinitrotoluene (DNT) is one of the priority pollutants

(1) commonly found in munitions wastes. It is recalcitrant to biological treatment in aerobic processes (2),

such as the activated sludge system, but can be degraded (3) in a sequential anaerobic/aerobic biosystem.

2,4-DNT is completely transformed to 2,4-diaminotoluene (DAT) with ethanol as the primary substrate in an

anaerobic reactor. Subsequently, 2,4-DAT is readily mineralized (3) in an aerobic reactor. 2,4-DNT can not be

transformed in the anaerobic reactor (4) without a primary substrate. In this study, the anaerobic biotransformation of 2,4-DNT with ethanol, methanol, acetic acid,

or hydrogen as primary substrate was investigated. The

effect of the primary substrate on the reductive transformation of 2,4-DNT was also studied.

2,4-DNT-transforming anaerobic cultures were acclimated with 2,4-DNT and ethanol, methanol, or acetic

acid as the feed organic substrates in three chemostats.

The concentrations of 2,4-DNT and the primary substrates in the feed to the three chemostats are listed in

Table 1. The chemical oxygen demand (COD) loading

for all three chemostats was the same. Minerals and

nutrients were added to the chemostat feed to support

bacteria growth. Na2S ⋅ 9H2O (50 mg/L) was added to

maintain a reducing environment in the chemostats. The

pH and the temperature in the chemostats were maintained constant at 7.2 and 35°C, respectively. The hydraulic retention time in the chemostats was 40 days.

2,4-DNT was completely biotransformed to 2,4-DAT in

all three chemostats. All the primary substrates (ethanol,

methanol, and acetic acid) were converted to methane

and carbon dioxide.

After steady-state operation was achieved in the chemostats, the mixed cultures from the chemostats were used

as the inocula for the batch tests to determine the kinetics of anaerobic biotransformation of 2,4-DNT with different primary substrates. The cultures were then

transferred into the batch reactors in an oxygen-free

anaerobic chamber at 35°C. The pH and the temperature in the batch reactors were maintained the same as

those in the chemostats. Different initial concentrations

of 2,4-DNT were used in the batch tests. To determine

the co-metabolic mechanism of the biotransformation of

2,4-DNT and the primary substrate, hydrogen was also

used as the primary substrate in the batch tests.

Results and Discussion

All the batch tests were run in duplicate, and the deviations of the results were less than 8 percent. The kinetics

of anaerobic biotransformation of 2,4-DNT with ethanol,

methanol, and acetic acid as the primary substrates are

illustrated in Figure 1 (a), (b), and (c), respectively.

2,4-DNT was completely biotransformed to 2,4-DAT via

4-amino-2-nitrotoluene (4-A-2-NT) or 2-amino-4-nitrotoluene (2-A-4-NT) under anaerobic conditions, regardless of the primary substrate. The rate of the

biotransformation of 2,4-DNT and the intermediates

(4-A-2-NT and 2-A-4-NT), however, was much higher in

the presence of ethanol than that in the presence of

either methanol or acetic acid. When ethanol was used

as the primary substrate, hydrogen was produced

during the acetogenesis of ethanol. The hydrogen

then served as the electron donor for the reduction of

2,4-DNT to 2,4-DAT. The bacteria also used ethanol for

their growth. When methanol or acetic acid was used as

Table 1. Concentrations of Substrates in Feed for the

Chemostats

Chemostat

Ethanol

Fed

Methanol

Fed

Acetic

Acid Fed

2,4-DNT, mg/L 91.7 91.7 91.7

Primary substrate, mg/L 500 696 978

104
Figure 1. Anaerobic biotransformation of 2,4-DNT with (a) ethanol, (b) methanol, or (c) acetic acid as the primary substrate.

105
the primary substrate, the substrates were used in the

biosystem to support the growth of the bacteria that

transformed 2,4-DNT to 2,4-DAT. Neither methanol nor

acetic acid was degraded until 2,4-DNT, 4-A-2-NT, and

2-A-4-NT were completely transformed to 2,4-DAT. The

hydrogen for the reductive transformation of 2,4-DNT

and the intermediates was probably from bacterial endogenous decay. 2,4-DNT was not biotransformed without a primary substrate (4) in both chemostat and batch

reactors. 2,4-DNT itself cannot support the growth of the

bacteria, and the primary substrate is necessary for

maintaining the biological activities to transform 2,4-

DNT to 2,4-DAT. The rate of the biotransformation of

2,4-DNT was very low in the initial stage of the process,

indicating that 2,4-DNT inhibited its own biotransformation. The presence of 2,4-DNT and its intermediates also

exhibited inhibition to the bioconversion of the primary

substrate (ethanol, methanol, or acetic acid). The higher

the initial concentration of 2,4-DNT, the longer was this

period of inhibition to the conversion of the primary

substrate. Ethanol, methanol, and acetic acid were rapidly converted by the bacteria after 2,4-DNT and its

biotransformation intermediates were completely transformed to 2,4-DAT.

To prove that hydrogen was the electron donor for the

reductive biotransformation of 2,4-DNT, the same batch

test was conducted with hydrogen as the primary substrate. The results are shown in Figure 2. In the control

reactors without 2,4-DNT, hydrogen with CO2 in the

reactors was immediately converted to methane (Figure

2a). When 2,4-DNT was initially present in the reactors,

hydrogen was first consumed for the biotransformation

of 2,4-DNT to 2,4-DAT, and for supporting the growth of

the bacteria. Methane was produced from the excess

hydrogen after 2,4-DNT, 4-A-2-NT, and 2-A-4-NT were

completely transformed to 2,4-DAT. This phenomenon

indicates that 2,4-DNT, 4-A-2-NT, and 2-A-4-NT also

inhibited the hydrogen-utilizing methanogenesis. The

higher the initial concentration of 2,4-DNT, the more

hydrogen was consumed for 2,4-DNT biotransformation

and the less hydrogen was left for methane production

(Figure 2b-f).

References

1. Keither, L.H., and W.A. Telliard. 1979. Priority pollutants I. A perspective view. Environ. Sci. Technol. 13:416-423.

2. McCormick, N.G., J.H. Cornell, and A.M. Kaplan. 1978. Identification of biotransformation products from 2,4-dinitrotoluene. Appl.

Environ. Microbiol. 35:945-948.

3. Berchtold, S.R., S.L. VanderLoop, M.T. Suidan, and S.W. Maloney.

1995. Treatment of 2,4-dinitrotoluene using a two-stage system:

Fluidized bed anaerobic GAC reactors and aerobic activated

sludge reactors. Water Environ. Res. In press.

4. Cheng, J., Y. Kanjo, M.T. Suidan, and A.D. Venosa. 1995. Anaerobic biotransformation of 2,4-dinitrotoluene with ethanol as primary

substrate: mutual effect of the substrates on their biotransformation. Submitted for publication in Wat. Res.

106
Figure 2. Anaerobic biotransformation of 2,4-DNT with H2 as the primary substrate.

107
Surfactants in Sediment Slurries: Partitioning Behavior and Effects on Apparent

Polychlorinated Biphenyl Solubilization

Jae-Woo Park, John F. Quensen, III, and Stephen A. Boyd

Michigan State University, East Lansing, Michigan

It is generally believed that the biodegradation of poorly

water soluble compounds in soil or sediment systems is

limited by low bioavailability due to strong sorption of the

compounds to natural organic matter. The use of surfactants to increase the apparent water solubility of such

contaminants has often been suggested as a way of

increasing their bioavailability to degrading microorganisms. A possible limitation of this approach is that solubility enhancement is much greater above the critical

micelle concentration (CMC) of the surfactant than below it, and these supra-CMCs are often toxic or inhibitory

to bacteria. A few reports, however, indicate that subCMC concentrations of surfactants may enhance the

anaerobic dechlorination of aromatic compounds. The

goals of our present research efforts are to determine if

sub-CMCs of surfactants can enhance the microbial

dechlorination of polychlorinated biphenyls (PCBs) and,

if so, by what mechanism(s). We have determined the

partitioning behavior of several surfactants in soil and

sediment slurries and their effects on PCB solubilization.

These experiments were undertaken to determine if an

increase in the apparent aqueous solubility of PCBs by

sub-CMCs of these surfactants is a plausible mechanism

for any observed enhancement of PCB dechlorination.

The sorption of four commercial nonionic surfactants

(Triton X-100, Triton X-405, Triton X-705, and Tween 80)

onto the Red Cedar River sediment used in our PCB

dechlorination assays was evaluated. Sorption isotherms were plotted, and Freundlich isotherms of the

form Cs=KCen were fitted to the experimental data

where Cs is the sorbed concentration of the surfactant

(mg/kg), Ce is the aqueous concentration of the surfactant (mg/L), and K and n are constants. K and n values

ranged from 1.193 × 10-4 to 1.009 × 10-3 and from 0.232

to 0.696, respectively. The Red Cedar River sediment

thus shows orders of magnitude less surfactant sorption

than has been reported for soils, as shown by the low K

value.

The distribution coefficients of three PCB congeners at

sub- and supra-CMC surfactant concentrations (up to

four times the CMC) were determined using [14C]labeled

PCBs. The aqueous-phase PCB concentrations increased at all surfactant concentrations tested compared with the sediment-water system without

surfactants. Notably, this included an increase in the

aqueous-phase concentrations of PCBs even at the

lowest surfactant concentration tested (0.05 times

CMC), especially for the inherently less soluble hexaand tetra-CBs by Tween 80. In fact, Tween 80 increased

the solubility of 2,2’,4,4’,5,5’-CB by a factor of 3.3 at

25 percent of its CMC, and by a factor of 6.3 at 75

percent of its CMC.

The low sorption of the surfactants by Red Cedar River

sediments has important consequences for PCB solubilization. Surfactant monomers sorbed to soils or sediments will increase the total organic matter content of

the solids and act as an additional sorptive phase. Consequently, if surfactants strongly sorb to the sediments,

they may actually decrease the aqueous phase concentration of nonionic compounds such as PCBs. When the

mass of sorbed surfactant is small, however, as in the

case of the Red Cedar River sediment, most of the

surfactant mass exists in the water, and the PCB solubilization effect of aqueous phase surfactant micelles

and monomers dominates the sorptive capabilities of

sediment-associated surfactant and native organic matter. Therefore, the aqueous phase concentration of PCB

increases even when relatively small amounts of surfactants are added to the system. While these solubility

enhancements are small relative to those that occur

above the CMCs of these surfactants, the increased

solubility may be enough to significantly increase the

rate of PCB dechlorination, especially for the more chlorinated and less water soluble congeners.

108
Partial Characterization of an Anaerobic, Aryl, and 

Alkyl Dehalogenating Microorganism

Xiaoming Zhang

National Research Council, Athens, Georgia

W. Jack Jones and John E. Rogers

U.S. Environmental Protection Agency, Athens, Georgia

Introduction

To better understand controls and pathways of anaerobic biotransformation of organic pollutants in contaminated environments, pure culture studies are beneficial.

To date, only a few strains of anaerobic dehalogenating

microorganisms have been isolated and characterized.

Among these, Desulfomonile tiedjei is probably the most

widely studied (1). In this study, we report the partial

characterization of an anaerobic bacterium capable of

both aryl and alkyl reductive dehalogenation.

Results and Discussion

An anaerobic bacterium, designated as strain XZ-1, was

isolated from freshwater pond sediment near Athens,

Georgia. Isolate XZ-1 is a sporeforming, motile rod capable of reductive dehalogenation of chlorophenols.

Electron acceptors, including sulfite, thiosulfate, and nitrate (but not sulfate), stimulated growth in the presence

of yeast extract and pyruvate. None of the following

supported growth or dehalogenation of chlorophenol

(CP): glucose, fructose, galactose, rhamnose, cellobiose, xylan, ribose, citrate, fumarate, acetate, peptone,

tryptone, casein hydrolysate, and casamino acids. The

addition of 1 mM carbon dioxide reduced the lag

time before growth. No growth was observed in the

presence of 4 percent air or higher. Growth was

completely inhibited by pentachlorophenol (PCP)

(≥ 32 μM), 2,3,4,5-tetraCP (≥ 8 μM), 3,4,5-triCP

(≥ 16 μM), 3,5-diCP (≥ 120 μM), 2,4-diCP (≥ 500 μM),

and 2-CP (≥ 4,000 μM). The generation time of isolate

XZ-1 was 1.8 hr at pH 7.5 (optimal) and 30°C.

Isolate XZ-1 removed ortho-chlorines from all orthochlorine-containing phenols tested (e.g., 2-CP and pentachlorophenol). Hydrogen, formate, ethanol, pyruvate,

and yeast extract served as electron donors for dehalogenation of CPs. Only pyruvate and yeast extract,

however, stimulated growth either in the absence or

presence of electron acceptors, including 3-chloro-4-hydroxyphenylacetate (an analog of ortho-CP). The aryl

dehalogenation activity was inducible, and induction

was inhibited by addition of chloramphenicol to cell

suspensions. Experiments with D2O demonstrated

that water was the exclusive proton source for aryl

dehalogenation of chlorophenols. Proton nuclear magnetic resonance (NMR) studies indicated that hydrogen

was incorporated at the same position where an orthochlorine was removed. Product solvent isotope effects

were 5.4 and 8.5 for dechlorination of 2,3-diCP and

2-CP, respectively. An increase in the assay temperature

reduced the product solvent isotope effect in 2,3-diCP

dechlorinations.

Cell suspensions of isolate XZ-1 also were capable of

reduction of 2,4,6-trinitrotoluene (TNT, 46 ppm) to

2,4,6-triaminotoluene via 2-amino-4,6-dinitrotoluene,

4-amino-2,6-dinitrotoluene, 2,6-diamino-4-nitrotoluene,

2,4-diamino-6-nitrotoluene, and several unidentified

intermediates. The TNT transformation pattern was different in aryl dehalogenation-induced cells and noninduced cells. The identified intermediates of TNT

reduction accumulated to lower levels in the induced

cells than in the noninduced cells. Addition of pyruvate

stimulated TNT transformation. Heat-treated cell suspensions exhibited only traces of TNT transformation

activity either with or without addition of pyruvate. Cell

suspensions of isolate XZ-1 also metabolized chloramphenicol in the presence of pyruvate. No intermediate(s)

of chloramphenicol transformation has been identified

to date.

Both noninduced and aryl dehalogenation-induced cell

suspensions of isolate XZ-1 dechlorinated tetrachloroethene to trichloroethene (TCE). A comparison of aryl

109
and alkyl dehalogenation rates in noninduced and induced cells suggests that at least two enzymes are

responsible for the two activities. Aryl dehalogenationinduced cells also slowly dechlorinated TCE to cis-1,

2-dichloroethene.

Additional studies are under way to identify the range of

transformation activities of dehalogenating isolate XZ-1

and to characterize further the physiology, nutrition, and

phylogeny of this anaerobe.

Reference

1. Mohn, W.W., J.M. Tiedje. 1992. Microbiol. Rev. 56:482.

110
Microbial Degradation of Petroleum Hydrocarbons in Unsaturated Soils:

The Mechanistic Importance of Water Potential and the Exopolymer Matrix

Patricia A. Holden

Department of Environmental Science, Policy, and Management, University of California,

Berkeley, California

James R. Hunt

Department of Civil Engineering, University of California, Berkeley, California

Mary K. Firestone

Department of Environmental Science, Policy, and Management, University of California,

Berkeley, California

Background

Total soil water potential, Ψ, is the potential energy per

unit volume of unsaturated soil water and is commonly

reported in -MPa (1). Matric water potential, Ψm, is the

largest component of Ψ in most soils and arises from the

interaction of soil water with soil surfaces. Matric potential determines water film thickness in soil, and thus

controls gas-phase mass transfer through soil pores and

solution-phase mass transfer through water films. Bacteria in biofilms are in equilibrium with water in their

environment, and adaptation to a given soil water potential or to a changing water potential condition during

wetting and drying will affect intrinsic bacterial physiology and biofilm characteristics (2). Because of its role in

both mass transfer and bacterial reaction rates, soil

water potential is an important environmental factor controlling petroleum biodegradation rates in unsaturated

soils.

Project Framework

Our biodegradation model for oil constituents at the

biofilm scale contains the following parameters that will

vary as a function of Ψ:

qm = intrinsic molar removal rate per area of

biofilm (moles/#-t)

Ks = intrinsic half saturation constant (moles/L3

)

Lb = total biofilm thickness (L)

De = effective diffusivity of contaminant through

biofilm (L2

/t)

ρm = number density of bacteria per mass of

biofilm (#/m)

ρb = density of biofilm or mass of biofilm per

biofilm volume (m/L3

)

Csat = aqueous solubility of petroleum hydrocarbon

(moles/L3

)

Experimental protocols include determining each parameter as a function of Ψ and determining the overall

removal rate in unsaturated soil as a function of Ψ. We

are also examining how physicochemical properties of

the bacterial matrix are altered with Ψ to effect hydrocarbon solubility and spreadability.

Biofilm Reactors and Preliminary Results

Phenanthrene, hexadecane, and methyl-decalin are the

selected test substrates representing the three major

classes of petroleum constituents. Pristane is the conservative tracer. Polyethylene glycol, a nonpermeating

solute with a molecular weight of 8,000 (PEG 8,000), is

used to set matric water potential in well-mixed and

biofilm culture systems.

Custom-designed biofilm reactors for developing

biofilms under unsaturated conditions have been constructed and are being tested using various growth substrates. Transmission electron micrographs taken

through biofilms grown under Ψ-controlled conditions

reveal architectural changes, specifically cell packing

and morphological, with Ψ. Preliminary diffusion studies

suggest that diffusional mass transfer through biofilms

is related to the Ψ-condition during growth.

111
References

1. Jury, W.A., W.R. Gardner, and W.H. Gardner. 1991. Soil physics,

5th ed. New York: John Wiley & Sons.

2. Harris, R.F. 1981. Effect of water potential on microbial growth and

activity. In: Parr, J.F., W.R. Gardner, and L.F. Elliott, eds. Water

potential relations in soil microbiology, SSSA special publication,

Vol. 11, Number 6, Madison, WI: Soil Science Society of America.

pp. 733-740.

112
Metabolic Indicators of Anaerobic In Situ Bioremediation of

Gasoline-Contaminated Aquifers

Harry R. Beller, Martin Reinhard, and Alfred M. Spormann

Department of Civil Engineering, Stanford University, Stanford, California

Bioremediation is one of a limited number of options for

restoring aquifers contaminated with the hazardous aromatic hydrocarbons that occur in unleaded gasoline,

such as benzene, toluene, ethylbenzene, and the xylenes (BTEX). Considering the cost and technical difficulty

associated with introducing oxygen into some aquifers,

in situ bioremediation using indigenous, anaerobic bacteria merits serious consideration for some contaminated sites. A major impediment to the acceptance of in

situ bioremediation is the difficulty of demonstrating that

decreases in the concentrations of BTEX in ground

water truly represent bacterial metabolism of these compounds rather than abiotic processes such as sorption,

dilution, or volatilization.

Work in our laboratory has included the characterization

of byproducts of alkylbenzene metabolism by pure and

mixed anaerobic cultures (1, 2). This research, which

has focused on sulfate-reducing cultures, has involved

the extensive use of gas chromatography/mass spectrometry for metabolite characterization. We have recently integrated such laboratory findings with field data

from a controlled-release experiment conducted at the

Seal Beach Naval Weapons Station in California.

Based on the concordance of laboratory studies of anaerobic bacteria and field observations from the aquifer

in Seal Beach, we propose a group of compounds including benzylsuccinic acid, benzylfumaric acid (or a

closely related isomer), and the o-, m-, and p-methyl

homologs of these compounds as biogeochemical indicators of in situ anaerobic alkylbenzene metabolism in

gasoline-contaminated aquifers. Under the controlled

conditions of the field study, a strong correspondence

was observed between the disappearance of alkylbenzenes from ground water over time and the appearance

of associated metabolic byproducts. This correspondence was both qualitative (i.e., only products specific

to the metabolism of toluene, o-xylene, and m-xylene

were observed, and only these three hydrocarbons were

depleted) and quantitative (i.e., metabolic byproduct

concentrations tended to increase as the associated

alkylbenzene concentrations decreased).

References

1. Beller, H.R., M. Reinhard, and D. GrbiŒ-GaliŒ. 1992. Metabolic

byproducts of anaerobic toluene degradation by sulfate-reducing

enrichment cultures. Appl. Environ. Microbiol. 58:3,192-3,195.

2. Beller, H.R. 1995. Anaerobic metabolism of toluene and other

aromatic compounds by sulfate-reducing soil bacteria. Ph.D. dissertation. Stanford University, Stanford, CA.

113
Contaminant Dissolution and Biodegradation in Soils Containing

Nonaqueous-Phase Organics

Larry E. Erickson, L.T. Fan, J. Patrick McDonald, George X. Yang, and Satish K. Santharam

Kansas State University, Manhattan, Kansas

Several models have been developed to describe the

dissolution, adsorption to soil, and biodegradation of

nonaqueous-phase contaminants (i.e., hydrocarbons) in

the subsurface, and remediation times have been estimated for various conditions (1-6). The significant ratelimiting factors determining the required bioremediation

time appear to be the rates of transport of the electron

acceptor or oxygen and organic contaminants in pores

and soil aggregates in the vicinity of the hydrocarbon

phase. The contaminants’ solubilities in the aqueous

phase affect their dissolution and transport. Contaminant dissolution and transport are more rapid than oxygen transport for more water-soluble compounds such

as toluene, and less rapid for less soluble compounds

such as pyrene. As long as the nonaqueous phase is

present, the higher the solubility of a compound, the

greater the extent of removal by pump-and-treat operations rather than by oxygen-limited biodegradation. The

sizes of aggregates and hydrocarbon blobs significantly

affect remediation time, which has been found to be

proportional to the square of the characteristic length of

the blob.

The available experimental data for pyrene and anthracene fit well with the results of simulation obtained with

one of the models. Besides dissolution, adsorption,

desorption, and biodegradation, this model takes into

account the hydrocarbon-phase size distribution; moreover, it expresses the rate of biodegradation by Monod

kinetics.

References

1. Yang, X., L.E. Erickson, and L.T. Fan. 1993. Transport properties

of toluene as a non-aqueous phase liquid in ground water. In:

Proceedings of the 8th Conference on Hazardous Waste Research. Manhattan, Kansas: Kansas State University. pp. 313-330.

2. McDonald, J.P., C.A. Baldwin, L.E. Erickson, and L.T. Fan. 1993.

Modeling bioremediation of soil aggregates with residual NAPL

saturation. In: Proceedings of the 8th Conference on Hazardous

Waste Research. Manhattan, Kansas: Kansas State University. pp.

346-365.

3. Gandhi, P., L.E. Erickson, and L.T. Fan. 1995. A simple method to

study the effectiveness of bioremediation aided, pump-and-treat

technology for aquifers contaminated by non-aqueous phase liquids, I. Single component systems. J. Haz. Mat. 39:49-68.

4. Gandhi, P., L.E. Erickson, and L.T. Fan. 1994. A simple method to

study the effectiveness of bioremediation aided, pump-and-treat

technology for aquifers contaminated by non-aqueous phase liquids, II. Multi-component systems. J. Haz. Mat. In press.

5. Yang, X., L.E. Erickson, and L.T. Fan. 1994. A study of dissolution

rate-limited bioremediation of soils contaminated by residual hydrocarbons. J. Haz. Mat. In press.

6. Santharam, S.K., L.E. Erickson, and L.T. Fan. 1994. Modeling the

fate of polynuclear aromatic hydrocarbons in the rhizosphere. In:

Proceedings of the 9th Annual Conference on Hazardous Waste

Remediation. Manhattan, Kansas: Kansas State University. pp.

333-350.

114
Protein Expression in Mycobacteria That Metabolize Polycyclic 

Aromatic Hydrocarbons

David E. Wennerstrom

University of Arkansas for Medical Sciences, Little Rock, Arkansas

Carl E. Cerniglia

National Center for Toxicological Research, Jefferson, Arkansas

Three species of mycobacteria have been isolated from

petroleum contaminated soil (Mycobacterium sp.

PYR-1) (1) or coal gassification sites (Mycobacterium

sp. PAH135 and M. gilvum) (2, 3). These organisms

have potential application in the bioremediation of polycyclic aromatic hydrocarbons (PAHs) because each can

mineralize various PAHs, including naphthalene, phenanthrene, pyrene, and fluoranthene. The present study

was initiated to investigate the molecular basis for the

degradation of PAHs by these species of mycobacteria.

To determine part of the physiological response of the

organisms to the presence of a metabolizable PAH in

the environment, we have analyzed the expression of

proteins by each organism in response to pyrene using

two-dimensional sodium dodecylsulfate-polyacrylamide

gel electrophoresis (SDS-PAGE). For each organism,

the pattern of separated proteins was distinct, and proteins increased in expression following addition of the

PAH. Major proteins increased in induced cells of Mycobacterium sp. PYR-1 had approximate masses of 105,

79, 53, 42, and 15 kDa. In comparison, three proteins

were induced in Mycobacterium sp. PAH135 (95, 70,

and 53 kDa) and in M. gilvum (72, 27, and 15 kDa). To

determine whether increased expression of these proteins is associated with metabolism of pyrene in Mycobacterium sp. PYR-1, uninduced cells were incubated

with the PAH for varying periods up to 8 hr, and the

amounts of pyrene metabolism and protein expression

were quantified by high-performance liquid chromatography (HPLC) analysis and densitometry of proteins

detected in two-dimensional SDS-PAGE gels, respectively. After a delay of about 1 hr, uninduced cells

metabolized all of the pyrene within 8 hr. The 79 kDa

protein, undetectable in uninduced cells, was expressed

at 1.2 percent of proteins within 2 hr and was fully

expressed at about 2 percent of total protein within 4 hr.

Partial characterization of this protein by N-terminal sequencing and hybridization of a synthetic oligonucleotide probe corresponding to the amino acid

sequence to BamIII-digested Mycobacterium sp. PYR-1

deoxyribonucleic acid (DNA) show that this protein is

similar to the katG gene product (catalese-peroxidase)

expressed in many other mycobacteria. Kinetics of increased expression of the 15 kDa protein followed those

for the 79 kDa protein. In contrast, the 42 kDa protein

was not increased until 6 hr and was not fully expressed

even at 8 hr after addition of pyrene. A variant of the

organism was isolated that failed to metabolize pyrene

and fluoranthene added to soft agar overlays or in liquid

cultures. The variant retained the ability to metabolize

naphthalene and phenanthrene. None of the proteins

studied was induced in this organism after exposure to

3 μg/mL pyrene for 24 hr. These results indicate that

additional components are required for metabolism of

pyrene and fluoranthene compared with those for metabolism of naphthalene and phenanthrene in Mycobacterium sp. PYR-1. Our results suggest that the proteins

studied are associated with metabolism of pyrene in

induced cells of this organism. These results provide

fundamental information about the proteins expressed

by these mycobacteria during PAH degradation. Clearly,

this information will be important for future application of

these mycobacterial strains as inoculants in the bioremediation of PAH-contaminated sites.

115
References

1. Heitkamp, M.A., W. Franklin, and C.E. Cerniglia. 1988. Microbial

metabolism of polycyclic aromatic hydrocarbons: Isolation and

characterization of pyrene-degrading bacterium. Appl. Environ. Microbiol. 54:2,549-2,555.

2. Grosser, R., D. Warshawsky, and J.R. Vestal. 1991. Indigenous

and enhanced mineralization of pyrene, benzo(a)pyrene, and carbazole in soils. Appl. Environ. Microbiol. 57:3,462-3,469.

3. Boldrin, B., A. Tiehm, and C. Fritzsche. 1993. Degradation of phenanthrene, fluorene, fluoranthene, and pyrene by a Mycobacterium

sp. Appl. Environ. Microbiol. 59:1,927-1,930.

116
Section 6

Hazardous Substance Research Centers

The Hazardous Substance Research Centers (HSRCs) conduct research on bioremediation

under the direction of ORD’s National Center for Extramural Research and Quality Assurance.

Research is sponsored by the following centers: the Northeast Hazardous Substance Research

Center (Regions 1 and 2), the Great Lakes and Mid-Atlantic Hazardous Substance Research Center

(Regions 3 and 5), the South/Southwest Hazardous Substance Research Center (Regions 4 and

6), the Great Plains and Rocky Mountain Hazardous Research Center (Regions 7 and 8), and the

Western Region Hazardous Substance Research Center (Regions 9 and 10).

The symposium’s poster session included presentations on the co-metabolic biodegradation kinetics of trichloroethylene in unsaturated soils; the effect of water potential on biodegradation kinetics

and population dynamics; developments in anaerobic and aerobic bioventing; developments in the

treatability protocol for co-metabolic bioventing; the environmental safety of commercial oil spill

bioremediation agents; the effectiveness of gas-phase bioremediation stimulating agents for unsaturated zone in situ bioremediation; protein expression of mycobacteria that metabolize polycyclic

aromatic hydrocarbons; a field evaluation of pneumatic fracturing enhanced bioremediation; and

the solids suspension characteristics related to slurry biotreatment performance.

117
Co-metabolic Biodegradation Kinetics of Trichloroethylene in Unsaturated Soils

Karen L. Skubal and Peter Adriaens

University of Michigan, Ann Arbor, Michigan

The ability of methanotrophic and heterotrophic bacteria

to aerobically transform chlorinated solvents is well established. Methane monooxygenase (MMO) and aryl

monooxygenase enzymes, produced by these microorganisms respectively during growth on primary substrates, catalyze the cooxidation and dehalogenation of

chlorinated ethenes including trichloroethylene (TCE)

and vinyl chloride. Bioventing may prove useful for

stimulating co-metabolism and achieving in situ remediation of vadose zone soils contaminated with chlorinated alkenes. This possibility motivated an investigation

of co-metabolic dechlorination by indigenous microbial

populations in soils collected from Wurtsmith Air Force

Base (AFB) in Oscoda, Michigan.

Contaminated aquifer and vadose regions at Wurtsmith

AFB contain perchloroethylene, TCE (up to 1,000 μg/L),

trans-dichloroethylene, vinyl chloride, dichlorobenzenes,

and benzene, toluene, ethylbenzene, and xylene (BTEX)

compounds. High methane concentrations have also

been detected in soil gas at the site, indicating potentially favorable conditions for methanotrophic bacteria.

Sandy soils from several depths are being characterized

and studied in aerobic batch microcosm systems at

room temperature to discern the relative importance of

methanotrophic and heterotrophic organisms, and to

optimize methods for their stimulation. Methanotrophs

are supplied with oxygen and methane, while heterotrophs are supported on toluene as the primary inducing substrate. A range of environmentally relevant

concentrations is studied, and following an acclimation

period TCE is added at approximately one-tenth the

level of primary substrate. The effect of soil moisture on

biodegradation kinetics is examined by comparing microcosms containing soil maintained at the local water

content of 4 percent to microcosms containing saturated

soil. In addition, substrate degradation by soil-derived

cultures is monitored in liquid medium without soil.

Bacterial growth on methane and toluene has been

stimulated, and ongoing work will evaluate optimum

primary to co-metabolic substrate ratios and elucidate

the effect of moisture content on TCE co-metabolism in

soil systems. Through development of a simple methodology for screening soils and microbial populations indigenous to a particular site, this study may clarify the

potential of bioventing to enhance chlorinated solvent

transformation in unsaturated zones containing mixed

wastes.

119
The Effect of Water Potential on Biodegradation Kinetics and 

Population Dynamics

Astrid Hillers and Peter Adriaens

Department of Civil and Environmental Engineering, University of Michigan, 

Ann Arbor, Michigan

Although bioventing is currently being applied in the

field, much remains to be learned about the underlying

parameters controlling biological degradation kinetics in

these systems. These parameters need to be systematically studied to improve modeling and design of

bioventing applications. In this investigation, the impact

of subsurface moisture content on biokinetic parameters

is studied, and the applicability of biological kinetics

obtained in saturated batch systems to the unsaturated

zone is evaluated. Specific emphasis is placed on studying the effects of water potential on oxygen availability,

microbial metabolism, and growth.

Mixed culture studies with indigenous microorganisms

derived from the unsaturated zone at the Wurtsmith field

site, Oscoda, Michigan, have been performed in batch

systems. No degradation of toluene was detected at a

field moisture of about 3 percent (by weight) even after

a month of incubation. Moisture contents between 12

and 16 percent moisture exhibited the fastest degradation of toluene. Differences in biodegradation kinetics

observed as a function of moisture content and independent of population shifts are being verified using

pure cultures of a toluene-degrading microorganism,

isolated from the same unsaturated soil samples.

Water potential, the thermodynamic variable expressing

water activity and therefore water availability for the

microorganisms, is used as the experimental variable

rather than the gravimetric moisture content. Varying

water contents of the soil as a result of drying due to

airflow in bioventing operations influence the different

components of the water potential in the soil matrix. The

osmotic and matric water potential components are

studied separately in their effect on bacterial growth,

energy production, and degradation kinetics. Bacteria

isolated from an unsaturated zone below Wurtsmith Air

Force Base are grown in liquid culture on toluate at

different concentrations of membrane diffusable solutes

(NaCl) and nondiffusable solutes (polyethyleneglycol,

PEG). PEG is used to simulate the effect of matric

potential independent of the effect of mass transfer limitations resulting from moisture changes in porous media. Salt additions to the liquid medium resulted in higher

growth rates of toluate degraders up to 0.2 M NaCl and

increased CO2 evolution. The amount of adenosine

triphosphate produced appeared to be independent of

the salt addition. Studies will be extended to assess

growth in homogenous solids of defined pore structure,

and mixed population studies will be performed to assess the separate effect of population shifts. The results

from these studies will serve as a model for water potential induced microbial stress in unsaturated soil horizons during bioventing.

120
Anaerobic-Aerobic Bioventing Development

Gregory D. Sayles

National Risk Management Research Laboratory,

U.S. Environmental Protection Agency, Cincinnati, Ohio

Munish Gupta and Makram T. Suidan

Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

Motivation

Surface spills and leaking pipelines and underground

storage tanks can result in vadose zone soils contaminated with hazardous chemicals. The vadose zone must

be cleaned to minimize contamination of ground water

and emissions of volatile organic chemicals to the atmosphere.

Aerobic bioventing is now a common approach to treating aerobically biodegradable contaminants (e.g., fuels)

in the vadose zone. Many highly chlorinated aromatics

and aliphatics can be destroyed microbiologically, most

rapidly by sequential anaerobic-aerobic treatment. Usually, the biochemical pathway providing the highest rate

for the initial steps of microbial destruction of the highly

chlorinated organic is anaerobic reductive dechlorination. Once partially dechlorinated, the resulting compounds typically degrade faster under aerobic, oxidizing

conditions. For example, perchloroethylene, wood treating wastes containing pentachlorophenol (PCP), polychlorinated biphenyls (PCBs), and many chlorinated

pesticides can be at least partially dechlorinated by microbial consortia under anaerobic conditions. The resulting

dechlorinated product often can be destroyed microbiologically under aerobic conditions.

For treatment of large volumes of soil at depth, the most

cost-effective approach is likely to be an in situ process

that takes advantage of the above well-known anaerobic

and aerobic biological activities. Our approach to this

challenge is to develop “anaerobic/aerobic bioventing.”

Approach

Anaerobic-aerobic bioventing requires development of

a new process: anaerobic bioventing. Conditions must

be established in the contaminated vadose zone to

induce anaerobic microbial biodegradation. Anaerobic

conditions may be established by injecting nitrogen into

the vadose zone to displace all oxygen. A volatile cosubstrate in the nitrogen stream may be needed to induce

the soil microorganisms to consume all the oxygen and

to establish a low oxidation-reduction potential (ORP).

The low ORP should induce dechlorination of the contaminant. The cosubstrates must be volatile and consumable by anaerobic microorganisms (e.g., ethanol,

acetone, hydrogen).

Once the contaminant is fully or mostly dechlorinated,

aerobic bioventing is initiated by injecting air into the

vadose zone. The aerobic microorganisms should then

complete the mineralization of the contaminant.

To move the technology to the field, the following questions are currently being addressed with pilot-scale

tests:

• Can venting with nitrogen at a low rate establish

adequate anaerobic conditions for dechlorination in

unsaturated soil? What is the cost to supply nitrogen

at the required flow rate?

• Can a cosubstrate introduced with the nitrogen

stream promote the dechlorination of target compounds such as PCP in unsaturated soil?

• What are the most effective volatile cosubstrates, biologically and economically, to promote anaerobic

dechlorination? (Several will be evaluated.)

• Will switching to an aerobic environment (by replacing nitrogen and the primary substrate with air in the

injection stream) promote mineralization of the

dechlorination byproducts?

• What classes of compounds are amenable to anaerobic/aerobic bioventing? (Several will be evaluated.)

121
• Will the addition of hydrogen gas (H2) to the nitrogen

stream aid in the development of reducing environments and promote dechlorination?

Initial experiments were conducted beginning in June

1995. The tests were conducted in pilot-scale soil columns. Two 4-ft long, 4-in. diameter glass columns were

built to simulate an in situ horizontal column of unsaturated soil. Gas sampling ports were placed along the

length of the column. The column was filled with sand

inoculated with secondary effluent to add biomass and

moisture. Essential inorganic nutrients were also added.

Nitrogen gas passed through an oxygen scrubber was

used to flush the columns of oxygen. At field scale, the

nitrogen could be supplied by tank or, perhaps more

economically, by separating it from atmospheric air using onsite molecular sieves. The first cosubstrates

evaluated were ethanol and acetone.

122
Development of Co-metabolic Bioventing: Laboratory Tests

Gregory D. Sayles

National Risk Management Research Laboratory,

U.S. Environmental Protection Agency, Cincinnati, Ohio

Alan D. Zaffiro and Jennifer S. Platt

IT Corporation, Cincinnati, Ohio

Background

The objective of this project is to accumulate the information necessary to write a protocol for determining

site-specific treatability of co-metabolic bioventing. “Cometabolic bioventing” is a bioventing-like process that

delivers air and a volatile co-metabolite to a vadose

zone contaminated with chlorinated solvent (e.g.,

trichloroethylene) to induce in situ biodegradation of the

contaminant. 

The process attempts to induce the following biologically

catalyzed (unbalanced) reactions:

Cosubtrate + O2 ⇒ CO2 + H2O

TCE + O2 + H2O ⇒ CO2 + HCl

where the cosubstrate is chosen because it stimulates

the productions of enzymes in the microbial culture that

oxidize trichloroethylene (TCE). Known volatile cosubstrates include methane, propane, butane, toluene, jet

fuel, gasoline, and isoprene. The co-metabolic bioventing system delivers the cosubstrate and oxygen from

gas injection wells to induce the in situ biodegradation

of the contaminant TCE.

The Remediation Technology Development Forum

Workgroup on In-Situ Bioremediation of Chlorinated

Solvents, a research team with members from industry

and government, plans to conduct field demonstrations

of three in situ bioremediation technologies for the

cleanup of chlorinated solvents at two sites. Laboratory

studies using site-contaminated soils are supporting the

field research. The three technologies are 1) co-metabolic

bioventing of TCE and dichloroethylene contaminated

vadose zone soils, 2) accelerated anaerobic biotreatment of chlorinated ethylene contaminated ground

water, and 3) intrinsic bioremediation of chlorinated ethylene contaminated aquifers. Dover Air Force Base

(AFB), in Delaware was selected as the first site.

One task under the co-metabolic bioventing project was

established to develop a protocol to test site-specific

treatability of co-metabolic bioventing. The U.S. Environmental Protection Agency’s National Risk Management Research Laboratory has assumed primary

responsibility for this task. Information from laboratory

testing using soil from Dover AFB and two to four other

sites will be used to generate the recommended protocol. The efficacy of the protocol will be evaluated by

comparing treatability test results with results of at least

two field demonstrations.

Approach

To simulate co-metabolic bioventing in a closed reactor,

a biodegradation test using soil from Dover AFB was

conducted in the following manner:

• Reactors were closed bottles containing 100 g of soil.

• Soil was unsaturated and not mixed.

• Cosubstrate (toluene or propane) was added to the

bottle. 

• No moisture, nutrients, or microorganisms were

added.

• Killed controls were established.

• Two sets of bottles were established:

– Reactors that were sacrificed in triplicate at various

times during the test.

– Reactors that were monitored by automated respirometry.

123
• Contaminant and cosubstrate loss with time were

monitored with the sacrificial reactors.

• Oxygen use and carbon dioxide production with time

were monitored by the respirometer.

The biodegradation test was conducted during June and

July 1995. Earlier tests conducted with clean soils

spiked with TCE and cosubstrates indicated that care

must be taken in the reactor design to minimize abiotic

losses of TCE. Bottles that expose the soil and its

atmosphere only to glass and Teflon showed significant

(greater than 40 percent) loss of TCE in killed controls

within a week. This loss was attributed to sorption into

the Teflon gaskets. We have redesigned the bottles to

maximize glass surfaces and virtually eliminate Teflon

surface area. Preliminary tests showed that these reactors minimized abiotic TCE loss very well.

124
Evaluating the Environmental Safety of Using Commercial Oil Spill

Bioremediation Agents

Jeffrey L. Kavanaugh

Center for Environmental Diagnostics and Bioremediation, University of West Florida,

Pensacola, Florida

C. Richard Cripe and Carol B. Daniels

Gulf Ecology Division, U.S. Environmental Protection Agency, Gulf Breeze, Florida

Rochelle Araujo

Ecosystems Research Division, U.S. Environmental Protection Agency, Athens, Georgia

Joe E. Lepo

Center for Environmental Diagnostics and Bioremediation, University of West Florida,

Pensacola, Florida

The use of commercial bioremediation agents (CBAs)

for reducing the ecological impact of oil spills raises

several risk assessment questions. The presence of

petroleum hydrocarbons may contribute some toxicity;

CBAs, with their associated chemical constituents (e.g.,

nutrients, dispersants, enzymes), microbes, and inert

ingredients, may add to this toxicity either directly or

indirectly through decreases in dissolved oxygen or increases in particulates. In addition, interaction of CBAs

with oil may have other environmental effects, either

through increasing the amount of petroleum hydrocarbons available to aquatic organisms (i.e., through biosurfactant activity) or by generation of toxic metabolites.

A related issue is whether use of a CBA could reduce

the toxicity of the oil (an efficacy issue).

A tiered approach, with increasing complexity, cost, and

effort, has been proposed to address the environmental

safety of CBA usage. Originally developed for assessing

effluents, 7-day chronic estimator tests using a fish

(Menidia beryllina) and a crustacean (Mysidopsis bahia)

were adapted to evaluate CBAs; the tests utilize endpoints of survival, growth, and, in the case of the mysids,

a measurement of egg production. Tier II evaluates the

toxicity of the CBA, alone and in the presence of a

water-soluble fraction of oil, to provide baseline information on CBA toxicity and potential synergism with petroleum hydrocarbons. Tier III examines effluents from

flow-through test systems that model a variety of aquatic

habitats (open water, beach, marsh) to assess toxicity

under more realistic conditions, where a CBA and oil are

allowed to interact.

Data are presented on the toxicity of a variety of CBAs

classified by vendors as microbial, nutrient, enzyme,

dispersant, and “other.” In the flow-through test systems,

the CBAs exhibited relatively low toxicity, either by themselves or in the presence of an artificially weathered oil.

During a particular period, an apparent interaction between one CBA and oil appeared to increase toxicity in

the marsh system. Toxicity reduction in the sand component of the beach test system could not be developed

into an efficacy endpoint because very small quantities

of oil produced measurable effects on a benthic

amphipod. 

125
UNIFAC Phase Equilibrium Modeling To Assess the Bioavailability of

Multicomponent Nonaqueous-Phase Liquids Containing 

Polycyclic Aromatic Hydrocarbons

Catherine A. Peters

Department of Civil Engineering and Operations Research, Princeton University, 

Princeton, New Jersey

This work is part of a project to evaluate bioremediation

of contaminants that are nonaqueous-phase liquid

(NAPL) mixtures of polycyclic aromatic hydrocarbons

(PAHs). This poster presents the first phase of this work,

aimed at gaining a thorough understanding of multicomponent NAPL/water-phase equilibria for very complex

mixtures. This provides basic information about maximum bioavailable concentrations of PAHs.

The thermodynamics of multicomponent NAPL/waterphase equilibria can be described with knowledge of the

mixture composition, NAPL-phase activity coefficients,

and pure solute aqueous solubilities. This analysis involves application of the UNIFAC model to predict

NAPL-phase activity coefficients for constituent compounds in four different tar materials for which detailed

composition data are available. This group contribution

method has proven to be useful for complex mixtures

such as coal tars because a mixture is represented by

a relatively small number of functional groups, making

thermodynamic analysis using excess Gibbs energy

models tractable. For this work, the molecular structures

of the uncharacterized portions of the tars are approximated through nonparametric regressions of functional

group characteristics with molecular weight. The UNIFAC model was found to predict nearly ideal behavior

for most tar constituents. The activity coefficients range

from 0.14 (quinoline) to 1.27 (ethylbenzene), but the

vast majority of the constituents are predicted to have

activity coefficients between 0.9 and 1.1.

These results provide a firm theoretical basis for making

an assumption of solution ideality for many tar constituents (i.e., Raoult’s law). The robustness of this conclusion is indicated through comparable results across

different tar materials, and through a sensitivity analysis

to the estimated characteristics of the uncharacterized

fractions. These results, in conjunction with laboratory

measurements of PAH biodegradation rates (for individual compounds and for multiple substrate NAPL systems), will eventually be integrated into a mathematical

model describing the rate of biotransformation of PAHcontaining NAPL contaminants and the dynamics of the

composition of the NAPL residual.

126
Field Evaluation of Pneumatic Fracturing Enhanced Bioremediation

Sankar N. Venkatraman

Department of Chemical and Biochemical Engineering, Rutgers University, 

Piscataway, New Jersey

Thomas M. Boland and John R. Schuring

Department of Civil and Environmental Engineering, New Jersey Institute of Technology,

Newark, New Jersey

David S. Kosson

Department of Chemical and Biochemical Engineering, Rutgers University, 

Piscataway, New Jersey

In situ bioremediation is often limited by the rate of

transport of nutrients and electron acceptors (e.g., oxygen, nitrate) to the microorganisms mediating the process, particularly in soil formations with moderate to low

permeability. To overcome these rate limitations, an investigation was conducted to integrate the process of

pneumatic fracturing with bioremediation. Pneumatic

fracturing is an innovative technology that uses highpressure air to create artificial fractures in contaminated

geologic formations, resulting in enhanced subsurface

air flow and transport rates. Following fracturing, the

pneumatic fracturing system can also be used to inject

electron acceptors and other biological amendments

directly into a formation to stimulate biodegradation. The

specific bioremediation process evaluated in this project

used amendment injections and low-rate in situ vapor

extraction to provide oxygen and other supplements,

which resulted in the formation of aerobic, denitrifying,

and methanogenic biodegradation zones, spatially distributed with increasing distance from the fracture interfaces. A “countercurrent” bioremediation process was

thus established with respect to the diffusion of contaminants towards the fracture interface.

A field pilot demonstration of the integrated technologies

was carried out at a gasoline refinery site over a

20-month period. Initial site characterization indicated

the presence of BTX at concentrations of up to 1,500

mg/kg soil, as well as other hydrocarbons. The soil at

the site was overconsolidated clayey silt with very low

permeability. The site was pneumatically fractured followed by periodic injections of subsurface amendments

over a period of 50 weeks. Results demonstrated that

fracturing increased subsurface permeability by an average of 36 times. Information gained from periodic

vapor sampling indicated that following subsurface injections, the production of carbon dioxide was enhanced

due to increased biological activity. Following a lag

phase, the methanogens became active, and an increase in methane production was observed. There was

no carbon dioxide or methane detected in the predemonstration vapor samples. The mass of carbon converted to carbon dioxide and methane was used as an

independent measure for the depletion of total carbon.

Based on this balance, the Cgenerated to Cbiodegraded ratio

was computed to be 3.3:1, indicating that other carbon

sources in gasoline also served as substrate and participated in the biodegradation process. After 1 year of

process operation and monitoring, soil samples obtained from the site indicated a 79-percent reduction in

soil-phase BTX concentrations, and over 85 percent of

the BTX reduction was attributed to biodegradation.

127
Solids Suspension Characteristics Related to Slurry Biotreatment Performances

J.-W. Jim Tzeng, Paul T. McCauley, and John A. Glaser

National Risk Management Research Laboratory, U.S. Environmental Protection Agency,

Cincinnati, Ohio

Introduction

Slurry biotreatment has been demonstrated to be an

effective process for bioremediation of contaminated

soils, sediments, and sludges (1). Solid-phase biotreatment such as composting or formal land treatment units

cannot compete with slurry treatment for the extent of

treatment in short timeframes. Slurry biotreatment has

been commonly conducted in reactor systems such as

agitated tanks or lined lagoons. Slurry systems provide

conditions of improved contact between the pollutant

and the microorganisms responsible for the desired

biotransformation. The extent of particle suspension by

agitation is a crucial factor controlling treatment efficiency. Power input from an impeller system dictates not

only the homogeneity of the slurry medium but also the

degree of particle suspension. Equally importantly, the

power input requirement is an important economic factor

in assessing the feasibility of bioslurry treatment for

particular solids to be treated (2). Very little attention is

given, however, to this important component of the treatment system. This work presents our current technical

efforts on identifying the conditions for optimal slurry

agitation for bioremediation of contaminated solids.

Results

Four flow regimes in terms of particle suspension characteristics have been identified; in increasing order of

impeller power input, they are: nonsuspension,

semisuspension, off-bottom suspension, and completesuspension regimes. Experimental results indicate that

unique relationships exist between the flow regimes and

power input. In addition, kinetic energy rather than impeller rotational speed dictates particle suspension dynamics in a slurry medium. A flow regime map (Figure

1) is constructed using power input as the primary parameter. At extremely low power inputs, particles remain

stationary and settle on the tank bottom. As the power

input exceeds a certain value, the upper layer of the

settled particles starts to become mobile. With a further

increase in power, the layer of settled particles reduces

in thickness, and eventually all particles are mobilized

with a portion of particles moving along the tank bottom.

Such a state corresponds to the minimum off-bottom

suspension as conventionally reported in the literature.

The suspended particles tend to fall back to the tank

bottom, however, due to insufficient momentum transferred from the liquid medium to particles. Dynamics of

particles in this regime can be described by an up-anddown motion, and the particle distribution is nonuniform

along the height of the slurry tank. An increase in the

impeller power input reduces the degree of nonuniformity, and the particle distribution becomes rather uniform as the power input exceeds the minimum

complete-circulation value.

Location of impellers (e.g., bottom clearance) greatly

affects the particle suspension. A substantial reduction

in power input required for both on-bottom and off-bottom particle suspension is obtained as an impeller is

placed near the tank bottom. The conventional design

of agitated tanks, with bottom clearance equal to the

Figure 1. Flow regime map of particle suspension in slurry agitation systems.

128
impeller diameter or tank radius, is inadequate for particle suspension applications.

Conclusion

Feasibility of slurry bioreactors for bioremediation of

contaminated solids depends on the energy consumption required to achieve adequate fluid mixing and particulate suspension in the slurry medium. Four particle

suspension regimes are identified with respect to the

power input to the slurry medium in an agitated tank.

Solids properties affecting the suspension behavior include 1) size and shape distribution, 2) density difference, and 3) solids loading. Operation under the

complete suspension regime for achieving maximum

uniformity of particle suspension may not be necessary

because the treatment efficiency may only be marginally

improved. This is because mass transfer resistance between particles and the bulk liquid phase is not the only

rate-limiting step in the soil-slurry treatment process.

References

1. U.S. EPA. 1990. Engineering bulletin: Slurry biodegradation.

EPA/540/2-90/016. Cincinnati, OH.

2. Muskett, M.J., and A.W. Nienow. 1988. Capital vs. running costs:

The economics of mixer selection. In: The Institution of Chemical

Engineers, ed. Fluid Mixing III.

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