United States
Environmental Protection
Agency
EPA 542-B-94-006 i/
June 1995
Solid Waste and Emergency Response (5102W)
&EPA
Innovative Site
Remediation
Technology
Bioremediation
Volume 1
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INNOVATIVE SITE
REMEDIATION TECHNOLOGY
BIOREMEDIATION
One of an Eight-Volume Series
Edited by
William C. Anderson, P.E., DEE
Executive Director, American Academy of Environmental Engineers
1995
Prepared by WASTECH®, a multiorganization cooperative project managed
by the American Academy of Environmental Engineers® with grant assistance
from the U.S. Environmental Protection Agency, the U.S. Department of
Defense, and the U.S. Department of Energy.
The following organizations participated in the preparation and review of
this volume:
(Air & Waste Management
Association
P.O. Box 2861
Pittsburgh, PA 15230
8"/ American Society of
i> Civil Engineers
345 East 47th Street
New York, NY 10017
American Academy of
Environmental Engineers0
130 Holiday Court, Suite 100
Annapolis, MD 21401
; Hazardous Waste Action
Coalition
1015 15th Street, N.W., Suite 802
Washington, D.C. 20005
American Institute of
Chemical Engineers
345 East 47th Street
New York, NY 10017
1 1 *5
\ w J Society for Industrial
^ = " Microbiology
3929 Old Lee Highway, Suite 92A
Fairfax, VA 22030
Water Environment
Federation
601 Wythe Street
Alexandria, VA 22314
Published under license from the American Academy of Environmental
Engineers®. © Copyright 1995 by the American Academy of Environmental
Engineers®.
-------
Library of Congress Cataloging in Publication Data
Innovative site remediation technology/ edited by William C. Anderson
288 p. 15.24 x 22.86cm.
Includes bibliographic references.
Contents: -- [1] Bioremediation — [2] Chemical treatment - [3] Soil washing/
soil flushing — [4] Stabilization/solidification -- [5] Solvent/chemical extraction —
[6] Thermal desorption — [7] Thermal destruction -- [8] Vacuum vapor extraction
1. Soil remediation. I. Anderson, William, C., 1943- . II. American Academy
of Environmental Engineers.
TD878.I55 1995 628.5'5 93-20786
ISBN 1-883767-01-6 (v. 1) ISBN 1-883767-05-9 (v. 5)
ISBN 1-883767-02-4 (v. 2) ISBN 1-883767-06-7 (v. 6)
ISBN 1-883767-03-2 (v. 3) ISBN 1-883767-07-5 (v. 7)
ISBN 1-883767-04-0 (v. 4) ISBN 1-883767-08-3 (v. 8)
Copyright 1995 by American Academy of Environmental Engineers. All Right:? Reserved.
Printed in the United States of America. Except as permitted under the United Sltates
Copyright Act of 1976, no part of this publication may be reproduced or distributed in any
form or means, or stored in a database or retrieval system, without the prior written
permission of the American Academy of Environmental Engineers.
The material presented in this publication has been prepared in accordance with
generally recognized engineering principles and practices and is for general
information only. This information should not be used without first securing
competent advice with respect to its suitability for any general or specific applica-
tion.
The contents of this publication are not intended to be and should not be:
construed as a standard of the American Academy of Environmental Engineers or of
any of the associated organizations mentioned in this publication and are not
intended for use as a reference in purchase specifications, contracts, regulalions,
statutes, or any other legal document.
No reference made in this publication to any specific method, product, process,
or service constitutes or implies an endorsement, recommendation, or warranty
thereof by the American Academy of Environmental Engineers or any such
associated organization.
Neither the American Academy of Environmental Engineers nor any of such
associated organizations or authors makes any representation or warranty of any
kind, whether express or implied, concerning the accuracy, suitability, or utility of
any information published herein and neither the American Academy of Environ-
mental Engineers nor any such associated organization or author shall be responsible
for any errors, omissions, or damages arising out of use of this information.
Book design by Lori Imhoff
Printed in the United States of America
WASTECH and the American Academy of Environmental Engineers are trademarks of the American
Academy of Environmental Engineers registered with the U.S. Patent and Trademark Office.
-------
CONTRIBUTORS
This monograph was prepared under the supervision of the WASTECH®
Steering Committee. The manuscript for the monograph was written by a task
group of experts in bioremediation and was, in turn, subjected to two peer reviews.
One review was conducted under the auspices of the Steering Committee and the
second by professional and technical organizations having substantial interest in the
subject.
PRINCIPAL AUTHORS
Calvin H. Ward, Ph.D., MPH, Task Group Chair
Foyt Family Chair of Engineering
Director, Energy and Environmental Systems Institute
Rice University
Raymond C. Loehr, Ph.D., P.E., DEE James C. Spain, Ph.D.
Hussein M. Altharthy Centennial Chair Chief, Environmental Biotechnology
Civil Engineering Department USAF Armstrong Laboratory
The University of Texas at Austin Tyndall Air Force Base, Fl.
Evan K- Nyei% P-E- John T- Wilson, Ph.D.
Vice President, Technical Resources Processes & Systems Research Division
Geraghty & Miller, Inc. Robert S. Kerr Environ. Research Lab
U.S. Environmental Protection Agency
Michael R. Piotrowski, Ph.D.
President Robert D. Norris, Ph.D.
Matrix Remedial Technologies, Inc. Technical Director, Bioremediation
Services
J. Michele Thomas, Ph.D. Eckenfelder, Inc.
Senior Research Associate
Department of Environmental Science
and Engineering
Rice University
In addition, the following made worthy contributions to the monograph:
Dolloff F. Bishop Maureen Leavitt
U.S. Environmental Protection Agency IT Corporation
Virginia R. Gordy, Ph.D. Harry Tsomides
Rice University Rice University
-------
REVIEWERS
The panel that reviewed the monograph under the auspices of the Project
Steering Committee was composed of:
George Pierce, Ph.D., Chair James J. Duffy, Ph.D.
Editor-in-Chief Director of Analytical Services
Society for Industrial Microbiology Oxychem
Cytec
James J. Ferris, Ph.D.
Richard A. Conway, P.E., DEE President, Federal Systems
Senior Corporate Fellow CH2M Hill Engineering
Union Carbide Corporation
Diane Saber, Ph.D.
Joseph J. Cooney, Ph.D. Environmental Services Division
Department Chair Fluor-Daniel
Environmental Sciences
University of Massachusetts at Boston
iv
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STEERING COMMITTEE
Frederick G. Pohland, Ph.D., P.E., DEE
Chair
Weidlein Professor of Environmental
Engineering
University of Pittsburgh
Domy Adriano, Ph.D.
Professor and Head
Biogeochemical Ecology Division
The University of Georgia
Representing, Soil Science Society of
America
William C. Anderson, P.E., DEE
Project Manager
Executive Director
American Academy of Environmental
Engineers
Colonel Frederick Boecher
Director of Risk Management
Chemical and Biological Defense
Command
U.S. Army
Representing, American Society of Civil
Engineers
Paul L. Busch, Ph.D., P.E., DEE
President and CEO
Malcolm Pirnie, Inc.
Representing, American Academy of
Environmental Engineers
Richard A. Conway, P.E., DEE
Senior Corporate Fellow
Union Carbide Corporation
Chair, Environmental Engineering
Committee
EPA Science Advisory Board
George Coyle
Division of Technical Innovation
Office of Technical Integration
Environmental Education Development
U.S. Department of Energy
Timothy B. Holbrook, P.E.
Engineering Manager
Groundwater Technology, Inc.
Representing, Air & Waste Management
Association
Walter W. Kovalick, Jr., Ph.D.
Director, Technology Innovation Office
Office of Solid Waste and Emergency
Response
U.S. Environmental Protection Agency
Joseph F. Lagnese, Jr., P.E., DEE
Private Consultant
Representing, Water Environment
Federation
Peter B. Lederman, Ph.D., P.E., DEE, P.P.
Center for Env. Engineering & Science
New Jersey Institute of Technology
Representing, American Institute of
Chemical Engineers
Raymond C. Loehr, Ph.D., P.E., DEE
H.M. Alharthy Centennial Chair and
Professor
Civil Engineering Department
University of Texas
Timothy Oppelt
Director, Risk Reduction Engineering
Laboratory
U.S. Environmental Protection Agency
David Patterson
Senior Technical Analyst
Waste Policy Institute
Representing, U.S. Department of Defense
George Pierce, Ph.D.
Editor-in-Chief
Journal of Microbiology
Manager, Bioremediation Technology
Development
Cytec Industries
Representing, Society of Industrial
Microbiology
Peter W. Tunnicliffe, P.E., DEE
Senior Vice President
Camp Dresser & McKee, Incorporated
Representing, Hazardous Waste Action
Coalition
Charles O. Velzy, P.E., DEE
Private Consultant
Representing, American Society of
Mechanical Engineers
Walter J. Weber, Jr., Ph.D., P.E., DEE
Earnest Boyce Distinguished Professor
University of Michigan
Representing, Hazardous Waste Research
Center
-------
REVIEWING ORGANIZATIONS
The following organizations contributed to the monograph's review and
acceptance by the professional community. The review process employed by each
organization is described in its acceptance statement. Individual reviewers are, or
are not, listed according to the instructions of each organization.
Air & Waste Management
Association
The Air & Waste Management
Association is a nonprofit technical and
educational organization with more than
14,000 members in more than fifty
countries. Founded in 1907, the
Association provides a neutral forum
where all viewpoints of an environmen-
tal management issue (technical,
scientific, economic, social, political,
and public health) receive equal
consideration.
This worldwide network represents
many disciplines: physical and social
sciences, health and medicine, engineer-
ing, law, and management. The
Association serves its membership by
promoting environmental responsibility
and providing technical and managerial
leadership in the fields of air and waste
management. Dedication to these
objectives enables the Association to
work towards its goal: a cleaner
environment.
Qualified reviewers were recruited
from the Waste Group of the Technical
Council. It was determined that the
monograph is technically sound and
publication is endorsed.
The reviewers were:
Paul Lear
OH Materials, Inc.
James Donnelly
Davy Environmental.
American Institute of
Chemical Engineers
The Environmental Division of the
American Institute of Chemical Engi-
neers has enlisted its members to review
the monograph. Based on lhat review
the Environmental Division endorses
the publication of the monograph.
American Society of Civil
Engineers
Qualified reviewers were recruited
from the Environmental Engineering
Division of ASCE and formed a Sub-
committee on WASTECH®. The mem-
bers of the Subcommittee have
reviewed the monograph and have
determined that it is acceptable for
publication.
The reviewers were:
Professor Katherine Banks
Department of Civil Engineering
Kansas State University
Jaakko Puhakka
Department of Civil Engineering
University of Washington
Professor H. David Stensel, Chair
Department of Civil Engineering
University of Washington
Hazardous Waste Action
Coalition
The Hazardous Waste Action Coali-
tion (HWAC) is an association dedi-
cated to promoting an understanding of
VI
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the state of the hazardous waste practice
and related business issues. Our mem-
ber firms are engineering and science
firms that employ nearly 75,000 of this
country's engineers, scientists, geolo-
gists, hydrogeologists, toxicologists,
chemists, biologists, and others who
solve hazardous waste problems as a
professional service. HWAC is pleased
to endorse the monograph as technically
sound.
The lead reviewer was:
Daniel Bonk
Baker Environmental, Inc.
Water Environment
Federation
The Water Environment Federa-
tion is a nonprofit educational orga-
nization composed of member and
affiliated associations throughout the
world. Since 1928, the Federation
has represented water quality spe-
cialists including engineers, scien-
tists, government officials, industrial
and municipal treatment plant opera-
tors, chemists, students, academic
and equipment manufacturers, and
distributors.
Qualified reviewers were re-
cruited from the Federation's Hazard-
ous Wastes, Industrial Wastes, and
Groundwater Committees. A list of
their names, titles, and business
affiliations can be found below.
It has been determined that the
document is technically sound and
publication is endorsed.
The reviewers were:
Terry E. Baxter
Assistant Professor
Civil & Environmental Engineering
Department
Northern Arizona University
Richard A. Bell
Project Manager
TRW, Inc.
Edward J. Bouwer
Professor of Environmental
Engineering
John Hopkins University
Christopher Englert
Dragun Corporation
Roger Hlavek
Chemical Engineer
Naval Air Warfare Center
Byung R. Kim
Principal Staff Engineer
Ford Research Laboratory
*L. Michael Szendrey
Vice President & General Manager
Biotechnical Processes, Inc.
George M. Wong-Chong*
Manager, Remediation Processes
ICF Kaiser Engineers, Inc.
*WEF lead reviewer
Society for Industrial
Microbiology
The Society for Industrial Microbiol-
ogy (SIM) is a nonprofit professional
association dedicated to the advance-
ment of microbiological sciences,
especially as they apply to industrial
products, biotechnology, materials, and
processes. Founded in 1949, SIM
promotes the exchange of scientific
information through its meetings and
publications, and serves as liaison
among the specialized fields of microbi-
ology. Membership in the Society is
extended to all scientists in the general
field of microbiology. Corporate
membership is available to industrial
companies interested in the aims of the
Society.
It has been determined that the
document is technically sound and
publication is endorsed.
VII
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ACKNOWLEDGMENTS
The WASTECH® project was conducted under a cooperative agreement
between the American Academy of Environmental Engineers® and the Office
of Solid Waste and Emergency Response, U.S. Environmental Protection
Agency. The substantial assistance of the staff of the Technology Innovation
Office was invaluable.
Financial support was provided by the U.S. Environmental Protection
Agency, Department of Defense, Department of Energy, and the American
Academy of Environmental Engineers®.
This multiorganization effort involving a large number of diverse profes-
sionals and substantial effort in coordinating meetings, facilitating communica-
tions, and editing and preparing multiple drafts was made possible by a
dedicated staff provided by the American Academy of Environmental Engi-
neers® consisting of:
Paul F. Peters
Assistant Project Manager & Managing Editor
Karen M. Tiemens
Editor
Susan C. Zarriello
Production Manager
J. Sanimi Olmo
Project Administrative Manager
Yolanda Y. Moulden
Staff Assistant
I. Patricia Violette
Staff Assistant
viii
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TABLE OF CONTENTS
CONTRIBUTORS III
ACKNOWLEDGMENTS viii
LIST OF TABLES xv
LIST OF FIGURES xviii
1.0 INTRODUCTION 1.1
1.1 Bioremediation 1.1
1.2 Development of the Monograph 1.2
1.2.1 Background 1.2
1.2.2 Process 1.3
1.3 Purpose 1.4
1.4 Objectives 1.4
1.5 Scope 1.5
1.6 Limitations 1.5
1.7 Organization 1.6
2.0 PROCESS SUMMARY 2.1
2.1 Fundamentals/Basic Science 2.1
2.2 Biogeochemistry and Biodegradation 2.2
2.3 Site Characterization Relevant to In Situ Bioremediation 2.3
2.4 Natural Bioattenuation of Hazardous Organic Compounds
in the Subsurface 2.3
2.5 Bioremediation Processes 2.4
2.5.1 In Situ Bioremediation Technologies 2.4
ix
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Table of Contents
2.5.1.1 Land Treatment 2.4
2.5.1.2 Bioventing 2.5
2.5.1.3 Air Sparging in the Unsaturated Zone 2.6
2.5.1.4 Liquid Delivery Systems 2.6
2.5.1.5 Air Sparging in the Saturated Zone 2.7
2.5.1.6 Potential Applications of In Situ
Bioremediation 2.8
2.5.1.7 Limitations of In Situ Bioremediation
Technologies 2.9
2.5.2 Ex Situ Bioremediation Technologies 2.9
2.5.2.1 Ex Situ Treatment of Contaminated Water 2.9
2.5.2.2 Ex Situ Treatment of Contaminated Soils/
Sediments 2.10
2.5.2.3 Biological Reactors for Contaminated Air 2.11
2.5.2.4 Potential Applications of Ex Situ
Technologies 2.12
2.5.2.5 Limitations of Ex Situ Bioremediation
Technologies 2.12
2.6 Process Evaluation 2.13
2.7 Technology Prognosis 2.14
3.0 PROCESS IDENTIFICATION AND DESCRIPTION 3.1
3.1 Overview of Bioremediation 3.1
3.2 Fundamentals and Basic Science 3.4
3.2.1 Microbial Ecology and Physiology 3.4
3.2.1.1 Organic Contaminants as Growth
Substrates for Microorganisms 3.5
3.2.1.2 Requirements for Microbial Growth
and Metabolism 3.6
3.2.1.3 Microbial Metabolism of Contaminants
that are not Growth Substrates 3.11
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Table of Contents
3.2.1.4 Sources of Microorganisms 3.14
3.2.1.5 Biodegradation of Synthetic Chemicals 3.17
3.2.2 Biogeochemistry and Biodegradation 3.25
3.2.2.1 Introduction 3.25
3.2.2.2 Oxidation-Reduction Potential and
Contaminant Biodegradation 3.26
3.2.2.3 Estimates of Biodegradation Using the
Expression of Electron Acceptor Demand 3.28
3.2.2.4 Nutrient Cycling and Availability 3.29
3.3 Site Characterization Relative to In Situ Bioremediation 3.31
3.3.1 Introduction 3.31
3.3.2 Overview of a Biofeasibility Assessment
Procedure 3.32
3.3.2.1 Site Characterization 3.33
3.3.2.2 Biotreatability Evaluation 3.35
3.3.3 Design Considerations for In Situ Bioremediation
of Aquifers 3.36
3.3.3.1 Regulatory Concerns 3.37
3.3.3.2 Geological Aspects 3.39
3.3.3.3 Hydrogeological Aspects 3.42
3.3.3.4 Geochemical Aspects 3.42
3.3.3.5 Biogeochemical Analyses 3.45
3.4 Natural Bioattenuation of Hazardous Organic
Compounds in the Subsurface 3.47
3.4.1 Patterns of Natural Bioattenuation 3.47
3.4.2 Aerobic Biotransformation of Easily Degraded
Compounds 3.48
3.4.3 Anaerobic Biodegradation of Contaminants as
Carbon or Energy Sources 3.51
XI
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Table of Contents
3.4.3.1 Sulfate-Reducing Conditions 3.52
3.4.3.2 Methanogenic Conditions 3.54
3.4.4 Biodegradation and Organic Contaminants as
Electron Acceptors 3.58
3.5 Bioremediation Processes 3.60
3.5.1 In Situ Bioremediation Technologies 3.64
3.5.1.1 Unsaturated Zone 3.64
3.5.1.2 Bioremediation Processes in the
Saturated Zone 3.92
3.5.1.3 Costs of In Situ Bioremediation
Technologies 3.117
3.5.2 Ex Situ Bioremediation Technologies 3.117
3.5.2.1 Ex Situ Treatment of Contaminated
Water 3.118
3.5.2.2 Ex Situ Treatment of Contaminated
Soils and Sediments 3.139
3.5.2.3 Biological Reactors for Contaminated Air 3.151
4.0 POTENTIAL APPLICATIONS 4.1
4.1 General Criteria 4.1
4.2 In Situ Bioremediation 4.3
4.3 Ex Situ Bioremediation 4.6
4.4 Biological Reactors for Contaminated Air 4.7
5.0 PROCESS EVALUATION 5.1
5.1 Basic Considerations 5.1
5.2 Process By-Products 5.2
5.3 Cleanup Levels Achievable and Duration of Treatment 5.3
5.4 Cost 5.5
xii
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Table of Contents
6.0 LIMITATIONS 6.1
6.1 General Limitations 6.1
6.2 Land Treatment Processes 6.2
6.3 Bioventing 6.2
6.4 Air Sparging 6.3
6.5 Liquid Delivery Processes 6.3
6.6 Soil-Pile Treatment 6.3
6.7 Slurry Reactors 6.4
6.8 Air Bioreactors 6.4
7.0 TECHNOLOGY PROGNOSIS 7.1
7.1 Application of Bioremediation 7.1
7.2 Process Improvements 7.2
7.3 Site Characterization 7.3
7.4 Site Closure Criteria 7.3
APPENDICES
A. Case Studies A.I
B. Glossary of Terms B.I
C. List of Acronyms C.I
D. List of References D.I
xiii
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LIST OF TABLES
Table Title Page
3.1 Examples of naturally-occurring organic compounds
amenable to bioremediation. 3.6
3.2 Microbial processes and redox potential 3.26
3.3 Degradation of/j-cresol under various redox conditions 3.27
3.4 Estimating biodegradation using the concentrations of
alternate electron acceptors or products found in
uncontaminated and ground contaminated groundwater,
respectively 3.29
3.5 Relative occurrence of benzene, trichloroethylene, and
tetrachloroethylene in water supply wells in California3.48
3.6 Rates of bioattenuation under sulfate-reducing conditions
in a plume from a gasoline spill 3.53
3.7 Comparison of kinetics of bioattenuation of an artificial
plume of deuterated aromatic organic compounds and a
full-scale plume under sulfate-reducing conditions 3.54
3.8 Kinetics of biodegradation of organic contaminants in
microcosms under methanogenic conditions 3.55
3.9 Comparison of bioattenuation of organic contaminants in a
plume of contaminated ground water and in microcosms
simulating the plume under methanogenic conditions 3.56
3.10 Bioattenuation of benzene, toluene, and xylenes in
microcosms and methanogenic ground water at an aviation
gasoline spill site 3.57
3.11 Rates of bioattenuation of alkylbenzenes in plumes from a
gasoline spill and a crude oil spill under methanogenic
conditions 3.58
xiv
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List of Tables
Table Title Page
3.12 Depletion of trichloroethylene (TCE) and cis-
dichloroethylene (DCE) in microcosms constructed with
core material from the TCE plume at Picatinny Arsenal,
New Jersey 3.60
3.13 Developmental status of in situ bioremediation 3.61
3.14 Developmental status of ex situ bioremediation 3.62
3.15 Land treatment design and operating factors 3.68
3.16 Comparison of biodegradation rates obtained by the in
situ respiration test with other studies 3.84
3.17 Potential times for contaminant reduction 3.85
3.18 Physical properties important to bioremediation processes 3.109
3.19 Costs of in situ bioremediation technologies 3.117
3.20 Suspended growth systems 3.130
3.21 Fixed film reactors 3.134
3.22 Characteristics of Different Types of Composting Systems 3.147
3.23 Treatment reactors costs for solid phase biological 3.150
4.1 Biodegradability of common contaminants 4.2
4.2 Process applicability assuming adequate biodegradability
of chemicals in the soil 4.4
A.I Mass of cells resulting from different inorganic nutrient
amendments A. 11
A.2 Seepage velocity of fronts of injected chloride, oxygen,
ammonia, and phosphate between the infiltration and
monitoring wells A. 12
xv
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List of Tables
Table Title Page
A.3 Depletion of alkylbenzenes in groundwater after 17 months
of infiltration of nutrients and oxygen A. 18
A.4 Changes in concentration of alkylbenzenes and total
petroleum hydrocarbons in core material during
bioremediation of an aquifer contaminated with aviation
gasoline A. 19
A.5 Concentrations of alkylbenzenes and total petroleum
hydrocarbons in core material from the most contaminated
interval, after 17 months of perfusion with mineral nutrients
and oxygen A.27
A.6 Hydrogen peroxide decomposition, nitrification, and potential
denitrification after infiltration with mineral nutrients and
hydrogen peroxide A.28
A.7 Initial bioassessment of site soils A.32
A.8 Bench-scale biotreatability study treatment scheme A.33
A.9 Bioremediation program operating record A.35
A.10 Summary of soil analytical results — 1988 season A.35
A. 11 Analysis of leachate collected in lysimeters over
two growing seasons A.36
A. 12 Summary of soil analytical results — 1989 season A.37
A. 13 Summary of soil analytical results — 1990 season A.38
xvi
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LIST OF FIGURES
Figure Title Page
3.1 Bioremedial system 3.2
3.2 Applicable bioremediation processes 3.3
3.3 Oxidation-reduction potentials of microbial electron
transfer reactions 3.9
3.4 Schematic of an in situ land treatment process 3.65
3.5 Bioventing design to maximize recovery of volatile
compounds 3.73
3.6 Bioventing design to optimize biodegradation using air
recovery only 3.74
3.7 Bioventing design to optimize biodegradation using air
injection only 3.74
3.8 Bioventing design to optimize biodegradation using air
recovery and injection 3.75
3.9 Well system for liquid delivery 3.95
3.10 Infiltration gallery for treatment of contamination in the
unsaturated zone or shallow contaminated aquifers 3.96
3.11 Groundwater sparging without optional vapor recovery 3.105
3.12 Groundwater sparging with optional in situ vapor stripping
for management of vapors 3.107
3.13a Air sparging in homogeneous aquifers 3.110
3.13b Air sparging in aquifers whh gravel/sand layer 3.111
3.13c Air sparging in the presence of clay lenses in upper portion
of aquifer 3.111
3.14 Plan view of staggered rows of air sparging well 3.114
3.15 Effluent organic concentrations with increasing bacterial
residence time 3.119
xvii
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List of Figures
Figure Title Page
3.16 Time effect on concentration found in a well 3.120
3.17 Aerated lagoon. 3.121
3.18 Activated sludge 3.122
3.19 Life cycle concentration from a well at the center of the
plume for an organic contaminant 3.123
3.20 Bacterial residence time with life cycle influent
concentration 3.125
3.21 Activated sludge, extended aeration, contact stabilization 3.128
3.22 Sequencing batch reactor treatment stages 3.129
3.23 Fixed-film bacterial growth 3.131
3.24 Trickling filter 3.132
3.25 Rotating biological contactor 3.133
3.26 Submerged fixed film 3.135
3.27 Low concentration submerged fixed film 3.136
3.28 Powered activated carbon treatment 3.137
3.29 Ex situ land treatment 3.142
3.30 Soil pile reactor 3.145
3.31 Ex situ bioremediation of volatile organic compounds 3.151
5.1 Biochemical process 5.2
A. 1 Schematic diagram of fuel spill and resulting plume A.5
A.2 Schematic diagram of well system A.8
A.3 Schedule of supply of oxygen or hydrogen peroxide to the
infiltration wells A. 13
xviii
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List of Figures
Figure Title Page
A.4 Breakthrough of oxygen and depletion of total
alkylbenzenes in a monitoring well in the most contaminated
depth interval seven feet from the infiltration wells A. 14
A.5 Breakthrough of oxygen and depletion of total
alkylbenzenes in a monitoring well in the most contaminated
depth interval 31 feet from the infiltration wells A. 15
A.6 Breakthrough of oxygen and depletion of total
alkylbenzenes in a monitoring well in the most contaminated
depth interval 50 feet from the infiltration wells A. 15
A.7 Breakthrough of oxygen and depletion of total
alkylbenzenes in a monitoring well in the most contaminated
depth interval 62 feet from the infiltration wells A. 16
A. 8 Breakthrough of oxygen and depletion of total
alkylbenzenes in a monitoring well in the most contaminated
depth interval 83 feet from the infiltration wells A. 17
A.9 Effect of DO and BTEX on the number of heterotrophs
and hydrocarbon-degraders in groundwater from BD-7B,
level 2 A.20
A. 10 Effect of DO and BTEX on the number of heterotrophs
and hydrocarbon-degraders in groundwater from BD-31,
level 2 A.20
A. 11 Effect of DO and BTEX on the number of heterotrophs
and hydrocarbon-degraders in groundwater from BD-50,
level 2 A.21
A. 12 Effect of DO and BTEX on the number of heterotrophs
and hydrocarbon-degraders in groundwater from BD-62,
level 2 A.22
xix
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List of Figures
Figure Title Page
A. 13 Effect of DO and BTEX on the number of heterotrophs
and hydrocarbon-degraders in groundwater from BD-83B,
level 2 A.23
A. 14 Effect of DO and BTEX on the number of heterotrophs
and hydrocarbon-degraders in groundwater from BD-108,
level 2 A.23
A. 15 Heterotrophs and hydrocarbon degraders in shallow and
deep subsurface cores at 31 ft A.24
A. 16 Heterotrophs and hydrocarbon degraders in shallow and
deep subsurface cores at 62 ft A.25
A. 17 Heterotrophs and hydrocarbon degraders in shallow and
deep subsurface cores at 108 ft A.25
A. 18 Catalase activity in shallow and deep subsurface cores 7,
31, 62, and 108 ft from the infiltration wells A.26
A. 19 Site configuration A.31
A.20 Biotreatability study total petroleum hydrocarbon
analysis A.34
A.21 Closure sampling record A.39
A.22 Closure sampling record A.39
xx
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Chapter 1
I
INTRODUCTION
This monograph on bioremediation is one of a series of eight on innova-
tive site and waste remediation technologies that are the culmination of a
multiorganization effort involving more than 100 experts over a two-year
period. It provides the experienced, practicing professional guidance on the
application of innovative processes considered ready for full-scale applica-
tion. Other monographs in this series address chemical treatment, soil wash-
ing/soil flushing, stabilization/solidification, solvent/chemical extraction,
thermal desorption, thermal destruction, and vacuum vapor extraction.
7. / Bioremediation
Bioremediation exploits the ability of certain microorganisms — het-
erotrophic bacteria and fungi — to degrade hazardous organic materials to
innocuous materials such as carbon dioxide, methane water, inorganic salts,
and biomass. Microorganisms may derive the carbon and energy required
for growth through biodegradation of organic contaminants, or, transform
more complex, synthetic chemicals through fortuitous cometabolism.
The processes discussed in this monograph fall into two categories: natu-
ral bioremediation and enhanced bioremediation. Natural bioremediation,
sometimes referred to as intrinsic bioremediation, depends on indigenous
microflora to degrade contaminants using only nutrients and electron accep-
tors available in situ. However, biodegradation rates will be less than opti-
mal if the microbes' nutritional and physiological requirements are not met.
Enhanced bioremediation technologies increase biodegradation rates by
supplying those nutrients, electron acceptors, or other factors that are rate
limiting.
1.1
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Introduction
Enhanced bioremediation can be used to degrade contaminants in situ or
ex situ. In situ and ex situ processes may be used to treat contaminated
liquids, solids, or air. Some examples of in situ processes include land
treatment, bioventing, liquid delivery, and air sparging. Ex situ technolo-
gies include slurry reactors, land treatment, composting, soil-piles, and
biofilters.
7.2 Development of the Monograph
1.2.1 Background
Acting upon its commitment to develop innovative treatment technolo-
gies for the remediation of hazardous waste sites and contaminated soils
and groundwater, the U.S. Environmental Protection Agency (EPA) estab-
lished the Technology Innovation Office (TIO) in the Office of Solid Waste
and Emergency Response in March, 1990. The mission assigned TIO was
to foster greater use of innovative technologies.
In October of that same year, TIO, in conjunction with the National Ad-
visory Council on Environmental Policy and Technology (NACEPT), con-
vened a workshop for representatives of consulting engineering firms, pro-
fessional societies, research organizations, and state agencies involved in
remediation. The workshop focused on defining the barriers that were im-
peding the application of innovative technologies in site remediation
projects. One of the major impediments identified was the lack of reliable
data on the performance, design parameters, and costs of innovative pro-
cesses.
The need for reliable information led TIO to approach the American
Academy of Environmental Engineers®. The Academy is a long-standing,
multidisciplinary environmental engineering professional society with
wide-ranging affiliations with the remediation and waste treatmenl: profes-
sional communities. By June 1991, an agreement in principle (later formal-
ized as a Cooperative Agreement) was reached. The Academy would man-
age a project to develop monographs describing the state of available inno-
vative remediation technologies. Financial support would be provided by
the EPA, U.S. Department of Defense (DOD), U.S. Department of Energy
1.2
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Chapter 1
(DOE), and the Academy. The goal of both TIO and the Academy was to
develop monographs providing reliable data that would be broadly recog-
nized and accepted by the professional community, thereby eliminating or
at least minimizing this impediment to the use of innovative technologies.
The Academy's strategy for achieving the goal was founded on a
multiorganization effort, WASTECH® (pronounced Waste Tech), which
joined in partnership the Air and Waste Management Association, the
American Institute of Chemical Engineers, the American Society of Civil
Engineers, the American Society of Mechanical Engineers, the Hazardous
Waste Action Coalition, the Society for Industrial Microbiology, and the
Water Environment Federation, together with the Academy, EPA, DOD,
and DOE. A Steering Committee composed of highly respected representa-
tives of these organizations having expertise in remediation technology
formulated the specific project objectives and process for developing the
monographs (see page iv for a listing of Steering Committee members).
By the end of 1991, the Steering Committee had organized the Project.
Preparation of the monograph began in earnest in January, 1992.
1.2.2 Process
The Steering Committee decided upon the technologies, or technological
areas, to be covered by each monograph, the monographs' general scope,
and the process for their development and appointed a task group composed
of five or more experts to write a manuscript for each monograph. The task
groups were appointed with a view to balancing the interests of the groups
principally concerned with the application of innovative site and waste
remediation technologies — industry, consulting engineers, research, aca-
deme, and government (see page iii for a listing of members of the
Bioremediation Task Group).
The Steering Committee called upon the task groups to examine and
analyze all pertinent information available, within the Project's financial
and time constraints. This included, but was not limited to, the comprehen-
sive data on remediation technologies compiled by EPA, the store of infor-
mation possessed by the task groups' members, that of other experts willing
to voluntarily contribute their knowledge, and information supplied by pro-
cess vendors.
1.3
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Introduction
To develop broad, consensus-based monographs, the Steering Committee
prescribed a twofold peer review of the first drafts. One review was con-
ducted by the Steering Committee itself, employing panels consisting of
two members of the Committee supplemented by at least four other experts
(See Reviewers, page iii, for the panel that reviewed this monograph). Si-
multaneous with the Steering Committee's review, each of the professional
and technical organizations represented in the Project reviewed those mono-
graphs addressing technologies in which it has substantial interesl: and com-
petence. Aided by a Symposium sponsored by the Academy in October,
1992, persons having interest in the technologies were encouraged to par-
ticipate in the organizations' review.
Comments resulting from both reviews were considered by the Task
Group, appropriate adjustments were made, and a second draft published.
The second draft was accepted by the Steering Committee and pairticipating
organizations. The statements of the organizations that formally reviewed
this monograph are presented under Reviewing Organizations on page v.
1.3 Purpose
The purpose of this monograph is to further the use of innovative
bioremediation site remediation and waste processing technologies, that is,
technologies not commonly applied, where their use can provide better,
more cost-effective performance than conventional methods. To this end,
the monograph documents the current state of a number of innovative
bioremediation technologies.
14 Objectives
The monograph's principal objective is to furnish guidance for experi-
enced, practicing professionals and users' project managers. The mono-
graph is intended, therefore, not to be prescriptive, but supportive. It is
intended to aid experienced professionals in applying their judgment in
deciding whether and how to apply the technologies addressed under the
particular circumstances confronted.
1.4
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Chapter 1
In addition, the monograph is intended to inform regulatory agency per-
sonnel and the public about the conditions under which the processes it
addresses are potentially applicable.
7.5 Scope
The monograph addresses innovative bioremediation technologies that
have been sufficiently developed so that they can be used in full-scale appli-
cations. It addresses all aspects of the technologies for which sufficient
data were available to the Bioremediation Task Group to describe and ex-
plain the technologies and assess their effectiveness, limitations, and poten-
tial applications. Laboratory- and pilot-scale studies were addressed, as
appropriate.
The monograph's primary focus is site remediation and waste treatment.
To the extent the information provided can also be applied to production
waste streams, it will provide the profession and users this additional ben-
efit. The monograph considers all waste matrices to which bioremediation
can be reasonably applied, such as soils, sludges, filter cake, air, and water.
Application of site remediation and waste treatment technology is site
specific and involves consideration of a number of matters besides alterna-
tive technologies. Among them are the following that are addressed only to
the extent that they are essential to understand the applications and limita-
tions of the technologies described:
• site investigations and assessments;
• planning, management, specifications, and procurement;
• regulatory requirements; and
• community acceptance of the technology.
1.6 Limitations
The information presented in this monograph has been prepared in accor-
dance with generally recognized engineering principles and practices and is
1.5
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Introduction
for general information only. This information should not be used without
first securing competent advice with respect to its suitability for any general
or specific application.
Readers are cautioned that the information presented is that which was
generally available during the period when the monograph was prepared.
Development of innovative site remediation and waste treatment technolo-
gies is ongoing. Accordingly, postpublication information may amplify,
alter, or render obsolete the information about the processes addressed.
This monograph is not intended to be and should not be construed as a
standard of any of the organizations associated with the WASTECH®
Project; nor does reference in this publication to any specific method, prod-
uct, process, or service constitute or imply an endorsement, recommenda-
tion, or warranty thereof.
7.7 Organization
This monograph and others in the series are organized under a uniform
outline intended to facilitate cross reference among them and comparison
of the technologies they address. Chapter 2.0, Process Summary, provides
an overview of all material presented. Chapter 3.0, Process Identification,
provides comprehensive information on the processes addressed. Each pro-
cess is fully analyzed in turn. The analysis includes a description of the
process (what it does and how it does it), its scientific basis, status of devel-
opment, environmental effects, pre- and posttreatment requirements, health
and safety considerations, design data, operational considerations, and com-
parative cost data to the extent available. Also addressed are process-
unique planning and management requirements and process variations.
Chapter 4.0, Potential Applications, Chapter 5.0, Process Evaluation, and
Chapter 6.0, Limitations, provide a synthesis of available information and
informed judgments on the processes. Each of these chapters addresses the
processes in the same order as they are described in Chapter 3.0. Chapter
7.0, Technology Prognosis, addresses the future use of bioremediation and
identifies elements of the processes that require further research and demon-
stration.
1.6
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Chapter 2
2
PROCESS SUMMARY
2.1 Fundamentals/Basic Science
Heterotrophic bacteria and fungi are the primary agents of decomposition
of natural organic matter in the biosphere. Some of these microorganisms
have the capability to utilize natural organic compounds — such as petro-
leum hydrocarbons, phenols, cresols, acetone, and cellulosic wastes — as
sources of carbon and energy. These and other naturally-occurring com-
pounds can be converted to carbon dioxide, methane, water, microbial bio-
mass, and by-products that are usually less complex than the parent mate-
rial. This process is a major component of municipal and industrial waste
treatment systems.
Microbial communities in soil typically have tremendous potential to
degrade a wide range of compounds. The highly successful application of
bioremediation to the treatment of gasoline leaks from underground storage
tanks is attributable to indigenous microorganisms.
To date, most bioremediation technologies deal with treatment of natural
organic compounds. A number of bioremediation strategies are used or are
being developed to treat synthetic chemicals, including pesticides,
chloroaromatic compounds, polychlorinatedbiphenyls (PCBs),
chlorophenols and chlorobenzoates, chloroaliphatic compounds,
nitroaromatic compounds, aniline, phthalates, dibenzodioxins and
dibenzofurans, methyl t-butyl ether, and metals. Some of these highly sub-
stituted compounds as well as some naturally-occurring compounds (poly-
cyclic aromatic hydrocarbons with four or more rings) might not be suitable
growth substrates, however, they may be degraded as the result of fortuitous
cometabolism. These fortuitous reactions stem from the broad substrate
specificity of some microbial enzymes. In this monograph, this type of
2.1
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Process Summary
transformation is termed "microbial metabolism of contaminants that are
not growth substrates." A variety of enzymatic reactions — oxidation,
hydrolysis, reductive dehalogenation, and reduction of nitro groups — cata-
lyze cometabolic processes.
The indigenous microbial community may not have the capability to
degrade specific synthetic chemicals of concern at a particular site. If
treatability studies show no degradation (or an extended delay before sig-
nificant degradation is achieved), inoculation with strains known to be ca-
pable of degrading the contaminant may be helpful. In a process known as
bioaugmentation, microbial strains are added that cannot use the contami-
nant as a growth substrate but, nevertheless, completely degrade the con-
taminant. Bioaugmentation with strains that cannot use the contaminant as
a growth substrate but, nevertheless, completely degrade the contaminant,
has proven successful in several laboratory applications while a few field
trials have been documented: a novel strain of Pseudomonas cepacia has
been used to degrade trichloroethylene; Phanerochaete chrysosporium bio-
degrades a wide range of organic compounds with nonspecific extracellular
peroxidases; and pentachlorophenol has been treated in soil bioreactors by
adding active biomass that has been grown on another substrate.
2.2 Biogeochemistry and Biodegradation
The effect that living organisms have on the geochemistry of the envi-
ronment is known as biogeochemistry. Biogeochemical processes control
the global cycling of carbon, nitrogen, phosphorus, and sulfur, as well as
some trace elements. Biodegradation of organic contaminants and biotrans-
formation of inorganic contaminants are influenced by a variety of bio-
geochemical processes and conditions in the environment:
• Oxidation-Reduction Potential - As an environment is converted
from oxidizing to reducing conditions, different microbial pro-
cesses and terminal electron acceptors are used for contaminant
biodegradation. In general, biodegradation rates tend to be
much lower under reducing conditions than under oxidizing
conditions. The change in redox potential can be used as an
indirect gauge of biodegradation that may have occurred in a
contaminated site.
2.2
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Chapter 2
Nutrient Cycling and Availability - When assessments of nutri-
ent requirements for contaminated environments are made, the
form (availability) of the nutrients in the matrix should be con-
sidered. Concentrations of carbon (e.g., total organic carbon,
dissolved organic carbon, total petroleum hydrocarbons, etc.),
nitrogen, and phosphorus are determined from soil/sediment
samples taken from the site. The nutrient amendment is formu-
lated according to the amount of carbon that can be biodegraded
when a known supply of electron acceptor is added.
2.3 Site Characterization Relevant to In
Situ Bioremediation
To develop a remedial plan, several studies should be conducted to de-
fine the key characteristics of a site:
• Site Characterization - evaluate the geochemical properties of
the site, the quality of the groundwater in contact with the con-
taminants, the flow properties of the contaminated matrix, and
the methods for measuring the contaminant concentration.
• Biofeasibility Evaluation - determine toxicity characteristics of
impacted soil/sediment and whether bioremediation processes
are already occurring. Laboratory treatability studies may be
conducted.
In addition, in situ bioremediation of aquifers involves several engineer-
ing considerations — hydrogeological analyses of the contaminated aquifer
and chemical analyses of the aquifer sediments and the groundwater.
2.4 Natural Bioattenuation of Hazardous
Organic Compounds in the Subsurface
Most organic contaminants enter the subsurface as an oily liquid, such as
a fuel spill or release of a chlorinated solvent. Groundwater moving
2.3
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Process Summary
through the release dissolves some of the contaminant, which then becomes
a plume of groundwater contamination. As the plume moves away from its
source, natural biological processes may attenuate the contamination in the
groundwater. Several studies have documented the phenomenon of natural
bioattenuation, and a mathematical model of aerobic bioremediation in
aquifers has been developed.
Aromatic organic compounds, such as the alkylbenzenes, certain simple
polyaromatic hydrocarbons (PAHs), and some nitrogen-containing hetero-
cyclic organic compounds can be degraded in groundwater in the absence of
oxygen. Anaerobic microorganisms may require months or years to adapt
to the contaminant. When oxygen is absent, nitrate, sulfate, carbonate, or
iron (III) can serve as the terminal electron acceptor. Halogenated organic
compounds can also serve as electron acceptors.
2.5 Bioremediation Processes
Contaminated liquid wastes, sludges, surface soils, subsurface sediments,
and air are all amenable to bioremediation. In situ processes treat contami-
nated soil or sediment where they exist; treatment of contaminated material
in an aboveground reactor or prepared bed comprise ex situ bioremediation
processes.
2.5.1 In Situ Bioremediation Technologies
Some in situ processes — land treatment, bioventing, and, in some cases,
air sparging — are appropriate for bioremediation in the unsaturated zone.
Other in situ processes — liquid delivery systems and air sparging — are
designed for bioremediation in the saturated zone.
2.5.1.1 Land Treatment
In situ land treatment is a managed treatment and disposal technology
that involves the 1) application of waste, sludge, or contaminated soil to
uncontaminated surface soils at a site and then tilling the applied material
into the surface soils or 2) tillage of contaminated surface soils. Land treat-
ment capitalizes on the natural assimilative capacity of the soil to decom-
pose and contain the contaminated material in the surface soil layer.
2.4
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Chapter 2
Site preparation activities vary according to site characteristics and regu-
latory requirements. Trees and rocks may have to be removed, drainage
ditches may be needed to intercept seasonally-high perched water table and
runoff, the soil pH may require adjustment, or the site may have to be con-
toured, terraced, or graded.
The waste material is applied uniformly to the prepared area and tilled
into a depth of 15 to 30 cm (6 to 12 in.). Nutrient applications may enhance
organic degradation rates initially, but may be unnecessary in subsequent
years.
Concentrations of total petroleum hydrocarbons may be reduced 90 to
99%. Specific organic contaminants also are degraded. Compounds with
higher aqueous solubilities have relatively shorter half-lives than less
soluble compounds. The average half-life of typical oil and grease from
refinery wastes and oily sludges ranges from 50 to 150 or more days. Much
effort is being expended to define degradation rates for volatile and
semivolatile compounds: BTEX — benzene, toluene, ethylbenzene, xy-
lenes, and PAHs.
2.5.1.2 Bioventing
Bioventing is the use of induced air movement through unsaturated soils,
with or without nutrient addition, to stimulate indigenous microorganisms
to convert organic contaminants, such as petroleum hydrocarbons, to less
hazardous substances, especially carbon dioxide and water. Many systems
that were designed as in situ vapor recovery systems for physical removal
of contaminants were subsequently found to stimulate biodegradation as
well.
Some systems use air injection wells alone or in conjunction with air
recovery wells. Bioventing has its most direct impact on the unsaturated
zone. Nutrients may become available to the bacteria through use of miner-
als present in the soils, addition of a nutrient-enriched leachate, or in the
case of nitrogen, through biological nitrogen fixation. Biodegradation rates
are enhanced by high moisture content.
Bioventing system designs incorporate some method of introducing oxy-
gen to the unsaturated soils and may include a means for nutrient addition.
The wells are placed in the contaminated zone and screened over some
interval, depending on the distribution of the contaminant, the soil type
distribution, and whether the surface is covered by an impermeable layer.
2.5
-------
Process Summary
The process removes and/or degrades a number of organic contaminants
that are biodegradable under aerobic conditions. Volatile, nonbiodegrad-
able constituents can be treated, but the cost of offgas treatment will be
increased. All constituents must be present at levels that are not toxic to the
microflora.
Where applicable, bioventing has the potential to be a low cost
remediation method. Because of the reduced need for offgas treatment and
lower volumes of air moved through the soils, bioventing should be less
costly than vapor stripping.
2.5.1.3 Air Sparging in the Unsaturated Zone1
Air sparging is a means of introducing air into the saturated zone to
transfer volatile compounds to the unsaturated zone for biodegradation. In
the unsaturated zone, air sparging involves the use of high air injection rates
and/or closely-spaced injection wells to transfer volatile compounds to the
unsaturated zone faster than they can be biodegraded in the saturated zone.
This way, a larger volume of soil and more microorganisms are available to
degrade the contaminant than would be available in the saturated z,one
alone.
Air injection flow rates must be carefully controlled to prevent transfer
of volatile constituents to the atmosphere. Some form of vapor recovery
and, possibly, groundwater protection may be necessary to prevent or limit
losses of contaminants.
2.5.1.4 Liquid Delivery Systems
Liquid delivery systems are used for bioremediation of contamination in
the saturated zone. The process is analogous to conventional wastewater
treatment in that a terminal electron acceptor and inorganic nutrients are
added to enhance contaminant degradation. In contrast to wastewater treat-
ment, which takes place in a bioreactor under controlled conditions, in situ
treatment is effected in the subsurface by indigenous microorganisms.
The liquid delivery technique has been used most often at sites contami-
nated with various types of petroleum hydrocarbons. A process variation
1 . See also Innovative Site Remediation Technology: Vacuum Vapor Extraction-
Ed.
2.6
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Chapter 2
has been tested at field scale to treat chlorinated aliphatic solvents, such as
trichloroethylene and less chlorinated ethylenes by simulating
methanotrophs with methane and oxygen.
Before subsurface bioremediation is initiated, an extensive site character-
ization must be conducted. Laboratory treatability assays can be conducted
using samples of contaminated sediment but not groundwater. Although
treatability assays provide information about biodegradation potential and
nutrient amendments that enhance it, extrapolations from the laboratory to
the field may not be exact. Biodegradation in situ is limited by the rate at
which oxygen is transferred to the contaminant-degrading microorganisms.
The delivery system (consisting of wells or trenches) is designed to cir-
culate adequate amounts of nutrients and oxygen through the zone of con-
tamination to maximize contaminant biodegradation. Injection wells or
trenches, through which nutrients and oxygen are added, are placed within
or close to the contaminated area. Groundwater extraction wells or
trenches may be included. Produced groundwater is extracted, treated
aboveground if necessary, and then disposed or amended with nutrients and
recirculated. The recirculation system is designed to hydraulically isolate
the target area and minimize contaminant migration out of the treatment
zone.
Oxygen is provided by sparging with air or pure oxygen or by adding
hydrogen peroxide to injected water. Because of the limited solubility of
oxygen in water, it is difficult to deliver large quantities of dissolved oxy-
gen to contaminated subsurface environments. Nitrate, sulfate, and salts of
iron (III), can act as alternate electron acceptors.
The cost of implementing a liquid delivery system depends on several
factors: type, amount, and extent of contamination, sediment characteris-
tics, and the source of oxygen.
2.5.1.5 Air Sparging in the Saturated Zone2
Air sparging in the saturated zone serves a two-fold purpose: it provides
oxygen, which acts as an electron acceptor for biodegradation in the aquifer
and the unsaturated zone, and it physically transfers volatile substances to
2 . See also Innovative Site Remediation Technology: Vacuum Vapor Extraction-
Ed.
2.7
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Process Summary
the unsaturated zone for capture by an in situ vapor recovery system. Dis-
solved oxygen is distributed through the aquifer by movement of air
bubbles, diffusion, and groundwater movement.
The degree to which biodegradation occurs (as opposed to physical trans-
fer to the unsaturated zone) depends on the characteristics of the contami-
nants, the site lithology, and the system design. Only minimal volatilization
of heavier hydrocarbon blends can be expected, whereas lighter petroleum
hydrocarbons will be both biodegraded and physically removed.
The sparging wells are operated intermittently with relatively long times
between air injection periods. This practice minimizes air stripping during
a lag period, thereby maximizing the contribution of biodegradation. The
off/on schedule can be developed by monitoring dissolved oxygen levels in
groundwater and volatile compounds in the unsaturated zone.
2.5.1.6 Potential Applications of In Situ Bioremediation
In general, the effectiveness of in situ technologies is dependent on con-
taminant biodegradability, contaminant concentration, and soil properties.
In situ methods require movement of air or water through undisturbed soils
to deliver the electron acceptor and/or nutrients. For this reason, in situ
processes are adversely affected when soil/sediment hydraulic conductivi-
ties are less than 10"4 cm/sec (3.3 x 10"6 ft/sec).
Although liquid delivery was the first in situ bioremedial approach for
treating subsurface contamination, air sparging has superseded this method
at most sites. Air sparging is less expensive, distributes oxygen across the
site more quickly, and is associated with fewer operational problems than
the liquid delivery method. Liquid delivery may be applicable in situations
where a pump-and-treat system is already in place or where site conditions,
such as fractured rock aquifers, aquifers with shallow water tables, or for-
mations with narrow saturated intervals, preclude air sparging. Also, if
control of plume migration is mandated, liquid delivery may be advanta-
geous.
Bioventing is most applicable where the depth-to-water exceeds ten feet
and the surficial soils do not require treatment or are being treated by an-
other method. Shallower soils and sites with shallower water tables can be
treated if the surface is capped. Bioventing systems can also remove
nonbiodegradable compounds, as well as contaminants that are difficult to
degrade, provided that offgas treatment is included in the process;.
2.8
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Chapter 2
2.5,1.7 Limitations of In Situ Bioremediation Technologies
Surface bioremediation methods, such as land treatment, use tillage to
aerate the soil. With this form of mixing and aeration, there is the potential
for volatile emissions. Whether such emissiions occur and whether thay are
a problem requires a site specific determination. Sufficient open space must
be available for land application of excavated soils.
Clayey materials with low permeabilities are less amenable to
bioventing.
The limitations of air sparging for bioremediation are not fully known at
this time. In most cases, vapor and groundwater recovery systems are re-
quired. The air may be channeled through permeable layers in heteroge-
neous soils with clay lenses, gravel stringers, etc., directing oxygen away
from the contaminated zone.
Liquid delivery processes are limited by the rate of groundwater recircu-
lation relative to the contamination level. Precipitation of nutrients, iron
precipitation, or formation of excess biomass may clog soil pores.
2.5.2 Ex Situ Bioremediation Technologies
Ex situ bioremediation technologies are those in which a waste that has
been removed from its point of origin is treated in a closed or open
bioreactor. Liquids, solids, and vapor are amenable to ex situ treatment.
Bioreactor design for aerobic treatment must solve two problems. First,
the bacteria must be in contact with the contaminants for extended periods
of time to complete the biochemical reactions. Secondly, the design must
ensure oxygen transfer to the bacteria. Energy requirements for oxygen
transfer are usually the main operating cost of a bioreactor, other than man-
power costs.
2.5.2.1 Ex Situ Treatment of Contaminated Water
Current designs for biological treatment of contaminated groundwater
are based on systems originally designed to treat wastewater. Bioreactors
for treating contaminated water can be separated into several main types:
• Suspended-growth reactors - The bacteria are grown in the water
and intimately mixed with the organic contaminants in the wa-
ter. Aerated lagoons or basins fall into this category. Oxygen is
2.9
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Process Summary
supplied with a surface aerator or air diffusers. Minimum resi-
dence time is about two days.
• Fixed-film reactors - Bacteria are grown on an inert support
medium within the reactor. The contaminated water passes over
the attached bacteria and forms a thin water film into which the
contaminants and oxygen diffuse. The bacteria degrade the
organic contaminants and waste by-products (CC^, t^O) dif-
fuse into the water film.
• Submerged fixed-film reactors - In this variation of fixed-film
reactors, the support medium is submerged in the water in the
reactor tank. The water is in constant contact with the bacterial
film, as opposed to passing through in thin water films.
• Reactors based on activated carbon - The combination of pow-
dered activated carbon and active bacteria increases the removal
capabilities of the treatment system. The activated carbon
adsorbs organic contaminants and acts as an attachment site for
bacteria. Another design, a fluidized-bed reactor, is basically a
submerged fixed-film system. The support medium consists of
small diameter particles. As water and air flow upwards through
the medium, the bed is fluidized. Recently, activated carbon has
been the main medium used in these systems.
• Miscellaneous designs - Currently, there are several anaerobic
reactor designs available. The main applications to date have
been in the food and beverage industries for treating wastewaters
with high concentrations of organic constituents.
2.5.2.2 Ex Situ Treatment of Contaminated Soils/Sediments
Ex situ biotreatment reactors for soil remediation fall into two main cat-
egories: slurry-phase treatment and solid-phase treatment.
Slurry-phase treatment involves maintaining contaminated soils or slud-
ges as an aqueous slurry. Solid-phase biotreatment includes land treatment,
soil-pile treatment, and composting. One of the major costs for all slurry
and solid-phase reactors is soil/sediment movement. Therefore, the most
economical use of bioreactors is when in situ treatment is not feasible and
excavation is required.
2.10
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Chapter 2
Either bioreactor vessels or lined lagoons may be used for slurry-phase
biotreatment. Basic features include aeration equipment, mechanical mix-
ing, and sometimes, an emissions control system. Soil or sludge is mixed
with nutrient-amended water to form an aqueous slurry. The system is
operated to maximize mass transfer rates and contact time between the con-
taminants and microorganisms.
The characteristics of solid-phase biotreatment processes follow:
• Ex situ land treatment - This process is sometimes known as a
prepared bed or on-site land-based bioremediation unit. An ex
situ unit operates under the same conditions as an in situ land
treatment unit with the possible addition of a leachate collection
system or liner to prevent migration and loss of contaminants.
• Soil-pile treatment - Two process variations exist: a water-based
system that delivers oxygen and nutrients by water movement
through the soil, and an aerated soil-pile system in which nutri-
ents are mixed with the soil when the pile is created and oxygen
is delivered through air pipes placed in the pile.
• Composting - Composting is a batch biological process used to
treat material with high concentrations of biodegradable organic
compounds. Waste destruction and conversion are achieved
with thermophilic, aerobic microorganisms that occur naturally
in decaying organic matter. However, high temperatures
(>50°C) rarely are achieved in composting hazardous waste.
Typical composting systems are the windrow, in-vessel, and
Beltsville.
2.5.2.3 Biological Reactors for Contaminated Air
Biofiltration is the biological removal of organic contaminants from gas
streams in a solid-phase reactor. The process is well-established in Europe
and Japan as an air pollution control technology.
A biofilter consists of one or more beds of biologically-active material,
primarily mixtures of compost, peat, or soil. Contaminated offgas is vented
through the filter, and air contaminants diffuse into the wet, biologically
active biofilm surrounding the filter particles. Complete biodegradation
results in production of CO , water, and microbial biomass. Oxidation of
2.11
-------
Process Summary
reduced sulfur compounds and chlorinated organic compounds also gener-
ates inorganic acids.
Biofilters are simple to operate and require little maintenance. The mi-
crobes need aerobic conditions and sufficient moisture. The filter medium
provides sites for adsorption of pollutants and attachment of microorgan-
isms and also supplies additional nutrients.
2.5.2.4 Potential Applications of Ex Situ Technologies
Use of ex situ technologies may be hampered by buildings or other struc-
tures on the site, the depth of the contamination, or location of contaminated
soils below the water table. Where excavation is feasible, aboveground
methods can be useful.
Contaminants that are more recalcitrant may be more effectively
remediated using slurry reactors in which environmental factors can be
controlled. For moderately-degradable compounds, particularly if con-
tained in large volumes of clayey soils, low temperature thermal desorption,
for volatile and semi-volatile compounds, may be more effective than soil-
piles.
Slurry reactors, especially if coupled with soil washing, can treat a wider
range of soils and contaminants than most other bioreactors. Liquid reac-
tors are suitable for treating groundwater recovered from pump-and-treat
systems; such reactors are particularly effective for treating soluble, non-
volatile, biodegradable organic contaminants.
Air bioreactors are used to treat offgas emissions from bioventing sys-
tems, in situ vapor recovery systems, soil reactors, and air stripper towers.
They are most applicable if the concentration of volatiles in the air phase is
moderate to low.
2.5.2.5 Limitations of Ex Situ Bioremediation Technologies
Soil permeability may limit oxygen transfer in soil-piles during treatment
of clayey soils. This problem may be overcome by adding bulking agents
or using slow-release oxygen compounds.
Slurry reactors require offgas capture if volatile constituents are present
above acceptable levels. This requirement may present a major challenge
for lagoons or other large systems. The efficacy of liquid reactor systems
2.12
-------
Chapter 2
may be limited by influent concentrations of organic contaminants that are
too high or low for cost-effective biodegradation or the need for constant
monitoring.
Use of air bioreactors may be limited by the size of the reactor required
to treat the air flow and mass of contaminants in the vapor phase. In some
cases, the size requirement of the reactor can be reduced by using another
method, such as a catalytic converter, during the first few weeks of opera-
tion when the concentration of volatile compounds is greatest. Alterna-
tively, the primary system could be made fully operational over a period of
several weeks.
2.6 Process Evaluation
Bioremediation processes depend on a living, biological system that runs
twenty-four hours a day, seven days a week. Biochemical processes oper-
ate best within optimum ranges of pH and temperature; the bacteria must be
supplied with material to enhance reaction rates (usually oxygen and nutri-
ents).
Most systems are designed to degrade the organic contaminants to CO2,
H2O and biomass, although persistant intermediates are typically produced.
Under anaerobic conditions, other by-products may result (anaerobic bio-
degradation of chlorinated aliphatic solvents produces lower substituted
chlorinated hydrocarbons and as final by-products, chloroethane or vinyl
chloride). Although such reactions occur naturally in the field,
bioremediation design should minimize the possibility that such unaccept-
able by-products would be formed.
The capital and operating costs of bioremediation depend on the type and
quantity of organic compounds present, site conditions, the volume of mate-
rial to be processed, and the site-specific remediation goals. The main di-
rect costs of bioremediation can be attributed to moving liquid or soils to
the reaction zone for the purpose of supplying oxygen to aerobic systems
and amending with nutrients.
2.13
-------
Process Summary
27 Technology Prognosis
The future of bioremediation hinges on several key issues. As these is-
sues are resolved, bioremediation will be used to treat more contaminants
under more demanding conditions.
• Regulatory concerns - Although recent changes in regulations
remove some restrictions on excavating, bioremediating, and
reusing soils on site, other regulations may restrict use of nutri-
ent amendments during in situ treatment. In the future, regula-
tions may also control the addition of genetically-altered micro-
organisms. Cleanup standards, particularly for total petroleum
hydrocarbons, may be overly conservative and difficult, if not
impossible, to achieve.
• Engineering considerations - Improvements related to delivery
of nutrients and/or electron acceptors will expand the conditions
under which bioremediation is cost-effective.
• New bioremediation approaches - White rot fungus, cometabolic
processes, anaerobic processes, methods for improving micro-
bial transport through porous media, use of vegetation, and ge-
netically-engineered microbes have been shown to be effective
in the laboratory and, in some cases, at pilot scale. Commercial-
ization will require additional effort.
• Additional data - Improvements are needed in methods of site
investigations, determination of nutrient and electron acceptor
requirements, biotoxicity avoidance, and materials handling.
Also needed are better methods for measuring contaminant con-
centrations and biological activity.
2.14
-------
Chapter 3
3
PROCESS IDENTIFICATION AND
DESCRIPTION
3. / Overview of Bioremediation
Bioremediation, for the purpose of this monograph, is the process by
which organic hazardous materials are biologically degraded, usually to
innocuous materials such as carbon dioxide, methane, water, inorganic
salts, biomass, and by-products that are less complex than the parent com-
pound. The process is basically an extension of the carbon cycle in which
organic and inorganic forms of carbon are cycled back and forth through
oxidation and reduction reactions. Figure 3.1 (on page 3.2) depicts the
concepts involved in bioremediation of organic contaminants. For
bioremediation in situ, biodegradation is effected by the indigenous microf-
lora. Ex situ treatment in bioreactors may entail inoculation with contami-
nant-degrading microorganisms.
The concept of bioremediation can be traced back centuries to such pro-
cesses as composting of organic wastes for soil conditioners and mulch
(Thomas, Ward et al. 1992). The technology has been expanded to include
the treatment of food wastes, agricultural wastes, and wastewater. More
recently, bioremediation concepts have been applied in treating hazardous
wastes and remediating contaminated soils and groundwater. Enhanced
bioremediation was used to treat industrial wastes as early as the 1940s
when the petroleum industry managed refinery wastes by land application
or treatment, which includes biological, as well as physical and chemical,
processes. Management of wastes by land application was unregulated at
that time.
3.1
-------
Process Identification and Description
Figure 3.1
Bioremedial System
Operational Conditions
• Nutrients • Delivery systems
• Electron acceptors for amendments
(O,, NO,-, CO2) • Bioavailability and
• pH Range degradation rates of
• Temperature impact contaminants
• Toxicity control
Outputs
• Gases
• Biomass
• Liquids
Bioremediation can be used to treat soil, water, and gases that contain biodegradable organic compounds Although
there are many bioremedial processes that can be considered, there are common factors that affect all such processes
These must be understood to obtain the desired performance The major factors are noted in the above schematic In
addition, there are outputs and residues that may require subsequent management
Bioremediation must be considered as a system with all of the inputs, outputs, and operational requirements carefully
considered. The difference in bioremediation processes, as applied, is due to the media (soil, water, gas) being
remediated, the site-specific operational requirements, and any outputs that must be managed.
The types of processes discussed in this monograph include natural and
enhanced bioremediation. Natural bioremediation is the process by which
contaminants are degraded by indigenous microorganisms that use available
nutrients and electron acceptors. As a result of natural bioremediation,
many contamination events may go unnoticed. When contamination is
detected, the concentrations of contaminants are usually lower than what
would be expected from the action of abiotic processes alone. Often, how-
ever, the rate of natural bioremediation is limited by nutrient and electron
acceptor availability, and the contaminants pose human health and environ-
mental risks. In these instances enhanced bioremediation is applied. En-
hanced bioremediation is the process by which the rate of contaminant deg-
radation is increased by adding an electron acceptor, nutrients, or other
countering factors that are limiting.
Enhanced bioremediation processes can be used to treat contamination in
situ or ex situ (figure 3.2 on page 3.3). In situ processes include treatment
of contaminants in the tillage zone (land treatment) and unsaturated and
saturated subsurface zones. Land treatment relies on the physical, chemi-
cal, and biological processes for attenuation of contamination. Following
are the desired results of land treatment:
3.2
-------
Chapter 3
Figure 3.2
Applicable Bioremediation Processes
PROCESSES BY MEDIA
EX SITU PROCESSES
• Prepared Bed (Soil/Sludges)
• Slurry Reactor (Soil/Sludges)
• Pile/Composting (Soil)
• Biofilter (Offgas VOC)
• Fluidized Reactor (Liquid)
• Wastewater Treatment Plant (Liquid)
Unsatu rated
(Vadose) Zone
• Contaminated Soils
Light non-aqueous phase
liquids (LNAPLs)
>ar Zone)
Saturated Zone
• Contaminated Soil
and Oroundwater
Dense non-aqueous
phase liquids (DNAPLs)
IN SITU PROCESSES
• Tillage
• Bioventing
Air Sparging
In-Situ Biological
• immobilization of the waste by the soil;
• stimulation of contaminant degradation by indigenous microor-
ganisms or by inoculation;
• minimization of volatilization and leaching of contaminants out
of the treatment zone; and
• control of surface water runoff (American Petroleum Institute
1983).
Technologies for treating contamination in the unsaturated and saturated
zones include bioventing, air sparging, and biofilters (figure 3.2).
Bioventing is the process by which contaminants in the unsaturated zone are
removed by volatilization and biodegradation as oxygen is supplied by
3.3
-------
Process Identification and Description
vacuum extraction and/or injection. The key objective in bioventing is to
maximize the rate of contaminant removal by volatilization and biodegrada-
tion or minimize cost of operation by promoting biodegradation versus
volatilization. Air sparging entails injecting air into the saturated zone to
enhance biodegradation of aquifer contaminants and/or transfer of volatile
groundwater contaminants to the unsaturated zone for treatment. By de-
signing systems to deliver limiting nutrients and a terminal electron accep-
tor in solution, contaminants in the saturated zone can be treated.
Ex situ processes include those for treating contamination in liquids,
solids, and air (see figure 3.2 on page 3.3). Processes for treatment of liq-
uids use suspended-growth, fixed-film, and submerged fixed-film
bioreactors. Each type of reactor is designed to maximize contact of micro-
organisms, contaminants, and required nutrients to increase rates of con-
taminant biodegradation. Solids can be treated using slurry phase pro-
cesses, land treatment, or composting/soil piles. Each of these processes is
managed in a manner designed to maximize rates of biodegradation and/or
minimize effluent concentration. Biofilters are composed of a solid phase,
such as compost, peat, or soil, through which a stream of contaminated air
is passed. Vapor-phase contaminants in the air stream are biodegraded
aerobically by microorganisms in the solid phase.
3.2 Fundamentals and Basic Science
3.2.1 Microbial Ecology and Physiology
The primary agents of bioremediation are the heterotrophic bacteria and
fungi, which derive the carbon and energy required for growth from organic
compounds. They are the primary agents of decomposition of natural or-
ganic matter in the biosphere and of organic pollutants that resemble the
natural substrates of these organisms. Organic contaminants may be de-
graded by a single type of microorganism, but more likely, by complex
mixed cultures, the components of which may require special environmen-
tal conditions. It is critical that the biotreatment system be designed with an
understanding of the biological processes involved. Understanding meta-
bolic pathways allows evaluation of the extent of biodegradation, intermedi-
3.4
-------
Chapter 3
ate metabolites that might accumulate, and requirements (such as electron
acceptors) that must be fulfilled for successful bioremediation.
The range of biodegradative processes available and the complexities of
the contaminants must be matched. Microorganisms with specialized meta-
bolic capabilities degrade only particular contaminants; complex microbial
communities possessing an array of biodegradative capabilities are site-
specific and cannot be considered as generic "biomass." Similarly, con-
taminants include a wide spectrum of functional groups and chemical prop-
erties and should not be considered as "organic loading."
3.2.1.1 Organic Contaminants as Growth Substrates for Micro-
organisms
The ideal application of bioremediation would be in treating natural or-
ganic compounds, such as petroleum hydrocarbons, phenols, cresols, ac-
etone, and cellulosic wastes, all of which can serve as growth substrates for
microorganisms (table 3.1 on page 3.6). The biodegradation pathways for
most of these compounds have been studied extensively and are well-de-
fined (Gibson 1984). They can be converted to carbon dioxide, water, and
microbial biomass with no accumulation of by-products or metabolites.
This process, called mineralization or complete biodegradation, is a major
component of the global carbon cycle. It is a self-sustaining process and
requires only the appropriate conditions of temperature, pH, moisture con-
tent, inorganic nutrients, and an electron acceptor. The biomass regenerates
itself through growth and requires no source of energy other than the or-
ganic compound. Such processes have formed the basis of municipal and
industrial waste treatment systems for many years. The processes have also
been used for bioremediation of creosote in wastes from wood treatment, as
well as petroleum hydrocarbons in refinery wastes, oil spills, and subsurface
material contaminated by fuels from leaking underground storage tanks.
In recent years an increasing number of synthetic organic (xenobiotic)
compounds have entered the environment and become subject to mineral-
ization. A variety of chloroaromatic and nitroaromatic compounds can be
readily mineralized if the appropriate microorganisms and conditions are
present. Indeed, many pesticides and detergents are now designed to be
mineralized in the environment in contrast to earlier products that were very
resistant to biodegradation.
3.5
-------
Process Identification and Description
Table 3.1
Examples of Naturally-Occurring Organic
Compounds Amenable to Bioremediation
Chemical Group Molecular Formula
Petroleum Hydrocarbons
Aliphatic (straight chain) C16H14 (hexadecane)
Aromatic (ring) C6H6 (benzene)
Phenols (hydroxylated ring) HOC6H^ (phenol)
Cresols (methylated phenol) HOC6H4CH3 (o~,m-, orp-cresol)
Acetone (ketone) CHjCOCH-,
Cellulose (polysacchande of repeating C6H10OS
glucose units)
3.2.1.2 Requirements for Microbial Growth and Metabolism
Harnessing a natural microbial community for use in bioremediauon re-
quires a good working knowledge of its nutritional and environmental re-
quirements. These requirements involve not only the basic parameters such
as temperature, pH, and inorganic nutrients common to all biological pro-
cesses, but also specific requirements for the degradation of a given pollut-
ant by a specific biochemical strategy. A brief discussion of general re-
quirements is provided below and is followed by a more detailed discussion
of several degradative processes of potential use in bioremediation.
3.2.1.2.1 Temperature Thermophilic bacteria can be useful for biodegra-
dation at temperatures ranging from 40° to 60°C (104° to 140°F), but so far
their use has been primarily limited to composting applications. Biodegra-
dation can also be detected at temperatures as low as 0° to 10°C (32° to
50°F); even though biodegradation rates are slower at low temperatures,
successful bioremediation at low temperatures has been reported. Tempera-
tures in the mesophilic range, between 10° and 40°C (50° to 104°F), are
more practical for field applications, and in some instances, contaminated
materials have been heated artificially to this range as a prerequisiite for
biotreatment. Within the mesophilic range, microbial growth rates and
biodegradation rates increase with increasing temperature. In the prediction
3.6
-------
Chapter 3
of rates, it must be borne in mind that the agents of biodegradation do not
constitute a simple catalyst, but a complex community of microorganisms.
Thus, a shift in temperature of a few degrees can cause dramatic changes in
the composition and function of the community. The specific population,
which is able to degrade the contaminant, may function over a narrow range
of temperatures or may be replaced by populations with different degrada-
tion kinetics or mechanisms. Therefore, the linear relationship that results
from a simple Arrhenius plot of reaction rate (Iog10 k) vs. temperature (1/T)
may not apply in many instances and can serve only as a general guideline;
prediction of changes in biodegradation rates as a function of temperature
must be specific to the particular site and situation. Avoiding temperature
fluctuations will improve the effectiveness of bioremediation.
3.2.1.2.2 Nutrients Inorganic nutrients, primarily nitrogen and phospho-
rus, are essential for all biological processes. Nitrogen can be provided in a
variety of forms such as nitrate, ammonium salts, and organic compounds,
such as urea. Several forms of inorganic phosphorus have been used. For in
situ processes, the choice is usually based on site geochemistry because of
potential interactions between phosphate and cations in the water or soil.
Treatability studies are usually conducted to determine nutrient require-
ments, and results are specific to the particular site and process. In early
bioremediation projects, carbon loading was determined and nitrogen and
phosphorus were provided in ratios known to be optimal for municipal
waste treatment systems. Such high nutrient levels are not necessary in
bioremediation applications where biomass is not lost from the system. For
example, in fixed-film systems (all in situ treatments involve fixed films)
most nutrients are recycled and retained in the system. In a well-designed
process, much of the organic contamination will be mineralized and biom-
ass accumulation and wasting will be minimized.
3.2.1.2.3 pH For bioremediation processes, the optimum pH will be site-
and process-specific and must be determined empirically during feasibility
studies; usually few problems are encountered within a pH range of 6 to
8.5. Because many biodegradation processes produce acids or bases, the
buffering capacity of the system must be sufficient, or neutralizing agents
must be added to maintain an optimum pH range. For example, degradation
of chlorinated solvents produces hydrochloric acid, and fermentation of
sugars produces organic acids.
3.7
-------
Process Identification and Description
3.2.1.2.4 Electron Acceptors Much of the energy for growth of microor-
ganisms is obtained during the transfer of electrons from organic substrates
to inorganic electron acceptors. Therefore, appropriate electron acceptors
are an absolute requirement for biodegradation, and provision of these elec-
tron acceptors often constitutes the greatest challenge in the design of in
situ bioremediation systems. The common electron acceptors are carbon
dioxide, sulfate, nitrate, and oxygen. In some instances, halogenated or-
ganic contaminants can serve as electron acceptors when they are used as
substrates for reductive dehalogenation. It is important to note that the
electron acceptors listed above are not interchangeable. Although some
bacteria can use either oxygen or nitrate as the terminal electron acceptor,
the microbial communities are distinctly different for each type of electron
acceptor. Their metabolic processes and their potentials for biodegradation
of pollutants are also very different. The presence of a given type of com-
munity will depend on the availability of the inorganic electron acceptors
and on the oxidation-reduction (redox) potential of the system. The redox
potential that supports oxygen-based metabolism is higher than that re-
quired for nitrate-based systems; sulfate- and carbon dioxide
(methanogenesis)-based systems require the lowest redox potentials and are
strongly inhibited by oxygen (see figure 3.3 on page 3.9). Microbial com-
munities change naturally as the availability of electron acceptors changes.
Bouwer (1992) provides an excellent discussion of the relationship between
redox potentials and biodegradation.
An additional consideration in applying oxygen-based systems; is the use
of molecular oxygen, not only as the terminal electron acceptor in metabo-
lism, but also for initial enzymatic oxidation of organic molecules. This is
particularly important in the case of hydrocarbons that are resistant to deg-
radation in the absence of oxygen. A few hydrocarbons such as toluene and
xylenes, and many partially-oxidized organic compounds, such as alcohols
and acids, can be mineralized under anoxic conditions. The rates of degra-
dation are slower than the corresponding aerobic processes, and the biomass
yields are lower because the anaerobic pathways yield less energy. Ali-
phatic hydrocarbons are not known to be biodegraded under anoxic condi-
tions. The oxygenase enzymes, which insert one or both atoms of molecu-
lar oxygen into an organic compound to yield hydroxyl groups, evolved to
attack naturally-occurring hydrocarbons. These enzymes can also catalyze
the initial oxidation of a variety of xenobiotic compounds, such as many of
the chloroaromatic, chloroaliphatic, and nitroaromatic compounds.
3.8
-------
Chapter 3
Figure 3.3
Oxidation-Reduction Potentials of Microbial Electron Transfer Reactions
Redox Potential
[mV]
-600-1
-AG°(pH7)
[kJ/mole-] pE°(pH7)
HS
1000-J
--20
-0
-+20
-+40
-+60
-+80
L + 100
L + 120
r -10
- -5
- 0
- +5
- +10
L- +15
Reprinted, by permission, from Chapter 1 Geochemistry and Biochemistry of Anaerobic Habitats by A J B. Zehnder
and W Stumm in Biology ol Anaerobic Microorganisms, 19, edited by A J B Zehnder, John Wiley and Sons,
Publisher Copyright 1988 by John Wiley and Sons, Inc
The use of nitrate as a terminal electron acceptor (denitrification and
nitrate reduction) shows considerable promise for in situ bioremediation of
aromatic hydrocarbons. Denitrification involves the reduction of NO3
through the following sequence: NO3~, NO2, NO, N2O, N2. Nitrate reduc-
tion involves the reduction of NO3 without the formation of N2O or Nr
Because nitrate is more soluble than oxygen in water, it may be useful for
subsurface bioremediation where transport of oxygen is limited. Low levels
of oxygen do not inhibit denitrification, and there is some evidence of syn-
ergy between the two processes. Furthermore, heterogeneous micro-envi-
ronments may exist and allow regions of low redox potential in close prox-
imity to aerobic regions in which denitrification may be inhibited. There-
3.9
-------
Process Identification and Description
fore, the addition of nitrate might allow conservation of oxygen during
treatment, but this possibility has not been rigorously proven.
Reductive dehalogenation reactions, which involve replacement of a
halogen with a hydrogen atom, seldom take place in the presence of oxy-
gen, but are favored under anoxic conditions. In some instances, such as
with chlorobenzoate, the reaction can be shown to be specific and to yield
energy (Mohn and Tiedje 1991). In others, such as reduction of polychlori-
nated biphenyls (PCBs) or chloroaliphatic compounds, the reactions are
poorly understood. The choice of a final electron acceptor dictates the
range of potential biotransformation reactions that can be expected in a
given system. This concept will be discussed in greater detail, and specific
examples will be provided.
3.2.1.2.5 Contaminant Bioavailability Several factors that affect the
interactions between organic contaminants and the microbes responsible for
their degradation pose major concerns in the design of a bioremediation
system. Microorganisms may have the metabolic capability to mineralize a
substance and yet fail to do so because it is insoluble, sorbed, or otherwise
unavailable to the cell. Biodegradation of polycyclic aromatic hydrocar-
bons (PAHs) is often limited because the contaminants are not available to
the biomass (Stucki and Alexander, 1987). Bioavailability is a particularly
important consideration in the design of in situ treatment systems in which
the contaminant may be localized as nonaqueous phase liquid (NAPL).
Some insoluble hydrocarbons can be degraded in aqueous solutions by mi-
crobes in direct contact with the surface of the hydrocarbon, whereas others
must be dissolved in the aqueous phase before appreciable biodegradation
will occur. Microorganisms often can produce surfactants to aid in the
solubilization of poorly soluble or immiscible substrates, and some consid-
eration has been given recently to the use of synthetic surfactants to aid the
process. The literature on the degradation of sorbed organic compounds is
inconclusive; some studies suggest that sorbed substances must desorb be-
fore they are biodegraded, and others suggest that some bacteria can take up
certain compounds directly from the sorbed state (Mihelcic et al. 1993).
Therefore, contaminant bioavailability must be considered for each site and
compound because there is little consensus about which factors influence
bioavailability and how bioavailability affects bioremediation.
3.10
-------
Chapter 3
Two other aspects of substrate availability are also important. Almost all
organic contaminants that are candidates for bioremediation are toxic at
very high concentrations. But, most of the higher molecular weight hydro-
carbons are not toxic at their solubility limits in water. Benzene, however,
is more toxic to microorganisms than other hydrocarbons and can only be
biodegraded within a certain concentration range. Fortunately, the aqueous
concentrations commonly encountered in petroleum-contaminated sites are
well below the toxic range. Toxicity might become an important consider-
ation only in sites contaminated with pure benzene. Chlorinated solvents
and nitroaromatic compounds also may be toxic at low concentrations. In
contrast, PCBs are relatively nontoxic to microorganisms. Toxicity also
depends on bioavailability and compounds that are sorbed are often less
toxic.
Microorganisms vary widely in their sensitivity to toxic organic com-
pounds; often, a microbial community can adapt and become resistant to
high concentrations of a toxic compound. In treatability studies, however, it
is very important to determine the toxicity of a waste to the specific organ-
isms able to degrade it and not to the general biomass. Thus, measurement
of biodegradation by a microbial community at a variety of contaminant
concentrations is a much more important measurement of toxicity than sur-
vival or plate counts.
One of the limitations of bioremediation is inherent in the nature of the
enzyme reactions involved. The kinetics of enzyme induction and substrate
binding and transport dictate that there will be a threshold substrate concen-
tration below which biodegradation rates will be negligible. Contaminants
initially present at very low concentrations may not be biodegraded at all.
Therefore, concentrations of contamination might be reduced sufficiently
by the time biodegradation stops. This must be carefully considered in
design of treatment systems and in extrapolation of biodegradation rates
measured at one concentration to predict rates at different concentrations.
3.2.1.3 Microbial Metabolism of Contaminants that are not
Growth Substrates
Almost all simple organic compounds are susceptible to some type of
transformation catalyzed by microbial enzymes. Structural complexity or
substitution with halogen, nitro, or other functional groups can produce
synthetic chemicals that cannot serve as growth substrates for microorgan-
3.11
-------
Process Identification and Description
isms. Some compounds can undergo the initial reactions in a degradative
pathway and be converted to intermediates that accumulate because they are
not susceptible to attack by subsequent enzymes in the pathway. Such
transformations can be useful if the contaminant is rendered less toxic or
more bioavailable. Because simple transformations do not yield energy, the
microorganisms involved will require an added organic substrate to support
growth and activity. Other compounds fail to serve as growth substrates
because they do not induce the synthesis of appropriate enzymes. This
class of compounds can be degraded if an inducer of the catabolic pathway
can be added.
These fortuitous reactions result from the broad substrate specificity of
some microbial enzymes and have been called cometabolism or cooxidation
in the past. But, because such processes can occasionally be shown to yield
carbon and energy and convert the contaminant to carbon dioxide (mineral-
ization), cometabolism seems to be less a distinct phenomenon. It is more
likely to prevail in natural microbial communities where low concentrations
of complex mixtures of organic compounds provide carbon and energy.
Growth of microorganisms on high concentrations of a single carbon source
is probably a laboratory artifact that bears little relation to natural ecosys-
tems. Because of the many types of transformations that can be considered
cometabolism or cooxidation, these reactions have been consolidated under
one category of "organic compounds that cannot serve as growth sub-
strates" for the purpose of this monograph. It should also be noted that
many synthetic compounds believed to be in this category have recently
been shown to serve as growth substrates for specific strains of bacteria.
Modern microbial strain construction and selection techniques are rapidly
providing new strains with novel capabilities.
3.2.1.3.1 Oxidation The most widely-used microbial transformations in-
volve the oxidation of organic compounds by the introduction of a hydroxyl
group derived from molecular oxygen. Such reactions usually are catalyzed
by nonspecific monooxygenase and dioxygenase enzymes that insert either
one or both atoms of molecular oxygen into the substrate. Such en2:ymes
catalyze the first step in the mineralization of the relatively inert hydrocar-
bons. Microbial cells grown on hydrocarbons, such as methane or toluene,
can oxidize a wide range of organic compounds. An important example is
the use of such reactions for the degradation of trichloroethylene, which
decomposes spontaneously after the addition of molecular oxygen. Alterna-
3.12
-------
Chapter 3
lively, hydroxyl groups can be introduced in nonpolar contaminants to in-
crease their solubility, but such an approach must be undertaken cautiously
because hydroxylation can increase the toxicity of many contaminants.
Oxygenase enzymes can also catalyze the displacement of a wide range of
aromatic substituents such as carboxyl, nitro, chloro, ether, and sulfonic
moieties.
3.2.1.3.2 Hydrolysis Microbial hydrolases, such as esterases, phos-
phatases, and Upases, can be used to detoxify or solubilize a variety of con-
taminants. Because hydrolysis of contaminants provides no energy for the
cells, they must be supplied with organic growth substrates. Alternatively,
cells can be immobilized and provided with low levels of a growth substrate
to support cell maintenance. Because hydrolases are usually stable extra-
cellular enzymes that do not require additional substances for their activity
(cofactors), they are also excellent candidates for use as immobilized en-
zymes. The most likely application of microbial hydrolases will be in treat-
ment of pesticide contamination, because many of the currently available
pesticides, such as parathion and diazinon, contain hydrolyzable ester link-
ages. For such compounds, hydrolysis eliminates or greatly reduces the
toxicity.
3.2.1.3.3 Reductive Dehalogenation Replacement of a halogen atom by a
proton (reductive dehalogenation) is catalyzed by a variety of anaerobic
microbial systems. An example is shown in the following equation:
CC14 (carbon tetrachloride) + 2H -> CHC13 (chloroform)+ HC1
Methanogenic systems, the most widely studied, can dehalogenate con-
taminants ranging from PCBs to tetrachloroethylene. The reactions seem to
be specific for certain isomers in some systems, and in a few cases, the
responsible organisms have been studied in pure culture (Fathepure and
Boyd 1988). The enzymology of the reactions is poorly understood at
present, but recent evidence indicates that the bacteria can derive energy
from the use of halogenated organic compounds as an electron sink.
Highly-chlorinated substrates tend to be more susceptible to reductive
dehalogenation than their lower chlorinated analogs. Therefore,
tetrachloroethylene is dehalogenated more rapidly than trichloroethylene,
which is, in turn, more readily attacked than vinyl chloride (Freedman and
Gossett 1989). The same sort of relationship applies to the isomeric
3.13
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Process Identification and Description
chlorophenols and the PCBs. This relationship is in direct contrast to that
of the oxygenase enzymes mentioned previously, in which more highly-
halogenated substrates are less susceptible to attack.
3.2.1.3.4 Reduction ofNitro Groups Nitro groups are the primary func-
tional groups of a variety of pesticides, dyes, and munitions. Many micro-
organisms synthesize enzymes that catalyze the reduction of nitio groups to
the amino level. The reactions with munitions, such as trinitrotoluene
(TNT), have been particularly well-studied (McCormick et al. 1978) as
have those with dinitrotoluenes (McCormick, Feherry, and Levinson 1976).
An example of a reduction of a nitro group follows:
C7H6N2O4 (dinitrotoluene) -» C?H10N2 (diaminotoluene)
The reductions can take place under aerobic, as well as anaerobic, condi-
tions and often yield metabolites resistant to further degradation. The reac-
tions studied to date seem to be nonspecific, and the microorganisms re-
sponsible require only a source of carbon and energy.
3.2.1.4 Sources of Microorganisms
3.2.1.4.1 Indigenous Microorganisms Microbial communities in soil typi-
cally have a tremendous potential for the degradation of a wide range of
natural chemicals. Thus, the "natural assimilative capacity" can lead to the
degradation of considerable amounts of contamination without any inter-
vention (see Section 3.4 on Natural Assimilative Capacity). This is particu-
larly true for petroleum hydrocarbons; the highly-successful application of
bioremediation to the treatment of gasoline from leaking underground stor-
age tanks has been carried out almost exclusively with indigenous microor-
ganisms. During the design of a bioremediation system, it is essential that
the presence of appropriate organisms be established. This can typically be
done during the treatability study when microbial communities from the
contaminated site are tested for their ability to remove the contaminant
under laboratory conditions. It is only necessary to measure the activity of
the microorganisms as shown by the degradation of the contaminant. Indi-
cators of microbial activity include numbers of contaminant-degrading or-
ganisms, measurements of contaminant degradation, C<2 uptake, and COa
evolution. Excessive amounts of time and money can be spent on enumera-
tion and identification of bacteria in samples from contaminated sites. Al-
3.14
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Chapter 3
though enumeration may be useful as a screening tool, the results seldom, if
ever, provide information useful in the subsequent design or operation of
the system. Low-microbial activity suggests that some environmental con-
dition is not conducive to rapid contaminant biodegradation and not that the
necessary microbes are absent. Occasionally, a low population of specific
organisms will produce a low initial degradation rate that will increase after
a lag or an acclimation period of a few days. During the acclimation period,
the population increases dramatically. If the results of the treatability study
indicate that the acclimation period is lengthy, it must be considered in the
design of the treatment system.
A wide range of microbial products are marketed for stimulating biodeg-
radation. They are designed to reduce the acclimation period, increase the
degradation rate, or increase the range of contaminants removed during
bioremediation. There is no evidence in peer-reviewed literature that any of
these products has been effective in rigorously-controlled studies. Natural
organic compounds, such as petroleum hydrocarbons, are degraded readily
by indigenous microbes, and added strains are not likely to compete suc-
cessfully with the indigenous community. There is substantial interest in
the development of specialized microbes, but rigorous experimentation will
be required to reveal a successful application.
3.2.1.4.2 Inoculation with Nonindigenous Strains The biodegradation
potential of indigenous microbial communities, discussed earlier, applies
largely to contamination with natural organic compounds. The potential for
degradation of synthetic chemicals may not be present in the indigenous
microbial community, particularly if the contamination is recent. Neverthe-
less, a wide range of synthetic chemicals can be biodegraded, although
microorganisms with the appropriate metabolic capabilities may not be
widely distributed in the field. This situation would be apparent during
treatability studies in which no degradation (or extended delays in the onset
of appreciable degradation) reveal the absence of strains able to degrade a
contaminant known to be degraded in other systems. In such instances,
inoculation with capable strains or starter cultures might prove helpful.
Startup of a bioreactor for removal of chloroaromatic or nitroaromatic con-
taminants, methylene chloride, pesticides, or creosote wastes may benefit
from inoculation with such strains. In situ subsurface applications would be
less likely to benefit from inoculation because transport of microorganisms
3.15
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Process Identification and Description
is limited in porous media. For inoculation to be useful, it must be demon-
strated that the degradative strains compete successfully with the indig-
enous microbial community during the treatability study.
3.2.1.4.3 Bioaugmentation In some instances, bioremediation can be car-
ried out by nonindigenous microorganisms that cannot use the contaminant
as a growth substrate (see Subsection 3.2.1.3 on Microbial Metabolism of
Contaminants that are not Growth Substrates). This specific process is
called bioaugmentation. The microorganisms must be added continuously
or supported on a secondary growth substrate. Delivery and maintenance of
the appropriate strains is the most daunting technical challenge in such sys-
tems. For example, pentachlorophenol has been treated successfully in soil
bioreactors by adding bacteria that completely degrade but do noi. use the
contaminant as substrate (Crawford and Mohn 1985). The system required
the regular addition of active biomass grown on another substrate.
Some bioremediation systems for contaminants that cannot serve as a
growth substrate will clearly require the addition of active biomass. Thus,
the addition of a novel constitutive strain of Pseudomonas cepacia capable
of degrading trichloroethylene (TCE), shows considerable potential for
bioremediation (Shields 1991). Because TCE cannot be used as the sole
source of carbon and energy by the P. cepacia, periodic renewal of the bio-
mass or addition of a secondary substrate growth will be necessary. Similar
considerations may apply to the use of specific methylotrophs, that are able
to degrade TCE. Other bioremediation strategies involving cometabolism
rely on stimulation of indigenous methylotrophs (Hazen 1991). Several
such approaches are in the development stages and may prove useful in the
future.
The white rot fungus, Phanerochaete chrysosporium, can biodegrade a
wide range of organic compounds by the use of nonspecific extracellular
peroxidases. It does not use the organic compounds as a source of carbon
and energy and must be grown on specific substrates. The lignin peroxi-
dase-mediated biotransformations only occur under conditions of nitrogen
limitation and in stationary phase.
Another form of bioaugmentation is the use of immobilized cells or en-
zymes. Contaminant-degrading organisms are immobilized in or on a solid
matrix such as alginate, hollow glass fibers, or porous diatomaceous earth
(Stroo 1992). The immobilization confers some protection from toxic lev-
3.16
-------
Chapter 3
els of contamination and microbial decay. Bioaugmentation has seen lim-
ited application for environmental cleanup, but has been very useful in the
chemical and pharmaceutical industries. It may provide advantages for
bioremediation with genetically-engineered organisms or for treatment of
waste streams toxic to microorganisms.
3.2,1.5 Biodegradation of Synthetic Chemicals
Most bioremediation to date has focused on treatment of natural organic
compounds, primarily petroleum hydrocarbons. Many synthetic organic
compounds can also be biodegraded or transformed by microorganisms, and
a number of strategies are being developed to take advantage of these capa-
bilities. This section briefly describes the current stage of development of
technologies for bioremediation of representative synthetic chemicals.
3.2.1.5.1 Pesticides For the last twenty years, pesticides have been de-
signed to degrade in the environment, and a considerable amount of infor-
mation is available on degradation kinetics in soil and water. Less informa-
tion is available on bioremediation of sites contaminated by accidental re-
leases where concentrations may be much higher than those in routine field
applications. Recently, the EPA's Risk Reduction Engineering Laboratory,
Cincinnati, Ohio, developed a Pesticide Treatability Data Base that will be
useful in evaluating the feasibility of bioremedial applications for treatment
of various pesticide wastes.
Most pesticides contain a chemical bond that can be hydrolyzed by mi-
crobes or abiotic reactions to yield harmless breakdown products. In some
instances, the products can serve as growth substrates for microorganisms,
which lead to "acclimation" or selection of a population of specific degrad-
ers able to degrade the pesticide at increased rates. For example, carbam-
ates, chlorophenoxyacetates, dinitrocresol, and some organophosphates can
serve as growth substrates for soil bacteria. A number of examples of this
phenomenon, as well as a good general discussion of pesticide biodegrada-
tion, are provided in Racke and Coats (1990). Any pesticide known to
serve as a growth substrate for bacteria or stimulate acclimation in soil
would be a good candidate for bioremediation.
Other pesticides offer no selective advantage to the microbes that cata-
lyze their degradation. For example, some compounds hydrolyzed by extra-
cellular enzymes yield no degradable products or yield products of no use to
3.17
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Process Identification and Description
the strain that catalyzed the hydrolysis. Other examples are the
organohalogen insecticides that are subject to reductive dehalogenation in
soil, but provide no advantage to specific microbes. In such situations,
acclimation is not observed and degradation rates are often proportional to
the total biomass. Recent discoveries that some anaerobes can derive en-
ergy from reductive dehalogenation (Mohn and Tiedje 1991) may render
this generalization obsolete.
3.2.1.5.2 Chloroaromatic Compounds Biodegradation of chloroaromatic
compounds has been studied extensively (Reineke and Knackmuss; 1988).
Many were thought to be resistant to biodegradation; however, bacteria able
to degrade all but the most complex molecules have been discovered during
the last twenty years. Chlorobenzenes are the simplest chloroaromatic com-
pounds and, until recently, were not considered to be biodegradable. Bacte-
ria have been isolated that are capable of growing on chlorobenzene
(Reineke and Knackmuss 1984; Nishino et al. 1992), 1,4-dichlorobenzene
(Schraaetal. 1986; Spain and Nishino 1987), 1,3-dichlorobenzene (deBont
et al. 1986), 1,2-dichlorobenzene (Haigler and Spain 1989), 1,2,4-
trichlorobenzene (van der Meer et al. 1987), and 1,2,4,5-tetrachlorobenzene
(Sander et al. 1991). The metabolic pathways for degradation of the iso-
meric chlorobenzenes are remarkably similar (Spain 1990) and lead to min-
eralization with release of the halogens as HC1. Chlorobenzene is the only
one of the compounds listed above that has been treated successfully by
bioremediation (Nishino et al. 1994). Preliminary results indicate that after
long-term contamination, indigenous strains are able to degrade chloroben-
zene under aerobic conditions (Nishino et al. 1992). Similar results have
been obtained in the laboratory with dichlorobenzenes (Spain unpublished),
suggesting that other isomers may exhibit similar biodegradative character-
istics. Laboratory and preliminary field work suggest that chlorobenzenes
are excellent candidates for bioremediation. Bacteria able to degrade chlo-
robenzenes may not be ubiquitous; therefore, recently contaminated sites
may require extended acclimation periods or inoculation with competent
strains. The design of the system must also include the capacity for neutral-
ization of the acid produced during chlorobenzene degradation if contami-
nant concentrations are high or if a vapor-phase system is used.
3.2.1.5.3 Polychlorobiphenyls (PCBs) Aerobic (Bedard et al. 1987),
anaerobic (Brown et al. 1987; Quensen, Tiedje and Boyd 1988), and combi-
3.18
-------
Chapter 3
nation systems (Bedard and Quensen, in press) have been described for the
degradation of PCBs. Anaerobic systems are most effective with the more
highly-chlorinated congeners whereas aerobic microorganisms work best
with the less-chlorinated congeners. No strains with the ability to use com-
plex PCBs as a growth substrate have been isolated. Therefore,
bioremediation of PCB-contaminated material is slow and complex. The
difficulties are offset, however, by the lack of alternative treatment options
for cleaning up large volumes of materials contaminated with low concen-
trations of PCBs. Pathways initiated by dioxygenase attack seem to be the
primary mode of aerobic degradation of PCBs. New bacterial strains are
being developed and tested in several laboratories. Their efficacy in the
field treatment of PCB-contaminated materials remains to be demonstrated.
3.2.1.5.4 Chlorophenols and Chlorobenzoates Oxidized chloroaromatic
compounds, such as the chlorobenzoates and chlorophenols, can often be
biodegraded by soil microorganisms. Unfortunately, chlorophenols are
toxic to bacteria at high concentrations and can inhibit their own biodegra-
dation. They can be degraded anaerobically via pathways initiated by re-
ductive dehalogenation (Boyd and Shelton 1984; Mikesell and Boyd 1986,
1988). Aerobic degradative pathways are initiated either by hydrolytic
removal of a halogen (Apajalahti and Salkinoja-Salonen 1986) or by
oxygenase attack on the ring and subsequent removal of the halogen after
ring fission (Reineke and Knackmuss 1988).
Although chlorophenols are biodegradable in laboratory studies and in
activated sludge, the bacteria responsible are not uniformly distributed.
Therefore, recent contamination may require long acclimation periods. In-
oculation with adapted strains, as previously noted, may be useful for chlo-
robenzenes. The toxicity of chlorophenols dictates that the bioremediation
system be designed to maintain low concentrations of the contaminant. The
release of HC1 requires provisions for neutralization of the medium during
the treatment of high concentrations.
Pentachlorophenol deserves special mention because of its widespread
use as a wood preservative. Aerobic biodegradation by bacteria is initiated
by a hydrolytic removal of the para chlorine to form tetrachlorohydro-
quinone (Steiert and Crawford 1986). Subsequent reductive dehalogen-
ations lead to ring fission substrates. Anaerobic degradation proceeds via
sequential reductive dehalogenations, leading to mineralization under
3.19
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Process Identification and Description
methanogenic conditions (Mikesell and Boyd 1986, 1988). Pentachlo-
rophenol has been the target of several successful bioremediation demon-
strations in the field (Stinson et al. 1991). In the United States,
bioaugmentation with a Flavobacterium sp. has proven successful
(Crawford and Mohn 1985). An alternative process involving mineraliza-
tion by indigenous strains or a consortium used only as an initial inoculum
has also been effective (Compeau, Mahaffey, and Patras 1991). In Europe,
an alternative process involving composting has been used (Valo and
Salkinoja-Salonen 1986). At wood treatment facilities, the contamination is
often a complex mixture including solvents, creosote, metals, and PAHs
that can inhibit microorganisms. Therefore, site-specific treatability studies
must be conducted for each situation.
3.2.1.5.5 Chloroaliphatic Compounds Chlorinated solvents are among
the most common contaminants at waste sites and in groundwater. They
are so volatile that they are seldom a problem in soil and surface water.
When their volatilization from a contaminated matrix is inhibited, they can
be very persistent because they are relatively stable and resist biodegrada-
tion. Chloroaliphatic compounds with only one or two chlorine substituents
can serve as growth substrates for microorganisms under appropriate condi-
tions. More heavily-chlorinated compounds can only be biodegraded by
microbes provided with an alternate growth substrate. Bioremediation in
the latter case is much more complex, but the magnitude of the problem and
the lack of alternative treatment options have generated a tremendous
amount of research on biodegradation of TCE and related solvents. An
excellent review on the biochemistry of TCE degradation has been pub-
lished (Ensley 1991).
Trichloroethylene was considered inert to bacterial degradation until the
discovery that cultures grown on methane (methylotrophs) could also oxi-
dize TCE (Wilson and Wilson 1985). The initial step in methane degrada-
tion is an oxidation catalyzed by methane monooxygenase. The enzyme is
very nonspecific and can also catalyze the oxidation of a wide variety of
organic compounds. Subsequent work (Oldenhuis et al. 1989; Tsien et al.
1989) has shown that Type II methylotrophs grown under conditions that
select the soluble form of methane monooxygenase have maximum activity
toward TCE.
3.20
-------
Chapter 3
Degradation of TCE by methylotrophs has been studied extensively in
the laboratory, at pilot scale, and in field demonstrations (Hazen 1991). A
major design difficulty results because methane is a competitive inhibitor of
TCE metabolism, and methane monooxygenase has a higher affinity for
methane than for TCE. A second problem lies in the irreversible loss of
activity of the enzyme during oxidation of TCE, and a third in that the en-
ergy required for TCE oxidation must be supplied externally. Methane
metabolism can supply the necessary energy, but its competition with TCE
makes it a poor choice as an energy source. Formate has been used with
some success in the laboratory to provide energy for the process (Oldenhuis
etal. 1989).
Trichloroethylene can also be oxidized by ammonia monooxygenase
(Arciero et al. 1989), isoprene-oxidizing enzymes (Ewers, Freire-Schroder,
and Knackmuss 1990), propane monooxygenase (Wackett et al. 1989),
toluene orf/io-monooxygenase (Shields et al. 1989), toluene para-
monooxygenase (Winter, Yen, and Ensley 1989), and toluene dioxygenase
(Zylstra, Wackett, and Gibson 1989). Each of the other systems mentioned
above require the presence of an appropriate inducer that may be a toxic
organic compound. Recently, a constitutive mutant of the Pseudomonas
cepacia strain containing the toluene ort/zo-monooxygenase was developed
(Shields 1991). This strain may prove useful, particularly in bioreactors,
because it requires no inducer for the toluene monooxygenase.
The products of TCE oxidation depend on the mechanism of the initial
oxidation. Monooxygenase attack produces TCE epoxide and chloral (Fox
et al. 1990), which decompose spontaneously to dichloroacetate, glyoxylate,
formate, and carbon monoxide. In contrast, dioxygenase attack has been
reported to initially yield TCE-dioxetane and 1,2-dihydroxy-TCE, which
rearrange to formate and glyoxylate (Li and Wackett 1992).
Reductive dehalogenation of chloroalkanes has been reviewed (Mohn
and Tiedje 1992). Under anaerobic conditions in the laboratory, TCE, 1,2-
dichloroethylene, and vinyl chloride can be converted to ethylene. Initial
studies indicated that the reductive dehalogenation required methanogenesis
(Freedman and Gossett 1989). Subsequent work has shown that the process
can also proceed in the absence of methanogenesis if sufficient methanol is
present (DiStefano, Gossett, and Zinder 1991). These studies indicate the
potential for bioremediation of this series of chlorinated solvents, but under
field conditions the process seldom goes to completion and intermediates,
3.21
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Process Identification and Description
such as vinyl chloride, often accumulate. A number of workers have sug-
gested that reductive dehalogenation might be used for the initial conver-
sion of highly-chlorinated solvents (such as perchloroethylene) to less chlo-
rinated intermediates which could then be degraded oxidatively (see, e.g.,
Fathepure and Vogel 1991).
Tetrachloromethane can be reductively dehalogenated (Egli et al. 1987),
but does not seem to be susceptible to oxidation by bacteria. Sirni larly,
chloroform can be degraded by reductive dehalogenation (Fathepure and
Vogel 1991). Dichloromethane can serve as a growth substrate under
methanogenic (Freedman and Gossett 1991) or aerobic (Brunner, Staub, and
Leisinger 1980) conditions. Therefore, it is an excellent candidate for
bioremediation. Pilot-scale studies and field applications will be discussed
in a later section.
Dichloroethane can be mineralized under aerobic conditions (van den
Wijngaard et al. 1992). Vinyl chloride can also serve as a growth substrate
under aerobic conditions for bacteria isolated from soil (Hartmans and
deBont 1992), but it is very volatile and would be difficult to treat in a
bioreactor.
3.2.1.5.6 Nitroaromatic Compounds Hydroxy- and carboxy-substituted
nitroaromatic chemicals are readily degraded by bacteria in soil and acti-
vated sludge. They serve as growth substrates for a variety of bacteria, and
most of the metabolic pathways have been worked out in detail (Spain and
Gibson 1991). The distribution of nitrophenol- and nitrobenzoate-degrading
strains in soil and water is often patchy, and extended acclimation periods
may be required before rapid biodegradation. Nitrobenzenes and
nitrotoluenes are much more resistant to biodegradation; however, nitroben-
zene, 4-nitrotoluene, 2,4-dinitrotoluene, and 1,3-dinitrobenzene can serve as
growth substrates for bacteria and are good candidates for bioremediation.
Three general mechanisms are used by bacteria for metabolism of aro-
matic nitro groups. Reduction to the amine level via hydroxylamitie seems
to be the most widespread mechanism. The resultant amino derivatives are
readily degraded, except in the case of the heavily substituted compounds,
such as TNT. Partial reduction to the hydroxylamine can lead to subse-
quent degradation by oxidative metabolism. Nitrobenzene and 4-
nitrobenzoate are degraded by this mechanism. Reduction of nitro groups
can take place under either aerobic or anaerobic conditions and seems to be
3.22
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Chapter 3
a relatively nonspecific process. Oxidative removal of the nitro group was
first described for 4-nitrophenol (Spain, Wyss, and Gibson 1979) and 2-
nitrophenol (Zeyer and Kearney 1984). The reaction involves the initial
replacement of the nitro group with a keto group, then a reduction to the
corresponding hydroxy derivative. Dioxygenase enzymes can remove the
nitro group from 1,3-dinitrobenzene (Dickel and Knackmuss 1991) or 2,4-
dinitrotoluene (Spanggord et al. 1991) and replace it with two adjacent
hydroxy 1 groups. Removal of the nitro group by an oxygenase mechanism
typically leads to mineralization of the parent molecule and results in the
formation of stoichiometric amounts of nitrite. A third mechanism of enzy-
matic attack on nitroaromatic compounds has been reported (Lenke, Rieger,
and Knackmuss 1992). Reduction of the ring of picric acid by a
Rhodococcus sp. yielded a hydride-Meisenheimer complex that was subse-
quently converted to 2,4-dinitrophenol with concomitant removal of one
mole of nitrite. This mechanism may eventually be applicable to the degra-
dation of other complex nitroaromatic compounds.
A considerable amount of research has been done on the biodegradation
of TNT. Reductive pathways leading to stable amines and dimers were
reported in early studies (McCormick, Feeherry, and Levinson 1976). Ad-
ditional studies have revealed the possibility of more extensive degradation
under anaerobic conditions (Roberts, Funk, and Korus 1992). Under aero-
bic conditions, TNT was shown to be mineralized slowly by Phanerochaete
chrysosporium (Fernando, Bumpus, and Aust 1990). Several pilot- and
field-scale bioremediation projects have been done with systems based on
composting (Williams, Ziegenfuss, and Sisk 1992), but little is known about
the mechanism of degradation or the final products of the process.
3.2.1.5.7 Aniline Although aniline is not a natural compound, it has been
released in the environment for many years because of its extensive use in
the chemical industry. It is readily biodegradable under aerobic conditions
by a variety of microorganisms. The initial reaction is a dioxygenase-cata-
lyzed removal of the amino group to form catechol, which is the substrate
for ring fission. Chloroanilines are more resistant to biodegradation and
bind readily to humic material in soil. Humic binding may lead to reduc-
tions in bioavailability, mobility, and toxicity of the compound (McCarthy
1989). In aquifer solids and groundwater, chloroanilines are slowly de-
graded by reductive dehalogenation under anaerobic conditions (Kuhn,
Townsend, and Suflita 1990). In pond sediments, an extended acclimation
3.23
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Process Identification and Description
period precedes reductive dehalogenation (Struijs and Rogers 1989). Under
aerobic conditions, a Moraxella sp. isolated from soil can use chloroanilines
as the sole source of carbon and nitrogen (Zeyer, Wasserfallen, and Timmis
1985).
3.2.1.5.8 Phthalates Although phthalates are biodegradable under aerobic
conditions in soil and water, extended acclimation periods are often re-
quired. They can serve as growth substrates for a variety of bacteria under
both aerobic and denitrifying conditions.
3.2.1.5.9 Dibenzodioxins and Dibenzofurans The stability of
dibenzodioxins and dibenzofurans is due to the presence of diarylether link-
ages, which are resistant to enzymatic attack. Tetrachlorodibenzodioxin,
coumaphos, and a number of pyrethroid insecticides contain such linkages
and are relatively persistent in the environment. Recently, bacteria able to
break the ether linkage have been isolated and studied in the laboratory.
The degradative mechanism involves angular dioxygenation followed by
cleavage of the unstable hemiacetal (Strubel et al. 1991). Dibenzofuran
(Fortnagel et al. 1990), coumaphos (Shelton and Somich 1988), and 3-
phenoxybenzoate (Topp and Akhtar 1991) can serve as growth substrates
for bacteria. A variety of other diphenyl ethers can be metabolized by bac-
teria grown on dibenzofuran, and several new strains with wider substrate
range are under development. Therefore, this class of compounds should be
considered as potential candidates for bioremediation. At present the
tetrachlorodibenzodioxins remain resistant to biodegradation.
3.2.1.5.10 Methyl t-Butyl Ether (MTBE) The gasoline additive, MTBE,
has been used extensively as an octane enhancer. Although there is little
evidence that MTBE is biodegraded in the field, a mixed culture has re-
cently been reported to degrade it in the laboratory (Salanitro et al. 1994).
The ether does not seem to interfere with degradation of other fuel compo-
nents.
3.2.1.5.11 Metals Microorganisms can catalyze a wide range of oxidation,
reduction, and methylation reactions involving metals. These reactions can
result in mobilization, immobilization, or volatilization of the metals. Such
reactions are well-documented and show considerable potential, but have
been little used in bioremediation because they do not destroy the metals.
3.24
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Chapter 3
Promising efforts are underway on the biological treatment of selenium-
contaminated soils (Thompson-Eagle and Frankenberger 1990). There has
also been some interest in the use of microorganisms and plants to sequester
or concentrate metals from dilute solutions, but the primary importance of
metals in bioremediation lies in their toxicity to microorganisms. Heavy
metals are used as biocides and can inhibit or kill the bacteria used in
biotreatment. Therefore, if they are present at toxic levels in mixed wastes,
they must be removed or their toxicity reduced prior to bioremediation.
3.2.2 Biogeochemistry and Biodegradation
3.2.2.1 Introduction
The effect that living organisms have on the geochemistry of the envi-
ronment is known as biogeochemistry. Biogeochemical processes control
the global cycling of the biologically-important elements — carbon, nitro-
gen, phosphorus, and sulfur — as well as the cycling of a variety of trace
elements. Biogeochemistry may be influenced by pH, temperature, ionic
strength, salinity, and UV light. The processes that influence a number of
aspects involved with hazardous materials are also an extension of nutrient
cycling by biogeochemical processes. Biogeochemical processes can:
• exert a strong influence on the fate and transport of many toxic
organic compounds in a variety of environmental media;
• strongly influence the redox state of a number of hazardous
metallic elements, thereby changing their mobility (Moore and
Ramamoorthy 1984) and bioavailability (McCarthy 1989); and
• influence the fate and transport of nonmetallic chemical species
that pose potential risks to human health and the environment
(e.g., nitrate and cyanide)(Keeney 1986; Bulger, Kehew, and
Nelson 1989).
In this section, the influences of biogeochemical conditions and pro-
cesses on biodegradation of organic contaminants and biotransformation of
inorganic contaminants in the environment will be explored. Knowledge of
the effects of biogeochemistry on biodegradation processes can be used to
evaluate relevant site characteristics to assess the feasibility of
bioremediation for treatment of a contaminated site.
3.25
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Process Identification and Description
3.2.2.2 Oxidation-Reduction Potential and Contaminant Bio-
degradation
Because more energy is derived from aerobic respiration than other mi-
crobial processes (figure 3.3 on page 3.9; table 3.2), oxygen is the preferred
electron acceptor, if present. The alternate electron acceptors are used more
or less sequentially in terms of their energy yields, although some simulta-
neous usage can occur. Once the concentration of oxygen becomes limit-
ing, between 0.1 and 1.0 mg/L, the microorganisms begin using nitrate and
continue using any oxygen that enters the region. Oxidized forms of iron
and manganese can serve as electron acceptors in microbial metabolism
before sulfate reduction is initiated, which can occur when nitrate is ex-
hausted. Ferric iron, but not manganese, can be used as an electron accep-
tor in aromatic hydrocarbon degradation. When present, iron and manga-
nese minerals solubilize concurrently with sulfate reduction. The reduced
forms of manganese and iron sufficiently scavenge oxygen so that the strict
anaerobes, such as some sulfate reducers and all methanogens, can de-
velop.
Many organic compounds (hazardous or otherwise) that enter oxidizing
matrices will undergo rapid and extensive biodegradation. This occurs as
long as the added substances are not toxic to the receiving microbial com-
munity, the microorganisms can readily develop the necessary enzymatic
capability to degrade the added substance, and as long as the matrix remains
oxidized. In general, biodegradation rates tend to be much lower under
Table 3.2
Microbial Processes and Redox Potential
Microbial Process
Aerobic respiration
Demtnfication
Manganese reduction
Iron reduction
Sulfate reduction
Methanogenesis
Electron Acceptor
Oxygen
Nitrate, nitrite,
nitrous oxide
Mn<+
Fe3t
S042-
C02
Products
H2O
N2
Mn2+
Fe2+
H,S
CH4
Eh (mV)
+810
+750
+396
-182
-220
-240
Adapted from Bouwer 1984, Schlesinger 1991
3.26
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Chapter 3
reducing than under oxidizing conditions (table 3.3). The difference in the
rates can be at least an order of magnitude (Godsy 1987). Furthermore,
biodegradation may be incomplete under reducing conditions, in some cases
resulting in the formation of metabolic by-products that are more toxic than
the parent compound (e.g., the reductive dehalogenation of 1,1-
dichloroethene to vinyl chloride under methanogenic conditions (Vogel,
Griddle, and McCarty 1987)).
Table 3.3
Degradation of p-Cresol Under Various Redox Conditions
Lag
Redox Conditions Biodegradabihty Time
Relative
Rate
Reference
Aerobic
Denitrifying
Sulfate-reducmg
Methanogenic
Hopper (1976, 1978)
Bossert and Young (1986)
Bak and Widdel (1986)
Smolenski and Suflita (1987)
Smolensk! and Suflita (1987)
Godsy, Goerlitz, and Ehrlich (1983)
Senior and Balba (1984)
From Suflita 1989
The fates of metallic and other inorganic hazardous substances are also
influenced by redox potential. Reducing conditions tend to result in a
chemical reduction of the introduced substance. For a number of metals,
reduced states tend to be more soluble and mobile than oxidized states.
Furthermore, studies suggest that if the metallic substances bind to humic
materials, both increases and decreases in mobility and toxicity may be
observed (for a review, see McCarthy 1989). If nitrate is the contaminant
of concern, reducing conditions may lead to nitrate removal (Bulger,
Kehew, and Nelson 1989). Oxidizing conditions tend to result in oxidation
of the introduced metallic or inorganic substance. The solubility and mobil-
ity of a number of metals tend to be reduced once the metal is in a more
oxidized form. Inorganic compounds, such as ammonia, can be oxidized
(Wilson 1992).
3.27
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Process Identification and Description
3.2.2.3 Estimates of Biodegradation Using the Expression of
Electron Acceptor Demand
The change in redox potential in environments contaminated with or-
ganic materials offers a relatively inexpensive tool that can be used as an
indirect gauge of the extent of biodegradation that may have occurred in a
contaminated site (Wilson, Armstrong, and Rifai 1993). The extent of con-
taminant biodegradation is estimated by first determining the concentrations
of electron acceptors that have been depleted in contaminated groundwater
but are present in uncontaminated groundwater very near to the contamina-
tion, and the concentrations of ferrous iron and methane which have accu-
mulated in contaminated groundwater. These values are used to estimate
the amount of contamination biodegraded using stoichiometry.
The following equations can be used to estimate contaminant biodegra-
dation based on the concentrations of alternate electron acceptors present in
groundwater where CH2 represents a generic alkane, such as pentane or
isooctane, and CH represents an aromatic hydrocarbon, such as benzene or
toluene:
Aerobic 4CH2 + 6O2 -> 4CO2 + 4H2O [1]
Aerobic 4CH + 5O2 -•> 4CO2 + 2H2O [2]
Denitrification 5CH + 4HNO3 -» 4N2 + 5CO2 + 2H2O [3]
Sulfate reduction 8CH + 5H2SO4 -> 4H2S + 8CO2 + 4H2O [4]
Iron reduction CH + 5Fe+3 + 5OH -> 5Fe+2 + CO2 + 3H2O [5]
Methanogenesis 2CH + 1.5H,O -> 0.75CH4 + 1.25CO2 [6]
To assess potential biodegradation in a hypothetical scenario, the follow-
ing groundwater quality parameters were used: (1)3 mg/L dissolved oxy-
gen, (2)10 mg/L nitrate as total nitrogen, and (3) 20 mg/L sulfate in uncon-
taminated groundwater near the impacted area and 40 mg/L ferrous iron and
24 mg/L methane (solubility of methane in cold water) in contaminated
groundwater from the impacted area. It is estimated that significant
amounts of hydrocarbon can be degraded by natural bioattenuation (table
3.4 on page 3.29). And it is important to note that only oxygen can be used
as an electron acceptor in both alkane and aromatic hydrocarbon biodegra-
dation. For aromatic hydrocarbon biodegradation, however,
methanogenesis, followed by denitrification, sulfate reduction, and iron
reduction, will be more important than oxygen.
3.28
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Chapter 3
Table 3.4
Estimating Biodegradation Using the Concentrations of Alternate
Electron Acceptors or Products Found in Uncontaminated and
Ground Contaminated Groundwater, Respectively
Equation
1
2
3
4
5
6
Electron Hydrocarbon
Acceptor/Product Consumed
L
per groun water
3mgO2
3mgO,
IV.SmgHNO,
20 mg H2SO4
40 mg Ferrous iron
24 mg Methane
0.9 mg
1 Omg
5.2 mg
42mg
1.8mg
32.5 mg
3.2.2.4 Nutrient Cycling and Availability
Biogeochemical processes affect nutrient availability and cycling, and,
because nutrients are important in biodegradation, they affect contaminant
biodegradability. Elements are being transformed constantly, taken up and
excreted by organisms and cycled between sequestered and bioavailable
forms. When assessments of nutrient requirements for contaminated envi-
ronments are made, the form (availability) of the nutrients in the matrix
should be considered.
An evaluation of nutrient relationships is usually conducted by collecting
soil or sediment samples and determining the concentrations of the various
forms of carbon (total organic carbon, dissolved organic carbon, and total
petroleum hydrocarbons), nitrogen (typically ammonia, nitrate, and nitrite),
and phosphorus (typically ortho-phosphate) using standard chemical proce-
dures and protocols. The nutrient amendment is determined first by calcu-
lating the amount of carbon that can be biodegraded when a known supply
of electron acceptor is added. Using this value, the amounts of nitrogen and
phosphorus required to degrade the carbon can be calculated using the
carbon:nitrogen:phosphorus (Redfield) ratios found in biological materials.
A common mistake made in bioremedial operations is to overestimate nutri-
ent requirements by using the amount of available carbon rather than the
3.29
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Process Identification and Description
amount of carbon that can be consumed, based on the supply of electron
acceptor. Addition of unnecessary nutrients is a waste of capital and time.
Ambient levels of nutrients and nutrient cycling from the biomass to the
water are considered when the nutrient amendment is formulated. The con-
centrations of carbon, nitrogen, and phosphorus in samples of water or soil
provides little insight into the dynamic nature of nutrient-contaminant rela-
tionships, which will be especially important in environments such as a
flowing groundwater or streams. Nitrogen and phosphorus may be cycling
constantly between the sequestered and bioavailable forms, and in a dy-
namic site, may be constantly delivered and removed. Furthermore, be-
cause measurements are usually made of the most biologically-available
forms of nitrogen (ammonia, nitrate, nitrite) and phosphorus (ortho-phos-
phate), incorrect conclusions can be drawn regarding nutrient balance and
availability because their concentrations are often below detectable limits.
Where nitrogen and phosphorus concentrations are below detection lim-
its, it is usually assumed that these nutrients are limiting. This assumption
may not be true, since organic forms of nitrogen, that is, those measured by
the total Kjeldahl nitrogen method, and phosphorus, as measured by the
total phosphorus method and the total reactive phosphorus method, etc.,
may not be readily available to the indigenous microorganisms, but, none-
theless, may represent important resource pools that can provide the mi-
crobes with the necessary nitrogen and phosphorus. These tests may esti-
mate supplies of these nutrients, but their predictive ability is unclear. In-
deed, there is evidence that additions of nutrients may actually impede con-
taminant biodegradation rates in some cases (Cunningham 1992). Because
of the dynamic nature of the soils and groundwater, however, only esti-
mates of nutrient concentrations in soils and water can be determined and
these values are used to determine nutrient formulations.
3.30
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Chapter 3
3.3 Site Characterization Relative to In
Situ Bioremediation
3.3.1 Introduction
An important determinant for successful application of bioremediation is
the physical nature of the contaminated matrix. Site features strongly influ-
ence whether bioremediation is feasible and which bioremediation
approach(es) will be effective.
Bioremediation is effective in matrices with high fluid-flow properties.
In the saturated zone, matrices with hydraulic conductivities >10~4 cm/sec
are most amenable to bioremediation through liquid delivery methods. (See
Subsection 3.5.1.2.1). In the unsaturated zone, matrices with hydraulic
conductivities as low as 10"5 to 106 cm/sec may be treated successfully
through bioventing (see Subsection 3.5.1.1.2), which can transport more
oxygen per unit volume than liquid delivery. But, these fluid-flow values
for treatability are not absolute; the higher the degree of contamination, the
greater the limitations imposed by fluid-flow properties.
Where this parameter is lower, the matrix may be more difficult to treat.
Matrices having high hydraulic conductivity can receive more air or liquid
flow per unit time than matrices with low values. Because air and liquids
contain the electron acceptor and possible nutrient amendments, longer
treatment will be required for matrices with low-flow properties than for
those with high-flow properties.
Matrices having low-hydraulic conductivity often contain clay. Con-
taminated clays are among the most difficult matrices to treat through
bioremediation because (1) contaminants sorb more strongly to clays than
sands and (2) the air or liquid flow rate through clays is lower than that for
sands. In addition to slow delivery of the electron acceptor and nutrients in
clays, bioavailability of the contaminant can be decreased by sorption to
clay particles (see Subsection 3.2.1.2.5 Contaminant Bioavailability). One
well-understood mechanism that can exclude contaminants from biodegra-
dation is the sorption of organic compounds to the 2:1 layered silicate clays
(Stotzky 1986).
A subsurface characteristic that will greatly impact the performance of
any in situ bioremediation approach is site heterogeneity. Although the
3.31
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Process Identification and Description
overall hydraulic conductivity of a site may render it amenable to in situ
treatment, contamination in zones of low hydraulic conductivity will be
more difficult to treat. Air or water that is injected into the subsurface may
not fully penetrate or may bypass these zones, leaving pockets of contami-
nation; however, treatment with air in the unsaturated zone will be less
hindered than treatment with water in the saturated zone. Further compli-
cating the treatment of heterogeneous sites is the fact that low permeability
zones generally have a relatively high capacity to sorb organic contami-
nants. High sorption capacity results in higher concentrations of contami-
nants that are more difficult to treat than the contamination found in high
permeability zones. The resulting uneven treatment will make it difficult to
assess remedial progress.
In some cases, a remedial system can be designed to overcome the prob-
lems associated with heterogeneous sites. If contamination has impacted
underlying geologic strata with varying permeabilities, it may be possible to
place injection points at appropriate depths to access the various affected
strata.
Because more permeable strata would not require as high pressures to
achieve appreciable injection rates compared to lower permeable strata, it
would be necessary to have separate injection systems for permeable versus
poorly permeable strata. In this way, the low-permeability strata system
can be pressurized appropriately.
The use of separate injection systems is better suited for treatment of
horizontal layers of different permeability rather than discontinous lenses.
Multiple discontinuous lenses will be the most difficult and expensive to
treat. In other cases, once the more permeable zone has been trealed below
the site specific clean-up goal, the residual contamination in the lower per-
meable zone may diffuse into the groundwater at a slow enough rate that
natural bioremediation (Section 3.4 Natural Bioattenuation of Hazardous
Organic Compounds in the Subsurface) may be sufficient to prevent migra-
tion.
3.3.2 Overview of a Biofeasibility Assessment Procedure
The key characteristics of a site that must be considered during the de-
velopment of a remedial plan are summarized in this section. This over-
view provides recommendations for studies that should be conducted during
the remedial investigation (RI) phase.
3.32
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Chapter 3
3.3.2.1 Site Characterization
The first task in site characterization is to acquire a thorough understand-
ing of the three-dimensional distribution of contaminant mass and the distri-
bution of properties at a site that will affect the rate and extent of
remediation. This task is important because the site condition is the ulti-
mate determinant of which treatment technologies can be effectively ap-
plied. A review of existing site assessment and investigation reports, re-
ports discussing local and regional geology, and interviews with site per-
sonnel may be helpful. By way of delineating the context of the contamina-
tion, the following should be determined:
• the geochemical properties at the site;
• the groundwater quality in contact with the contaminants;
• the fluid-flow properties of the contaminated matrix; and
• methods for measuring the concentration of contaminants.
The geochemical properties include soil, sediment and rock types; pH
(EPA Method 9045); nutrient concentrations (total Kjeldahl nitrogen, EPA
Method 351.2; total phosphorus, EPA Method 6010); and total organic
carbon concentration (American Society of Agronomy Method 90-2).
Knowledge of these characteristics is important because materials added to
or produced in the matrix during bioremediation can react with matrix con-
stituents. For instance, geological materials consisting of limestone
(CaCO3) or dolomite (MgCO3 and CaCO3) naturally buffer the CO2 pro-
duced during biodegradation, thus maintaining a neutral pH in the treatment
zone. These materials also may precipitate some of the added nutrients as
salts, which may limit fluid transport through the treatment area. In con-
trast, matrices composed of highly-weathered materials, which contain little
or no calcium and magnesium, will not be able to buffer the CO2 produced
during biodegradation. In highly-weathered materials, pH values can reach
levels as low as 4. In addition, materials consisting of clay-sized particles
tend to trap oily-phase material, particularly above the water table.
The quality of the groundwater in contact with the contaminants is im-
portant. The groundwater contains the nutrients and electron acceptors that
will be available to the microorganisms and affects the buffering capacity of
the treatment zone. Hard water, which contains high concentrations of
cations, will have a high capacity to buffer the CO2 produced during
bioremediation and maintain a neutral pH in the treatment zone; however,
3.33
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Process Identification and Description
these cations may precipitate out with added nutrients, reducing fluid flow.
In aquifers with soft water, pH values may be low because the buffering
capacity is low.
The fluid-flow properties of a matrix will affect the transport of air or
water through it. The fluid-flow properties in the unsaturated and saturated
zones are determined by measuring the pneumatic and hydraulic conductiv-
ity, respectively.
The horizontal hydraulic conductivity can be measured using a slug,
pump, or tracer test (Thomas, Marlow et al. 1992). The type of test will
depend on the economics and the information desired. While slug tests are
relatively inexpensive, the hydraulic conductivity within only a short radius
of the well is obtained. Pump and tracer tests are expensive, but provide a
more extensive regional estimate of fluid flow. While pump and tracer tests
can be used in any matrix, the slug test is used in those matrices that are
moderately permeable-to-tight. The vertical hydraulic conductivity can be
determined using two wells placed in the same location where one is
screened above the other. A tracer is added to one well and its break-
through into the second well is measured.
The pneumatic conductivity is measured using two wells. Air is injected
into one well and extracted from the second (Cho and DiGiulio 1992). Af-
ter stabilizing the well pressure, the soil-air distribution between the wells is
determined.
Fluid-flow properties should be determined at multiple locations because
of the inherent heterogeneity of subsurface formations. Variations in verti-
cal and horizontal hydraulic conductivity can be better defined by determin-
ing the vertical stratigraphy at various locations. The stratigraphy can be
determined by examination of corings during well installation.
The method used to measure the contamination is also important. For
many types of contamination, the bulk of the material is trapped as an oily
phase, and only small amounts are dissolved. If based on analysis of
groundwater only and not analyses of soil and sediments, estimates of con-
tamination will be low. Borings are made through the contaminated to the
uncontaminated material in the contaminant source area, (e.g., underground
tank, pipeline rupture) and representative samples are collected at depths
that exhibit signs of contamination. Color changes in the samples can be
used as indicators of contamination when compared to uncontaminated
3.34
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Chapter 3
samples of the same soil type and from the same formation (material con-
taminated with petroleum compounds will be blue/gray or green, whereas
uncontaminated material will be light brown to red). Through analyses of
samples of soil or sediment, the following determinations should be made:
• the concentrations of lighter-than-water contaminants (LNAPLs)
in the "smear" zone at the fluctuating interface between the va-
dose and saturated zones beneath and downgradient from the
contaminant source;
• the concentrations of sorbed heavier-than-water contaminants
(DNAPLs) at intervals across the vertical extent of the affected
aquifer beneath and downgradient from the contaminant source;
and
• the concentrations of contaminants in the vadose zone.
3.3.2.2 Biotreatability Evaluation
The first step in the evaluation is to determine whether or not microbial
processes that will be used during remediation are already occurring at
some limited rate (natural bioattenuation). This evaluation can be made by
conducting an in situ soil-gas survey in the unsaturated zone and/or by de-
termining groundwater quality. Depending on the requirements of the situ-
ation, these tests may be used instead of a laboratory biotreatability study.
For the soil-gas survey, soil-gas samples from the contaminated and
uncontaminated zones are collected and analyzed for oxygen and CO2.
Concentrations of oxygen of less than 21% (atmospheric is =21%) and CO2
greater than 1% (atmospheric is =0.03%) in the affected zone would suggest
contaminant biodegradation, oxygen consumption, and CO2 production by
the indigenous microorganisms. The CO2 may be produced as a result of
aerobic and anaerobic processes. For groundwater, the presence or absence
of the electron acceptors (oxygen, nitrate, nitrite, and sulfate) in contami-
nated and uncontaminated samples may indicate that biodegradation is oc-
curring (Piotrowski 1989). In uncontaminated groundwater, there is usually
>1 mg/L O2, nitrogen is present as nitrate and not ammonia, sulfate is
present, and iron and methane are absent. The presence of reduced iron,
>0.1 mg/L methane, H2S, and <0.5 mg/L O2 in contaminated water would
suggest natural bioattenuation is occurring. Groundwater is sampled in a
series of wells upgradient and downgradient of the spill. At many sites,
existing monitoring wells can be used.
3.35
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Process Identification and Description
Care must be exercised in measuring O2 in groundwater. Groundwater
containing a layer of free-floating product should not be sampled, to avoid
fouling the dissolved oxygen probe. Sampling procedures, such as bailing
or pumping a well, can easily reaerate the water. The method chosen for
sampling groundwater from a well will be specific to that well and depend
on the depth of the water table and the hydraulic conductivity of the aquifer.
An indirect indicator of the absence of O2 would be the presence of reduced
iron and/or methane.
Laboratory treatability studies may be conducted. To determine the rate
and extent of contaminant biodegradation, samples of sediment or soil are
subjected to batch studies using varying nutrient amendments based on the
site characterization. However, the fact that laboratory-derived biodegrada-
tion data may not be the same as those in the field must be considered. The
results of these studies are used to determine the final nutrient formulation
required to achieve the most rapid and cost-effective rate of remediation. In
addition, these studies can be used to determine whether the prospective
nutrient amendment is compatible with the subsurface material. Care must
be taken to avoid precipitating the nutrients and plugging the aquifer. In
materials containing clay, potassium rather than sodium salts are used to
prevent swelling of the clay. When the groundwater is hard, tripolyphos-
phate instead of orthophosphate is used, because it will solubilize rather
than precipitate iron, calcium, and magnesium. In aquifers high in iron,
added oxygen can oxidize and thus precipitate the iron.
Although microbial numbers and types have been determined in early
bioremediation projects (Raymond et al. 1975; 1978), they are riot neces-
sary to predict the feasibility of bioremediation. Biofeasibility tests should
be conducted to measure microbial activity (biodegradation) rather than
microbial numbers.
3.3.3 Design Considerations for In Situ Bioremediation of
Aquifers
/
Regulatory concerns and site characteristics must be considered during
the development of the remedial design. Regulatory concerns are the driv-
ing mechanism for aquifer restoration and often have a marked impact on
the bioremedial treatment ultimately approved for the site. Failure to ad-
dress regulatory concerns early in the remedial design process often results
in regulatory resistance or outright rejection. The geology, hydrogeology,
3.36
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Chapter 3
geochemistry, and biogeochemistry of the site will also affect system de-
sign.
3.3.3.1 Regulatory Concerns
Among the most important site aspects that can strongly influence the
design of an in situ bioremdiation system is the the regulatory concern re-
garding the contaminated groundwater. A contaminated groundwater
plume located upgradient from a municipal drinking water source or sensi-
tive aquatic habitat will elicit more concern than a plume located in an unin-
habited, less-sensitive area.
Sites Upgradient from Sensitive Areas. Concerns over plume migra-
tion toward a sensitive area typically result in a regulatory mandate that
plume control be adopted as a part of the remedial strategy. Consequently,
some form of pump-and-treat system is installed as a hedge against further
plume migration.
The aboveground treatment for the extracted groundwater must reduce
contaminant levels to concentrations that meet regulatory standards. Often,
some form of "rough" treatment is used to cost-effectively reduce the con-
taminant concentration and then the partially-treated groundwater is passed
through a "polishing" step (such as activated carbon) to consistently pro-
duce groundwater containing acceptable contaminant concentrations.
Aboveground biotreatment options (such as those discussed in Sections
2.5.2.1 and 3.5.2.1) are now commonly used for the rough treatment step.
Historically, the treated groundwater has been considered waste that
must be disposed. Generally, the water is discharged to a body of water
under a National Pollutant Discharge Elimination System (NPDES) Permit
or to a publically-owned treatment works (POTW) under a local permit. In
both cases, the pollutant concentrations must not exceed regulatory stan-
dards.
Instead of disposing of the treated groundwater, all or a portion of the
water can be amended with biostimulating additives and returned to the
treatment area as described in Sections 2.5.2.1 and 3.5.2.1, Liquid Deliv-
ery. Moreover, if the aboveground treatment process includes a biologi-
cally-based "roughing" step, biological "preconditioning" of the water may
provide an added stimulating effect by supplying nutrients and microorgan-
isms adapted to degrade the contaminants. However, the transport of mi-
3.37
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Process Identification and Description
croorganisms for the purpose of contaminant biodegradation in the subsur-
face has not been demonstrated.
As long as the groundwater extraction rate is equal to or greater than the
liquid delivery rate, the hydraulic control of plume migration should be
maintained (closed loop). McCarty et al. (1989) estimated that coupling a
pump-and-treat system with an in situ bioremediation program can reduce
the overall time for aquifer restoration by as much as one-half or more.
Assuming re-injection of the treated and amended groundwaleir is ap-
proved, a regulatory constraint may be converted into a substantial remedial
benefit. More and more states are recognizing the potential benefits of
allowing re-injection of treated groundwater for aquifer biorestoration.
States vary in their regulatory stances towards injection of treated
groundwater. A number of states will not allow it at all. In these cases,
injection of appropriately amended clean groundwater, surface water, or
dechlorinated municipal water may be allowed for in situ bioremediation.
In other states (e.g. Texas), treated groundwater can be used for re-injec-
tion as long as the treatment process consistently produces water containing
contaminant constituents below applicable regulatory goals. Finally, in
Michigan, re-injection of untreated groundwater has been permitted in a
number of cases. In most situations involving groundwater reinjection, it
will likely be necessary to ensure that hydraulic control of plume migration
will be maintained during re-injection so that off-site plume migration does
not occur.
An interesting development in this regard was a study conducted at a site
in Michigan by a partnership (termed "CoBioReM") involving a state regu-
latory agency, the oil and gas industry, and two universities in a 1991 study.
Contaminated groundwater was extracted, amended with biostimulants, and
returned to the subsurface without aboveground treatment. After 7 months
of treatment, 90% reduction in BETX in groundwater was observed. The
success of this project raises the possibility that aboveground trealment of
extracted groundwater may not always be required for closed-loop, in situ
bioremediation systems. This change in regulatory attitude could result in a
substantial reduction in costs.
Sites Not Upgradient from Sensitive Areas. At those sites where sen-
sitive downgradient receptors are not threatened (i.e., sites located in flat
areas with minimal natural groundwater flux or areas at considerable dis-
3.38
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Chapter 3
tance from sensitive receptors), plume control may not be required, espe-
cially if a nonliquid delivery approach such as air sparging (Sections 2.5.1.5
and 3.5.1.2.3) was implemented. Because groundwater is not extracted, the
costs associated with air sparging are usually less than those associated with
closed-loop delivery systems. Nevertheless, liquid delivery without
groundwater capture (if permitted) or closed-loop treatment might be ad-
vantageous for flat sites with minimal natural groundwater flux as a means
to enhance the rate of aquifer restoration by maximizing mass transport and
mixing in the contaminated zone.
3.3.3.2 Geological Aspects
The physical aspects of the aquifer represent the environmental condi-
tions under which the in situ treatment system must operate. If available,
information on regional geology should be reviewed to gather insight into
potential subsurface conditions. Also, if available, boring and monitoring
well logs from investigatory activities at the site or nearby should be re-
viewed for information concerning grain-size distributions of aquifer sedi-
ments, stratigraphic variability, mineral composition, and the presence of
bedrock.
If new borings are made as part of the remedial investigation, sediment
samples from the aquifer should be collected from key aquifer strata for
grain size and mineralogical analyses. These sediment samples will also be
suitable for geochemical (Section 3.3.3.4) and biogeochemical (Section
3.3.3.5) analyses. Geological factors must be evaluated and taken into con-
sideration to install a bioremedial system that will have a reasonable prob-
ability of success under site-specific conditions.
Grain-Size Distribution. The grain-size distribution of aquifer sedi-
ments influences pore-space volume and distribution and consequently, the
aggregate permeability of the aquifer sediments. Basically, the finer the
grain sizes, the more difficult it will be to achieve biostimulation. Problems
have been reported for aquifer sediments with permeabilities less than or
equal to 3.3 x 10'3 cm/sec (Lee et al., 1987). Furthermore, the in situ
bioremediation program itself can reduce aquifer permeability. When treat-
ing fine-grained sediments, means to evaluate changes in aquifer permeabil-
ity should be included (i.e. injection pressure changes, groundwater level
fluctuation).
3.39
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Process Identification and Description
The grain-size distribution of the aquifer sediments should be determined
or a preliminary characterization of sediment grain size may be conducted
by inspecting existing boring or monitoring well logs.
If the results of geological and hydrological (see section 3.3.3.3) tests
indicate adequate permeability, the data can be used in commercially-avail-
able aquifer-response models to evaluate various configurations for injec-
tion/extraction wells or trenches to provide a basis for remedial design
(Piotrowski et al. 1993). Injection wells have been successfully used in
predominately sand and gravel aquifers to biogeochemically influence ex-
tended regions of aquifers (e.g., > 200 feet in the downgradient direction
over depths ranging to greater than 80 feet below ground surface)
(Piotrowski, 1991; Piotrowski et al. 1994).
In low-permeability aquifers, trench systems may be necessary to extract
and deliver groundwater at appreciable rates and across extended aquifer
regions. While the rate of groundwater flux across a treatment zone can be
modified to some extent in tight aquifers, its maximum rate (and the rate of
biorestoration) will ultimately be constrained by the zone's aggregate per-
meability. Several vendors offer biodegradable slurries that allow ready
installation of infiltration/extraction trenches and associated conveyance
piping below the water table (Piotrowski et al. 1993). The costs associated
with trench installation increase substantially with depth. At some sites,
cost may favor injection/extraction wells over trenches even though
trenches may produce a more rapid rate of biorestoration across extended
aquifer regions.
Stratigraphic Variability. Stratigraphic variability can strongly influ-
ence the performance of an in situ bioremediation system. For example, it
can result in channeling of groundwater through permeable zones such that
the amendments used for in situ biostimulation may bypass the less perme-
able, yet contaminated zones. Channeling may especially occur in alluvial
aquifers which can be comprised of lenses of markedly different
permeabilities. The challenge at these sites is to devise a system that will
deliver amendments to the most contaminated zones without substantial
loss to permeable regions.
Although fine-grained lenses may adsorb organic contaminants and in-
hibit groundwater penetration, the fine-grained materials also tend to inhibit
contaminant penetration into the lens. Consequently, a fine-grained lens
may contain a veneer of contamination that may be treatable by passing
3.40
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Chapter 3
groundwater amended with biostimulants over its surfaces. Promoting bio-
degradation of the veneer of contamination may render the lens an inconse-
quential source of groundwater pollution.
Mineral Composition of Sediments. The nature of the mineral compo-
sition of the aquifer sediments can influence both the physical and chemical
aspects of a bioremedial design. Chemical aspects of the mineral content of
aquifer sediments are discussed in Section 3.3.3.4. If sediment samples
consist of more than 30 percent clay, it may be advisable to have portions of
the samples analyzed for clay content using X-ray diffraction. The 2:1-
layered silicate clays (e.g. illite, smectite, montmorillonite) tend to allow
penetration of hydrophobic organic compounds into the clay-lattice struc-
ture which can interfere with microbial access to the compounds (for a re-
view see pages 118-122 in Alexander 1994). In contrast, Ill-layered sili-
cate clays (e.g. kaolinite) do not allow penetration of organic contaminants.
If 2:l-layered clays are common, application of a surfactant may be re-
quired to enhance bioavailability of hydrophobic organic compounds
(Leavitt et al. 1992; Bonin et al. 1994). However, the surfactant should be
nontoxic, biodegradable, and not inhibit contaminant biodegradation.
Moreover, hydraulic control of the contaminant plume may be required to
prevent contaminant migration resulting from an enhancement of its mobil-
ity by surfactant application.
The Presence Of Bedrock. Those aquifers that contain bedrock pose
unique challenges to bioremedial design. Bedrock fractures represent
groundwater conduits for contaminant transport whose pathways are often
difficult to map or predict. Improper selection of the sites of the injection/
extraction components can easily render an in situ bioremediation system
ineffective. A case study of bioremediation of bedorck was published by
Bell and Hoffman (1991).
One method used to address this design complication is to apply a dis-
ruptive force to artificially increase the bedrock fractures, which increases
aquifer permeability for enhanced biorestoration. Attempts have been made
to increase permeability using hydraulic (Davis-Hoover et al. 1991; Vesper
et al. 1994) and pneumatic (Schuring 1993) fracturing. However, care must
be taken that the fracturing does not enhance contaminant mobility.
3.41
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Process Identification and Description
3.3.3.3 Hydrogeological Aspects
The ability to move groundwater through a contaminated zone: can be
vital to the success of an in situ bioremediation system. Information con-
cerning groundwater flow can be obtained by conducting standard
hydrogeological tests (e.g. draw-down tests, slug tests). These tests are
often used to develop pump-and-treat remedial systems. Often overlooked
during the design of a closed-loop in situ bioremediation system is the abil-
ity of the contaminant source area to accept the treated and amended
groundwater. This information can be obtained by conducting standard
percolation tests in the source area (Piotrowski et al. 1993).
As mentioned in Section 3.3.3.1, the information from draw-down and
percolation tests can be used in commercially-available groundwater-re-
sponse models to evaluate various extraction and delivery configurations to
provide a basis for a closed-loop design. As discussed further in Section
3.3.3.4, a draw-down test also provides the opportunity to collect informa-
tion on groundwater chemistry and contaminant concentration that can be
important to the design of an aboveground water treatment system.
Finally, air sparging can produce localized mounding in the surface of
the water table in the vicinity of the sparging well. The mounding repre-
sents a groundwater gradient that can increase groundwater flow and could
elicit regulatory concern that enhanced contaminant migration may result.
Therefore, groundwater level measurements during a pilot test of air
sparging may be useful to evaluate this possibility and provide an indication
if plume control may be warranted.
3.3.3.4 Geochemical Aspects
The geochemistry of the aquifer sediments are also important for
bioremedial design. Topics discussed in this section include sediment con-
taminant concentrations, considerations of iron-rich aquifers, evaluations of
sediment redox capacities, and chemical analyses during extended aquifer
draw-down tests.
Sediment Contaminant Concentrations. Of utmost importance to
regulatory and remedial design is the concentration of adsorbed contamina-
tion. The adsorbed contaminants (which can exist as organic coatings,
microglobules, or free product) represent long-term sources of groundwater
contamination. If the adsorbed contaminants are not treated along with the
3.42
-------
Chapter 3
groundwater, extended treatment times will be required. Therefore, an
evaluation of the adsorbed contaminant concentrations in and around the
contaminant source area should be performed during the remedial investiga-
tion.
For organic liquids with specific gravities < 1.0 g/cc (light nonaqueous
phase liquid, LNAPL), typically the upper strata of the aquifer will be im-
pacted. For organic liquids with specific gravities > 1.0 g/cc (dense
nonaqueous phase liquid, DNAPL), extended vertical sections of aquifer
strata may be impacted, especially in the vicinity of the contaminant source
area. During the advancement of soil or well borings, it is advisable to
collect samples from those strata likely to be most contaminated and ana-
lyze for contaminants. When analyses of contaminant concentration and
nutrients (e.g. Kjeldahl nitrogen, phosphorus, total organic carbon) are
coupled, potential nutrient requirements for in situ contaminant biodegrada-
tion can be assessed.
Considerations for Iron-Rich Aquifers. Iron-rich sediments under
reducing conditions can generate groundwater enriched with dissolved fer-
rous iron. The presence of dissolved iron can affect the operation of both
above- and below-ground components of an in situ bioremediation system.
If an oxidative approach is used, conversion of the soluble ferrous to the
insoluble ferric species can result in excessive formation of iron precipitates
and floes that can clog pore spaces in aboveground filters, pump-intake
screens, and between aquifer sediment particles.
Iron precipitation interferes with groundwater flow. The operation plan
for the bioremedial system at such sites may require frequent changes of
abovegound filters and in situ pump screens and the frequent application of
chemical or physical means to break-up or dislodge iron floes in the forma-
tion. Although iron precipitation may be an operational concern in the early
stages of oxic treatment of an iron-rich, contaminated aquifer, once oxic
conditions develop throughout the contaminated zone, iron will tend to
precipitate and remain in the aquifer. Consequently, iron precipitation
problems in the aboveground components will abate.
Finally, the presence of iron can influence the selection of the form of
phosphorus fertilizer used in phosphorus-deficient aquifers. Brown and
Norris (1994) suggest the use of tripolyphosphate to reduce the potential for
iron precipitation.
3.43
-------
Process Identification and Description
Sediment Redox Capacities. The oxidation-reduction (redox) capaci-
ties of aquifer sediments may resist efforts to adjust redox conclilions in the
aquifer (Barcelona and Holm 1991). Appreciable quantities of an oxidant
may be consumed by abiotic oxidations of the reduced species before mea-
surable changes in groundwater redox are observed. This possibility must
be taken into consideration during the selection of the oxidant for the in situ
bioremedial system.
If anticipated or known that the aquifer sediments will consume a large
amount of the oxidant in abiotic redox reactions, it may be advisable to
initially use high concentrations of the oxidant, hydrogen peroxide, to
quickly satisfy the abiotic oxygen demand. Once the abiotic demand has
been met, lower peroxide concentrations or alternate oxygen sources may
be used to reduce operational costs. Although high concentrations of hy-
drogen peroxide may sterilize regions of the aquifer close to the injection
point, recolonization probably occurs after peroxide addition is terminated.
Additional discussion of means to oxidize aquifers is presented in Section
3.3.3.5.
Chemical Analyses During Extended Aquifer Draw-Down Tests. As
mentioned in Section 3.3.3.3, important information on groundwater chem-
istry and contaminant concentration may be obtained during an extended
aquifer draw-down test. Many times the aboveground treatment system is
sized based on the results of point-in-time measurements of groundwater
contaminants. Because the concentration of organic contaminant in ground-
water declines as more and more water is extracted, aboveground systems
based on samples collected during the early phase of pumping may be over-
sized. To avoid sizing problems, groundwater samples should be collected
at intervals during the course of an extended draw-down test (e.g at 12 and
24 hours during a 24-hour test) and analyzed for contaminant concentration
(Piotrowski et al. 1993).
In addition to contaminant concentration, the groundwater samples
should be analyzed for dissolved solids, alkalinity, hardness, anions, cat-
ions, and total Kjeldahl nitrogen. Hard, alkaline water with a higli concen-
tration of dissolved solids may cause scaling problems in system compo-
nents. If so, design measures can be implemented to reduce the scaling
problem.
The results of anion, cation, and total Kjeldahl nitrogen analyses can be
used to assess ionic balance and nutrient relationships (iron, ammonia, and
3.44
-------
Chapter 3
potassium concentrations can be measured in the cation scan; nitrate, nitrite,
and phosphate can be measured in the anion scan). The results of the analy-
ses in conjunction with sediment analyses may indicate that some form of
nutrient addition may benefit the in situ bioremedial process.
3.3.3.5 Biogeochemical Analyses
The results of biogeochemical analyses of groundwater and aquifer sedi-
ments can be used to assess the potential for successful bioremediation.
When oxygen is used as the terminal electron acceptor, the means by which
oxygen is delivered will depend on the redox of the aquifer.
Groundwater Analyses. In addition to the chemical analyses discussed
in Section 3.3.3.4, groundwater may be analyzed for dissolved oxygen con-
centrations and other indicators of microbial activity. Because microorgan-
isms are usually attached to surfaces rather than free in the aqueous phase,
the most accurate assessment of microbial numbers and biodegradation
potential should be determined using samples of subsurface material.
Oxygen concentrations can be monitored in wells located upgradient
from (uncontaminated) and within the plume. If dissolved oxygen is mea-
sured in a transect of monitoring wells extending from upgradient areas into
the plume, a typical pattern is observed. Upgradient of the plume, dissolved
oxygen is detectable (2-8 mg/L), whereas within the plume dissolved oxy-
gen concentrations are < Img/L. The absence of dissolved oxygen along
the transect is a result of aerobic biodegradation of contaminants
(Piotrowski 1989; Piotrowski 1991). Although simple, this method has
been used successfully at a Superfund site to convince EPA to allow a pilot
study of in situ bioremediation (Piotrowski 1989).
Other indicators of microbial activity that can be measured in groundwa-
ter include the concentrations of ferrous iron, methane, and electron accep-
tors other than oxygen (see Section 3.2.2.3, Estimates of Biodegradation
Using the Expression of Electron Acceptor Demand).
Sediment Analyses. Samples for microbial analyses should be incu-
bated at the in situ temperature of the groundwater of the site. Various
enumeration and activity measurements are available, such as total counts,
viable counts, counts of contaminant-specific degraders, and radiolabelled
biodegradation studies. A broad range of methods has been summarized
(Kemp et al. 1993). Caution should be exercised in extrapolating the results
3.45
-------
Process Identification and Description
of laboratory experiments to the field. The results of in situ pilot studies are
the best indicators of remedial time requirements.
Means to Deliver Oxygen. Because of the redox capacities of sedi-
ments (Section 3.3.3.4), the means used to meet chemical and biological
oxygen demands of the contaminated subsurface will influence the remedial
design and costs.
Hydrogen peroxide has been used successfully to supply oxygen to con-
taminated aquifers. Although relatively expensive, it may be especially
useful for quickly eliminating abiotic oxygen demand in reducing aquifers.
Also, ozone has been used. An added benefit of these oxidants is that they
can also chemically oxidized organic contaminants.
Direct oxygen supply with air, oxygen generators, or liquid oxygen have
been used successfully. The oxygen is delivered to the groundwater by
sparging aboveground or in situ (Brown and Jasiulewicz 1992). Unfortu-
nately, these forms of oxygen addition are not very efficient. An alternative
approach, in situ delivery of oxygen microbubbles (colloidal gas aphrons),
holds promise as a more efficient means to deliver oxygen (Michaelson and
Lofti, 1990).
Finally, another relatively new approach involves using semi-permeable
tubules which are charged internally with oxygen gas. As groundwater
flows over the surface of the tubules, oxygen molecules diffuse across the
membranes and directly enter the groundwater without bubble formation
(Semmens et al. 1991). These systems are 100 percent efficient in oxygen
delivery. Field tests of aboveground systems indicate that dissolved oxygen
concentrations of >40 mg/L can be created, even at water flow rates of 250
gallons per minute (Piotrowski et al. 1994).
In summary, the extent of reducing conditions in the contaminated aqui-
fer will be the driving force in the selection of the means to deliver oxygen.
At some sites, it may be advisable to use a series of oxygen sources (i.e.,
hydrogen peroxide or ozone followed by air or oxygen sparging) over the
course of remediation to convert the aquifer to oxic conditions and stimu-
late elevated rates of aerobic biodegradation of the organic contaminants.
3.46
-------
Chapter 3
3.4 Natural Bioattenuation of Hazardous
Organic Compounds in the Subsurface
3.4.1 Patterns of Natural Bioattenuation
Occasionally organic contaminants enter the subsurface as spills of liq-
uids miscible in water, or as true solutions in water. Examples include eth-
ylene glycol, used as a de-icing agent for aircraft, and organic solvents con-
taining oxygen, such as acetone or butanol, used in the pharmaceutical in-
dustry. Such releases tend to move with the flow of groundwater and be
flushed away from the point of release.
Most organic contaminants, however, enter the subsurface as an oily
liquid, such as a fuel spill or a release of a chlorinated solvent. Groundwa-
ter moving through the material dissolves a small portion of the contami-
nant, which becomes a plume of groundwater contamination. Because the
contaminant mass in the oily material is much greater than that dissolved in
the groundwater, the spill can continue to maintain the plume more or less
indefinitely. As the plume moves away from its source, natural biological
processes may attenuate the contamination in the groundwater.
There are three patterns of natural biotransformation of hazardous or-
ganic compounds in the subsurface. A wide variety of organic materials are
degraded easily with oxygen or nitrate as the electron acceptor. Under
aerobic and dentrifying conditions, microbial populations quickly adapt and
reach high densities. These conditions result in the first pattern, in which
the rate of biodegradation quickly becomes limited by the rate of supply of
some nutrient, not by the microbial capacity to degrade the contaminant.
Some organic contaminants can also be degraded in the absence of oxy-
gen. In the absence of oxygen, their degradation follows a second pattern;
the rate of degradation usually is limited by the reaction rate of the active
microorganisms. Reaction rate is related to substrate concentration by a
hyperbolic function, and is best described by Monod or Michaelis-Menton
kinetics. Biodegradation under methanogenic, sulfate-reducing, and iron-
reducing conditions follow this pattern. Supplies of carbonate and iron
minerals usually are not limiting, but if the supply of sulfate is depleted,
then biodegradation under sulfate-reducing conditions would start to follow
the first pattern.
3,47
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Process Identification and Description
In the third pattern, the organic compound serves as an electron acceptor,
instead of electron donor or carbon source, its common role. Under this
pattern, the rate of transformation of the organic compound is controlled by
the electron acceptor demand exerted by other organic compounds and by
competition from other, more conventional electron acceptors.
3.4.2 Aerobic Biotransformation of Easily Degraded Com-
pounds
Hadley and Armstrong (1991) compared the number of water supply
wells in California that were contaminated with benzene, which is easily
degraded in aerobic groundwater, with the number contaminated with
trichloroethylene or tetrachloroethylene, which are not easily degraded
aerobically. Because of the ubiquity of underground storage tanlcs that leak
gasoline, one might expect that more wells would be contaminated with
benzene than with the chlorinated solvents; however, the solvents were
encountered more frequently, and at higher concentrations (table 3.5). This
finding suggests that natural biodegradation was removing the benzene
from aerobic groundwaters in California.
Wilson et al. (1985) used microcosms to estimate the potential for
biotransformation of naphthalene, methylnaphthalenes, dibenzofuran, and
fluorene in a plume of contaminated groundwater originating from a dis-
posal lagoon for wood-preserving wastes. The contaminants were not bio-
degraded in uncontaminated aquifer material from the site; however, rapid
Table 3.5
Relative Occurrence of Benzene,Trichloroethylene,and
Tetrachloroethylene in Water Supply Wells in California
Number of Median Range of
Compound Wells Concentration Concentrations
(ug/L)
Benzene
Trichloroethylene
Tetrachloroethylene
9
188
199
0.2
3.2
1.9
0.
0.
0.
1- 1.1
1-538
1 - 166
Hadley and Armstrong 1991
3.48
-------
Chapter 3
degradation was detected in aerobic aquifer material collected from the
margin of the plume. First-order rate constants were on the order of 1.0%
per week. The plume had persisted for many years, which was inconsistent
with the rapid kinetics of biodegradation in oxygenated groundwater. An
examination of the geochemistry of the groundwater revealed that the
plume was persistent in anaerobic water, but the organic contaminants were
entirely depleted in water from the plume that had been oxygenated by ad-
mixing with uncontaminated groundwater.
As a consequence of microbial adaptation in the aquifer, the rate of bio-
degradation of organic contaminants was not limited by the metabolic activ-
ity of the microorganisms. The rate of biodegradation was limited by the
rate at which oxygen was supplied to the plume. The pH was low (<4), and
there was no evidence of anaerobic biodegradation.
Borden et al. (1989) confirmed the laboratory microcosm observations
by conducting injection and withdrawal tests in the plume of creosote con-
tamination. There was minimal degradation of the PAHs when the water
injected into the aquifer did not contain oxygen; however, rapid and exten-
sive degradation of these compounds was detected when oxygenated
groundwater was injected. Oxygen was not limiting for biodegradation at
concentrations as low as 0.7 mg/L. The injection and withdrawal tests were
conducted over a period of a few days. On this time scale, the minimum
concentration of total PAHs that could be achieved was 30 to 70 (J-g/L. But,
the same compounds were undetectable in monitoring wells in the natu-
rally-bioremediated portion of the plume. Presumably, much lower concen-
trations can be achieved with a longer residence time.
Mixing processes in aquifers that can blend oxygen into a plume include
dispersion and diffusion. Dispersion is proportional to advective flow in the
aquifer; the extent of mixing is proportional to the distance the plume
moves. Diffusion is proportional to residence time of the plume in the aqui-
fer. Borden and Bedient (1986) constructed a mathematical model,
BIOPLUME, of aerobic bioremediation in aquifer that assumed an adapted
microbial population was present, that oxygen was required for biodegrada-
tion, and that oxygen transport was the rate-limiting step. The rate of
reaeration is dependent on the saturated thickness of the aquifer containing
hydrocarbon contamination and the vertical dispersion coefficient. These
parameters were used to estimate a site-specific pseudo first-order
3.49
-------
Process Identification and Description
reaeration constant, which predicted an apparent first-order biodegradation
rate constant of 25% per year.
The model adequately simulated the plume of contamination from the
creosote waste lagoon, which was forty years old (Borden et al. 1986). The
plume had significantly attenuated after traveling 400 m (437 yd), requiring
approximately twenty years. Simulations suggested that the plurne had
already reached its maximum extent.
Chiang et al. (1989) used BIOPLUMEII, a commercial version of the
model available from Rice University in Houston, Texas, to evaluate the
natural degradative characteristics of a plume of alkylbenzenes originating
from the flare pit of a natural gas plant. Microbes in the aquifer had
adapted to degrade the contaminants. As a consequence of adaptation, oxy-
gen was absent at locations where alkylbenzenes were present. At locations
where oxygen was present, alkylbenzenes were absent. BIOPLUME II
adequately predicted the concentrations of alkylbenzenes in monitoring
wells at the site. The apparent first-order biodegradation rate constant was
35% per year.
Barker, Patrick, and Major (1987) created a plume of alkylbenzenes in a
sandy water-table aquifer in Canada, then monitored the attenuation of the
plume as it moved through the natural gradient. The uncontaminated water
in the aquifer was oxygenated. The plume contained benzene, toluene, o-
xylene, and chloride as a conservative tracer. The plume initially contained
a concentration of 7.6 mg/L alkylbenzenes in a volume of 1,800 L. The
initial plume was approximately 3 m (9.8 ft) long in the direction of ground-
water flow. The plume attenuated in 1.2 years after moving 30 m (98.4 ft)
through the aquifer.
The alkylbenzenes were removed at the same rate without a lag period.
Initial removals were apparently zero-order. As the plume segregated into
regions with higher and lower hydraulic conductivity, the rate of removal of
the suite of alkylbenzenes became pseudo first-order. There was good cor-
relation between the field data and laboratory microcosm studies, but the
authors concluded that it was fortuitous.
Klecka et al. (1990) described a plume of phenolic compounds and PAHs
originating from buried wastes from a charcoal manufacturing plant. The
aquifer was aerobic, except for the region that contained the plume of con-
tamination. The plume attenuated within 60 to 100 m (66 to 109 yd) of its
3.50
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Chapter 3
source. Groundwater would require 0.7 to 1.4 years to move 100 m (109
yd). A one-dimensional model, called BIO1D, that predicts biodegradation
rates along a flow path in an aquifer was used to estimate the relative im-
portance of sorption and biodegradation for contaminant removal.
The contribution of sorption to contaminant removal was determined to
be negligible through a sensitivity analysis performed with laboratory sorp-
tion data. Microcosm studies, however, showed that microbes in the aquifer
were capable of removing the organic contaminants at rates that would
explain the attenuation of the plume. Although the reaction kinetics in the
microcosms corresponded to the rate of attenuation in the field, the authors
concluded that the "biological reaction rates [in the plume] are controlled
by the availability of dissolved oxygen in the zone of contamination."
3.4.3 Anaerobic Biodegradation of Contaminants as Carbon
or Energy Sources
It is now well established that aromatic organic compounds, such as the
alkylbenzenes, certain simple PAHs, and some nitrogen-containing hetero-
cyclic organic compounds, can be degraded in groundwater in the absence
of oxygen (Grbic-Galic 1990). The aromatic compounds are oxidized first
to phenols or organic acids, then transformed to long-chain volatile fatty
acids, which are finally metabolized to methane and carbon dioxide. Adap-
tation of the microorganisms to degrade the contaminants is slow, requiring
months to years. Destruction of the hazardous contaminants, such as the
alkylbenzenes, is associated with accumulation of fatty acids, production of
methane, solubilization of iron, and reduction of nitrate and sulfate
(Cozzarelli, Eganhouse, and Baedecker 1990; Wilson et al. 1990).
When oxygen is absent, then nitrate, sulfate, carbonate, and iron III can
serve as terminal electron acceptors (see Subsections 3.2.1 and 3.2.2).
Plumes of aromatic compounds do not contain nitrate (John Wilson, per-
sonal experience; see B. H. Wilson et al. 1990 for an illustration), indicating
that microbial adaptation to use nitrate as a terminal electron acceptor oc-
curs readily, and ambient concentrations of nitrate are quickly consumed.
Natural attenuation of plumes of aromatic contaminants through nitrate
respiration should be similar to oxidative biodegradation.
Water table aquifers, particularly in agricultural areas, contain consider-
able electron-accepting capacity in the form of nitrate. A typical water
3.51
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Process Identification and Description
table aquifer may contain 5 mg/L of dissolved oxygen and 5 mg/L nitrate-
nitrogen. There is, however, potentially three times the electron-accepting
capacity in the nitrate as in oxygen. To the authors' knowledge, the poten-
tial contribution of nitrate has not been factored into models of the natural
aerobic attenuation of organic contaminants.
3.4.3.1 Sulfate-Reducing Conditions
Plumes of aromatic compounds commonly contain sulfate, and, depend-
ing on its concentration, sulfate can also be an important electron acceptor.
Acton and Barker (1992) monitored the attenuation of benzene, toluene, o-
xylene, m-xylene, ethylbenzene, and 1,2,4- trimethylbenzene in a forced-
gradient injection experiment in a plume from an active landfill owned by
the City of North Bay, Ontario, Canada. Leachate was collected from the
plume, amended with approximately 200 ug/L of the six organic com-
pounds, bromide as a tracer, and 12 mg/L sulfate, and then re-injected into
the aquifer. After a 30-day lag period, toluene and m-xylene were de-
graded, but the other organic compounds were not.
Thierrin et al. (1992a, b) described the bioattenuation of a plume from a
gasoline spill in Perth, Australia, under sulfate-reducing conditions. A
plume of benzene in groundwater extended at least 420 m (459 yd) from the
spill, while toluene was completely degraded within 200 m (219 yd). Un-
contaminated groundwater in the aquifer contained more than 20 mg/L of
sulfate. The concentration of sulfate in groundwater from the plume was
below that detected in uncontaminated groundwater from the site (Thierrin
et al. 1992b). In an experiment, 400 L of groundwater was extracted from
the plume and amended with bromide and a solution of fully deuterated
benzene, toluene, p_-xylene, and naphthalene as tracers of contaminant trans-
port and biodegradation; the amended groundwater then was injected back
into the plume 80 m (87 yd) from the source area. The deuterated com-
pounds could be distinguished from the natural aromatic compounds with
good precision and sensitivity using mass spectrometry.
The concentration of the bromide and deuterated aromatic compounds
was monitored over time in a series of cluster wells that sampled water
above, within, and below the plume (table 3.6 on page 3.53). The data are
from a series of vertically-stacked cluster wells down gradient of the spill
area.
3.52
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Chapter 3
Table 3.6
Rates of Bioattenuation Under Sulfate-Reducing
Conditions in a Plume From a Gasoline Spill
Depth Below Water Table
Travel Time of Bromide Benzene Toluene p-Xylene Naphthalene
IIICICIM vvttfkS ptik-cm ituiuvcu j^ti wcei^
Slightly oxygenated water at the upper edge of the contaminant plume
0.75/>10 above above above
plume plume plume
1.0/6.2 6.5 >81 >32
Anoxic/Sulfate-reducing Zone
1.25/65
1.5/58
1.75/54
2.0/5.5
2.25/5.8
2.5/5.9
2.75/5 9
3.0/>10
3.0
1.2
none
none
none
14
20
below
plume
35
6.6
7.0
5.1
none
3.6
8.3
below
plume
1.7
none
3.2
2.1
1.6
2.5
9
below
plume
above
plume
>32
>40
6.6
14
15
12
18
30
below
plume
Thiernnelal 1992b
Total attenuation was corrected for sorption and dilution to estimate
bioattenuation. Rates of removal were greatest at the margins, perhaps
because the supply of sulfate and oxygen there was greater. In the interior
of the plume, the attenuation of benzene was actually slightly less than that
of bromide.
Thierrin et al. (1992a) compared the bioattenuation of the tracer plume of
deuterated aromatic compounds to the attenuation required to fit a model of
the full-scale plume. These two independent assessments of the behavior of
aromatic compounds were in close agreement for benzene and toluene
(table 3.7 on page 3.54).
Benzene was not removed in the heart of the plume. The rates of
bioattenuation of toluene, the xylenes, a trimethylbenzene, and naphthalene
were adequate to degrade these compounds within a few hundred meters of
travel (table 3.7 on page 3.54).
3.53
-------
Process Identification and Description
Table 3.7
Comparison of Kinetics of Bioattenuation of an Artificial Plume of
Deuterated Aromatic Organic Compounds and a Full-Scale Plume
Under Surrate-Reducing Conditions
Tracer test with deuterated Required to make computer
Organic compound organics in the field model fit field data
—percent per week depleted—
Benzene
Toluene
Ethylbenzene
o-Xylene
p-Xylene
m+p-Xy\ene
1 ,3,5-Trimethylbenzene
Naphthalene
<06
48
no data
no data
2.2
no data
no data
15
<06
4.0
2.1
3.9
not resolved
2.9
2.7
3.0
Thiernnetal. 1992a
3.4.3.2 Methanogenic Conditions
The important electron sinks in many anaerobic plumes are the iron III
and mixed valence iron minerals in the aquifer matrix (Lovley 1991) and
bicarbonate in carbonate/bicarbonate-buffered groundwaters. In the case of
carbonate and iron III, the supply of electron acceptor is not limiting for
contaminant biodegradation. The metabolic activity of the microorganisms
becomes the rate-limiting step. As a consequence, the rate of reaction is
controlled by the density of active organisms and by the concentration of
metabolizable compounds.
Godsy and his associates in the U.S. Geological Survey have studied
methanogenic transformation of a series of phenols in a plume originating
from a disposal lagoon for wood-creosoting wastes. Microcosm studies
were used to estimate the kinetic constants of degradation of four phenols
(Godsy, Goerlitz, and Grbic-Galic 1992b).
The concentrations of phenols at which microbial growth is one-half
maximum (Ks) were low, near 1 mg/L (table 3.8 on page 3.55). The maxi-
mum growth rate (\i) was slow in comparison with that of other microorgan-
isms. The yield coefficients (transformation of substrate to biomass, Y)
3.54
-------
Chapter 3
were low, and the rate of self-consumption (Kd) was low or undetectable.
Further, there seemed to be an upper limit to the microbial biomass in the
microcosms (Godsy, Goerlitz, and Grbic-Galic 1992b). Although Monod
growth kinetics predicted that biomass would reach much higher concentra-
tions, the final biomass in the microcosms ranged from 0.030 to 0.044 mg/
L. Apparently, there is some density-dependent constraint, such as preda-
tors of the phenol-degrading microorganisms, that limits the development of
biomass.
Table 3.8
Kinetics of Biodegradation of Organic Contaminants in
Microcosms Under Methanogenic Conditions
Compound
Phenol
2-Methylphenol
3-Methylphenol
4-Methylphenol
Umax
(per day)
Oil
0.044
0.10
0.10
Ks
(mg/L)
1.3
0.25
0.55
3.3
Y
(mg biomass/
mg phenol)
0.004
0.003
0002
0042
Kd
(per day)
0.001
0.002
0.000
0.000
Adapted from Godsy et al. 1992b
These observations support a provisional paradigm for the function of
anaerobic biodegradation in aquifers in which the concentration of active
biomass is controlled by some density-dependent constraint. Under this
scenerio, the density of active organisms is uniform throughout the plume
and is independent of the concentration of aromatic contaminants. The
degradation of an individual aromatic contaminant will be zero-order at
concentrations above 1 mg/L and first-order at lower concentrations.
Therefore, it should be possible to model bioattenuation of aromatic con-
taminants with a simple one-dimensional model, such as BIO ID, calibrated
from microcosm studies or from field estimates of contaminant attenuation.
Godsy, Goerlitz, and Grbic-Galic (1992a) found a high degree of correla-
tion between the behavior of laboratory microcosms and the field-scale
3.55
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Process Identification and Description
plume of creosote contamination mentioned earlier (table 3.9). The time
taken for water to move a certain distance from the source along a flow path
was compared to the residence time in the microcosm.
Table 3.9
Comparison of Bioattenuation of Organic Contaminants in a
Plume of Contaminated Groundwater and in Microcosms
Simulating the Plume Under Methanogenic Conditions
Behavior
C3 to C6 volatile fatty acid converted to
acetate and tpethane
Benzoic acid degraded
Phenol degraded
Quinolme and isoquinolme degraded
2-,3-,4-Methylphenol degraded
Quinoline and isoqumoline degraded
Residence in
Microcosm
(days)
OtoSO
50 to 99
OtoSO
100 to 200
100 to 180
Residence in Plurne
(m along flow path)
OtoSO
53 to 98
53 to 98
53 to 125
produced
The average seepage velocity in the plume was 10m per day
Plumes of alkylbenzenes from fuel spills behave the same way. Wilson
et al. (1990) described the methanogenic bioattenuation of benzene, toluene,
and xylenes (BTX) in groundwater contaminated by a spill of aviation gaso-
line. Attenuation of total BTX was measured quarterly over a 4-year period
in wells along a flow path running through the centerline of the plume.
Removal followed first-order kinetics, varying from 10 to 34% per week.
When a purge well field went on line, one of the monitoring wells was iso-
lated from the source area. Alkylbenzene concentrations dropped rapidly in
the isolated well. Interestingly, toluene disappeared more than twice as
rapidly as benzene. Shortly after the BTX compounds disappeared in the
isolated well, core material was acquired from that region for a laboratory
microcosm study. The kinetics of bioattenuation were very similar for in
situ measurements along the flow path, in stagnant water near the isolated
monitoring well, and in the microcosm study (table 3.10).
The independent field-scale estimates of bioattenuation were in agree-
ment within a factor of three, and those of the microcosm, within a factor of
3.56
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Chapter 3
ten. The seepage velocity was known within a factor of two.
Cozzarelli, Eganhouse, and Baedecker (1990) reported anaerobic
bioattenuation of alkylbenzenes in a plume of contamination from a spill of
crude oil in Minnesota. Methane was detected in the groundwater and long-
chain volatile fatty acids accumulated. Bioattenuation in monitoring wells
in a transect along the centerline of the plume was reported. Toluene and
o-xylene disappeared without a lag period, indicating that the microorgan-
isms had adapted to degrade the compounds. Both were depleted within 12
m (40 ft) of the source area. Ethylbenzene degradation began after toluene
Table 3.10
Bioattenuation of Benzene,Toluene, and Xylenes in Microcosms
and Methanogenic Groundwater at an Aviation Gasoline Spill Site
Compound Microcosms Flow path Isolated well
percent per week
Benzene
Toluene
m+/>-Xylene
o-Xylene
All Xylenes
50 5
30 130
40
50
3
17
47
10
and o-xylene had disappeared. Benzene was depleted without a lag period
in the presence of toluene and o-xylene; however, the rate of benzene deple-
tion slowed greatly after these compounds disappeared. The average seep-
age velocity in the plume is 0.1 m/day (0.3 ft/day) (personal communica-
tion, Philip Bennett, U. of Texas at Austin). This value was used to express
the attenuation between wells as a first-order rate constant (table 3.11 on
page 3.58).
When the plume mixed with oxygenated groundwater, elevating the
dissolved oxygen concentration above 1.0 mg/L, all the aromatic hydrocar-
bons disappeared. This occurred within 100 m (109 yd) of the spill.
3.57
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Process Identification and Description
Wilson, Kampbell, and Armstrong (1993) studied bioattenuation of aro-
matic hydrocarbons in a plume produced from gasoline leaking from a un-
derground storage tank. Methane was detected in the groundwater and
acetate accumulated. Bioattenuation of alkylbenzenes was estimated be-
tween the spill area and monitoring wells located 100 days down gradient
and 200 days downgradient. The rates of anaerobic bioattenuation were
very similar to those rates in the crude oil spill (table 3.11) and the aviation
gasoline spill (table 3.10 on page 3.57).
3.4.4 Biodegradation and Organic Contaminants as Elec-
tron Acceptors
In the absence of oxygen, halogenated organic compounds can serve as
electron acceptors. The most important example of this process is the se-
quential reductive dehalogenation of tetrachloroethylene to TCE, to cis or
frans-dichloroethylene, to vinyl chloride, and finally to ethylene, ethane,
and methane (Vogel and McCarty 1985; de Bruin et al. 1992).
This process is well-documented qualitatively, but there have been few
full-scale studies of the kinetics of dehalogenation. Researchers from the
Trenton, New Jersey Office of the U.S. Geological Survey and the R.S.
Kerr Laboratory of the U.S. EPA compared the kinetics of TCE dechlorina-
tion in microcosms to the rate of dechlorination in a large plume of TCE at
the Picatinny Arsenal in New Jersey (Imbrigiotta et al. 1991).
Table 3.11
Rates of Bioattenuation of Alkylbenzenes in Plumes from a Gasoline
Spill and a Crude Oil Spill Under Methanogenic Conditions
Compound
Crude Oil Spill
Gasoline Spill
(0 to 100 days from
source)
Gasoline Spill
(100 to 200 days
from source)
percent per week
Benzene
Toluene
o-Xylene
p-Xylene
m-Xylene
Ethylbenzene
12
50
40
19
13
21
16
15
13
12
3
26
8
9
8
13
3.58
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Chapter 3
Trichloroethylene was discharged from a plating shop to a shallow dry
well from 1973 to 1978. A presumed source area of nonaqueous phase
liquids produced a plume of TCE, cw-dichloroethylene, and vinyl chloride.
The plume extended 488 m (1,600 ft) from the source area to the point of
groundwater discharge to surface drainage. Although the average residence
time for water was 0.7 to 1.5 years (Voronin 1991), the concentration of
TCE in the plume did not decline significantly within 4.6 years after TCE
discharge to the aquifer ceased (Imbrigiotta et al. 1991); this indicates that
the groundwater plume was stable and sustained by dissolution of oily
phase liquids in the source area.
The highest concentrations of TCE were in the water table aquifer, just
above the confining layer at the mid-point between the source and point of
discharge. The concentration of TCE declined an order of magnitude from
the area of highest concentration to the point of discharge to surface water.
First-order rate constants for TCE attenuation, estimated from average seep-
age velocities in the plume, ranged from 3 to 9% per week. Laboratory data
from Wilson (1988), Wilson et al. (1991), and Ehlke et al. (1991) are pre-
sented in table 3.12 (on page 3.60). The rates of TCE disappearance in the
microcosms varied, but, in general, were consistent with the rate of attenua-
tion at field scale. On the other hand, c/s-dichloroethylene (DCE) only
disappeared in microcosms constructed from regions of the aquifer contain-
ing significant amounts of it.
A variety of organic compounds can serve as electron donors for biologi-
cal reductive dechlorination (DiStefano, Gossett, and Zinder 1992). Some
microcosms were amended with volatile fatty acids, toluene, and p-cresol in
an attempt to stimulate reductive dechlorination (table 3.12 on page 3.60)
(Wilson et al, 1991; Ehlke et al. 1991). These compounds actually sup-
pressed reductive dechlorination in some experimental treatments. To date,
the chemical reducing agent that facilitates biological reductive dechlorina-
tion in the TCE plume at Picatinny Arsenal has not been identified; how-
ever, the process resulted in a substantial destruction of contaminant mass
before the TCE discharge to surface water.
3.59
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Process Identification and Description
Table 3.12
Depletion of Trichloroethylene (TCE) and cis-Dichloroethylene
(DCE) in Microcosms Constructed With Core Material From
the TCE Plume at Picatinny Arsenal, New Jersey
TCE with DCE with
TCE Alone Amendments DCE Alone Amendments
Incubation
(Weeks) —(percent per week)-—
Near source area, coarse sand and fine gravel
40 53 39
14 <0.1
Near mid point, the area of highest TCE and cis-DCE concentration, fine sands and silts
25 0.9 0 6
14 18, <01 33
Near the point of discharge to surface drainage, fine sands and silts.
40 5.8 11
25 4.6
14 60
Uncontaminated material near point of discharge to surface water, medium sands.
25 5.1
14 0.5
3.5 Bioremediation Processes
The following physical processes can be used for the biological
remediation of liquid wastes, sludges, and contaminated surface soils, sub-
surface sediments, and air:
• in situ treatment of contaminated soil or sediment; and
• ex situ treatment of contaminated material in an aboveground
reactor or prepared bed. The remediated materials are disposed
of at an appropriate site.
In situ processes include land treatment, bioventing, liquid delivery, and
air sparging; their developmental status is shown in Table 3.13 (on page
3.61). Ex situ processes include treatment in aboveground reactors, land
treatment, composting, soil piles, and biofilters; their developmental status
is shown in Table 3.14 (on page 3.62).
The selection of the biological treatment process is based on the physical
and hydrogeological characteristics of the site, the chemical nature of the
3.60
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Chapter 3
waste, and the clean-up levels required. For example, in situ or ex situ land
treatment or composting generally might be feasible for bioremediation of
soil contaminated with diesel fuel; however, because of land constraints or
the presence of a seasonally-high water table in a particular site, in situ land
treatment may not be feasible. A number of summaries, reports, and texts
describe the use of these processes in treating diverse hazardous and non-
hazardous wastes and contaminated soils (American Petroleum Institute
1983; Brown, Evans, and Frentup 1983; Sims, Sims, and Matthews 1989;
Overcash and Pal 1979; Loehr and Malina 1986; US EPA 1990; Thomas,
Ward et al. 1992; Nyer 1992).
Soil, sediment, or a bulking material is used in each of these three pro-
cesses to facilitate the bioremediation activity. Soil and sediment provide
nutrients, microorganisms, and buffering capacity for in situ and ex situ
land treatment systems, and bulking material increases porosity to facilitate
oxygen transfer in composting. In many cases, the soil or sediment is also
being bioremediated. Generally, a mixture of contaminants is treated. The
complex bonding forces exhibited by various soil fractions, particularly
Table 3.13
Developmental Status of In Situ Bioremediation Technologies
Developmental Status
Technology
Natural bioattenuation
Waste Treated
Petroleum hydrocarbons
Chlorinated solvents
Laboratory/
Developing
Demonstrated
in Field Trial
+
Proven
Full Scale
Land treatment Petroleum hydrocarbons
PAHs; sludges; contaminated soils
Bioventing Petroleum hydrocarbons
Chlorinated solvents
Liquid delivery Petroleum hydrocarbons
(aerobic) Nonchlonnated solvents
Chlorinated solvents
Pesticides
Liquid delivery Monoaromatic hydrocarbons
(anaerobic) Chlorinated solvents; PAHs
Air sparging Petroleum hydrocarbons
Chlorinated solvents
3.61
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Process Identification and Description
Table 3.14
Developmental Status of Ex Situ Bioremediation Technologies
Developmental Status
Laboratory/ Demonstrated Proven
Technology Waste Treated Developing in Pilol Study Full Scale
Suspended growth Aromatic and aliphatic hydrocarbons; mono- and
Dichlorinated aromatic and aliphatic hydrocarbons
(>50 mg/L)
Fixed-film Petroleum hydrocarbons
Submerged fixed-film Aromatic and aliphatic hydrocarbons
Activated carbon-based Pesticides; volatile organic compounds; low
Concentrations of chlorinated organic compounds
Slurry-phase PAHs; chlorinated solvents, munitions; pesticides
Land treatment Petroleum hydrocarbons; PAHs
Soil-pile Petroleum hydrocarbons, PAHs; nonchlonnated
Solvents; chlorobenzenes
Composting Munitions
Biofilter Hydrocarbons; chlorinated alkanes, alkenes
clays and organic matter, can affect the treatability of organic compounds in
soils (see Subsection 3.2.1.2.5 Contaminant Unavailability and Section 3.3
Site Characterization Relevant to In Situ Bioremediation). Therefore, an
understanding of soil characteristics is important to the effective application
of these processes.
Other parameters that affect the performance of bioremediation systems
include the following:
• Equalization. For ex situ bioremediation processes, equalization
minimizes variations in the contaminant load. This is important
because biological treatment is sensitive to variations in organic
loadings.
• Nutrient Management. Nutrient management is important be-
cause an insufficient amount of nutrients will slow removal of
organic compounds. The principal inorganic nutrients are nitro-
gen and phosphorus. Trace amounts of potassium, calcium,
3.62
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Chapter 3
sulfur, magnesium, iron, and manganese are also required for
optimum biological growth.
• Oxygen Supply and Aeration. An adequate supply of oxygen is
critical to an environment in which aerobic organisms can grow
and metabolize the organic material. The oxygen can be pro-
vided as atmospheric oxygen or in the form of oxygen-supplying
compounds (e.g., peroxides). In composting, diffusion of oxy-
gen is related to the physical and chemical characteristics of the
compost. If the supply of oxygen does not keep pace with the
needs of the organisms, it will become a limiting factor. Me-
chanical aeration, mixing, or the addition of oxygen-containing
compounds are effective solutions.
• Temperature. Biological growth can occur within a wide range
of temperatures, although most microorganisms are active pri-
marily between 10° and 35°C (50° to 95°F). The rate of bio-
chemical reactions in cells increases with temperature up to a
maximum, above which the rate of activity declines as enzyme
denaturation occurs and organisms either die or become less
active.
• pH Control. In general, neutral or slightly alkaline pH levels
favor biological growth. The optimum pH range for most organ-
isms found in biological treatment systems is between 6.0 and
8.0. Treatment effectiveness is generally not affected by
changes within this range; however, pH levels outside of this
range can lower treatment performance. The pH of any waste
remediated in these processes should be monitored and adjusted
during pretreatment, such as during the feed preparation for
composting.
• Microorganisms. The nature and quantities of organic com-
pounds will affect their biodegradability. In all biological treat-
ment systems, the organisms are naturally subjected to a selec-
tion process in which the organisms capable of efficient biodeg-
radation under the given circumstances increase their numbers.
(See Subsection 3.2.1 Microbial Ecology and Physiology)
The complex nature of wastes, sludges, and contaminated soils presents
a unique challenge in terms of successful remediation and minimization of
3.63
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Process Identification and Description
impacts on other parts of the environment. When determining the applica-
bility of bioremedial technologies, treatment test results from a series of
actual site samples, representative of the variety of site conditions, should
be evaluated (see Section 3.3, Site Characterization Relative to In Situ
Bioremediation, Introduction). In evaluating the potential effectiveness of a
technology, the entire treatment train configuration should be considered,
including:
• materials handling and preprocessing;
• contaminant destruction, physical transfer, or immobilization;
and
• effluent, emission, or residue treatment technologies.
Where possible, the results of previous studies and successful treatment
applications should be consulted to aid in the evaluation.
3.5.1 In Situ Bioremediation Technologies
3.5.1.1 Unsaturated Zone
3.5.1.1.1 Land Treatment Introduction. In situ land treatment is a man-
aged treatment and disposal technology that entails the application of waste,
sludge, or contaminated soil to uncontaminated surface soils at a site and
then tilling or plowing the material into the surface soils. This process may
also be applied directly to surface soils that have been contaminated by
chemical or waste spills. The design and operation of a land treatment fa-
cility is based on sound scientific and engineering principles, as well as on
extensive, practical field experience.
The objectives of in situ land treatment are:
• biological degradation of organic waste constituents;
• immobilization of inorganic waste constituents; and
• avoiding bioaccumulation of waste constituents that may be
detrimental to human health and the environment.
Land treatment uses the assimilative capacity of the soil to decompose
and contain the contaminated material in the surface soil layer. The mecha-
3.64
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Chapter 3
nisms of transformation of contaminated soils are illustrated in figure 3.4.
Many studies (American Petroleum Institute 1983; Brown, Evans, and
Frentrup 1983; Loehr, Martin, and Neuhauser 1992; Loehr et al. 1990) have
shown that oil, metals, and other constituents of environmental concern are
successfully treated or immobilized by land treatment systems.
Land treatment has been used to treat the following:
• wastes from coal gasification and liquefaction, food processing,
leather tanning, paper and pulp production, petroleum refining,
and wood-preserving industries;
• sludges and contaminated soils at Superfund sites; and
• municipal wastewater and sludge.
The process has been used under a wide range of hydrogeologic condi-
tions and in the major climatic regions of the United States, Europe, and
Canada. Land treatment usually requires minimal management compared
with other bioremediation techniques, but does require aeration of the soil,
Figure 3.4
Schematic of an In Situ Land Treatment Process
Volatile and Gaseous
Emission
Biological and
Chemical
Reactions
Treatment Zone
(up to 5 ft)
A
Precipitation
IX / Kun-un ana
/ /Contaminated Runoff Control /fl
/ / Materials Uniformly .//
/ / Incorporated in the Soil ^/
t
1
^ Tillage Depth
I (Typically 6 to 1 2 Inches)
\
Percolation of Soil Water
3.65
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Process Identification and Description
maintenance of optimum soil characteristics (i.e., pH and nutrient balance),
and available land.
The soil and waste mixture is managed in a manner that:
• enhances immobilization of waste by soil;
• stimulates degradation of waste by indigenous micro flora;
• minimizes volatilization and leaching of waste out of the treat-
ment area; and
• controls surface water runoff.
The mechanisms of immobilization and degradation include sorption,
hydrolysis, photolysis, chemical degradation, and biodegradation. Factors
that affect biodegradation in land treatment include the type and concentra-
tion of the waste, presence of waste-degrading organisms, pH, temperature,
and the availability of oxygen, water, and nutrients. Usually, the indig-
enous soil microflora is stimulated to degrade the wastes; however, micro-
organisms that have adapted to degrade the contaminants may be: added if
contaminant-degrading activity is absent in the waste, sludge, or contami-
nated soil.
Land treatment uses the assimilative capacity of the soil to decompose
and contain the applied waste in the surface soil layer (usually the top 15 to
30 cm (6 to 12 in.)). This soil layer is referred to as the zone of incorpora-
tion (ZOI). The ZOI and the underlying soils, where additional treatment
and immobilization of the applied waste constituents occur, constitute the
treatment zone.
Since few organic substances are completely resistant to biodegradation,
many organic contaminants can be treated by an in situ land treatment pro-
cess. But, numerous organic contaminants require a long time to be
remediated because they have chemical structures that resist biodegrada-
tion. Such compounds include high molecular weight cross-linked poly-
mers, highly branched hydrocarbons, and halogenated compounds. Some
of these recalcitrant materials are naturally produced (e.g., coal, tannins,
and humic substances), and some are synthetic (e.g., polyethylene). Other
compounds are somewhat more susceptible to biodegradation yet are usu-
ally classified as nonbiodegradable because of the long time required for
appreciable biodegradation to occur. Fortunately, remediation times are
long in land treatment units and even seemingly recalcitrant organic
3.66
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Chapter 3
conpounds in the applied wastes and sludges frequently are degraded and
immobilized.
Organic compounds that are barely water-soluble present difficult bio-
degradation problems for microorganisms since the organisms and their
enzymes must function in an aqueous solution. Therefore, some organic
contaminants whose chemical structures should permit biodegradation are
slow to decompose because of limited water solubility.
Metals and other inorganic constituents typically are immobilized in the
soil and are not degraded or otherwise removed from the site; however,
their redox states may be altered. Wastes with high-metal content, such as
electroplating sludges, are not prime candidates for land treatment.
Table 3.15 (on page 3.68) arrays the important waste, site, and opera-
tional factors that need to be considered in the design and operation of an
effective in situ land treatment unit. Other major factors include soil mois-
ture and pH and available nutrients. Water is essential for microbial activ-
ity, and nonsporeforming microorganisms will die or remain inactive at
very low water concentrations. Alternatively, excessive water levels do not
directly harm microorganisms, although they do retard aerobic metabolism
by minimizing oxygen transfer in the soil. The aerobic biodegradation of
organics in soil is greatest when the moisture content is 50 to 70% of the
soil field capacity (Bartha and Dibble 1979). Wet soils also limit site access
for waste application and tilling operations.
Although anaerobic degradation occurs in soil, it should be limited for
effective land treatment because anaerobic biodegradation is slower and
less complete for most organic contaminants. In addition, most metals are
more water soluble in a reduced state and more susceptible to leaching.
The optimum pH for soil biodegradation lies between 6 and 8; however,
effective biodegradation can occur outside this range. In many land treat-
ment operations, soil pH is kept above 6 to avoid metals migration, as well
as to optimize microbial activity. Low soil pH can be modified by adding
agricultural lime to the treatment soils, raising the pH to the appropriate
range. Microbial growth and metabolism can be reduced by abrupt pH
changes; therefore, soil pH should be modified cautiously.
3.67
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Process Identification and Description
Table 3.15
Land Treatment Design and Operating Factors
Waste Characteristics
Physical Composition Salts
Orgamcs Nutrients
Metals pH
Site Factors
Soil Characteristics Climate
Topography Temperature
Soil texture Precipitation
Soil moisture content
Cation Exchange Capacity Hydrogeology
Soil pH Depth to seasonally high water table
Soil microorganisms Depth to usable aquifer
Nutnents Proximity to surface water
Operational Factors
Waste Application Soil Amendments
Organic loading Nutrients
Hydraulic loading Moisture
Frequency of application pH control
Method of application
Monitoring
Storm Water Management Waste Characteristics
Run-on/Runoff control Soil
Leachate
Waste Incorporation Vegetation (if grown)
Depth of incorporation Runoff (if any)
Frequency of cultivation
Available oxygen is critical for any form of aerobic bioremediation. Soil
microorganisms use oxygen transferred to soil water from the atmosphere.
Oxygen availability is a function of:
• the amount of void space in the soil;
• the partial pressure of oxygen in the soil atmosphere;
• the oxygen transfer rate from soil atmosphere to soil water; and
• the rate at which soil microorganisms are using the available
oxygen.
3.68
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Chapter 3
It is possible to maintain aerobic conditions in a land treatment unit by:
• keeping the soil unsaturated;
• moderate tilling;
• avoiding unnecessary compaction (heavy trucks, etc.); and
• limiting the loading of rapidly-biodegradable matter so that oxy-
gen demand does not exceed the oxygen transfer rate.
Soil texture and permeability also affect aerobic conditions. Sand or
loam is conducive to aerobic bioremediation, while heavy clay is not.
Sandy soils permit rapid infiltration of liquids and oxygen, but also may
allow pollutant migration. Clayey soils provide better containment, but can
reduce infiltration of irrigation water and precipitation. Clayey soils also
reduce air exchange, potentially causing anaerobic conditions.
Site Preparation and Equipment. Site preparation for in situ land treat-
ment includes: (1) removing site vegetation, (2) separating the sites into
several plots, and (3) constructing an elevated roadway/dike around the
facility to control surface run-on and runoff and permit site access. Many
land treatment sites are graded to promote surface drainage and collect run-
off. Slopes in excess of 5% are generally not recommended because of
erosion and runoff control problems. A 1 to 2% grade is common, provid-
ing controlled runoff and preventing ponding of added water.
Site preparation activities vary depending on the existing site characteris-
tics and regulatory requirements. Following are some examples:
• removing trees and rocks for site access and ease of tilling;
• digging drainage ditches to intercept seasonally-high perched
water table and runoff;
• adjusting soil pH for low pH soil; and
• contouring, terracing, and grading to intercept and divert off-site
run-on, contain on-site runoff, and provide access to the site.
During operation, the land treatment unit may need fertilization and pH
control (by liming of the soil). The amount and frequency of lime applica-
tion depends on the pH and the buffering capacity of the soils, but typical
ranges are from 1.8 to 3.6 tonne (2 to 4 ton) of lime per acre every 1 to 3
years on acid soils to maintain near neutral soil pH (American Petroleum
3.69
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Process Identification and Description
Institute 1983). Generally, soil monitoring data are used to determine the
lime requirement.
Nutrient applications (nitrogen, in particular) may enhance organic deg-
radation rates in the initial period, but may not be necessary in subsequent
years. Over-application of nitrogen fertilizers may result in excessive
leaching of soluble nitrates. Other nutrients, in addition to those applied,
are available at land treatment sites. Many agricultural soils have a high
reserve of phosphorus. Mineralization of organic nitrogen in the soil, the
applied waste, and sludges may provide a significant nitrogen input and can
recycle applied nutrients after repeated waste applications.
The waste, sludge, or contaminated soil is applied uniformly and tilled in
at a usual depth of 15 to 30 cm (6 to 12 in.), although depths of up to 46 cm
(18 in) are practical with heavy tilling equipment. The cost of the equip-
ment and the available land area will affect the selection of the tilling depth.
After the application, distribution, and incorporation of the waste, additional
tilling generally is performed to:
• increase the available area for soil microbial contact with con-
taminants;
• maintain aerobic and homogenous conditions; and
• maintain a loose, moist, and well-mixed soil to maximize or-
ganic contaminant decomposition.
Performance. For hydrocarbon-contaminated wastes and soils, removal
of total petroleum hydrocarbons (TPHs), as measured by gas chromatogra-
phy (GC) methods, will be in the range of 90 to 99%. In some instances,
treatability studies have indicated removals of the following nature: PAH
such as naphthalene, acenaphthene, and fluorene - 80-95%; higher ring
PAH - 30-70%. Field studies have achieved removals in the same range
when proper conditions for biodegradation have existed. Loss rates for
organic constituents are typically described by the "half-life" of the organic
compound. Compounds with higher aqueous solubilities have relatively
shorter half-lives than less soluble compounds. The average half-fife for
typical oil and grease found in refinery wastes and oily sludges in land
treatment units ranges from 50 to 150 or more days. Removal or loss rates
vary widely and depend on the type of hydrocarbons and site-specific fac-
tors affecting the rates of biological activity.
3.70
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Chapter 3
Most of the effort in defining degradation rates for hydrocarbon-contami-
nated materials has been spent evaluating loss rates for volatile and
semivolatile compounds: BTEX — benzene, toluene, ethylbenzene, xy-
lenes—and PAHs. The PAHs are of interest because several are suspected
carcinogens. The BTEX compounds are removed (lost) rapidly by a combi-
nation of volatilization and biodegradation; the half-lives are generally less
than a week.
The primary mechanisms influencing the fate of PAHs are volatilization,
adsorption/desorption, and biological oxidation. In solids (e.g., soil and
sludge), no significant volatilization or biological oxidation of PAHs will
occur unless they desorb from the solid phase. Therefore, PAH adsorption/
desorption will be the rate-limiting step in the fate of PAHs. Only the two-
and three-ring PAHs have a high volatilization potential. The more com-
plex PAHs can be biodegraded, but have long half-lives of weeks to
months.
The owner or operator of a land treatment unit must establish a program
to ensure that organic and inorganic contaminants applied are degraded,
transformed, or immobilized within the treatment zone. Federal and state
regulations define the principal components of an in situ land treatment
facility as:
• the wastes to be applied;
• the design and operating measures necessary to maximize degra-
dation, transformation, and immobilization of the waste constitu-
ents; and
• an unsaturated zone monitoring program.
Before applying waste, the owner or operator must demonstrate that
constituents can be completely degraded, transformed, or immobilized in
the treatment zone. Generally, a treatment demonstration is required to
establish that the operating practices at the site will protect human health
and the environment, considering the characteristics of the waste, soil, and
climate at the site. The treatment demonstration will be used to determine
permit requirements and operating principles. The U.S. EPA has published
a permit guidance manual governing the treatment demonstration (US EPA
1986).
In situ land treatment units also must have a groundwater monitoring
program to detect and correct any groundwater contamination. The permit
3.71
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Process Identification and Description
guidance manual (US EPA 1986) prescribe the requirements for runoff and
run-on controls, growth of food chain crops on such sites, closure and post-
closure care, and record keeping.
Advantages/Disadvantages. One advantage of in situ land treatment is
that the organic and inorganic contaminants are reduced in concentration
and/or immobilized and not simply disposed. Other advantages include low
capital and operational costs, low-to-moderate manpower and maintenance
requirements, and no long-term liability ensuing from off-site disposal. If
the unit is overloaded or poorly operated, however, organic and inorganic
compounds might leach below the land treatment unit. Other disiad vantages
are the need for a large land area and limited control over volati le organic
loss to the atmosphere.
3.5.1.1.2 Bioventing Introduction. Bioventing is the use of induced air
movement through unsaturated soils, with or without nutrient addition, to
reduce soil contamination through biodegradation (Hinchee 1993; Van Eyk
and Vreeken 1989b). The process stimulates the indigenous microorgan-
isms to convert organic contaminants, such as petroleum hydrocarbons, to
less hazardous substances, especially carbon dioxide and water.
Most bioventing systems have used air recovery wells (see figure 3.5 on
page 3.73), such as those used in in situ vapor recovery systems, to move
air and thus provide oxygen for the indigenous bacteria. Many systems that
were designed as in situ vapor recovery systems to physically remove the
contaminants were found to stimulate biodegradation as well; these were
subsequently relabeled as bioventing systems (Bennedsen, Scott, and
Hartley 1987).
Some systems have been designed with air injection wells either alone
(figure 3.6 on page 3.74) or in conjunction with air recovery wells (figure
3.7 on page 3.74). The use of air injection can be as effective as air recov-
ery in terms of providing air for biodegradation. Air injection should not be
used near buildings because it could result in exposure or explosion haz-
ards. It can be used to treat heavy petroleum blends in areas away from
buildings and utility trenches, and in conjunction with an air recovery sys-
tem.
Systems that are designed primarily to promote biodegradation, as op-
posed to physical removal, will have the air recovery wells located outside
the most heavily-contaminated area (figure 3.8 on page 3.75) and incorpo-
3.72
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Chapter 3
Figure 3.5
Bioventing Design to Maximize Recovery of Volatile Compounds
Ground Surface
Water Table
rate significantly lower air flows than systems designed for physical re-
moval. This approach has the advantage of requiring less offgas treatment
because it focuses on destroying the contaminants rather than transferring
them to a different medium.
Nutrient addition is not always incorporated in bioventing systems. In
some cases it is not necessary, and in others it may be very difficult to ac-
complish. Where the soil types and site infrastructure permit, aqueous nu-
trient solutions can be percolated from the surface. In low-transmissive
soils or where surface access is restricted, percolation of nutrients may be
difficult or infeasible. Attempts to introduce gaseous phase nitrogen
sources have not yet been successfully demonstrated.
Scientific Basis. The scientific basis for this technology has the follow-
ing aspects:
• the metabolic capabilities and requirements of microorganisms;
• the dynamics of air flow through unsaturated soils;
• the adsorption and desorption of the organic contaminants on
soils;
3.73
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Process Identification and Description
Figure 3.6
Bioventing Design to Optimize Biodegradation Using Air Injection Only
_From
Compressor
Ground Surface
Water Table
Figure 3.7
Bioventing Design to Optimize Biodegradation
Using Air Recovery and Injection
rTo Blower
(Low Flow
Water Table
3.74
-------
Chapter 3
• the transport of nutrients through soils; and
• the volatility of the organic contaminants.
The microbiology of bioventing is basically the same as that applied
during and understood from years of biological treatment of wastewaters
and land treatment applications. The practical basis for implementation of
the technology has resulted from development of in situ vapor recovery and
bioremediation soil-pile technologies.
As explained in Subsection 3.2.1, Microbial Ecology and Physiology, a
variety of common aerobic soil bacteria can use a variety of classes of or-
ganic compounds as food and energy sources. Other microorganisms may
produce enzymes capable of at least partially degrading some compounds,
while utilizing different classes of compounds as their primary food and
energy sources.
The stoichiometric oxygen requirements for the conversion of hydrocar-
bons to carbon dioxide is approximately 3 kg (6.6 Ib) of oxygen for 1 kg
(2.2 Ib) of biodegradable organic matter. In wastewater treatment systems,
actual oxygen demand is more typically 0.8 to 1.7 kg (1.8 to 3.7 Ib) of oxy-
Figure 3.8
Bioventing Design Optimize Biodegradation Using Air Recovery Only
. To Blower
(Low Flow)
Ground Surface
Water Table
3.75
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Process Identification and Description
gen per kg (Ib) of degradable carbon (Eckenfelder 1967). The actual oxy-
gen requirement depends on the degree of mineralization of the contaminant
and the efficiency of the oxygen supply system. If local anaerobic condi-
tions exist, anaerobic degradation of the intermediate products may serve to
lower the effective oxygen demand.
Nutrient requirements other than nitrogen and phosphorus are relatively
small and in nearly all cases can be met by the existing minerals in the soils.
Nitrogen and phosphorus requirements for cell production have been esti-
mated at 4.3 kg (9.5 Ib) of nitrogen and 0.6 kg (0.13 Ib) of phosphorus per
100 kg (220 Ib) of biological oxygen demand (BOD). Actual demand in
any system will depend on the extent of direct oxidation of the organic mat-
ter versus cell growth, rate of recycle of nutrients from dead cells, and exist-
ing sources of nutrients including bacteriological nitrogen fixation.
The dynamics of air flow through soils are well understood, but have not
been studied as thoroughly as those of groundwater flow (Wilson, Clarke,
and Clarke 1988). The permeability of soils to air is two to three orders of
magnitude greater than the soils' permeability to water (Wilson and Ward
1986). As a result, pressure differences will cause air to flow through the
interstitial spaces between the soil pores, provided the spaces are not filled
with water or other liquids. The ease of flow is dependent on the soil per-
meability and moisture content, and thus air will flow more readil y through
sands than through clays. Diffusion of oxygen into pore spaces is; important
and contributes substantially to the distribution of oxygen. The air flow
through soils has been thought to follow Darcy's law and has been modeled
and predicted on this basis. Recent work by Wilson et al. (1992) suggests
that Darcy's law does not hold in many cases (where the Reynold's number
exceeds four or, perhaps, one). Several models for in situ vapor recovery
have been developed and used for the physical removal of volatiles from
unsaturated soils (Van Eyk and Vreeken 1989a).
Because air is easier to move through soils than water and because air
has a higher oxygen-carrying capacity than water, it is possible to deliver a
relatively large amount of oxygen through soils even of moderate or low
permeability. For instance, a very modest air flow rate of 0.3 mVimin (10
ftVmin) will introduce 100 kg (220 Ib) of oxygen per day or 38,000 kg
(83,790 Ibs) per year, equivalent to enough oxygen to convert hydrocarbons
to carbon dioxide at a rate of 15.9 kg (35 Ib) per day or 5,670 kg (12,500 Ib)
per year.
3.76
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Chapter 3
Organic compounds with a vapor pressure greater than approximately 0.1
mm at normal atmospheric conditions are sufficiently volatile to be recov-
ered through in situ vapor recovery under favorable conditions. The rate at
which a particular compound can be volatilized and moved through a par-
ticular soil depends on whether the compound is present as a free phase,
adsorbed to the soil particles or humates associated with the soils, or dis-
solved. The tendency of organic compounds, which are neither adsorbed
within the soil particles or dissolved, to volatilize can be predicted from
their vapor pressures. Volatilization of compounds that are in the dissolved
phase can be predicted from the Henry's Law constant. Use of either pa-
rameter alone can be misleading; in general, the Henry's Law constant is
highly preferred. An Apparent Henry's Law constant, or lumped partition-
ing coefficient, which must be determined experimentally, may be preferred
for use in computer models of contaminant attenuation (Wilson, Clarke, and
Clarke 1988).
Once volatilized, compounds can readsorb as they move through the soil
matrices. Adsorption and readsorption on soil particles brings the volatile
organic compounds in contact with the soil bacteria, permitting biodegrada-
tion. Ostendorf and Kampbell (1990) have documented a hydrocarbon-
vapor transport model that takes into account soil and organic constituent
properties, biodegradation rates based on Michaelis-Menton kinetics, and
air flow through unsaturated soils. The model was tested in the laboratory
and at two test sites and resulting data were compared to data generated by
Marley and Hoag (1984) and by Thorton and Wootan (1982). The model's
data suggested that volatile constituents of aviation gasoline can be reduced
effectively in soil containing adapted microorganisms and sufficient nutri-
ents.
Nutrients may become available to the bacteria through minerals present
in the soils, addition of a nutrient-enriched leachate, or in the case of nitro-
gen, through biological nitrogen fixation (DuPont 1992a). Attempts to add
nitrogen as ammonia gas have not been successful (R.E. Hinchee, private
communication). Ammonia gas is adsorbed to soils within a few inches of
the injection point. This rapid adsorption has been taken advantage of by
farmers who introduce the gas a few inches below the surface of the soils.
Typically, loss of ammonia vapors to the atmosphere cannot be detected
readily.
3.77
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Process Identification and Description
The feasibility of nutrient addition to unsaturated soils is dependent on
both access to the surface and on the affinity of the soils for the individual
nutrient components. If the surface is covered, nutrient addition will be
more cumbersome. Nutrients are usually added as an aqueous solution that
is allowed to percolate through the soils. Addition of a nutrient solution
may be followed by addition of unamended water or by fortuitous rainfall.
The most available nutrient sources, such as ammonium and phosphate, can
be adsorbed by soil particles; thus, their movement may be retarded through
soils in an aqueous solution. Nitrate has little affinity for soils and, there-
fore, its movement in the unsaturated zone is retarded only by biological
consumption and dilution with residual moisture. The degree of adsorption
of ammonium and phosphate sources is greater in clayey or silty soils than
in sandy soils. Among the common phosphate sources, orthophosphate is
typically adsorbed less than are the polymeric phosphates such as tripoly-
phosphate (Norris and Subramanyam 1992) and trimetaphosphate
(Aggarwal, Means, and Hinchee 1992). While adsorption retards the move-
ment of nutrients through the soils, it also causes the nutrients to be retained
on the soil surfaces where they will be available for the bacteria.
Nitrogen fixation under anaerobic conditions has been observed.
Whether or not nitrogen fixation can provide significant quantities of nitro-
gen for contaminant biodegradation at a spill site is unknown. A recent
study by DuPont (1992b) concluded that nitrogen fixation requires the pres-
ence of a readily-metabolizable substrate, such as glucose, before the pro-
cess can provide a significant source of nitrogen in bioventing applications.
Others have speculated that nitrogen fixation may be significant at old spill
sites, and two recent field studies indicated that nutrient addition is not al-
ways necessary or beneficial. These studies do show, however, that mois-
ture management is important (DuPont 1992b).
In summary, what is known about metabolic processes and requirements,
movement of air through soils, and transport of inorganic nutrients, supports
the concept of bioventing as a remediation process for unsaturated soils
containing petroleum hydrocarbons and other biodegradable constituents.
The actual need for nutrients under field conditions is not yet well under-
stood, and some available field and laboratory data appear contradictory.
The issue is further complicated by the difficulty of distributing nutrients
under some conditions.
3.78
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Chapter 3
System Designs. All system designs incorporate some method of intro-
ducing oxygen to the unsaturated soils and may include a method of nutri-
ent addition. Most systems have used wells to recover or inject air; how-
ever, for shallow water table sites, trenches may be preferable, provided that
the excavated soils will not be listed as hazardous waste. Figure 3.6 (on
page 3.74) shows a typical design for a vapor extraction system using wells.
The wells are placed in the contaminated zone and screened over some
interval depending on the distribution of the contaminant, the soil type dis-
tribution, and whether the surface is covered by an impermeable layer. In
uncovered sites, the greatest radius of influence of a well, particularly
within the capillary zone, is effected with wells that are screened over a
narrow interval located just above the water table (or impermeable zone) or
near the bottom of the contaminated interval. For covered sites, the well
can be screened over a longer interval and extend closer to the surface.
When the contamination extends into the saturated zone, the ability to
supply oxygen much more rapidly and inexpensively through air movement
(as opposed to water circulation) should be considered. If the water table is
at least 3.1 m (10 ft) below the surface, the well screen may extend below
the water table to take advantage of the exposure of additional soils during
periods when the water table is lowered. The water table can be lowered by
pumping out groundwater in the vicinity of the treatment area. In highly-
permeable soils, very large volumes of water will need to be recovered and
treated. In formations with a high-clay content, the volume of water re-
quired to dewater the zone of interest will be much smaller than in more
permeable aquifers; however, the soils might not drain rapidly enough to
permit sufficient permeability for air movement into the exposed soils. The
applicability of dewatering needs to be evaluated by hydrogelogists experi-
enced in dewatering sites.
If the primary purpose of a system is to promote biodegradation rather
than physical removal of volatile constituents, the air extraction wells
should be located on the periphery of the contaminated area as shown in
figure 3.8 (on page 3.75). This will increase the flow path of the air and
incorporate a larger volume of soil into the treatment zone. Maximizing the
volume of soil in the treatment zone makes more bacteria and native nutri-
ents available for contaminant biodegradation.
Locating the wells on the periphery of the treatment zone and not cen-
trally located within the zone will increase and possibly double the number
3.79
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Process Identification and Description
of wells required to sufficiently aerate the treatment zone. However, plac-
ing wells in the contaminated zone will result in offgassing of contaminants
that will require secondary treatment. Installation of more wells is less
expensive than secondary treatment of volatilized contaminants.
Using low air extraction rates, volatile contaminants generally will have
more time to be biodegraded before being physically removed from the
subsurface. Furthermore, constituents volatilized in the more highly-con-
taminated areas may be readsorbed and biodegraded as they pass through
the less contaminated areas as shown in figure 3.8 (on page 3.75).
As with vapor extraction systems, it may be advisable to operate the
system intermittently. Because oxygen demand decreases as biodegrada-
tion proceeds, the rate of oxygen supply can be reduced continually as long
as at least 2% oxygen is maintained in the contaminated zone (DuPont
1992a). Once the contaminants have been degraded to an appropriate level,
the rate of oxygen supply can be reduced by cycling the recovery wells or
by operating the recovery wells on a rotating schedule. The latter approach
can result in the use of a smaller blower and reduced demand for offgas
treatment. Rotating wells and adjusting airflows also reduces the chance
that dead spots, regions of little or no flow, will occur.
Air injection wells can be used in conjunction with air recovery wells.
Air can be injected into either the saturated or unsaturated zone. Because
air injection may cause volatile constituents to migrate away from the injec-
tion point, these systems require more care in their design than those in
which air is simply extracted.
Injecting air into the unsaturated zone may provide the greatest benefit
by screening the injection wells in areas of low permeability, while screen-
ing the air extraction wells in the more permeable zones.
Soil moisture levels are an important factor. Biodegradation rates are
enhanced by high-moisture content. But, water-saturated soils inhibit the
distribution of air and, therefore, oxygen. One rule of thumb is to maintain
moisture levels between 40 and 60 percent of field saturation. Designing
and operating bioventing systems to maintain low-air flows minimizes loss
of moisture. In fine soils, it may be advantageous to remove some of the
moisture initially to allow adequate movement and distribution of oxygen.
Despite the well-documented requirements for nutrients in biological
wastewater treatment and in laboratory treatability studies, etc., the need for
3.80
-------
Chapter 3
nutrients in bioventing still is not well understood. The maximum amount
of nutrients that should be required is a ratio of 100:10:1 or 2 of C:N:P.
Because nutrients are recycled and not all of the carbon is used for cell
growth, the requirements are less than the above by a least one-half and
probably by greater than one-fourth. Furthermore, nutrient addition is not
always readily effected. Currently, it is effected through the addition of
aqueous solutions of nitrogen- and/or phosphorus-containing compounds,
which also add moisture. Thus, nutrient and moisture maintenance must be
coordinated.
At sites where the surface above the contaminated zone is not covered,
dilute nutrient solutions or granular nutrient blends can be added to the
surface and allowed to percolate through the soils. Additional water can be
supplied through a combination of spraying and fortuitous rainfall to trans-
port the nutrients deeper into the formation.
At sites where the surface is covered with asphalt or cement, nutrients
may be added just below the surface provided there is a sufficient gravel or
sand layer below the asphalt or cement. When a gravel or sand base is not
present or where contamination is present at depth, nutrients and moisture
may be added through a series of wells or trenches. If nutrients are added
beneath or close to a building, care must be taken not to adversely affect the
load-bearing capabilities of the soils.
The selection of nutrient sources is also important. Since little is known
about nutrient requirements during bioventing, it is difficult to predict with
certainty which nitrogen and phosphorus sources will be most effective in
stimulating microbial activity. It is clear, however, that nitrate percolation
into the formation will be less retarded by adsorption on soils than ammo-
nium percolation. Furthermore, orthophosphate will be retarded less than
tripolyphosphate, and, in most soils, orthophosphate will be retarded more
than ammonium. Clearly, it will require less water to transport nitrate to a
given depth than will be required to transport ammonium. On the other
hand, some sorption of the nutrients to the soil may be beneficial. In some
cases, nitrate might be flushed through the soils and unavailable to the bac-
teria. In addition, nitrate may be transported into groundwater and reach
levels that exceed drinking water standards (10 mg/L as nitrogen).
Flushing nutrients through the soils also creates some potential for trans-
porting contaminants from the unsaturated to the saturated zone. If the
saturated zone is already contaminated with the same constituents, whatever
3.81
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Process Identification and Description
remediation system is implemented in the saturated zone should accommo-
date the organic contaminants transported into that zone. Operation of the
air recovery system before nutrient addition should reduce the potential for
vertical transport of contaminant constituents by oxidizing and/or physically
removing the most mobile constituents.
Impact. Bioventing has its most direct impact on the unsaturated zone.
The process will remove and/or degrade a number of organic contaminants.
Nutrients that are added, but not utilized by the microorganisms, will cause
some increase in nutrient concentration. Some changes in inorganic species
may occur, such as an increase in the oxidation state of metals such as iron.
Depending on how the system is operated, soil moisture may increase or
decrease.
Bioventing may prevent or minimize contamination of groundwater in
removing the soluble constituents from the unsaturated zone and preventing
their migration into the saturated zone. Bioventing also may remove and/or
degrade constituents within the capillary zone and exposed soils when the
water table is low.
Air quality can be negatively impacted if large quantities of volatile con-
stituents are discharged to the atmosphere. Operating at low air flow rates
and/or use of offgas treatment protects air quality. In some cases, espe-
cially where there are significant concentrations of volatile compounds near
the surface, the bioventing process will reduce the mass of volatiles reach-
ing the atmosphere.
Pre- and Posttreatment Requirements. There are no special pre- or post-
treatment requirements other than site delineation, post-treatment sampling,
and getting permits.
Design Data, Unit Sizing. To design an effective bioventing system, it is
necessary to determine the distribution of the contamination, whether the
contaminants are biodegradable and/or sufficiently volatile, the soil proper-
ties and depth to water, and the site infrastructure.
Unfortunately, there is never as much site characterization data as de-
sired. As a result of the inherent complexity of subsurface conditions and
the cost of thoroughly defining them, it is usually necessary to design some
flexibility into the system. However, it is necessary to have sufficient data
to achieve the goals of the design effort.
3.82
-------
Chapter 3
The identification of all constituents (or mixtures) of interest is necessary
in order to know whether the process has a reasonable chance of succeed-
ing. The biodegradability and/or ventability of all constituents of interest
has to be established from previous experience, published data, or labora-
tory treatability studies.
The mass of contaminants present in the area to be addressed must be
estimated in order to anticipate the potential nutrient and offgas treatment
requirements, as well as to estimate remediation times.
The soil types and distribution and the depth to water must be deter-
mined in order to conceptualize how a system might be designed and to
determine whether there is sufficient potential to warrant pursuing this op-
tion. Detailed designs typically require field tests be conducted. The com-
plexity of the field test will depend on the site conditions, including the
volume of soil to be treated. The larger, more complex sites justify more
extensive pilot studies than do small sites. In small sites, such as service
stations, a large portion of the contaminated zone is located close to the tank
pit and/or underground utility trenches. Consequently, a test at any one
location would not be representative of any other location. It may be best in
such cases to design the air extraction wells in a conservative fashion and
be prepared to add a few additional air extraction wells as necessary.
Soil gas surveys can be conducted initially to provide a general sense of
the distribution of volatiles and biodegradable substances. Traditional soil
gas testing can be augmented to provide an approximation of soil perme-
ability and thus the ease of providing oxygen through aeration. Early use of
soil gas techniques may improve the efficiency and lower the cost of soil
boring to delineate the contamination.
Laboratory tests can be conducted to determine the feasibility of
bioventing, estimate the rate and extent of biodegradation, and provide data
for use in designing engineering tests. Plate counts, pH, and results of
respirometer tests, often are used to identify conditions inhibiting biodegra-
dation. Microcosm studies have been conducted to determine the rate and
extent of biodegradation, the effect of added nutrients on soil permeability,
and the nutrient requirements of the microorganisms. Unfortunately, these
tests frequently do not enable accurate prediction of field results. Tests to
determine the ease of nutrient percolation through the soils can provide
qualitative information on the ease of nutrient transport and can determine
the nutrient sources that will move most readily through the soils.
3.83
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Process Identification and Description
In situ respirometry tests can determine the existence of ongoing biodeg-
radation (Hinchee and Ong 1992). Such tests may include the use of tracer
gases or the equivalent of aquifer pump tests to determine the potential for
induced air flow through pilot-test areas. The former can be as simple as
analyzing soil gas samples in both contaminated and uncontaminated zones
for oxygen and carbon dioxide levels. Low-oxygen levels and high-carbon
dioxide levels suggest that biodegradation is occurring. Potential rate data
can be obtained by monitoring soil gases in conjunction with intermittent
air recovery. The data in table 3.16 indicate that hydrocarbon biodegrada-
tion rates can vary from 0.2 to 20 nig/day per kg of soil (2.0 x 10"7 to 2.0 x
Table 3.16
Comparison of Biodegradation Rates Obtained by the ir Situ
Respiration Test with Other Studies
Site
Various (8
locations)
HillAFB,
Utah
Tyndall
AFB, Florida
Netherlands
Netherlands
Undefined
Undefined
Undefined
New Zealand
Scale of
Application
In situ
respiration
tests
Full-scale.
2 years
Field pilot,
1 year and
in situ
respiration
tests
Undefined
Field pilot,
1 year
Full-scale
Full-scale
Full-scale
Pilot-scale/
Full-scale
Respiration
Contaminants
Various
JP-4 Jet Fuel
JP-4 Jet Fuel
Undefined
Diesel
Gasoline and
Diesel
Diesel
Fuel Oil
Diesel
Spent Oil
Estimated
Rates Biodegradation
(% 02/hour) Rates
0 02 - 0.99 0.4-
1 9 mg/kg/day
up to 0 52 up to
10 mg/kg/day
01-10 2-
20 mg/kg/day
0 1 - 0.26 2 -
5 mg/kg/day
0 42 8 mg/kg/day
50 kg/well/day
100
kg/well/day
60 kg/well/day
0 2 - 20
mg/kg/day
References
Hinchee and
Ong 1992
Hinchee et
al 1991
Miller 1990
Urhngs et al.
1990
van Eyk and
Vreeken 1989b
Ely and
Heffner
1988
Ely and
Heffner 1988
Ely and
Heffner 1988
Hogg et al.
1992
a Rates reported by Hinchee et al (1991) were first-order with respect to oxygen, for comparison purposes, these
have been converted to zero-order with respect to hydrocarbons at an assumed oxygen concentiation of 10%
b Rates reported as oxygen consumption rates, these have been converted to hydrocarbon degradation rates
assuming a 3 1 oxygen-to-hydrocarbon ratio
c Units are in kilograms of hydrocarbon degraded per 30 standard ft3per mm (scfm) extraction vent well per day
3.84
-------
Chapter 3
105 Ibs/day per Ib of soil). This type of rate data can be used to estimate
potential remediation times based on the initial contaminant levels. Table
3.17 lists calculations made in this manner along with calculations based on
stoichiometric oxygen requirements and air-flow rates.
Table 3.17
Potential Times for Contaminant Reduction1-2
Initial
Concentration
(ug/kg)
500
1,000
5,000
10,000
20,000
50,000
10 mg/kg/day
Reaction Rate
40
90
490
990
1,990
4,990
1 Pore Volume
Per Day
Utilization1
Days
20
45
245
495
995
2495
3 Pore Volumes
Per Day
Utilization3
8.7
17
82
165
332
832
1 Assumes regulatory approved final concentration of 100 ppm, no abiotic loses, and a 3 1 oxygen-to-carbon
utilization
2 From Hmchee and Ong 1992
3 Based on utilization of all oxygen in indicated volume of air
Oxygen concentration is a more reliable indicator than carbon dioxide
concentration because of the complex behavior of carbon dioxide with re-
spect to adsorption by calcium minerals and solution/dissolution from
groundwater and soil moisture. Carbon-13 measurements can distinguish
between carbon dioxide resulting from biodegradation of petroleum hydro-
carbons and that released by dissolution of minerals.
Pilot and feasibility tests are generally conducted using one or more air
extraction wells and several air monitoring points located at various direc-
tions and distances from the extraction well and at varying depths. Pressure
changes in the monitoring points are determined as a function of air flow
and vacuum applied to the extraction well(s). These data can be used to
calculate soil permeabilities and plotted manually to estimate the radii of
influence of extraction wells. Preferably, a computerized modeling pro-
3.85
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Process Identification and Description
gram is used to determine the air-flow paths and radii of influence of the
tested extraction wells based on pilot-test data. A treatability protocol de-
veloped by the U.S. Air Force and tested at over 100 sites is available
(Hinchee et al. 1992).
Information for Consideration. Information required to evaluate and
employ bioventing includes the properties of all contaminants of interest,
the site hydrogeology, the presence of any underground utilities, foundation
designs, regulatory requirements, including permissible soil concentrations
and air discharge limits for the constituents of interest, and permit require-
ments.
The primary contaminants of interest must be biodegradable under aero-
bic conditions. Volatile, nonbiodegradable constituents can be treated, but
will increase the cost of offgas treatment and may require larger blowers or
different well configurations and, possibly, longer treatment times. It is
important that all constituents be at levels below that which would be toxic
to the microorganisms. If compounds that are neither biodegradable nor
ventable are also present above the regulatory levels or other goals,
bioventing alone will not be effective.
The type of soil and its permeability affect the ease of air flow through
the soils. Bioventing is much more difficult in soils with low permeability,
such as clayey soils, particularly when moisture levels are sufficient to fill
nearly all of the pore space. (To some extent, operation of an air recovery
system will remove moisture and provide better air distribution.) Shallow
soils (less than 10 ft deep) can experience significant temperature variation
during the year. Thus, the rate of biodegradation in these soils due to
bioventing can also vary with time during the year. The average annual rate
can be maximized by employing soil-warming techniques such as; percola-
tion of warm water through the contaminated zone, or by burying heating
tape (Leeson et al. 1994).
The depth to the water table or other gas impermeable layer will affect
the radius of influence of an air extraction well and the decision to use verti-
cal wells or trenches. Unless the surface is covered with cement or asphalt,
a depth to water of at least 3.1 m (10 ft) is necessary to avoid using a large
number of wells or vapor recovery trenches. When the surface is covered, a
single recovery well may have a very large radius of influence, provided
that the cover has maintained its integrity.
3.86
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Chapter 3
The more stringent the cleanup goal, the larger the area that must be
treated and the longer the system must be operated. This is particularly true
for sites contaminated with high levels of heavier petroleum hydrocarbons.
Typically, lighter hydrocarbons can more readily be converted than heavier
hydrocarbons.
Technology Variations. Many variations of bioventing can be devised to
move air through the contaminated soil. These include air recovery wells
and trenches combined with air injection wells, trenches, and sparge points.
Injected air can be recycled from the air recovery blower with or without
offgas treatment. Examples of some generic approaches are shown in table
3.16 ( on page 3.84) and in figures 3.5 to 3.8 (on pages 3.73 to 3.75).
One should take advantage of site conditions rather than try to forcefit a
familiar system on a site. For instance, air recovery wells may not be ap-
propriate for sites with an open surface and a shallow water table. Alterna-
tives should be considered, including air sparging to provide oxygen or, if
the contamination is shallow enough, land treatment or excavation with
some form of treatment such as a soil pile. Selection of excavation should
take into account the possibility of uncontrolled releases of VOCs.
For soils contaminated with nonvolatile biodegradable constituents, air
injection through wells, trenches, or sparging may be remedial options and
could represent the lowest cost alternative. Air injection systems may also
be considered where there is sufficient soil above the contamination to act
as a biofilter.
Cost. Where applicable, bioventing has the potential to be a low-cost
remediation method. The cost of bioventing will depend not only on the
site location and conditions and cleanup goals, but also on the particular
design employed. Despite its similarities to in situ vapor stripping, the cost
of bioventing should be significantly less because the need for offgas treat-
ment is reduced and the total volume of air moved through the soils is
lower. For bioventing of heavier petroleum hydrocarbons, particularly if
substantially weathered, offgas treatment may not be required at all.
It is assumed here that a bioventing system design will maximize biodeg-
radation relative to physical removal. Thus, systems should provide air at
rates only slightly exceeding those required to allow the maximum rate of
respiration. The designs should also place wells outside the areas of high
contamination to increase the amount of soil available to support biodegra-
3.87
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Process Identification and Description
dation. These measures, while promoting in situ bioremediation, limit case
history information because most bioventing systems have been designed as
in situ vapor recovery systems with biodegradation as an unqualified side
benefit.
It has been estimated that venting of gasoline requires approximately 100
L of air per g (1,660 ft3/lb)of gasoline. Assuming a 20% efficiency, the
system would have to operate long enough to provide 500 L of air per g
(8,000 ftVlb) of gasoline. Stoichiometric calculations indicate that approxi-
mately 10 L (0.3532 ft3) of air is required to provide sufficient oxygen to
convert 1 g (0.002 Ib) of gasoline to carbon dioxide and water. Assuming a
20% efficiency of the introduced oxygen, 50 L of air per g (800 ftVlb) of
gasoline would be required under a bioventing mode. Many assumptions
underlie these calculations, but they serve to illustrate that the demands on
the air recovery system will be much less for bioventing than for vapor
recovery systems.
In a companion monograph', the costs for in situ vapor recovery technol-
ogy are estimated for a hypothetical site. For vapor recovery, total costs are
highest for systems designed for the greatest rate of remediation be:cause of
the high-capital costs. Costs decrease rapidly as the design de-emphasizes
speed and becomes less capital intensive, and then gradually increase as
long-term operation and maintenance costs begin to offset lower capital
costs. Bioventing systems resemble the less capital intensive vapor recov-
ery systems, except that offgas treatment capital and operating costs gener-
ally will be less.
When considering costs, it is also assumed that site delineation has been
completed and that groundwater dewatering is not required. The costs for
bioventing can be divided into engineering, capital, and operating/monitor-
ing. Engineering costs include field tests to determine the radius of influ-
ence and air-flow characteristics of the soils. This allows the determination
of well spacing and screening intervals. Field tests using existing wells and
soil gas probes can cost under $10,000 for small homogeneous sites. Costs
can reach $100,000 for larger, more complex sites where field respirometry
and biodegradation assays are conducted.
1. See Innovative Site Remediation Technology: Vacuum Vapor Extraction—Ed.
3.88
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Chapter 3
System installation will depend on the depth of soils being treated and
drilling conditions. Well installation costs will range from $2,000 to $5,000
per well for most sites. Blowers, controllers, electrical, equipment building,
security, etc. should range from $20,000 to $50,000. Costs will be higher
for tighter soils because the well spacing will have to be closer than for
sands and gravel.
A significant cost element is the offgas treatment system. Hinchee
(1993) estimated that bioventing costs might increase from $13/m3($10/
yd3)to as high as $52 to $78/m3 ($40 to $60/yd3) if offgas treatment is in-
cluded. Offgas treatment can range from a few thousand dollars for two
carbon canisters to $100,000 for a catalytic incinerator. If offgas treatment
is included, monitoring costs will also increase.
Operating costs can be minimal if the monitoring program is managed
well. High quality equipment and proper design will minimize maintenance
and travel to the site for routine monitoring. Monitoring can be done prima-
rily with portable equipment to measure flow rates, pressures, and concen-
trations of oxygen, carbon dioxide, and organic vapors. This approach will
limit laboratory costs to those for periodic confirmatory sampling for or-
ganic vapors. Soil sampling is another cost. Overall operation and mainte-
nance costs, depending on the size of the operation, may range from
$10,000 to $50,000 per year.
Case histories with the most reliable and thorough technical information
do not usually provide the best cost information. Cost information from
specific sites is not always in a form that allows direct cost comparison of
completed projects. One recent report (Struttman and Holderman 1992)
indicated costs of $45,000 to $53,000 for treatment of gasoline-contami-
nated soils over a 20 m (60 ft) radius to a depth of 6 to 10 m (20 to 30 ft) or
a volume of 1,200 m3 (1,600 yd3). The system was designed with four re-
covery wells placed outside the "hot spots". The apparent air recovery rate
of 4.8 mVmin (170 ftVmin) appears to be much higher than needed to pro-
vide oxygen for biodegradation. Offgas treatment was excluded under a
permit exemption. The data from the site were not sufficient to determine
how much gasoline removal was due to biodegradation or the extent of
remediation that was achieved. It is not clear whether extended operations,
which would lead to additional costs, were required.
As with many in situ remediation projects, economy of scale is an impor-
tant consideration. In the above example, costs were approximately $39/m3
3.89
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Process Identification and Description
($30/yd3). At another site, approximately 16,000 m3 (21,000 yd3) of soil
were contaminated with an average of 5,000 ppm of diesel fuel (Downey,
Hall, and Miller 1992). Soils consisting largely of silty sand, bul
interbedded with clayey soils, were contaminated to 21 m (70 ft). Initial
pilot tests were conducted leading to an estimated cost of treatment of
$200,000, or $13/m3($10/yd3).
Summary. At this time, the basic principles of bioventing and under-
standing of what the gross factors that affect bioventing are well under-
stood. As a result, good conceptual designs can be provided. Modeling of
in situ vapor recovery provides a good basis for locating and designing air
recovery wells. But, many critical questions remain, including the follow-
ing:
• When is nutrient addition required? Does nitrogen fixation play
a significant role?
• What is the relationship between respirometry results and actual
bioventing treatment rates?
• What are the limits of concentrations of various types of con-
taminants that can be treated?
• To what levels can various contaminants be treated?
• What are the effects of cold weather?
• What is the effect of various venting rates on biodegradation?
• What are the limits of low-permeability soils?
• Can PAHs (or other compounds) be too tightly bound to soils to
degrade?
• How easily can gas-phase constituents be reabsorbed and de-
graded?
• What are the optimum soil moisture concentrations in different
soil types?
In addition, many questions about how system design and operating con-
ditions affect the final outcome remain. For example, one can imagine a
scenario where managing a site to maximize biodegradation over physical
recovery can lead to nutrient deprivation that would not have occurred if the
system were managed to increase the amount of physical removal. Given
3.90
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Chapter 3
the documented differences between nutrient demands in laboratory tests
and field tests, many questions will need to be answered in the field.
As with most remediation technologies, knowledge about bioventing is
limited in part because much of the available information is too highly com-
partmentalized among a large number of organizations and individuals.
This might be addressed by a clearinghouse to collect information in an
appropriate form.
3.5.1.1.3 Air sparging2 In air sparging, air is introduced into the saturated
zone to transfer volatiles to the unsaturated zone for biodegradation. Wells
screened below the water table are used to inject air into the aquifer. As the
air flows radially upward and outward in a series of air channels, volatile
constituents are transferred to the unsaturated zone where they can be
adsorbed by the soil and biodegraded by the indigenous microorganisms.
Air sparging in the unsaturated zone entails the use of high-air injection
rates and/or closely-spaced injection wells to transfer volatile compounds to
the unsaturated zone relatively faster than they are biodegraded in the satu-
rated zone. By including the unsaturated zone in the treatment area, a larger
volume of soil and potentially a greater number of microorganisms will be
available for contaminant biodegradation than those contained in the satu-
rated zone alone. At sites where the unsaturated zone consists of fairly
permeable sands and is uncapped, it will be much easier to provide nutrients
to the unsaturated than to the saturated zone.
It is necessary to control carefully the air injection flow rate to prevent
transfer of volatile constituents to the atmosphere. In practice, it is likely
that regulatory agencies or prudence will dictate some form of vapor recov-
ery or other controls to prevent or limit losses to the atmosphere and, possi-
bly, groundwater control as well.
Air sparging has been demonstrated at the U.S. EPA and Coast Guard
demonstration project at Traverse City, Michigan, however, the data have
not been reported yet. Air sparging in the unsaturated zone has not been
documented or discussed in the cited reports on air sparging in the saturated
zone. During sparging in the saturated zone, however, some portion of the
2. See Innovative Site Remediation Technology: Vacuum Vapor Extraction—Ed.
3.91
-------
Process Identification and Description
volatile components probably are transferred to the unsaturated zone and
biodegraded.
3.5.1.2 Bioremediation Processes in the Saturated Zone
3.5.1.2.1 Liquid Delivery Introduction. R. L. Raymond and coworkers
developed and patented the first bioremedial technique for treating aquifers
contaminated with gasoline (Raymond 1974; Raymond, Jamison, and
Hudson 1975; Raymond et al. 1978). The process is analogous to some
conventional wastewater treatments in that a terminal electron acceptor,
usually oxygen, and inorganic nutrients are added to enhance conlaminant
degradation in a bioreactor (Thomas and Ward 1989). But, unlike wastewa-
ter treatment, which takes place in a confined and easily-managed
bioreactor, in situ treatment takes place in an unconfined and sometimes
unpredictable bioreactor, the subsurface. Indigenous subsurface microor-
ganisms are stimulated to degrade the contaminants.
Site Characterization. Before subsurface bioremediaton is initiated,
extensive site characterization must be conducted (see Section 3.3 Site
Characterization Relative to In Situ Bioremediation). The most important
characteristic of a site is the presence of contaminant-degrading microor-
ganisms. The relative number of these organisms will depend on the sedi-
ment characteristics, the type and concentration of contamination, and nutri-
ents (Thomas and Ward 1992). Contaminant-degrading microorganisms
have been detected in most samples of groundwater and sediments contami-
nated with petroleum compounds(see Subsection 3.2.1, Microbial Ecology
and Physiology). Although increases in the number of hydrocarbon-degrad-
ing microorganisms can be detected in a matter of days after a petroleum
spill, degraders of synthetic contaminants may require longer periods or
may not develop at all. Low numbers of contaminant-degrading organisms
may result from contaminant toxicity (Phelps et al. 1988) and low-nutrient
availability; low numbers often are associated with sediments having high-
clay content (nontransmissive) (Sinclair and Ghiorse 1989; Phelps et al.
1988).
To determine the feasibility of bioremediation, biodegradation
(treatability) assays can be conducted in the laboratory using samples of
contaminated sediment but not groundwater. The results of treatability tests
using groundwater from unpurged wells may be misleading because
groundwater collected from wells may contain non-native microorganisms
3.92
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Chapter 3
(Thomas, Lee, and Ward 1987). The potential for contaminant biodegrada-
tion and the nutrients required to maximize those rates are determined. Be-
cause of the highly carbonaceous nature of organic wastes, inorganic nutri-
ent additions may be required. Studies are conducted to determine the types
and concentrations of inorganic nutrients that maximize contaminant bio-
degradation rates; however, the native fertility of the subsurface should be
considered before nutrient amendments are formulated. The macronutrients
required are nitrogen and phosphorus, although additions of micronutrients,
such as magnesium and manganese, may be helpful.
Although treatability assays provide information about biodegradation
potential and nutrient amendments that enhance it, extrapolations from the
laboratory to the field may not be exact, especially for rates. In situ biodeg-
radation is limited by the rate at which oxygen is transferred to the contami-
nant-degrading microorgansims. In laboratory microcosms in which oxy-
gen is not limiting, biodegradation is limited by the intrinsic metabolic rate
of the microorganisms.
Computer models also have been used to assess the feasibility of
bioremediation at sites contaminated with readily-biodegradable com-
pounds, such as gasoline components (Norris 1993). In such instances,
modeling the treatability of a site may be more cost-effective than conduct-
ing laboratory studies.
In addition to determining the nutrient requirements for the contaminant-
degrading microorganisms, laboratory studies are conducted to determine
the compatibility of the added electron acceptor and nutrients with the
groundwater and subsurface materials. Nutrients that precipitate or com-
plex with subsurface materials will not be bioavailable and may clog the
aquifer (Thomas and Ward 1989). Well screens may be clogged by miner-
als that precipitate because the chemistry of the injection water and ground-
water is different (Norris 1993). Injection of oxygenated water into aquifers
containing low levels of dissolved oxygen may oxidize and precipitate iron
and other metals. Orthophosphates, which are often added as sources of
nutrients, will form calcium, magnesium, and iron precipitates which may
clog the aquifer. Use of sodium phosphates can cause clays to swell and
decrease the hydraulic conductivity of the formation.
Precautions can be taken to prevent nutrient incompatibility (Norris
1993). To avoid precipitation problems, water to be injected or re-injected
can be filtered to remove metals and formation-compatable nutrients can be
3.93
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Process Identification and Description
used. By using potassium, rather than sodium phosphates, in formations
having high-clay content, problems associated with swelling clays can be
avoided. Precipitation of calcium, magnesium, and iron with phosphates
can be avoided through use of tripolyphosphate at equimolar or greater
ratios, which will complex and solubilize these metals.
A thorough investigation of the hydrogeology and contaminant distribu-
tion is also required. Briefly, parameters such as hydraulic conductivity,
depth to the water table, rate and direction of groundwater flow, specific
yield of the aquifer, and types and concentrations of contaminant!? are deter-
mined. Analysis of sediments provides a better indication of the contami-
nant regime than does analysis of groundwater, since NAPLs and sorbed
contaminants will not be detected in aqueous samples. Because the subsur-
face is heterogeneous, many parameters may be determined at multiple
locations.
Contaminants sorbed to sediments and dissolved in groundwater are
treated. For saturated sediments, physical removal of the contaminant
source is more cost-effective and faster than biological treatment (Hurlburt
1987). Therefore, a free product phase should be removed before: in situ
bioremediation is implemented in saturated materials. Free product recov-
ery, however, usually can remove less than 50% of the contamination in the
subsurface.
System Design. Using information collected from the site characteriza-
tion such as hydraulic conductivity, aquifer thickness, dispersivity, depth to
the water table, and concentrations of contamination and oxygen, the well
system can be designed with (Rifai et al. 1988) or without the use of com-
puter models. Models also are used to predict the time required for
remediation, based on the oxygen requirements for contaminant biodegrada-
tion and the rate of oxygen transport through the subsurface (Norris 1993).
The delivery system is designed to circulate adequate amounts of nutri-
ents and oxygen through the zone of contamination to maximize contami-
nant biodegradation. Injection wells or trenches, through which nutrients
and oxygen are added, are placed within or close to the contaminated area
(figure 3.9 on page 3.95). Groundwater extraction wells or trenches may or
may not be included, depending on the presence or absence of down-gradi-
ent receptors of concern. Produced groundwater is extracted, treated
aboveground if necessary, and then disposed of in an environmentally-
sound manner or amended with nutrients and recirculated. If the water is
3.94
-------
Chapter 3
Flow"
Figure 3.9
\Atell System for Liquid Delivery
Mixing Tank:
Nutrients
Surface •HBhnBB
Injection
Well
Water Table V
Groundwater j
1
,
Air Pump
-/nfl
F~
V
t
Produced water can be
p recirculated or treated at the
umP surface using other remedial
zfTffTI1 °Py'°"S' 7-^
Producing
Well
V
- Sparger Device
Nutrient and Oxygen-Rich
Groundwater
recirculated, removal of biomass and carbon may be required to avoid clog-
ging the aquifer at the re-injection point. Many states require permits to
inject any water into an aquifer, however, permits for injection of treated or
clean water generally have not been difficult to obtain. Some agencies have
limited the amount of nitrogen sources being injected into the aquifer to
lOmg/L as nitrogen.
A recirculating system is designed to hydraulically isolate the target area
and minimize contaminant migration out of the treatment zone. Some sys-
tems recirculate 100% of the produced water, whereas others use municipal
water or uncontaminated groundwater from the site; in the latter case, the
produced water must be disposed in an environmentally-sound manner.
Other systems recirculate only a portion of the produced water to control
groundwater flow through the isolated area when necessary. Where the
unsaturated zone requires treatment or the aquifer is shallow, nutrients and
oxygen may be added through infiltration galleries instead of injection wells
(figure 3.10 on page 3.96). This variation allows the oxygen and nutrients
to reach microorganisms in the unsaturated as well as in the saturated zone.
3.95
-------
Process Identification and Description
Figure 3.10
Infiltration Gallery for Treatment of Contamination in the
Unsaturated Zone or Shallow, Contaminated Aquifers
Infiltration Gallery
Contaminant Source
Unsaturated Zone
Groundwater
Flow
Monitoring wells must be installed in the treatment area and used to
assess the progress of remediation. Samples of groundwater are collected
before, during, and after bioremediation and analyzed for concentrations of
oxygen, contaminants, nutrients, tracers, pH, and microorganisms. The
monitoring program must be tailored to address several basic issues: regula-
tory requirements, monetary constraints, and hydrogeological aspects of the
site. The first and second issues are self-explanatory. For the third issue, if
groundwater flow rate is appreciable, it may be useful to have frequent
monitoring early on to evaluate the possibility that the in situ system may
produce quick results. On the other hand, when groundwater flow is slow,
monitoring efforts can be spread out over time because it can be anticipated
that not much will happen for a while.
Oxygen is added by sparging with air or pure oxygen at the bottom of an
injection well, or by using hydrogen peroxide. This system can be modified
so that air is added through spargers located directly in the aquifer (see
Subsection 3.5.1.2.3, Air Sparging (Bioremedial Processes in the Saturated
Zone)). The problem with using oxygen in an aqueous solution is its lim-
3.96
-------
Chapter 3
ited solubility in water. Depending on temperature, sparging with air or
pure oxygen can yield about 8 and 40 mg dissolved O2/L, respectively.
An alternate source of oxygen is hydrogen peroxide (H2O2), which is
infinitely soluble in water and dissociates to form H2O and 1/2 O2. Al-
though H2O2 can supply large quantities of O2, the oxidant can be toxic to
microorganisms at concentrations as low as 100 mg/L (Texas Research
Institute 1982). To minimize toxicity, H2O2 is added initially at concentra-
tions as low as 50 mg/L and increased step-wise to concentrations as high
as 1,000 mg/L.
Although H2O2 potentially can deliver more oxygen than sparging with
air or pure oxygen, the efficiency of delivering oxygen using the oxidant
has varied (Norris 1993). If the rate of H2O2 decomposition is much greater
than the rate of oxygen consumption, oxygen offgas may plug the aquifer
and deplete the oxygen available to the microorganisms. In rare cases, the
decomposition rate may be too slow and limit biodegradation. The rate of
H2O2 decomposition is enhanced by the presence of organic matter, catalase
(biological catalyst), and metals such as iron, copper, manganese, and chro-
mium. In one field study, H2O2 was decomposed excessively by catalase-
producing bacteria located in infiltration galleries (Spain et al. 1989). De-
composition rates can be enhanced by using catalysts such as chelated met-
als, or retarded by pretreating the aquifer with agents, such as inorganic
phosphates that are also used as nutrients, that will inactivate the naturally-
occurring catalysts (Raymond et al. 1986).
Although H2O2 can deliver more oxygen than sparging with air or pure
oxygen, the ability to supply an electron acceptor for biodegradation using
peroxide is still limited by the maximum theoretical solubility of oxygen in
water (40 mg/L). An alternate electron acceptor that is more soluble in
water than oxygen is nitrate (see Subsections 3.2.1, Microbial Ecology and
Physiology; 3.2.2, Biogeochemistry and Biodegradation; and 3.5.1.2.2,
Alternate Electron Acceptors); depending on the source, the solubility of
nitrate is in the g to kg/L range. But, federal standards on the limits for
nitrate in drinking water (10 mg/L as nitrogen and 45 mg/L as nitrate) will
constrain the amount of nitrate that can be added to the subsurface.
Nitrate has been shown to serve as the electron acceptor for some aro-
matic hydrocarbons (Kuhn et al. 1988; Mihelcic and Luthy 1988). Nitrate
has been used in field-scale trials in biorestoration of aquifers contaminated
3.97
-------
Process Identification and Description
with petroleum hydrocarbons (Batterman 1986; Hutchins, Downs et al.
1991). (See also Subsection 3.5.1.2.2, Alternate Electron Acceptors).
Performance Monitoring. Monitoring is critical because of the dynamic
nature of a bioremedial operation. Parameters that often are monitored
include contaminant concentrations in groundwater and sediments, dis-
solved oxygen, inorganic nutrients, pH, and H2O2. Monitoring is important
especially in the early phases of bioremediation to identify problems and
the need to modify the system design and operation.
Interpretation of the monitoring results can be misleading if trie param-
eters are analyzed singly rather than integratively. Analysis of groundwater
for contaminant and nutrient concentrations is informative of progress,
although it is not always indicative of removal of contaminants s orbed and
entrained in the sediment. To accurately assess contaminant removal, sedi-
ment samples must be collected and analyzed periodically during treatment.
During treatment, oxygen and/or H2O2 usually are not detected in samples
of groundwater; the appearance of oxygen and/or H2O2 or an increase in
their concentration (breakthrough) suggests that contaminants have been
biodegraded in that area (Piotrowski et al. 1984). However, at sites with
high contaminant concentrations and oxygen demand, rapid oxygen break-
through may indicate flow through preferential flowpaths and non-uniform
treatment.
Monitoring weekly changes in microbial numbers in groundwater as a
measure of bioremedial progress can be misleading because of the dynamic
nature of microbial communities in situ. Environmental parameters such as
rainfall, predators, and microbial competition will affect population size.
An initial decrease in microbial numbers after injection of water into the
subsurface may be a result of dilution. But, a precipitous decline in micro-
bial numbers and/or an increase in contaminant concentration may indicate
inhibition of contaminant-degrading microorganisms.
Applications. The majority of sites that have been bioremediated using
the liquid delivery technique have been contaminated with light nonaqueous
phase liquids such as commercial blends of petroleum hydrocarbons. The
lighter material (e.g., gasoline), containing the low-molecular weight, more
soluble compounds, is more easily bioremediated than is the heavier mate-
rial (e.g., coal tar), containing the high-molecular weight, less soluble com-
pounds. Because high-molecular weight hydrocarbons are sparingly
soluble and sorb to sediments, they are less bioavailable and more difficult
3.98
-------
Chapter 3
to treat (Brubaker and Stroo 1992). Liquid delivery in the saturated zone
has been demonstrated in the treatment of aviation fuel at Traverse City,
Michigan (Wilson, Armstrong, and Rifai 1993). For dense nonaqueous
phase liquids, contaminant distribution limits the applicability of liquid
delivery. Unlike light nonaqueous phase liquids, the contamination is un-
evenly distributed rather than confined to a relatively small portion of the
saturated zone that can be more easily treated.
In addition, the liquid delivery process functions best in formations with
hydraulic conductivity of at least 10"4 cm/sec. For those formations with
lower K values, transport of an electron acceptor and essential nutrients
through the contaminated zone will require longer periods and be more
costly. In addition, there is a greater chance of plugging the formation in
the low-transmissive materials in which clays and fines often are found .
A variation of the process has been used to treat the chlorinated aliphatic
solvents, such as trichloroethylene and less chlorinated ethylenes (Semprini
et al. 1990). The process involves stimulating the growth of indigenous
methanotrophs (methane oxidizers) with methane and oxygen. The
methanotrophs metabolize methane using the enzyme, methane mono-
oxygenase, which is nonspecific and also oxidizes the chlorinated ethenes.
Advantages/Disadvantages. All bioremediation processes have the po-
tential to destroy the contamination rather than transfer it to another part of
the environment; however, the most obvious and important advantage of the
liquid delivery system is that treatment occurs in situ and obviates excava-
tion and transportation of contaminated sediments. In contrast to pump-
and-treat methods, which merely remove dissolved contaminants, in situ
bioremediation processes can also treat contamination sorbed and entrained
in the sediment. In addition, liquid delivery is faster than pump-and-treat.
Disadvantages are specific to the site. Bioremediation may be inhibited
by the presence of toxic concentrations of contaminants. As previously
discussed, formations with K values less than 10"4 cm/sec will require more
time for treatment and are more easily plugged than those with greater K
values. In addition, it may be difficult to obtain discharge permits to dis-
pose of the produced groundwater that is not recirculated (Norris 1993).
Costs. Estimates of the cost for implementing liquid delivery will be
specific to the site (Norris 1993). The type, amount, and extent of contami-
nation will be important factors. In addition, the sediment characteristics
3.99
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Process Identification and Description
will affect the system designed which must assure an adequate supply of
oxygen and nutrients to the contaminant-degrading microorganisms. The
source of oxygen will affect cost; although H2O2 is more expensive than
oxygen, overall cost may be reduced when using the oxidant if the time
required to treat the contaminants is reduced.
3.5.1.2.2 Alternate Electron Acceptors Introduction. Because of the
limited solubility of oxygen in water, it is difficult to deliver large quantities
of dissolved oxygen to contaminated subsurface environments. A variety of
oxy-anions can subsitute for oxygen and allow microbial degradation of
organic contaminants. Practical alternate acceptors include nitrate, sulfate,
and salts of iron III.
Nitrate is highly soluble in water, has a high electron-accepting capacity
for its mass, does not sorb appreciably to subsurface materials, and is not
toxic to microorganisms. By using nitrate, large quantities of electron-
accepting capacity are easily contained in groundwater circulated through a
spill. But, nitrate is expensive and toxic to humans. Although nitrate is
very soluble in water, the usual end product of nitrate reduction, N2, is
poorly water soluble. If N2 accumulates, bubbles may form that exclude
water from the pore spaces and decrease the hydraulic conductivity of the
subsurface material.
Sulfate is also highly soluble in water, has a high electron-accepting
capacity for its mass, and does not sorb appreciably. It is inexpensive and is
not toxic to microorganisms. But, sulfide, the end product of sulfate reduc-
tion, is toxic to both humans and microorganisms.
Iron III salts are slightly soluble in water and have a low electron-accept-
ing capacity expressed on a weight basis. It may be possible to distribute
iron through contaminated material as a colloidal suspension in groundwa-
ter, but, this has not been demonstrated. More practical applications may
involve mechanically blending iron minerals or iron salts with contaminated
material. Iron II, the end product of iron reduction, is not particularly toxic
at concentrations that would be expected during bioremediation.
Nitrate. In the early 1980s, Batterman (1986) used a mixture of nitrate
and oxygen to remediate a large spill of heating oil in the upper Rhine Val-
ley. Benzene and toluene were completely removed, but the xylenes were
more persistent. This remains the most successful large-scale application of
an alternate electron acceptor for remediation of a fuel spill. Hutchins,
3.100
-------
Chapter 3
Downs et al. (1991) conducted a detailed study of the use of nitrate to
remediate a JP-4 jet fuel spill in Michigan. Benzene, toluene, ethylbenzene,
and all three xylenes were removed from groundwater circulated through
the spill. Core analyses also revealed that these compounds were removed
from the residual oily phase hydrocarbons left behind after active
remediation ceased. Microcosm studies conducted with core material from
the site confirmed removal of toluene, ethylbenzene, m- and/7-xylene under
strict denitrifying conditions (Hutchins, Sewell et al. 1991). o-Xylene,
however, only degraded when one of the other alkylbenzenes was present.
Benzene was not degraded at all, whether present as the sole carbon source,
or in combination with the alkylbenzenes.
Removal of benzene in the field may have resulted from low concentra-
tions of oxygen (0.5 to 1.5 mg/L) in the groundwater circulated through the
spill (Hutchins, Downs et al. 1991). But, the precise role of oxygen has not
been defined. The alkylbenzenes can be degraded under strictly anoxic
conditions through pathways involving oxidation of the alkyl substituent.
These pathways are not available to benzene. Oxygen may be required for
oxygenases that initiate the metabolism of benzene.
Lemon, Barbara, and Barker (1988) studied the efficacy of nitrate for
bioreclamation of an artificial plume of alkylbenzenes in Ontario, Canada.
Under strictly denitrifying conditions, only toluene was removed. The aqui-
fer had previously been aerobic and there was little time for anaerobic accli-
mation to occur at the time the experiment was conducted.
Hutchins and Wilson (1991) noted that removal of alkylbenzenes in the
JP-4 jet fuel spill was a zero-order process. At 10°C (50°F), toluene was
removed at a rate of 0.2 mg/L/day, ethylbenzene at 0.13 mg/L/day, m- or p-
xylene at 0.14 mg/L day, o-xylene at 0.13 mg/L/day, and 1,2,4-
trimethylbenzene at 0.073 mg/L/day.
Nitrate also has been used to reclaim a gasoline spill in California
(Sheehan et al. 1988). From 95% to 98% of the purgeable alkylbenzenes in
groundwater were removed.
Sulfate and Iron III. As of this writing, neither sulfate nor iron III has
been used as an electron acceptor in a field-scale remediation, although both
sulfate and iron III have important roles in natural bioattenuation. Recent
work in Australia demonstrated that ambient concentrations of sulfate (20 to
100 mg/L) were responsible for removal of toluene, ethylbenzene, the xy-
3.101
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Process Identification and Description
lenes, and 1,3,5 trimethylbenzene from a plume of contamination from a
leaking underground storage tank (Thierrin et al. 1992a). Benzene was not
removed. Lovley and Phillips (1988) have reviewed the importance of iron
reduction in the natural carbon cycle. An organism has been isolated that
degrades toluene, phenol, and p-cresol with amorphous iron III as; sole elec-
tron acceptor (Lovley and Lonergan 1990).
Sulfate is much less expensive than nitrate. Concerns about sulfide tox-
icity are probably responsible for the lack of interest in sulfate as an alter-
nate electron acceptor. Recent laboratory work suggests a practical solution
to the sulfide toxicity problem. Beller, Grbic-Galic, and Reinhard (1992)
studied microcosms and enrichment cultures prepared from soil that had
been contaminated with fuel hydocarbons for a long period. Sulfate reduc-
tion supported removal of toluene in the microcosms and enrichment cul-
tures. Addition of iron III significantly stimulated the removal of toluene
by sulfate reduction. The degradation of toluene, reduction of iron III, and
removal of sulfate were consistent with formation of elemental sulfur, iron
II sulfide, and iron III carbonate, but there was no free sulfide. If iron min-
erals are a significant component of the aquifer matrix, it may be possible to
use sulfate effectively as an alternate electron acceptor.
The Question of Benzene. Complete mineralization of benzene under
anaerobic conditions was recently documented in samples of sediment from
Seal Beach, California (Edwards and Grbic-Galic 1992). The electron ac-
ceptor was not confirmed, but was probably sulfate. The experimental sys-
tem contained 1,900 mg/L sulfate, no oxygen or nitrate. Benzene was not
degraded in material that also contained other alkylbenzenes. Benzene
degradation in the absence of oxygen is possible, but it is also rare and un-
predictable at the present time. In the few cases where it can be inferred
from field data, and the one case where it is confirmed with a laboratory
study, it was associated with sites that have been exposed to petroleum
hydrocarbons for years or decades. For the present, engineering designs to
bioremediate benzene in situ should consider adding oxygen in addition to
the alternate electron acceptor.
3.5.1,2.3 Air Sparging Introduction. Air sparging provides oxygen as an
electron acceptor for biodegradation and physically removes volatile sub-
stances from the unsaturated zone. Air is forced into the aquifer through
well points or wells screened beneath the water table. Air moves radially
outward and upward from the point of injection, resulting in increased lev-
3.102
-------
Chapter 3
els of dissolved oxygen and transfer of volatiles to the unsaturated zone.
Dissolved oxygen is distributed through the aquifer through diffusion and
the movement of air, and groundwater movement. Volatiles reaching the
unsaturated zone are typically captured using in situ vapor recovery with
offgas treatment. A recent field study has demonstrated the use of the un-
saturated zone as a biofilter, thus avoiding the need for expensive offgas
treatment as discussed elsewhere in this monograph.
The way in which air moves through saturated soils is not well under-
stood. Some authors, in early publications, described air flow during
sparging as a series of bubbles. Currently, it is believed that bubbles may
only be present in significant quantities in coarse gravels. It is more likely
that in most soils, air will flow through discreet channels. The spacing
between discrete continuous flow channels is not well understood, but is
critical to the effectiveness of air sparging, especially for physical removal
of volatile constituents.
Therefore, air sparging can be thought of as functioning through one or
more of the three following mechanisms:
• biodegradation in the aquifer;
• biodegradation in the unsaturated zone; and
• physical transfer to the unsaturated zone for capture by an in situ
vapor recovery system.
The relative importance of the mechanisms will depend on the properties
of the contaminants, the site conditions, and the system design. This sec-
tion discusses systems which emphasize biooxidation in the aquifer. Sys-
tems designed to emphasize biooxidation in the unsaturated zone are dis-
cussed elsewhere in this monograph. A comprehensive discussion of air
sparging as a means of promoting physical transfer can be found in the
companion monograph on in situ vapor recovery.3
The use of air sparging for biodegradation in the saturated zone has
evolved through efforts to improve upon the liquid delivery system. The
liquid delivery system supplies oxygen by sparging air or pure oxygen at
the bottom of injection wells. Because the low solubility of air in water
limited the rate of oxygen delivery, hydrogen peroxide, which is infinitely
3. See Innovative Site Remediation Technology: Vacuum Vapor Extraction—Ed.
3.103
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Process Identification and Description
soluble in water, was used as an alternate source of oxygen. Practical con-
siderations have limited hydrogen peroxide concentrations to the 100 to
1,000 mg/L range. Using hydrogen peroxide, greater rates of ox ygen deliv-
ery were achieved than were possible by sparging into the injection water
and have led to successful bioremediation of aquifers. However, there are
problems with the use of hydrogen peroxide other than its relatively high
cost (see section on Liquid Delivery in the Saturated Zone) (Norris 1993;
Lowes 1991). Sparging air directly into the aquifer offers a potentially
technically preferable and more cost-effective method of oxygen introduc-
tion than the liquid delivery system.
A much larger volume of oxygen can be introduced into the aquifer by
air sparging than by other technologies. Very close to the air channels, it
may be possible to maintain dissolved oxygen near saturation. However, it
is not yet understood how much dissolved oxygen is maintained at points
midway between air channels. Further, there is currently insufficient infor-
mation to predict the fraction of the oxygen introduced below the water
table that is transferred into the aqueous or dissolved phase.
Bioremediation systems using air sparging would use many more and more
closely-spaced air sparging wells compared to the number and spiacing of
injection wells used for a hydrogen peroxide system. The belief of many
practitioners is that this will result in a greater oxygen availability across
the contaminated zone.
Furthermore, direct air sparging introduces oxygen across the entire con-
taminated zone during the treatment period, rather than relying on oxygen
transport through the contaminated zone. Thus, air sparging will more
likely result in remediation of all portions of the site at similar rates, as
opposed to a degradation front moving across the site during liquid deliv-
ery.
System Design. Sparging wells are placed at short intervals located near
the bottom of the aquifer (figure 3.11 on page 3.105). Air is supplied at a
sufficient pressure to overcome the head pressure of groundwater and sur-
face tension. The outward and upward movement of air through the satu-
rated soil matrix has the following effects:
• as the air bubbles move through the groundwater, the elevated
pressure of the air bubbles induces air and, therefore, oxygen to
be dissolved into the aqueous phase. Diffusion accelerates the
3.104
-------
Chapter 3
Figure 3.11
Groundwater Sparging Without Optional Vapor Recovery
Compressed Air
movement of oxygen and mitigates the effects of channeling of
the air flow;
while air moving through the soil pore spaces creates turbulence,
improving soil and water contact and, thereby, increasing the
rate of dissolution of the contaminants, the greatest amount of
mixing is thought to occur during the transient stage following
startup and shut-down;
the movement of air through the groundwater acts as a crude air
stripper, transferring dissolved compounds with sufficiently
large Henry's Law constants to the unsaturated zone;
adsorbed-phase volatile compounds can be transferred to the air
stream and carried to the unsaturated zone;
the temporary mound observed in air sparging is caused by pres-
sure and the displacement of water from pores that become air
filled; and
3.105
-------
Process Identification and Description
• physical displacement of free-phase substances, especially dense
nonaqueous phase liquids (DNAPLs), may occur at very high air
flow rates.
The degree to which biodegradation will occur (versus physical transfer
to the unsaturated zone) will depend on the characteristics of the
contaminant(s), the site lithology, and the design of the system. For heavier
hydrocarbon blends, volatilization will be minimal regardless of the system
design and site conditions. For the lighter petroleum hydrocarbons, both
biodegradation and physical removal processes should be considered in the
system design and operating conditions. In this section, systems whose
objective is to maximize contaminant reduction via biodegradation in the
saturated zone will be discussed.
The design and implementation of air sparging systems for
bioremediation requires a basic understanding of biodegradation, volatiliza-
tion of organic compounds, and air movement; however, the effects of air
movement through saturated soils is currently not well understood. Al-
though biodegradation is the goal of the process, it also is necessary to con-
trol volatile organic compounds that reach the unsaturated zone through in
situ vapor recovery systems (figure 3.12 on page 3.107). Because of the
heterogeneous soil conditions in most aquifers, it is typically necessary to
conduct field tests and use other site information to engineer a safe and
reliable remediation system.
Bacteria require nutrients, especially nitrogen and phosphorus, to grow
and thereby degrade organic contaminants. Accordingly, it is generally
accepted that nutrient addition is necessary for most bioremediation pro-
cesses. But, the results of recent bioventing field tests have raised questions
concerning the necessity and benefits of adding nutrients for bioremediation
of the unsaturated zone (Downey and Guest 1991; DuPont 1992a; Miller
1990; Hinchee et al. 1991). Nutrient addition may not be necessary in un-
saturated zone treatment. It has been speculated that nutrient addition is
more important for bioremediation of the heavier, more recalcitrant com-
pounds on which bacteria grow slowly, than for remediation of the; more
readily-degradable substances on which bacteria grow faster (McKenna and
Heath 1976). At this time, there is not enough information to determine
whether nutrient addition is necessary. Until there is better understanding,
conceptual designs of these systems should provide for nutrient addition.
The decision to employ nutrient addition can be decided during pilot testing
or implementation.
3.106
-------
Chapter 3
Because addition of air to the saturated zone displaces water, it is com-
mon to use a groundwater capture system to prevent migration of dissolved
contaminants. A portion of this captured water can be used to inject dis-
solved nutrients into the aquifer. Another benefit of groundwater recovery
and re-injection is the movement of water between air-flow channels, thus
increasing the movement of dissolved oxygen containing water. In shallow
aquifers, nutrients can be percolated from the surface. A large body of
experience exists on distribution of nutrients into aquifers for
bioremediation exists. There have been few concerted efforts, however, to
select nutrient blends for specific conditions. See also Subsection 3.5.1.1.2,
Bioventing, concerning nutrient addition.
While air sparging has been used for the physical removal of volatiles
from the unsaturated zone, as discussed by Brown and Jasinlewicz (1992)
and Brown and Fraxedas (1992), the movement of air through saturated
soils is far from fully understood. Several field tests have been performed
that established radii of influence under specific site conditions. One labo-
Figure3.12
Groundwater Sparging With Optional In Situ Vapor
Stripping for Management of Vapors
Compressed Air
Ambient Air
Vacuum Extraction Well Bat I Contaminated Soil|
3.107
-------
Process Identification and Description
ratory study that was conducted to determine how the radius is affected by
flow rates (Wilson, Clarke, and Clarke 1992) indicated that the radius of
influence was approximately equal to the saturated thickness; channeling
within the test apparatus soils was also observed. Other laboratory studies
also demonstrated the formation of channels and indicated that the radius of
influence increased with increased heterogeneity of soil particles. One
recent paper developed a mathematical model for correlating radii of influ-
ence with the distribution of elevated pressure in the vadose zone (Wilson,
Clarke, and Clarke 1988). Rough rules for estimating radii of influence of
sparging wells predict that the air bubbles will reach the groundwater sur-
face at a horizontal distance from the injection point equal to the depth be-
low the water table at which the air is introduced. This assumes that there
are no low-permeability lenses that impede upward movement of the air and
does not consider diffusion of oxygen, which is important for biodegrada-
tion but not for physical removal of volatile compounds.
Despite the relatively undeveloped status of the use of air sparging for
bioremediation, the information and experiences derived from air sparging
for physical removal of volatile compounds and experiences with more
traditional aquifer bioremediation can be used to intelligently implement
sparging to treat aquifers. As with all in situ technologies, it is necessary to
have a good understanding of the environmental conditions and the proper-
ties of the constituent contaminants. It is advisable, however, to carefully
monitor performance to verify that the intended results are occurring or to
identify corrective measures that can be taken.
The mechanism for treating compounds in the saturated zone under air
sparging conditions will depend on the relative ease of biodegrading the
compounds compared to the ease of stripping them from the aquifer. The
latter will depend on the compounds' water solubility, vapor pressure, and
physical state within the aquifer (dissolved, free phase, or adsorbed). Dis-
solved compounds will be most easily stripped if they have a low-water
solubility and a high-vapor pressure. The tendency to be stripped from the
dissolved phase is most directly related to a compound's Henry's Law con-
stant (table 3.18 on page 3.109). For example, acetone, which has a higher
vapor pressure than benzene or TCE, has a significantly lower Henry's Law
constant because it is more soluble in water and thus less easily lost to the
vapor phase. Compounds with a Henry's Law constant greater than IxlO"3
mVmole will have a significant tendency to transfer to the vapor phase.
Physical properties that will affect the disposition of many compounds of
3.108
-------
Chapter 3
Table 3.18
Physical Properties Important to Bioremediation Processes
Compound
Benzene
TCE
Acetone
Butanol
Phenol
Naphthalene
Pyrene
Diethylhexylphthalate
Phenanthrene
Solubility
(mg/L)
1,791
1,100
Misc
77,000
87,000
31
01
03
09
LogKow
2.13
2.40
-024
0.88
1 46
3.28
520
5.11
4.46
Vapor
Pressure
(mm Hg)
95
69
231
7
0.5
02
6.9 x 10-7
5 0 x lO'6
6.8 x 10-"
Henry's Law
Constant
(mVmole)
5 4 x lO'3
lOx ID'2
3.7 x 10'5
56x 10-6
4 x 10-7
4 6 x 10-4
1.1 x 10-5
1.1 xlO'5
3.9 x 10-5
From Howard 1989,1990, Montgomery and Wilkins 1990
environmental concern can be found in several handbooks (Howard 1989,
1990; Montgomery and Wilkins 1990; Verschueren 1983).
Many compounds with relatively high Henry's Law constants, such as
benzene, are also readily biodegradable under aerobic conditions. Com-
pounds with low Henry's Law constants, such as phenol, will readily un-
dergo biodegradation, but will not be easily removed through volatilization.
Thus, there is more flexibility in bioremediating the former compounds
with relatively high Henry's Law constants, for which designs can take
advantage of their volatility and ease of biodegradation.
Following are the site characteristics that are important for air sparging:
• contaminant identification, including properties;
• contaminant levels and distribution in each phase;
• depth to groundwater;
• thickness of aquifer;
• soil types and distribution;
• surface conditions;
• utility trenches;
3.109
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Process Identification and Description
• presence of receptors; and
• groundwater flow rate and direction.
Identification of the contaminants is important so that the properties of
the compounds can be evaluated for sparging suitability. The contaminant
levels and distribution are needed for predicting oxygen and, possibly, nu-
trient requirements and evaluating potential for inhibition.
The hydrogeological information is important because the site
hydrogeology will determine whether air can be injected at sufficient rates,
the flow patterns that will result, and the design of the air injection system.
Air will obviously be more readily introduced into coarse than fine soils.
Air injected into soils will tend to move through the more permeable zones.
This can lead to channeling and result in areas of the contaminated zone
that are not treated. If air is injected below a confining layer or clay lens,
the air will move laterally before reaching the water table.
Figures 3.13a to 3.13c show the effect of site hydrogeology on air-flow
patterns and, therefore, on the systems' design and performance. In figure
3.13a, the soils are relatively homogeneous. Air travels radially from the
sparging well and reaches the vadose zone at a distance approximately
equal to that from the water table to the top of the sparging well screen. In
figure 3.13b (on page 3.111), the soils are relatively fine, except for a
coarse sand layer located near the bottom of the saturate interval. In this
type of matrix, the air flows preferentially through the coarse layer, signifi-
Figure3.13a
Air Sparging in Homogeneous Aquifers
Compressed Air
3.110
-------
Chapter 3
Figure 3.13b
Air Sparging in Aquifers With Gravel/Sand Layer
Compressed Air
- Sand and Gravel
cantly increasing the lateral movement of air. But, the injected air may not
pass through soils above this layer. Similarly, in figure 3.13c, less perme-
able soils near the top of the aquifer cause lateral movement of the air be-
fore it can reach the vadose zone.
The effect on oxygen distribution attributed to stratified layers in the
saturated zone was observed during a treatability test of an air sparging
system. A 0.3 m (1 ft) thick gravel and sand layer near the bottom of a 2 m
(6 ft) thick saturated interval consisting of silt with some sand resulted in
dissolved oxygen levels increasing from 1.2 mg/L to 5.5 mg/L at a distance
Figure 3.13c
Air Sparging in the Presence of Clay Lenses in Upper Portion of Aquifer
Compressed Air
3.111
-------
Process Identification and Description
of 10 m (32 ft) in less than 2 hours (Norris and Clarke 1991). Dissolved
oxygen levels were increased at a distance equal to at least five limes the
saturated interval above the sparging well screen.
Depth to groundwater is important in treating aquifers where volatile
components are present. Regardless of the design and operating param-
eters, some volatile compounds will be carried into the unsaturaled zone.
When the water is very shallow, it is difficult to design an efficient vapor
recovery system unless the surface is capped. Furthermore, the period dur-
ing which contaminants travel through and can be biodegraded in a thin
vadose zone is short. On the other hand, shallow water tables may allow
nutrients to be introduced to the water table through percolation.
For air sparging, the radius of influence is typically defined as; the great-
est distance at which the air reaches the upper surface of the aquiifer. When
air sparging is used to promote biodegradation, the radius of influence of
interest is the greatest distance at which significant increases in dissolved
oxygen can be attained. The radius of influence for air sparging is a useful
quantity in that it is the minimum radius of influence for sparging to pro-
mote biodegradation.
A conceptual design for air sparging can be developed by assuming that
the radius of influence of air reaching the water surface is approximately
equal to the saturated interval to a depth of 9 m (30 ft). At greater depths,
uncertainties concerning air flow increase. The radius over which dissolved
oxygen levels can be increased can be significantly larger than the radius
for physical removal. This information can provide an indication of feasi-
bility. But, since soils are typically heterogeneous and there has been lim-
ited documented experience with air sparging systems, it is necessary to
conduct field studies to determine the radius of influence at each site.
For aquifers contaminated with nonvolatile constituents only, iit is suffi-
cient to determine the radius over which increased dissolved oxygen will be
attained. For several reasons, dissolved oxygen measurements should be
used as semiquantitative data. This can be accomplished using a sparging
well and groundwater wells or drive points. The sparging well should be
installed close to the bottom of the aquifer. Continuous split-spoon samples
should be taken and carefully logged, unless nearby boreholes or wells have
already been sampled in this manner. Wells or drive points should be in-
stalled along a straight line at distances equal to one-half, one, two, four,
and eight times the distance between the top of the air sparging well screen
3.112
-------
Chapter 3
and the water table. At least one additional well should be installed along a
line perpendicular to the line of monitoring points. Air should be injected
into the sparging well sequentially at three increasing pressures and flow
rates. The dissolved oxygen levels should be determined at decreasing time
intervals after each change in operating conditions. Sparging rates can be
increased without interruption of the air flow; however, air flow can be
terminated for several hours before testing a higher rate of air flow to allow
air to be dispersed uniformly throughout the treatment zone. The latter
method is preferred if channeling is likely to be a problem.
In aquifers where volatile compounds are present, it is also necessary to
monitor the unsaturated zone for radius of influence and for these com-
pounds. In addition to one or more sparging wells and several monitoring
points in the saturated zone, monitoring points in the unsaturated zone and,
typically, a vapor recovery system will be required. Monitoring points in
the unsaturated zone can be wells screened across both the saturated and
unsaturated zone, or separate monitoring points in the unsaturated zone. In
shallow aquifers, soil-gas monitoring probes can be used to sample the
unsaturated zone. The low cost and ease in installing these probes allows
extra soil-gas sampling points to be inserted as the test proceeds to more
precisely locate the radius of influence.
The radius of influence has been estimated by measuring increases in air
pressure at probes located above the water table. The drawback of this
approach is that increases in air pressure will extend past the point where air
reaches the water table. A paper by Wilson et al. (1992) describes a math-
ematical model that uses air pressure readings from several probes to esti-
mate the actual radius of influence. Another approach is to add a tracer gas,
such as helium or sulfur hexafluoride, to the sparged air and monitor the
discharge of an operating vapor recovery well located proximate to the air
injection well; the vapor recovery well must be operated at a flow rate ex-
ceeding that of the sparging well (Norris and Mutch 1991). Air reaching
the unsaturated zone will contain low levels of the tracer gas. The gas that
reaches the unsaturated zone will be captured by the vapor recovery well.
Thus, any monitoring point from which tracer gas is detected must be
within the region in which the air bubbles reached the water table.
For very shallow aquifers, it may be possible to observe air in vertically
inserted pipes that extend a few centimeters into the water table. This
method gives the most direct indication of air reaching the water table and,
3.113
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Process Identification and Description
because of the ease of installation, several points can be installed before or
during the field test.
Some test protocols call for measuring the radius of influence sequen-
tially during air sparging only, air recovery only, or air sparging and air
recovery together (Brown 1993). Radii of influence induced by both air
sparging and vapor recovery and the concentrations of volatile constituents
in the recovered air are monitored. This approach provides design data for
placing air recovery and air sparging wells and an indication of the transfer
of volatile constituents to the unsaturated zone.
Several design strategies are possible. Systems can be designed to mini-
mize remediation times by maximizing both biodegradation and physical
removal at a cost of increased capital and operating costs (especially for
offgas treatment). Other systems can be designed to minimize costs by
favoring biodegradation over physical removal, which is likely to require
longer to reach clean-up goals.
In one design, sparging wells are placed in staggered rows (figure 3.14),
commonly referred to as a "five spot." Staggering adjacent rows improves
the overlap between wells. The wells should be spaced at 70 to 80% of the
radius of influence to provide complete overlap. To increase the probability
of all areas of the site being remediated in the same time frame, spacing
should be closer in the most highly-contaminated areas.
In formations free of rubble and large stones, it may be possible to use
one or more horizontally installed wells. This approach was used at the
Savannah River Project for stripping chlorinated solvents from the saturated
Figure 3.14
Plan View of Staggered Rows of Air Sparging Well
3.114
-------
Chapter 3
zone and in an ongoing test to inject air and methane to promote
cometabolism of the residual chlorinated solvents (Schroeder et al. 1992).
The design called for the addition of methane, as well as air and nutrients,
to the aquifer to promote degradation of chlorinated solvents through
cometabolism.
For sites with volatile constituents, it is likely that an air recovery system
will be needed unless biodegradation can be completed in the unsaturated
zone. The same type of air recovery systems as discussed in Subsection
3.5.1.1.2, Bioventing, can be used with the same strategies for either physi-
cal removal or biodegradation as the primary mechanism for removal of
volatile compounds from the unsaturated zone.
In most instances, the radius of influence of air recovery wells will be
greater than that of sparging wells, particularly for capped sites. Thus,
fewer air recovery wells than air sparging wells will be normally incorpo-
rated into the design. The total flow of the air recovery wells, however,
must be greater than the total flow of the air sparging wells. Spacing and
location of recovery wells relative to air sparging wells will depend on the
concentration of volatile compounds and whether the intent is to enhance
biodegradation in the unsaturated zone or treat the contaminants at the sur-
face.
In many instances, it may be necessary to capture groundwater to prevent
migration of dissolved constituents during remediation. Designs for typical
pump-and-treat systems should suffice, taking into account that the air
sparging system may, in some geological conditions, increase the lateral
movement of the groundwater.
Nutrient addition may or may not be needed and feasible. If nutrient
addition is required and is very difficult or infeasible, it will probably be
best to design and operate the system to promote physical removal of vola-
tile compounds rather than biodegradation.
For shallow aquifers, particularly those with sandy soils, nutrients may
be most effectively added by percolating from the surface. In general, re-
covered groundwater can be amended with nutrients and then re-injected.
System Operation. Operation of air sparging systems to maximize the
biodegradation contribution requires that adequate nutrient levels are
present while initially injecting oxygen for a short duration with relatively
long times between air injection periods. This minimizes air stripping dur-
3.115
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Process Identification and Description
ing a lag period and allows sparging wells to be operated intermittently. To
be efficient, the wells must be operated at sufficient pressure to have ad-
equate radii of influence. Air is then introduced at much faster rates than
can be used for biodegradation. Thus, it is only necessary to operate the
wells for short periods (e.g., 0.5 to 1 hr) between relatively long inoperative
periods (e.g., 12 to 48 hr). The off/on schedule can be developed based on
dissolved oxygen levels in groundwater and volatile compounds in the un-
saturated zone.
A fortuitous effect of the off/on operating schedule is the change in the
groundwater elevations (Brown 1993). During sparging, the groundwater
level rises an increment in the vicinity of the sparging well, in addition to
the rise resulting from air extraction. After a short period of operation or
when sparging is interrupted, the water table drops below the equi librium
level that existed before sparging and then rises to the equilibrium level.
The soils near the water table in the vicinity of the sparging well alternate
between being saturated and unsaturated; thus this interval may experience
the greatest benefit as a result of increased oxygen flow through unsaturated
as well as saturated soil and maintaining high levels of moisture. As a re-
sult, these soils are subject to the highest levels of physical removal.
During operations, measurements should include:
• concentrations of carbon dioxide, oxygen, and volatile com-
pounds in the vadose zone;
• contaminant concentrations, dissolved oxygen (and maybe car-
bon dioxide), and pH in groundwater;
• groundwater levels during and between air injection; and
• flow rates and air pressure of the air sparging and vapor recov-
ery systems.
Insufficient information has been reported about air sparging to establish
rules of good practice. But, it is essential that persons involved in the de-
sign and implementation of this technology understand the principles in-
volved and follow a generally conservative approach, especially with re-
spect to the potential for uncontrolled migration of volatile compounds
within the unsaturated zone and for migration of dissolved constituents.
Some level of pilot testing is required in developing remedial designs, and
designs should be flexible so that additional components can be added to
the system.
3.116
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Chapter 3
This technology does not present any unique health and safety require-
ments other than that of insuring that vapor and dissolved phase migration
are monitored and controlled, if necessary. However, great care does need
to be taken with regard to migration of vapors into buildings and utility
trenches. Otherwise, following the regulations prescribed under the Occu-
pational Safety and Health Act (OSHA) should be adequate.
3.5.1.3 Costs of In Situ Bioremediation Technologies
The cost of in situ bioremediation is site- and contaminant-specific. The
types and concentrations of contaminants, the volume of contaminated ma-
terial, permeability, soil characteristics, clean-up standards, depth to water
table, monitoring requirements, and site location will affect cost (Norris,
1993). Other factors that are specific to the technology will also affect cost
(Table 3.19).
Table 3.19
Costs of In Situ Bioremediation Technologies
Technology
Land treatment
Bioventmg
Liquid delivery
Air sparging
Waste
Petroleum hydrocarbons;
contaminated soil; sludge
Petroleum hydrocarbons
Petroleum hydrocarbons
Petroleum hydrocarbons
Technology-Specific
Factors Affecting Cost
liner; runoff requirement
- offgas treatment
+ offgas treatment
oxygen source
Cost ($/m3)
$10 - 80
$10-20
$52 - 78
$50 - 200
$25 - 150
US EPA 1991
3.5.2 Ex Situ Bioremediation Technologies
Ex situ bioremediation technologies are those in which a waste that has
been removed from its point of origin is treated in a closed or open
bioreactor. The types of wastes that can be treated include liquids, solids,
and air. Although a bioreactor is commonly thought of as a vessel in which
3.117
-------
Process Identification and Description
the waste is treated in a controlled manner, there are other forms, including
lined lagoons, soil-piles, composting piles, and soil (bio) filters. All
bioreactors, however, are designed and managed to maximize the biological
reaction in an economical manner. The reactor design solves particular
problems that are encountered with the contaminated wastes under consid-
eration.
There are two basic problems to be solved in any design of an aerobic
bioreactor. The first is that of contact between the bacteria and the organic
contaminants. The second is that of oxygen transfer to the bacteria. The
various bioreactor designs that are available can be compared by the way
they solve these two problems. Other criteria also will be used to establish
the advantages and disadvantages of specific reactor designs, but bacterial
contact and oxygen transfer are the two functions that are common to all
reactor designs.
Bacterial contact amounts to more than merely mixing the bacteria with
the organic contaminants. The goal of the biological reaction is to destroy a
maximum amount of the contaminants and to leave a minimum concentra-
tion of the contaminants. To achieve these goals, the bacteria must be put
in contact with the contaminants and given extended periods to complete
the biochemical reactions. In other words, the bacteria must have a long
residence time in the reactor. Figure 3.15 (on page 3.119) shows the rela-
tionship between the effluent organic concentration and the residence time
of the bacteria.
Oxygen transfer does not affect the performance of the reactor design as
long as a minimum oxygen concentration is maintained. Oxygen transfer is
related mainly to the cost of biological treatment. Energy for oxygen trans-
fer is usually a main operating cost of a bioreactor, exceeded only by labor.
3.5,2,1 Ex Situ Treatment of Contaminated Water4
Many designs for biological treatment of contaminated groundwater are
based on systems originally designed for treating relatively high organic
strength wastewater. It is inappropriate to assume that if "activated sludge"
cannot be used successfully to treat the lower organic loads associated with
groundwater contamination problems then biological treatment will not
4. Reprinted by permission of Van Nostrand Reinhold Publishers from •'Groundwa-
ter Treatment Technology," Second edition, by E. Nyer. Copyrighl 1992 by Van
Nostrand Reinhold Publishers.
3.118
-------
Chapter 3
Figure 3.15
Effluent Organic Concentrations With Increasing
Bacterial Residence Time
S =
Effluent organic concentration with
increasing liquid residence time in
an aerated lagoon
Increasing
Liquid Residence Time
work. If a biological reaction is possible, then biological reactors can be
designed in such a way that problems specific to groundwater treatment are
addressed. Several designs have recently been developed specifically for
low concentrations. Similar problems must be overcome when treating
liquid waste streams emanating from pond remediations.
Contaminant concentrations can follow three patterns over the life of a
project (figure 3.16 on page 3.120). First, there is the constant concentra-
tion exhibited by a leachate. If the source of contamination is not removed,
the source will replace the contaminants as fast as they can be removed with
the groundwater system, and the concentration will remain constant.
The second possible pattern arises when the contamination plume is
being drawn toward a groundwater removal system, such as a municipal
drinking water well system. In this situation, contaminant concentrations
increase over time. The well is originally clean, but becomes more con-
taminated as the plume is drawn toward the well.
The final pattern is that incident to remediation. If the original source of
contamination is removed, the concentration of the contaminants will de-
crease over time. The reduction in contaminant concentration is a result of
3.119
-------
Process Identification and Description
Figure 3.16
Time Effect on Concentration Found in a Well
Leachate
Increasing
Time
retardation, natural chemical and biochemical reactions, and dilution by the
surrounding groundwater.
A bioreactor designed for groundwater must be able to treat the contami-
nants over the entire life cycle of the remediation. Simply designing the
system to treat the initial concentration is not sound engineering design.
Bioreactors that treat contaminated water can be separated into five main
groups: suspended-growth reactors, fixed-film reactors, reactors based on
activated carbon, submerged fixed-film systems, and miscellaneous designs.
In suspended-growth reactors, the bacteria are grown in the water and inti-
mately mixed with the organic contaminants in the water. In a fixed-film
system, bacteria are grown on an inert support medium within the reactor
and the water containing the contaminants passes over the film of bacteria.
Submerged fixed-film designs place the medium below the water level, and
activated carbon can serve a dual role of adsorption and support media. The
miscellaneous designs are the many special designs developed during the
last several years.
Costs of ex situ treatment of contaminated groundwater cannot be esti-
mated easily. The nature of the contaminated groundwater will determine
which type of ex situ treatment is used and the costs will be highly contami-
3.120
-------
Chapter 3
nant and vendor specific, depending on the equipment required to build the
reactor.
3.5.2.1.1 Suspended-Growth Reactors The simplest bioreactor system for
treatment of groundwater is an aerated lagoon or basin. An existing pond or
tank can be used as the reactor. In some cases, portable swimming pools
have been used as the aeration tank. Figure 3.17 shows the configuration of
an aerated lagoon. The contaminated groundwater enters the aerated vessel
and bacteria in the reactor degrade the contaminants and create new biom-
ass. The liquid residence time in the reactor, which is equal to the bacterial
residence time, must be sufficient for the bacteria to reproduce.
Oxygen is supplied to the tank by a surface aerator or air diffusers. Suf-
ficient power must be supplied to provide an adequate oxygen concentra-
tion, 2 mg/L, and/or to keep the tank completely mixed. With reactors hav-
ing low residence time, oxygen supply is usually the controlling factor for
bioreactor performance, whereas mixing is usually the controlling factor for
those having long residence time. A 2-day residence time is the minimum
recommended to maintain low concentrations of effluent contaminants, and
two days is about the time required for the bacteria to reproduce and replace
the biomass lost in the effluent. At lower residence times, the bacteria are
washed out of the reactor such that insufficient numbers of bacteria remain
for efficient removal of contaminants. At longer residence times, total flow
Figure 3.17
Aerated Lagoon
o, N, P
CO,
^ T '
Groundwater
Organic
L_1__J
Aerated lagoon
Clean
Water
Bacteria
3.121
-------
Process Identification and Description
and the cost of power to mix the tank will be important operational factors.
The main operating cost (other than personnel) for any aerobic biological
treatment system is that of the power required for oxygen supply or mixing.
These costs will limit the size of the reactor.
Two problems usually arise with the use of an aerated lagoon design.
First, the extent of treatment, usually 50 to 70% of the biodegradable or-
ganic content, is controlled by the limited residence time of the bacteria.
The residence time is limited because the bacteria that are created in the
reaction will be in the water when it leaves the reactor. Second, the pres-
ence of bacteria in the effluent also can elicit regulatory concerns. A clari-
fier can be added to the system to remove the biological solids; however,
bacteria grown in an aerated lagoon do not settle readily.
These problems can be solved by separating the liquid residence time
from the bacterial residence time. Figure 3.18 shows that by adding a clari-
fier to remove the bacterial solids from the water stream and returning them
to the aerated reactor, the bacterial residence time is independent of the
Figure 3.18
Bacterial Residence Time With Life Cycle Influent Concentration
60
50
40
30
I 20
10
MLSS = 3,000 mg/L
^^,
MLSS = 1,500 mg/L
500 1000 1500
Influent Organic Concentration (mg/L)
3.122
-------
Chapter 3
liquid retention time. From figure 3.18, the liquid residence time (RL) is
calculated as:
RL = V/Q
The Bacterial Residence Time (RB) is calculated as figure 3.19:
X*V
B~QW,XR+(Q-QW)XE
[7]
[8]
The bacterial residence time is governed by the wasting of the settled
bacteria from the clarifier and by the uncontrolled loss of bacteria in the
clarifier effluent. Bacteria returned to the aeration tank are called activated
sludge (figure 3.19).
The activated sludge process is the most widely-used method of biologi-
cal treatment in wastewater treatment. The basic advantages are
• the process produces low-effluent concentrations from water
containing high concentrations of organic compounds;
• the system can treat many organic contaminants at the same
time; and
Figure 3.19
Activated Sludge
x,v
Q + QR
Q-QW
Q = Flow
QB = Recycle Flow
Qw = Sludge Wastage Flow
X = Mixed Liquor Suspended Solids
XR = Clarifier Underflow Solids Concentration
XE = Effluent Solids Concentration
V = Volume of Aeration Basin
S = Organic Concentration
3.123
-------
Process Identification and Description
• the same equipment can be used to treat a variety of i nfluent
conditions (with equalization of the contaminant load to prevent
overloading the system).
The main disadvantages are:
• the cost of manpower to keep the system adjusted to the influent
conditions;
• the relative cost of oxygen transfer compared to fixed-film sys-
tems; and
• the critical need to keep the bacteria in a growth stage: in which
their settling tendencies are at a maximum.
The activated sludge process can remove 85 to 95% of the biodegradable
organic content from the influent, and 99+% of specific organic com-
pounds, depending on influent concentrations. Effluent concentrations of
10 to 30 mg/L of biochemical oxygen demand (BOD), a general measure-
ment of biodegradable organic material in wastewater, can be expected with
a properly operated system. Effluent concentrations of specific compounds
will vary. For example, effluent phenol concentrations from an activated
sludge system can be as low as 0.01 mg/L.
Certain compounds (e.g., sugars and alcohols) will degrade very quickly
in a biological system. Other compounds may require longer retention
times to degrade. The more readily a compound can be assimilated by the
bacteria, the faster and more efficiently the bacteria can incorporate the
compound into new biomass. Another approach to understanding retention
time is to realize that the bacteria must first remove the easily degradable
contaminants before the enzymes necessary to degrade the refractory com-
pounds are induced. In the design of the treatment plant, this can be repre-
sented by the Food to Microorganism ratio, F/M. From figure 3.19 (on page
3.123), the formula would be:
F/M = Q*S/V*X [9]
In the activated sludge process, all of the compounds are being degraded
at the same time in a completely-mixed tank. The bacteria residence time
(RB) or F/M models can be used to understand how to accomplish this con-
current degradation. One of the main advantages of the activated sludge
process is that the bacterial residence time and the F/M ratio can be con-
trolled to accommodate the degradation of a variety of compounds that have
different degradation rates.
3.124
-------
Chapter 3
The main problem with the activated sludge design is the critical need to
keep the bacteria in a form in which they readily settle. If the bacteria do
not settle properly, the clarifier will not be able to remove them from the
water stream. If the bacteria are not separated from the water stream and
returned to the aeration basin, the whole process fails. This does not mean
that the bacteria have lost their ability to degrade the contaminants. How-
ever, without being able to separate the bacterial residence time from the
liquid residence time, the net result is that the activated sludge process is
converted to a simple aerated lagoon with lower removal rates.
To maintain the settling properties, the environment in which the bacteria
grow must be maintained relatively constant, and the bacteria must be
grown to the proper sludge age that promotes flocculation. In groundwater
treatment, the influent concentration of contaminants has very little varia-
tion on a day-to-day basis. There is normally no need for equalization (con-
trolling the influent concentration of organic compounds to prevent over-
loading the system) as in wastewater treatment. The main problems using
the activated sludge process in treating groundwater are the changing con-
centration of the contaminants during the project life cycle and growing the
bacteria in their flocculant stage during the entire project. For illustration,
assume figure 3.20 represents the influent life cycle concentration. The
Figure 3.20
Life Cycle Concentration From a Well at the Center
of the Plume for an Organic Contaminant
1500
1000
500
00
6 8 10
Time, Years
3.125
-------
Process Identification and Description
flow (Q) will be 94,625 L/day (25,000 gpd) for the entire life of the project.
Also assume that all other environmental parameters are acceptable for
biological treatment. Figure 3.19 (on page 3.123) will again represent the
activated sludge treatment system. To keep the bacteria in a growth phase
in which they settle properly, the bacteria should have a Residence Time
(RB) between 5 and 20 days. The following scenario describes what hap-
pens to the sludge age during the life of the project.
Assume:
Mixed Liquor Suspended Solids (MLSS)(X) = 3,000 mg/L
Yield Coefficient (Y) = 0.25 g/g (0.25 Ib/lb)
Volume of the Aeration Tank (V) =151,400 L (40,000 gal)
RB = (X*V)/(Q*S*Y)
For Year 1,S = 1500 mg/L
RB= 12.8 days
For Year 3, S = 1200 mg/L
RB = 16 days
For Year 5, S = 600 mg/L
RB = 32 days
And For Year 7, S = 300 mg/L
RB = 64 days
As can be seen from these data, the system will maintain the proper
sludge age for about 4 years. After this time, the bacterial residence time
will be too high, the bacteria will lose their settling properties, and the clari-
fier will not be able to separate the bacteria from the treated water. Once
the clarifier fails, the system will not be able to maintain a high concentra-
tion of bacteria in the aeration basin. At this point, the system will no
longer remove a high percentage of the incoming organic contaminants.
One solution to this problem is to lower the MLSS concentration. Figure
3.18 (on page 3.122) summarizes the bacterial residence times for the treat-
ment system at MLSS levels of 3,000 mg/L and 1,500 mg/L. Lowering the
MLSS concentration does extend the useful life of the treatment system, but
the system still fails before the cleanup can be completed. In addition, there
is a lower limit to the MLSS. The MLSS concentration entering the clari-
3.126
-------
Chapter 3
fier must be about 1,250 mg/L or above to ensure proper settling. Bacteria
rely on flocculation in order to settle and a critical biomass is required to
ensure enough contact between the flocculating particles.
Another method to extend the useful life of the system is to divide the
aeration basin into two or more tanks. In our example, two 75,700 L
(20,000 gallon) tanks, instead of the one 151,400 L (40,000 gallon) tank,
could be used. Assuming 1,500 mg/L MLSS, at year 6, one aeration basin
could be shut down. This would effectively halve the bacterial residence
time in the system at a steady MLSS. An added advantage of this method
would be that half of the blowers could also be shut down. Not only would
the system last longer, but it would also cost less to run in the final years of
operation.
Of course, an activated sludge system designed in this manner has limita-
tions. The final few years of the cleanup will still create a very long sludge
age. The actual design may have to include different unit operations to
clean up the groundwater over the entire life cycle. Further, even with an
optimal design, the activated sludge system will still require a relatively
high level of operator attention to properly maintain the operation of the
system.
Two other equipment designs using suspended growth bacteria are the
extended aeration and contact stabilization designs (see figure 3.21 on page
3.128). Although neither of these designs has a particular advantage for
groundwater treatment, both have been widely used in wastewater treat-
ment.
The only difference between the extended aeration and activated sludge
designs is that the reaction chamber (aeration basin) is larger in the ex-
tended aeration design, which extends the time that the bacteria are aerated.
The performance of the extended aeration process is more stable (the larger
tanks serve as internal equalization) than that of activated sludge and as a
result, produces less waste sludge.
Most "packaged plants" that can be purchased from vendors are extended
aeration designs. Therefore, the reader has a good chance of encountering
the proposed application of this design for treatment of groundwater. When
used in treating groundwater, the extended aeration design will have the
same limitations as the activated sludge design.
3.127
-------
Process Identification and Description
Figure 3.21
Activated Sludge, Extended Aeration,Contact Stabilization
Completely Mixed Activated Sludge
Waste ^ ^
t
W»* > %
t
U Effluent •CT
Biomass
Recycle
Extended Aeration
1 Effluent ^ —
w^fm i
biomass
Recycle
•777 Clarified Effluent ^
Excess Biomass ^
— 7 Clarified Effluent __
Excess Biomass w
Wai
frste I tik
Contact/Organic Uptake
Contact Stabilization
Effluent
Clarified Effluent ,
Biomass
Recycle
Excess Biomass
Stabilization/Organic Digestion
The contact stabilization process is widely used for high concentrations
of easily-degradable organic compounds. The waste comes into contact
with the bacteria in a small aeration tank (figure 3.21). The bacteria quickly
assimilate the organic contaminants without digesting them and, with the
organic contaminants, are removed in a clarifier and sent to a large aeration
tank. The bacteria digest (stabilize) the contaminants in this tank. After the
contaminants are digested, the bacteria are returned to the contact tank and
the cycle starts again. The main purpose of this design is to save space and
subsequent operating costs. It is unlikely that the reader will encounter this
design in a groundwater treatment system; however, it may be mentioned in
a feasibility study and knowledge of its basic design will be necessary.
3.128
-------
Chapter 3
One final method for applying suspended-growth bacteria is the sequenc-
ing batch reactor, which has been applied widely to industrial waste during
the last several years. It also has been used to treat landfill leachate and for
pond remediations. Following are the main steps in the operation (see fig-
ure 3.22):
Figure 3.22
Sequencing Batch Reactor Treatment Stages
Fill
React
Settle
Draw
Idle
1. Fill the reaction tank with the contaminated water while main-
taining full aeration;
2. Once the tank is full, bacteria completely digest the organic
contaminants;
3. Stop aeration and subsequent mixing, and allow the bacteria to
settle;
4. Decant the clean water and discharge; and
5. Start the cycle again.
The sequencing batch reactor design often uses two tanks operated in
parallel. While one reactor is accepting water, the other reactor is going
through the subsequent steps of digestion, settling, and decanting. The
reactors switch back and forth to maintain a constant influent flow. The
3.129
-------
Process Identification and Description
advantages of the sequencing batch reactor are simplicity of operation and
flexibility under a variety of influent conditions. The main disadvantage for
groundwater treatment would be operation with low concentrations of influ-
ent organic contaminants. See table 3.20 for a summary of the advantages
and disadvantages of suspended-growth designs.
Table 3.20
Suspended Growth Systems
Advantages Disadvantages
Intimate contact between biomass and waste Completely dependent on clarifier performance
Several methods available for adjusting performance High operation attention required
Very low concentration of specific orgamcs in effluent
Large-scale system relatively inexpensive
3.5,2.1.2 Fixed-Film Reactors Another way to use bacteria in an
aboveground treatment system is to set up a fixed-film biological unit. In
fixed-film systems, an inert support medium with a large surface area is
placed in the reactor. Bacteria naturally attach and grow to form a biofilm
on any surface provided them (figure 3.23 on page 3.131). The contami-
nated water enters the tank and forms a thin film over the attached bacteria
(fixed-film) into which the contaminants diffuse. The bacteria degrade the
organic contaminants and the waste by-products (CO2, H2O) diffuse into
the water film. Oxygen from the atmosphere diffuses through the water
film and into the bacterial fixed-film. There are four important advantages
in the fixed-film systems:
• the bacteria can be maintained at a high concentration without
the need of a clarifier;
• oxygen can be supplied at low cost;
• the system can tolerate some variation in organic load and the
intermittent presence of toxic chemicals; and
• the system is easily operated.
3.130
-------
Chapter 3
Figure 3.23
Fixed-Film Bacterial Growth
Subsurface Anaerobic
Microbes
A fixed-film system requires less operator attention than an activated
sludge system. Bacteria will grow attached to the medium and remove
organic contaminants from the water over a wide range of operating condi-
tions. When there are too many bacteria in the fixed-film, the bacteria will
slough off and leave the reactor with the water. A clarifier can then be used
to remove the biological solids before final discharge of the effluent.
The two main types of fixed-film reactors are trickling filters and rotat-
ing biological contactors (RBC). Figure 3.24 (on page 3.132) illustrates a
trickling filter design. The contaminated water is pumped to the top of the
reactor and distributed over the medium. The water is broken up into thin
films and trickles down through the medium.
Several types of inert support media can be used in a trickling filter.
Originally, small 8 to 13 cm (3 to 5 in.) diameter rocks were used to support
the bacterial population. Because of the low surface area per unit volume of
rock and the low oxygen transfer capacity that resulted from the small void
space, only a small bacterial mass developed. Plastic media have replaced
rocks in recent years. The two main categories of plastic media are dumped
3.131
-------
Process Identification and Description
Figure 3.24
Trickling Filter
Waste .
\/I\/]\/|\
Flow Distributor
Inert Packing
\f
Excess
Biomass
packing and stacked packing. Dumped packing is the same type of plastic
medium used in packed-tower air strippers and usually is used as a replace-
ment for rocks in existing trickling filters and small new systems. Stacked
packing comes in large bricks and usually is applied to large systems.
Figure 3.25 (on page 3.133) illustrates an RBC design. The system is
characterized by an elongated tank or tank series containing cylinders of
plastic media, which are rotated longitudinally, as the contaminated water
passes through the tank(s). In this system, the water enters one end of the
tank. The medium first rotates down into the water where the contaminants
come into contact with the bacteria. The medium then rotates up into the
atmosphere, after which a thin film of water forms on the medium, and the
oxygen transfers through the thin film of water to the bacteria. Trie RBC is
probably the most energy-efficient oxygen transfer method for biological
systems.
3.132
-------
Chapter 3
Figure 3.25
Rotating Biological Contactor
\V5iste
Rotating
Disks
Clarified
Effluent
Excess
Biomass
There are several technical disadvantages in the fixed-film reactors.
Fixed-film reactors are plug-flow reactors. The water comes in at one end,
passes by the bacterial film, and exits the other end of the reactor. As a
result, the influent end of the reactor is subjected to the high concentration
of the influent contaminant. In completely-mixed reactors, the influent is
immediately mixed with the contents of the tank. The influent contamina-
tion may be toxic, or pockets of high concentrations of material may be
found as the groundwater is recovered from the aquifer. The bacteria in the
fixed-film reactor will be subjected to the full concentration. The effluent
water can be recycled to minimize this effect, but this adds to the cost of
operation.
Another problem with fixed-film reactors is that they will not remove as
high a percentage of the influent contaminants as an activated-sludge sys-
tem. Removal of particular contaminants is very important in groundwater
treatment. General removal of organic compounds will be important, de-
pending upon the final disposal of the water. The design engineer can ex-
pect 75 to 90% BOD removal and 85 to 95% removal of a specific organic
compound. The lower the influent concentration, the lower the percentage
removal that can be expected. Table 3.21 (on page 3.134) summarizes the
advantages and disadvantages of the fixed-film systems.
3.133
-------
Process Identification and Description
Table 3.21
Fixed Film Reactors
Advantages Disadvantages
Low operator attention Plug flow
Retention of slow growing bacterial populations Limited operation at high influent concentrations
Low cost oxygen transfer Hard to adjust operation
Resistent to shock loads
Can be shut down for up to a few weeks and
then restart
3.5.2.1.3 Submerged Fixed-Film Reactors A relatively new biological
design is a combination of suspended-growth and fixed-film reactor de-
signs, generally referred to as submerged fixed-film reactors (figure 3.26 on
page 3.135). In these units, the plastic medium is submerged in trie water in
the reactor tank. The bacteria grow on the plastic medium as in a fixed-film
system; however, the water is in constant contact with the film, as opposed
to passing through in thin films.
There are two main ways in which the submerged fixed-film design can
be used. In the first design, used for many years in wastewater treatment
(Nyer and Ziegler 1983), the reactor is designed for completely mixed op-
eration and for handling influents with high concentrations of organic con-
taminants in the influent. The medium is aerated from below, and as it is
released, the air pushes water in front of it as it rises, creating an air-lift
pumping action. With sufficient air, the biological reactor tank will be
completely mixed. This design mode can accommodate an influent with an
organic content ranging from 50 to 5,000 mg/L.
The main advantages of this design are ease of operation and high-qual-
ity performance. Submerged fixed-film reactors can perform as well as
activated sludge units; however, they do not depend on a clarifier for main-
taining the bacteria in the reaction tank. This design accomodates a large
variety of operating conditions and requires little attention by operators.
The submerged fixed-film unit combines the advantages of the suspended-
growth and fixed-film systems without the disadvantages of either.
3.134
-------
Chapter 3
The main disadvantages of the submerged fixed-film design are the high
cost of oxygen transfer and lack of scalability. Because of the nature of the
design, there is a natural height limitation to the tank and, therefore, oxygen
cannot be released at an optimum depth. The second problem lies in the
scaling of the unit. Suspended-growth and fixed-film units become more
economical as the systems get larger. Because the tank and the medium
both get larger in direct relationship to the size of the system, the sub-
merged fixed-film reactor does not provide an economy of scale for larger
systems.
Neither of these disadvantages has a large effect on groundwater applica-
tions. First, the cost of oxygen transfer is a small part of the total cost of a
groundwater biological treatment system. Second, most groundwater treat-
ment systems are relatively small and the cost advantage of large-scale
systems does not apply.
A second mode in which the submerged fixed-film units can be applied
to groundwater is in a low-concentration design. Submerged fixed-film
systems can be designed to treat influent concentrations as low as 1 to 20
mg/L, a very important consideration for groundwater applications. High
concentrations (greater than 50 mg/L) are rarely found in groundwater, and
Figure 3.26
Submerged Fixed Film
\\fcste
Water Level
Aeratii
ion
3.135
-------
Process Identification and Description
when they are, life-cycle design considerations reduce the concentration
below 50 mg/L in a short period of time.
Figure 3.27 illustrates the low-concentration design of a submerged
fixed-film unit. The basic design is the same as the original submerged
fixed-film design. Plastic media are submerged below the water level in the
reactor tank. The low-concentration reactor uses small amounts of air and a
plug-flow pattern. The water enters the top of the tank and is distributed
across the media. The water flows down through the media and exits the
bottom through a collection system. The air is released below the media.
Very small amounts of air are used due to the low oxygen demand in a low-
concentration reactor and the need to maintain a nonmixed state in the reac-
tor.
Even under these conditions, the low concentration of organic contami-
nants in the influent is not sufficient to maintain microbial growth in the
reactor. Normally, influent concentrations of less than 20 mg/L will result
in a rate of bacterial decay that is faster than bacterial growth. Therefore,
the low-concentration reactor must operate in a decay mode, not in the nor-
mal growth mode of biological treatment systems. In the decay mode, bac-
teria usually are grown on the fixed-film using a synthetic feed source until
the population size stabilizes. Then, the synthetic feed is removed and the
Figure 3.27
Low Concentration Submerged Fixed Film
Waste
Flow Distributor
Inert Support
.eration /
3.136
-------
Chapter 3
low concentration of influent is conveyed through the system. Under these
conditions, the bacteria slowly decay. With proper design and operation,
the decay period can last between 6 months to 1 year before regrowth is
required.
There are more than 20 low-concentration reactor units currently operat-
ing full scale in the field. Compounds that have been treated in this reactor
design range from acetone and methylethylketone to benzene and chlo-
robenzene. Reactors are also currently being used to treat groundwater
containing tetrahydrofuran and f-butanol.
3.5.2.1.4 Reactors Based On Activated Carbon There are several other
types of designs that have been used to treat organic contaminants in
groundwater. Two of the most popular in recent years are the powdered
activated carbon treatment (PACT) units and fluidized-bed reactors.
The PACT unit is basically an activated sludge treatment system with
powdered-activated carbon maintained in the reactor (figure 3.28). The
combination of powdered-activated carbon and active bacteria increases the
removal capabilities of the treatment system in comparison with systems
using either alone. The powdered-activated carbon can remove slowly
degrading or nondegrading organic material from the water while the bacte-
ria can attach to the powdered-activated carbon and consume the contami-
nants adsorbed to the carbon. The system has been used widely to treat
Figure 3.28
Powered Activated Carbon Treatment
Powdered Activated Carbon
Wiste
Recycle
Clarified
Effluent
Excess Biomass
and Carbon
3.137
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Process Identification and Description
hazardous organic waste. While it has been used mainly in the wastewater
field, the PACT system is now being used to treat contaminated groundwa-
ter. The main advantage of the system is that it can treat a large variety of
organic compounds. The main disadvantage is that it is basically an acti-
vated-sludge design, and will suffer the same limitations.
A fluidized-bed reactor is basically a submerged fixed-film system. The
fluidized-bed design has been applied recently to a number of groundwater
remediation problems, including treating tank bottom wastes, produced
water brines, and groundwater contaminated with BTEX, aliphati c hydro-
carbons, and PAHs from gas and oil production activities. The granular
activated carbon fluidized-bed reactor has been able to consistantly provide
removal of 99+% for BTEX, and 2- to 4-ring PAHs for all these streams at
liquid residence times of minutes and organic loading rates of 1 - 5 kg
COD/m3-d. The technology has been applied at over 40 full-scale and pilot-
scale installations within the United States. Currently, the technology is
being demonstrated for the removal of chlorinated solvents by operating the
system in an anaerobic mode, with detention times under an hour, In this
design, the support medium consists of small-diameter particles. Water and
air flow in an upflow pattern through the medium, fluidizing the bed. Typi-
cal media are sand, activated carbon, glass beads, and other small particles.
Recently, activated carbon has been the main medium applied in fluidized-
bed reactors. The activated carbon once again fulfills the dual purpose of
adsorbing organic contaminants and acting as an attachment site for bacte-
ria.
The fluidized-bed design can also be run in high- and low-concentration
modes. If the influent organic concentration is high (50 to 5,000 mg/L),
high ratios of recirculation are used with pure oxygen as the oxygen source.
When influent organic concentrations are low, recirculation is kept to a
minimum and the liquid residence time in the reactor is kept short (5 to 30
min). The fluidized-bed design has recently been applied to several full-
scale groundwater cleanups.
3.5.2.1.5 Miscellaneous Reactors Another possible approach to treatment
of contaminated groundwater is anaerobic reactors. All the current labora-
tory work in this area indicates that bacteria can be used to degrade chlori-
nated hydrocarbons. Once these reactions are well understood, and the
required environment defined, full-scale reactors will be possible. Several
3.138
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Chapter 3
anaerobic reactor designs are currently available, mainly used to treat
wastewater with high concentrations of organic constituents. The main
areas of application are the food and beverage industries. Full-scale data on
the application of anaerobic processes to groundwater are limited. But, as
the laboratory research leads to pilot-scale applications and the technology
is advanced, anaerobic reactors should be considered as an alternative treat-
ment.
Several other types of biological reactors are also available, and new
designs are constantly being developed. But with the exception of the low
concentration submerged fixed-film design and the activated carbon fluid-
ized-bed design, there are no other units that are currently being specifically
designed for treatment of groundwater. Again, it is very important to un-
derstand that just because activated sludge is not a viable design for treat-
ment of groundwater, it does not mean that biological treatment is not pos-
sible. Any new design for groundwater treatment will have to address the
specific problems of groundwater. The main problem areas are life-cycle
design, treatment of low concentration of organic contaminants, and port-
ability. The designer will have to understand the advantages and disadvan-
tages of each biological reactor and exploit them to the best advantage dur-
ing the life cycle of the project.
3.5.2.2 Ex Situ Treatment of Contaminated Soils and Sediments
Contaminated soils and/or sediments are frequently present at a site
where the groundwater quality has deteriorated. Bioremediation of soils
may be performed either in situ or ex situ. Some sites are not suited to in
situ treatment because of hydrogeological conditions at the site or the char-
acteristics of the wastes. At such sites, soils may be remediated using ex
situ treatment. This section addresses the various types of bioreactors used
in ex situ treatment.
Ex situ biotreatment reactors for soil remediation fall into two main cat-
egories, slurry-phase treatment and solid-phase treatment. In slurry-phase
treatment, contaminated soils or sludges are maintained as an aqueous
slurry. Solid-phase biotreatment relies on principles applied in agriculture
and in the biocycling of natural compounds; its process options include land
treatment, soil-pile treatment, and composting.
There are advantages and disadvantages to each of these designs, but all
of the reactors follow the basic concepts of any biological reactor. Their
3.139
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Process Identification and Description
main purpose is to maintain an environment in which the organic contami-
nants and bacteria can react under optimum conditions. The bacterial re-
quirements of environment (O2, pH, temperature, etc.) and nutrient addition
(NH3, PO4, etc.) must be met by all reactors. The solid-phase reactors have
one more requirement. They must also maintain the proper moisture con-
tent.
3.5.2.2.1 Slurry-Phase Biotreatment Slurry-phase biotreatment may be
performed either in bioreactor vessels or in lined lagoons, but the basic
units include aeration and mechanical mixing equipment and, sometimes,
an emissions-control system. Depending on the setting, slurry-phase
bioremediation may be compared with the activated-sludge process or aer-
ated lagoon treatment. In either application, soil or sludge and nutrient-
amended water are combined to form an aqueous slurry. Mixing must be
sufficient to keep the solids in suspension and oxygen must be supplied
throughout the slurry matrix to promote aerobic microbial activity.
Bioslurry reactors are operated so as to maximize mass-transfer rates and
contact time between the contaminants and microorganisms.
Oversized material must be removed. The first step in the treatment
process is to slurry the soil or sludge to be treated, which is passed through
a trommel screen to remove gravel and debris with diameters larger than
0.64 cm (0.25 in.). More water may then be added to obtain the desired
slurry density before bioslurry treatment. Maximum treatment efficiencies
are generally obtained with soil slurries containing 30 to 50% dry solids by
weight, although difficulty in maintaining the solids in suspension limits the
acceptable slurry solids content range to 20 to 30% for some bioslurry reac-
tors (Stroo 1991; Brox and Hanify 1989).
Three pilot-scale demonstrations of slurry-phase treatment, each lasting
60 days, were performed at a site contaminated with oil-refining wastes.
The contaminants of concern were PAHs and oil and grease (Stroo 1989).
The treatment vessels ranged from a 64,345 L (17,000 gal) reactor to a
2,800,000 L (750,000 gal) aerated lagoon. The reactor solids loadings
ranged from 5% to 30% (i.e., the slurries consisted of 5 to 30% dry solids
by weight). Data pertaining to the oil and grease treatment results were not
reported, however, reductions in total PAH concentrations ranging from 76
to 92% and in carcinogenic PAHs (i.e., 5-ring and 6-ring PAHs) concentra-
tions ranging from 25 to 89% were effected. The highest removal efficien-
3.140
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Chapter 3
cies were expected in the reactors with the lowest solids loadings, but this
was not the case. In fact, the highest total and carcinogenic PAH reductions
were obtained in the lagoon with 30% solids loading (Brox and Hanify
1989).
Slurry systems have an economic and technical advantage when the con-
taminated solids already contain a high-moisture content. Lagoon bottoms
are a prime example. In addition to the physical advantages of not having
to add water, the reactors actually perform two functions: normal degrada-
tion of the contaminants and reduction in volume, accomplished through
direct degradation, and by breaking of the emulsion with the subsequent
release of soils and water. Even when the soils require further treatment,
the reduction in volume can still economically justify slurry treatment.
3.5.2.2.2 Ex Situ Land Treatment Ex situ land treatment is sometimes
known as prepared-bed or on-site, land-based bioremediation treatment or
land farming. The process involves spreading wastes over the surface to
enhance natural microbial degradation of contaminants. Land treatment
was the first method used for bioremediation of soils and sludges and has
been successfully applied by the petroleum industry in the managed dis-
posal of petroleum refinery wastes for decades (Hildebrandt and Wilson
1990).
An ex situ land treatment unit operates under the same conditions as in
situ land treatment unit, with the possible addition of a leachate collection
system installed beneath the unit. The collected leachate may be returned to
the bed for additional treatment or treated by an on-site wastewater treat-
ment system and then returned to the unit (leachates having a high-content
of specific chemicals), or sent to an off-site wastewater treatment facility, if
available. An ex situ land treatment unit provides better control of any
leachate and, therefore, overcomes one of the potential disadvantages of an
in situ land treatment unit.
A lined land treatment facility may be required for treatment of certain
wastes, sludges, and contaminated soils and their constituents. The liner
prevents migration of contaminants to underlying soils and/or groundwater.
The subgrade construction of a lined facility is similar to the subgrade con-
struction of an unlined facility, except that either a clayey soil of a specified
compacted permeability or a synthetic (generally high-density polyethylene
(HOPE)) material is used. The liner, which prevents downward migration
3.141
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Process Identification and Description
of water, is placed across the subgrade and the sides of the berms. A
leachate collection system is used in combination with the liner system to
control surface drainage.
Leachate collection systems include a drainage bed, usually made of
sand, generally 30 cm (12 in.) thick and sloped to a sump for discharge
(figure 3.29). The sump is filled with cobbles (usually of 15.2 cm (6 in.)
diameter), with HOPE or PVC discharge pipes to either a retention pond or
an approved discharge facility; however, the leachate could be recycled
back to the land treatment unit, particularly if water is needed for moisture
control. Generally, at least 30 cm of sand or a permeable soil should be
placed on top of the leachate collection system. This sand or soil acts to
protect the leachate collection system. About 30 to 46 cm (12 to 18 in.) of
native topsoil is then placed on the sand and acts as the initial zone of incor-
poration. Once the topsoil is in place, the contaminated soil is applied and
incorporated in the same manner as that of an unlined facility.
Figure 3.29
Ex Situ Land Treatment
Air Emissions
Air Emissions
Control System
X
Leachate Control System
Leachate
3.142
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Chapter 3
At ex situ land treatment units, the sludge or contaminated soil is com-
monly applied to the existing soil or prior soil-waste mixtures in lifts. A
rule of thumb is to apply wastes, sludges, or contaminated soils on the ex
situ unit only as deep as available tilling equipment can incorporate it. An-
other application, or lift, is applied only after the previously-applied mate-
rial has been bioremediated to desired clean-up requirements. The unit
includes a spray irrigation system for application of nutrients and inoculum,
if desired, and for control of the soil moisture content. These items can also
be added by the same soil mixing equipment that is used for oxygen trans-
fer. A greenhouse top and an air-management system may also be included
if containment of volatile emissions is required. The contaminated soil is
tilled regularly (daily, every other day, or weekly depending upon oxygen
requirements) to promote homogenization of the soil and increase the oxy-
gen available to the indigenous microorganisms. An ex situ unit can be
used for many years.
The ex situ treatment site should be selected so as to minimize earth
moving and grading and the volume of potentially-contaminated runoff and
reduce treatment time to minimize the number and depth of lifts to be
treated. Other important considerations include proximity to irrigation
water, power sources, and an approved water discharge location, if needed.
The treatment unit usually is constructed adjacent to the site requiring
bioremediation to minimize transportation costs and to provide better tech-
nical and management control of the process. Except for the fact that the
on-site unit is constructed aboveground, the fundamentals and operation of
the ex situ unit are the same as those of the in situ soil unit.
Ex situ land treatment units have been used successfully to remediate (1)
contaminated soils at spill sites, (2) industrial wastes and residues, and (3)
soils at surface impoundments and lagoons that are being closed. It is com-
monly used today to treat soil contaminated with petroleum and wood-
preserving wastes. At a pilot-scale land treatment facility constructed at an
oil gasification site, successful bioremediation of approximately 4,600 m3
(6,000 yd3) of soil contaminated with coal tar was achieved. The contami-
nated soil was placed in a bed to a depth of 0.6 m (2 ft) and regularly tilled
and irrigated. Results achieved in a 4-month treatment period included a
73% reduction in BTX concentrations, a 36% reduction in the concentration
of oil and grease, and an 86% reduction in the concentration of total PAHs.
Two-ring and 3-ring PAH concentrations decreased by 92%; 4-ring PAH
3.143
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Process Identification and Description
concentrations decreased by 80%; and 5-ring PAH concentrations decreased
by 65% (Hutzler et al. 1989).
Bioremediation of soil contaminated with wood-preserving wastes was
successfully demonstrated in another pilot study. The contaminants moni-
tored were pentachlorophenol (PCP) and creosote, which is a complex mix-
ture containing a number of PAHs. During a 5-month period, the concen-
tration of PCP was reduced by 95%, while reductions in PAH concentra-
tions ranging from 50 to 75% were achieved (Linkenheil and Patnode
1987).
3.5.2.2.3 Soil-Pile Treatment There are two types of soil-pile reactors.
One delivers oxygen and nutrients by water movement through the soil
(water-based). In the other, the air-reactor design, the nutrients are mixed
in with the soil when the pile is created, and oxygen is delivered by air
movement through the soil.
The water-based soil-pile treatment is identical to ex situ land treatment,
except that the soil is not tilled. The contaminated soil is spread on a lined
treatment bed equipped with a drainage collection system. An irrigation
system is used to deliver a constant flow of a solution containing nutrients
and an inoculum, if desired. The collected irrigation stream drains to a
sump, from which it may be conveyed to a liquid-phase bioreactor. The
bioreactor effluent is channeled back through the irrigation system. The
soil-pile system may be totally enclosed if volatile emissions control is
necessary.
Like ex situ land treatment, the first soil-pile treatment was successfully
applied to soils contaminated with petroleum and wood-preserving wastes.
At one wood-preserving site, a water-based soil-pile reactor was con-
structed within an existing Resource Conservation and Recovery Act
(RCRA) impoundment area. Approximately 918 m3 (1200 yd3) of sludge
and contaminated soils were mixed with an equal volume of native soils and
spread in a 15 cm (6-in.) layer in the reactor. The soil-pile was irrigated
daily to maintain the desired moisture content within the bed.
The contaminants monitored were benzene-extractable (BE) hydrocar-
bons and 16 PAHs. Concentrations of BE hydrocarbons were reduced by
69% over the first year of operation, with most of the reduction occurring
during the first 4 months (i.e., May through September). Average removal
rates of 95% were achieved in the first year for the 2-ring and 3-ring PAHs.
3.144
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Chapter 3
The average total PAH reduction rate was 90%, while the rate for 4- and 5-
ring PAHs was 72%. As for the BE hydrocarbons, the greatest PAH reduc-
tions were achieved in the warm-weather months. These results were ex-
pected since the rate of biological degradation would be greatest during
warmer weather. Further contaminant reductions achieved over the winter
months were slight (Williams and Ziegenfuss 1989; Bourquin 1989;
Golueke and Diaz 1989).
The air reactor design is more versatile than the water-based design,
which is limited in size and oxygen transfer because of the reliance on wa-
ter movement. The air reactor can be larger and handle higher concentra-
tions.
The air soil-pile is constructed in a similar fashion to the water design
(figure 3.30). An impermeable layer is placed on the ground. Then the soil
(mixed with nutrients and inoculum, if required) is placed directly on the
liner. There is no need for a leachate collection system since water is not
added. Air pipes, placed in the soil as the pile is created, are used to deliver
and/or collect air. The spacing of the pipes is dependent upon the perme-
ability of the soil. The pipes are connected to the vacuum side of a blower.
The exhaust air can be treated if required. The pile can be any size, but is
Figure 3.30
Soil Pile Reactor
Mains & Laterals
Piped Back to Blower Top Ljner
Stormwater and
Leachate Collection
Trenches to Sump
3.145
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Process Identification and Description
usually limited to a maximum cross-section of 3.7 x.6 m (12 x 20 ft) height
and width. A plastic liner also is placed on top of the pile to keep rainwater
out and to aid the proper air movement through the pile.
3.5.2.2.4 Composting Introduction. Composting is similar to the process
used for composting of leaves, garbage, and food-processing residues. Bio-
degradation of the organic contaminants occurs within the compost matrix,
which consists of the contaminated material mixed with organic carbon
sources and bulking agents, such as straw, bark, or wood chips. Al the
completion of a bioremediation composting process, the treated material
must be disposed of in an environmentally-sound manner. Both degrada-
tion and immobilization, however, have occurred in this process, and the
composted material should not cause surface or groundwater problems at its
ultimate disposal site.
The essential elements of composting are the same as for any
bioremediation process:
• moisture;
• aeration;
• acclimated organisms;
• satisfactory carbon-nitrogen-phosphorus balance; and
• nontoxic conditions.
Following are the characteristics of contaminated soils or residues con-
sidered suitable for composting:
• constituents able to be eliminated by volatilization or degrada-
tion, or immobilized in the system;
• a low amount of free liquid so that an aerobic condition can be
maintained;
• a high ratio of inert solids to biodegradable organic compounds;
and
• a mixture that can be easily broken up by mechanical turning
and/or is porous so as to allow air to move through the
composting solids.
Typical composting systems that can be used for bioremediation are the
windrow, in-vessel, and Beltsville systems (table 3.22 on page 3.147). The
3.146
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Chapter 3
Table 3.22
Characteristics of Different Types of Composting Systems
TVpe
Windrow
Beltsville
In-vessel
Open/Closed
Open
Open
Open/Closed
Aeration
Turning pile
Distribution system
Distribution system
Examples of Wastes Treated
Waslewater sludge, food
processing wastes, manures
Sludge, wastewater sludge,
other organic wastes
Wastewater sludge, food
processing waste
Reprinted with permission of the Academic Press from 'Encyclopedia of Microbiology, Volume 1" by J.M. Thomas, C.H.
Ward. R.L. Raymond. J.J. Wilson, and R.C. Loehr. Copyright 1992 by Academic Press.
windrow system is an open system in which the compost is stacked in elon-
gated piles (windrows) and mechanically turned for aeration. In-vessel
composting entails placing the compost inside an open or closed reactor in
which aeration is achieved by mechanical mixing and/or blowers. The
Beltsville system is an open pile with an air distribution system under the
pile. Air is drawn through the pile from the atmosphere and exhausted
through a blower, generally, to an air pollution control system. Bulking
agents are commonly added to increase the porosity and assist the flow of
air to maintain aerobic conditions. For bioremediation composting systems,
adequate operational control, as well as control of all emissions such as
leachate and offgases, is important.
Operational Considerations. Composting is typically a batch biological
process used to treat material with high concentrations of biodegradable
organic compounds. Waste destruction and conversion are achieved by
thermophilic aerobic microorganisms that occur naturally in decaying or-
ganic matter. The basic objective of this process is to maximize microbial
activity, while decreasing waste volume, odors, and aqueous side streams.
The elements of effective composting are adequate nutrients, satisfactory
pH and temperature, nontoxic conditions, adequate oxygen, and satisfactory
moisture conditions.
For some hydrocarbon-contaminated soils, the addition of municipal
sludge or agricultural materials may be required to increase the temperature
3.147
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Process Identification and Description
of the pile in order to increase the metabolic rate of the microorganisms
and, thereby, increase rates of contaminant biodegradation.
Composting is a four-step process operated so as to achieve a high de-
gree of treatment or remediation. The first step entails mixing waste or
contaminated soil with a bulking agent to enhance aerobic conditions. In-
creasing porosity enhances oxygen transfer and decreases the moisture con-
tent of the mass. In the second step, air, and possibly heat, is introduced to
the system to effect the temperature conditions necessary for aerobic de-
composition of the organic compounds. The third step is curing in which
biodegradation is effected. The fourth step entails separation of the bulking
agent from the remediated material for reuse.
All composting processes require staging and treatment areas. These
areas are typically lined with concrete to permit ease in handling materials.
Area requirements are specific to the particular process, with the windrow
system requiring the greatest area for soil treatment.
Temperatures within the compost should range between ambient and
50°C (120°F). This range should be maintained to maximize
bioremediation, but the higher temperatures may not be possible with
wastes or contaminated soils containing low concentrations of biodegrad-
able organic compounds.
The windrow entails the deposition of wastes in 0.9 to 4.3 m (3 to 14 ft)
wide windrows to permit biological decomposition. Microbial activity is
facilitated by periodic turning of the windrows to aerate the mass and re-
lease excess heat, which can be detrimental to microbial growth. Fans may
be used to create an induced draft through the pile. Bioremediation may be
complete in approximately 6 to 8 weeks, depending on the type of organic
compounds to be remediated.
All windrows should consist of at least 40% solids to maximize windrow
stability and assure proper air space for aeration. Modifications to the mass
can be made by adding bulking agents. A leachate runoff colleclion and
treatment unit should be part of the windrow system, since relatively small
amounts of water are removed by other mechanisms.
In static-pile composting, air is drawn through a static pile when a
vacuum is generated by perforated pipes beneath the waste. Drawing air
through the pile enhances aeration and reduces odor and volatile organic
emissions.
3.148
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Chapter 3
A modification to this process involves static-pile aeration with pressure
ventilation and controlled heat removal. Heat output, temperature, ventila-
tion, and water removal can be monitored for process optimization. An air
compressor draws a vacuum at the bottom of the pile and returns the air at
the top, providing aeration. Makeup air is added, as necessary, as oxygen is
depleted. Nutrients and water can be provided through an irrigation system
located at the top of the pile. Air addition should be adequate to maintain
internal oxygen levels in the 5 to 15% range.
Conventional windrow composting requires a hard, dry surface (usually
concrete) and large tractors to turn the piles. Static and aerated piles in-
clude perforated pipes installed beneath the piles with blowers used to cre-
ate a vacuum and offgas treatment. Conventional windrows and static piles
may have a roof placed over the piles to reduce precipitation infiltration.
Materials that are amenable to the composting bioremediation process
include sewage sludge, soils contaminated with diesel fuel and similar pe-
troleum products, and wastes and residues from brewing, antibiotic, fermen-
tation, food processing, mineral oil, and munitions operations. Bulking
agents may be added to increase the porosity and facilitate aeration. Mate-
rials used as bulking agents include fibrous plant material, wood chips, and
bark. For compounds that are difficult to biodegrade, such as those found
in munitions residue, the waste can be mixed with highly biodegradable
material. This material serves as a carbon source for the microorganisms
while the compounds are biodegraded through cometabolism. Materials
used for this purpose should be inexpensive; examples include food pro-
cessing wastes, manure, and plant material.
Two field-scale studies were performed to investigate the feasibility of
using composting to treat explosives and propellant-contaminated sedi-
ments at two Army ammunition plants. Contaminants of concern at one site
included 2,4,6-trinitrotoluene (TNT), hexahydro-l,3,5-trinitro-l,3,5-triazine
(RDX), and octahydro-l,3,5,7-tetranitro-l,3,5,7-tetraazocine (HMX). The
contaminant of concern at the other site was nitrocellulose (NC) (Williams
and Ziegenfuss 1989). Four compost piles of approximately 9 m3 (12 yd3)
each were constructed at each site. The piles were placed on 8-inch thick
concrete pads which drained to a sump. The compost mixture consisted of
cow or horse manure, straw, alfalfa, horse feed, and contaminated sediment
and was moistened with water when it was placed on the concrete pads.
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Process Identification and Description
Each compost pile contained a network of perforated pipes through which
air was drawn to aerate the piles.
The compost period lasted 22 weeks at the first site. During that time the
following results were achieved: reductions in TNT concentrations ranging
from 99.6 to 99.9%; reductions in RDX concentrations ranging from 94.8 to
99.1%; and reductions in HMX concentrations ranging from 86.9 to 96.5%.
At the second site, reductions in NC concentrations ranging from 91 to 98%
were achieved during a 14-week composting period (Williams and
Ziegenfuss 1989). Although composting was effective in reducing the con-
centrations of these compounds, by-products more toxic than the parent
materials may be produced.
3.5.2.2.5 Cost Comparison of Ex Situ Bioremedlation Technologies for
Contaminated Soils Estimated per-ton costs for the four ex situ
bioremediation technologies are presented in table 3.23. All unit costs are
based on the assumption that the waste treated is contaminated soil having a
unit weight of 1,762 kg/m3 (110 lbs/ft3). Ex situ land treatment and
composting are attractive because of their low costs, which are in the ranges
of $39 to $88/tonne ($35 to 80/ton) and $44 to $ 110/tonne ($40 to 100/
ton), respectively (Lynch and Genes 1988; Environmental Protection
Agency 1988; Williams and Ziegenfuss 1989). Soil-pile treatment costs are
in the $99 to $110/tonne ($90 to $100/ton) range, while slurry-phase treat-
ment costs range from $88 to $165/tonne ($80 to $150/ton) (Bourquin
Table 3.23
Treatment Reactor Costs for Solid Phase Biological Methods
Process Cost
Off-site Disposal in Permitted Hazardous Waste Landfill $200 to $300/ton plus transportation costs
Off-site Incineration in Permitted Facility $300 to $1200/ton plus transportation costs
Engineered Land-Farm Treatment $3 5 to $ 100/ton
Soil Pile Treatment $50to$100/ton
Composting $50 to $70/ton
Bioslurry Reactor Treatment $80 to $ 150/ton
3.150
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Chapter 3
1989; Golueke and Diaz 1989). All of the ex situ bioremediation technol-
ogy costs compared favorably with the higher costs of off-site landfill dis-
posal ($220 to $330/tonne ($200 to $300/ton)) and off-site incineration
($330 to $2,200/tonne ($300 to $2,000/ton)). The landfill disposal and
incineration costs shown do not include transportation costs, which are
substantial. For example, transportation costs for a 9,072 kg (20,000 Ib)
truckload of contaminated soil hauled by a qualified hazardous materials
transporter range from $2.50 to $3.50/loaded mile.
One of the major costs of all solid-phase reactor operations is soil move-
ment. The best way to save money on a remediation project is to incorpo-
rate the biological reactor into a project that is already moving the soil. For
example, when buried drums are removed from soil, the biological system
can be placed as each section is handled. The biosystem can then treat the
contaminated soils at essentially no additional cost for soil movement.
3.5.2.3 Biological Reactors for Contaminated Air
3.5.2.3.1 Introduction. Biofiltration is the biological removal of organic
contaminants in gas streams in a solid-phase reactor (figure 3.31). This
process is well established in Europe and Japan where it has been used as an
air pollution control technology in successfully controlling odors, volatile
Figure 3.31
Ex Situ Bioremediation of Volatile Organic Compounds
Soil, Slurry and
Wastewater Treatment
tt
Off gases
L
Bioventing
Air Sparging •
3.151
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Process Identification and Description
organic compounds (VOCs), and air-toxic compounds. Biofiltration has
economic and other advantages over existing air pollution control methods,
particularly if applied to offgas streams that contain only low concentrations
(typically less than 1,000 ppm as methane) of air pollutants that are easily
biodegraded (Leson and Winer 1991).
A biofilter also has been known as a "bodenfilter". A biofilter consists
of one or more beds, typically 1 m (3.3 ft) in height, of biologically-active
material, primarily mixtures based on compost, peat, or soil. Contaminated
offgas is vented through the filter. The air contaminants diffuse into the
wet, biologically-active biofilm that surrounds the filter particles. Aerobic
degradation of the pollutants occurs in the biofilm if microorganisms are
present that can metabolize them. End products from the complete biodeg-
radation of air contaminants are CO2, water, and microbial biomass. The
oxidation of reduced sulfur compounds and chlorinated organic compounds
also generates inorganic acids.
Compost produced from municipal waste, wood chips, bark, or leaves
has been the basis of filter material used in recent applications in Europe,
although peat and heather mixtures have also been used. Biofilters built in
the U. S. have been mostly "soil beds" in which biologically-active mineral
soils were used as filter materials.
Most biofilters have been constructed as open, single-bed systems, al-
though enclosed, multi-story units have been considered.
In the U. S., the first systematic research on the biofiltration of H,S was
conducted in the early 1960s (Carlson and Leiser 1966). This work in-
cluded the successful installation of several soil filters at a wastewater treat-
ment plant near Seattle and demonstrated that biodegradation, rather than
sorption, accounted for the odor removal. Other successful soil bed appli-
cations in the U. S. include the control of odors from rendering plants
(Prokop and Bohn 1985) and the destruction of propane and butane released
from an aerosol can filling operation (Kampbell et al. 1987).
Since the early 1980s, biofiltration has increasingly been used in Ger-
many to control VOCs and air toxics emitted from industrial facilities such
as chemical plants, foundries, print shops, and coating operations. This
development was brought about primarily by new federal regulations
(Leson and Winer 1991).
3.152
-------
Chapter 3
Biofilters have been shown to be suitable for many uses. They have
been employed mostly to remove odors from waste gases in chemical and
pharmaceutical manufacturing and food processing, as well as in storage
tank vents and fuel- and solvent-handling systems (Bohn and Bohn 1988).
These biofilters can remove and safely dispose of approximately 99% of
slightly volatile and easily-biodegradable organic compounds, such as alde-
hydes, SO2, NOx, and H2S, and about 90% of volatile gases such as meth-
ane, propane, and carbon monoxide. In addition, they remove liquids and
solid particulates from the gas streams.
Biofilters differ from other air pollution control methods in several ways:
• contaminants are both adsorbed and oxidized;
• suitable reaction times allow complete treatment of the organic
compounds; and
• both organic and inorganic gaseous pollutants are removed.
The factors limiting biofilter treatment are the biodegradability of the
organic contaminants and the permeability and chemistry of media. Be-
cause these factors vary widely, the design of biofilters is specific to the
particular site.
Because microbial degradation of odorous compounds and VOCs is the
primary removal mechanism within a biofilter, the filter material also must
provide the proper environment for microbial growth. Microorganisms
require both aerobic conditions and adequate moisture. The filter should
also contain materials on which the microbes can feed to ensure that the
microbial population can survive if a shutdown of the entire system should
occur for any length of time.
Biofilters are simple to operate and require little maintenance. The only
maintenance required is merely periodic monitoring of the moisture content
and pH of the filter material. The microorganisms most effective for odor
removal are sensitive to acid conditions. The pH of the filter should be
measured frequently at several locations and depths. If the pH drops below
the recommended level, a base, such as lime, should be added to restore
neutral pH conditions.
3.5.2.3.2 Operational Considerations. Humidity control in the biofilter is
important. Too little moisture results in dry zones and loss of microbial
activity. Too much moisture results in the development of anaerobic zones
3.153
-------
Process Identification and Description
and, consequently, poor effluent quality and odor production. A moisture
content between 40 to 60% by weight is considered optimal (Leson and
Winer 1991). The type of distribution network is determined by two crite-
ria: the need to distribute the gas uniformly and the power required. Three
types of systems can be used:
• perforated pipe;
• pressure chamber systems; and
• cinder block systems.
In perforated pipe systems, the bed is underlaid with a network of perfo-
rated pipes set in a gravel bed. In pressure chamber systems, a large pres-
sure chamber at the bottom of the bed supplies and distributes air. In the
cinder block system, air distribution is accomplished through prefabricated,
slotted, concrete blocks. The blocks provide both aeration and drainage
systems.
The following are ideal properties of biofilter media:
• high adsorption capacity;
• low pressure drop;
• high nutrient content;
• pH buffering capacity;
• adequate moisture content; and
• temperature between 25°to 35°C (77° to 95°F).
The filter media provide sites for adsorption of pollutants and for attach-
ment of microorganisms. The media also are a source of additional nutri-
ents so that microbial activity is not limited by nutrient availability.
Other operating parameters of concern are residence time and loading
rates. The removal efficiency is highly dependent on the residence time,
which represents the average time that a gas molecule spends inside the
filter bed. Recommended residence times for odor control vary between 5
and 30 seconds. Gas-loading rates for processes utilizing compost filter
beds are typically 1.5 to 3 mVmVmin (5 to 10 ft3/ft2/min) (Leson and Winer
1991).
For any given offgas, the biofilter size required for the desired rate of
removal of specific contaminants depends on the hourly contaminant load
3.154
-------
Chapter 3
(in grams per hour) as compared to the degradation capacity of the filter
material for a specific constituent — usually given in terms of grams per
hour per cubic meter of biofilter material. Degradation rates for common
air pollutants can vary widely and depend predominantly on the type of
pollutant and the biological and physical characteristics of the filter mate-
rial. Typical rates for easily-degradable VOCs, such as alcohols, ketones
and many aliphatic and aromatic hydrocarbons, range between 50 and 100
g/m3 per hour(3 x 103 to 6 x 10'3 lb/ft3 per hour). Higher chlorinated or-
ganic compounds have lower rates of degradation, with the degradation rate
decreasing with increasing chlorine substitution (Leson and Winer 1991).
Gas-flow rates affect the filter size required to provide the degradation
capacity for a given pollutant load. Since offgases from industrial sources
often contain a variety of organics, and since degradation rates will depend
on the offgas concentration of the target pollutant, pilot testing on a partial
offgas stream from a source is usually conducted to determine the required
size and operational conditions for a full-scale system.
3.5.2.3.3 Costs. Installation costs for soil beds range from $5-6 per mVh,
and depend on the piping required to distribute the gases throughout the soil
(Bohn and Bohn, 1988). The only operating cost is from the power (0.4W
per mVh) required to overcome the 2-3 in. water gauge backpressure.
3.155
-------
Chapter 4
4
POTENTIAL APPLICATIONS
See also Appendix A for case studies of bioremediation and Subsection
3.2.1, Microbial Ecology and Physiology, which discusses the applicability
of bioremediation to classes of compounds.
4.1 General Criteria
The usefulness of bioremediation will depend on the limits of contami-
nant biodegradability and the physical constraints to creating the necessary
conditions to achieve an acceptable rate and extent of contaminant biodeg-
radation. Table 4.1 (on page 4.2) summarizes the potential biodegradability
of some common, naturally-occurring and xenobiotic contaminants.
Bioremediation can be used to treat sludges and water from impound-
ments, sediments, and groundwater within the saturated zone, soils within
the unsaturated zone, excavated soils, recovered groundwater, and gases
extracted from in situ soils or aboveground reactors. Remediation of a spe-
cific site can entail one bioremediation process, several in combination, or
one or more in conjunction with other treatment and remedial processes.
For instance, a number of sites have been treated using a groundwater recir-
culation system to treat the saturated soils and groundwater, a bioventing
system to treat the unsaturated soils, and an aboveground soil reactor to
treat excavated soils. Such systems could have easily incorporated a liquid
phase bioreactor to treat the recovered groundwater as well.
The applicability of the in situ and ex situ bioremediation processes, as a
function of site characteristics, type of contamination, and performance, is
summarized in table 4.2 (on page 4.4). Some processes are marked with
two abbreviations to indicate that specific site conditions will result in a
range of effectiveness for that type of process. Primary processes are the
4.1
-------
Potential Applications
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4.2
-------
Chapter 4
main technologies used to treat the contamination. When necessary, sec-
ondary processes are used to treat the emissions and/or effluents generated
by the primary processes.
Some or all of the soils at some sites may not be amenable to any ex situ
technologies because excavation is precluded by the location of buildings or
other structures, the depth of the contamination, or location of contaminated
soils below the water table. For this type of site, in situ bioremediation may
be useful.
4.2 In Situ Bioremediation
Parameters that adversely affect performance of all in situ processes
include soil and sediment hydraulic conductivities (K) that are less than 10~4
cm/sec (3.28 x 10"6 ft/sec) and high organic matter content (see table 4.2 on
page 4.4). In matrices with low values of K, it is more difficult to deliver
oxygen and nutrients rapidly and pore spaces are plugged more readily than
in those that are more transmissive. In addition, contaminants may sorb to
organic matter and become less available for biodegradation and physical
removal. Other parameters are more specific to the particular process. The
effectiveness of a bioremediation process at a particular site will depend on
other parameters that affect site operation and process performance, such as
site and contaminant characteristics, the remediation goals, and regulatory
requirements. In general, many bioremedial processes can be considered as
treatment options when the contaminants are low in concentration and
readily biodegradable, and when the affected matrices have K values of 10"4
cm/sec (3.28 x 10"6 ft/sec) or greater. Conversely, sites with high concen-
trations of relatively difficult to degrade contaminants and less transmissive
sediments will have limited or no bioremediation options.
Although liquid delivery was the first in situ bioremedial approach de-
veloped for treating subsurface contamination, air sparging probably re-
places this method at most sites where control of plume migration is not
required. Air sparging is less expensive, distributes oxygen across the en-
tire site faster, and presents fewer operational problems than the liquid de-
livery method. But, liquid delivery will still be used at sites where air
sparging is not applicable, where a pump-and-treat system already exists
4.3
-------
Potential Applications
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4.4
-------
Chapter 4
and liquid delivery is an easy addition and where control of plume migra-
tion is required. Sites at which air sparging would not be applicable include
fractured rock aquifers, aquifers with shallow water tables (unless the sur-
face is capped), and formations with narrow saturated intervals.
Unsaturated soils contaminated with biodegradable substrates generally
can be treated by bioventing whether or not the contaminants are volatile.
Bioventing is most effective where the depth-to-water exceeds ten feet and
the surficial soils (upper 0.6 m (2 ft)) do not require treatment or are being
treated by other methods such as land treatment. If the surface of the con-
taminated site is capped, bioventing can be used to treat shallower soils and
sites with shallower water tables. When the depth-to-water is less than ten
feet and the surface is not capped, well spacing will have to be quite close
and thus capital expenditures great. Bioventing systems can also remove
nonbiodegradable contaminants and those that are more difficult to degrade
such as chlorinated solvents, through physical removal, provided that
offgases are appropriately treated.
Bioventing is clearly more appropriate than ex situ methods, including
land treatment, soil-pile treatment, slurry bioreactors, low-temperature ther-
mal desorption, or incineration where excavation is not feasible. Where
excavation is feasible, aboveground methods will be more useful than
bioventing if the contaminants are difficult to treat. For moderately degrad-
able compounds, particularly at sites containing large volumes of contami-
nated clayey soils, low-temperature thermal desorption may be a better
remedial option than bioventing, providing excavation is feasible.
Saturated soils and groundwater can be treated in situ using variations of
the liquid delivery method or sparging if the site is sufficiently transmis-
sive. Hydraulic conductivities of 10"4 cm/sec (3.28 x 10"6 ft/sec)or greater
are generally acceptable for these processes, although sites with lower K
values may be treated if the contaminant load is light. Where applicable, in
situ bioremediation will probably be the method of choice for treating aqui-
fers contaminated with biodegradable compounds. The only other com-
monly applied technology is pump-and-treat which generally requires long
times to meet remediation goals, except for sites that contain very soluble
compounds. If timing is a concern in treating an aquifer contaminated with
volatile contaminants, air sparging to transfer the volatile compounds to the
unsaturated zone followed by capture with an air recovery system may be
more efficient than the Raymond process; however, there will be additional
costs associated with the offgas treatment.
4.5
-------
Potential Applications
In general, the effectiveness of in situ methods is more dependent than ex
situ methods on contaminant biodegradability, contaminant concentration,
and subsurface conditions. Although the transfer of nutrients and an elec-
tron acceptor to contaminant-degrading microorganisms can be easily
achieved in ex situ processes, in situ methods require the movement of air
or water through undisturbed soils to deliver the electron acceptor and/or
nutrients.
4.3 Ex Situ Bioremediation
Excavated soils and sludges can be bioremediated with ex silu systems.
Soils are most commonly treated with land treatment or soil piles. The
method of choice depends on the available space for treatment, the charac-
teristics of the contaminated soils, and the biodegradability of the contami-
nants. More intrinsically recalcitrant contaminants may be more effectively
remediated using slurry reactors. For moderately degradable compounds,
particularly for large volumes of clayey soils, low-temperature thermal
desorption may be more effective than soil piles. In some instances, sani-
tary landfill tipping fees may be low enough to make disposal of nonhazard-
ous soils the least costly alternative.
Slurry reactors, particularly when combined with soil washing can be
used to treat a wider range of soils and contaminants than most other
bioreactors. The addition and transfer of nutrients and an electron acceptor
to contaminant-degrading microorganisms is more easily controlled in
slurry reactors than in in situ processes.
Groundwater recovered from pump-and-treat systems, including aquifer
bioremediation systems, and discharge water from bioreactors can be
treated in liquid reactors. Nonbiodegradable constituents in the water to be
treated can be removed in a polishing step using activated carbon or chemi-
cal treatment. Liquid-phase bioreactors are particularly effective for biode-
gradable organic contaminants that are also soluble and nonvolatile; such
compounds are poorly removed by either air strippers or activated carbon.
4.6
-------
Chapter 4
4.4 Biological Reactors for Contaminated
Air
Air bioreactors can be used to treat the offgas emissions from bioventing
systems, in situ vapor recovery systems, soil reactors, and air-stripper tow-
ers. Air bioreactors are most applicable when the concentration of volatiles
in the air phase is moderate-to-low. For systems designed for the rapid
physical removal of volatile contaminants, air bioreactors might be used
after the rate of physical removal has diminished.
4.7
-------
Chapter 5
5
PROCESS EVALUATION
See figure 5.1 (on page 5.2) for a schematic summary of an aerobic bio-
chemical process. In this example, benzene (C6H6) is used by the bacteria
as a sole source of carbon and energy for growth. To complete the reaction,
the bacteria require a terminal electron acceptor (in this case oxygen) and
major nutrients (nitrogen and phosphorus), as well as other minor nutrients
commonly found in soil. In addition, the environment must be conducive to
the growth of the benzene-degrading organisms (suitable pH and tempera-
ture). As a result of the benzene biodegradation, by-products (CO2, H2O)
and new bacteria will be produced.
5.1 Basic Considerations
Proper bioremediation design involves creating the correct environment
for the bacteria, and supplying the bacteria with the material that is limiting
their rate of reaction (usually O2 and nutrients). Figure 5.1 (on page 5.2)
represents an aerobic design, but other biochemical reactions can be repre-
sented by a simple switch of the electron acceptor. The same basic require-
ments apply to any biological process, whether the reactor is aboveground,
in situ, or consists of water or soils that are being treated.
The key to the design of an ex situ liquid-phase reactor is to contain the
bacteria in the reaction zone. In solid-phase reactors, the delivery and ex-
pense of supplying oxygen and nutrients and maintaining the correct envi-
ronment are the keys to design. For in situ remediation designs, the key is
delivery of oxygen and sometimes nutrients to the bacteria in the zone of
contamination and, to a lesser extent, the expense of the oxygen or electron
acceptor that must be supplied. The design concerns for all of these pro-
cesses center around the practical problems in maintaining the correct envi-
ronment and delivering the needed material to the bacteria.
5.1
-------
Process Evaluation
FigureS.l
Biochemical Process
CO, H2O
Neutral pH D.O>1.0mg/L
BACTERIUM
Adequate
mperature Water
Bacteria
N, P
The key consideration in biological treatment is that the process is a
living system that is run 24 hours a day, 7 days a week. The system cannot
be turned on and off. Furthermore, biochemical processes must have a
start-up period. A certain amount of time is required for the bacteria to
adapt to degrade the contaminants. The last key consideration is that living
systems do not respond well to severe changes in environmental conditions.
Biochemical processes operate best within optimum ranges of pH and tem-
perature.
5.2 Process By-Products
The by-products that result from a biochemical process depend on the
original organic contaminants that are being degraded, the environment in
which the bacteria are grown, and the terminal electron acceptor used by the
bacteria. Most remediation designs strive to degrade the organic material to
CO2 and H2O. New bacteria will also be produced during degradation, and
the new bacteria can be considered as a by-product of the reaction, When
oxygen is the terminal electron acceptor and the organic compound is de-
5.2
-------
Chapter 5
gradable, then the only by-products are CO2, H2O, bacteria, and depending
on the substrate, inorganic substituents (e.g. Cl' and NO2~). But, under
anaerobic conditions, other by-products can be produced. Probably the best
known of these by-products comes from anaerobic degradation of chlori-
nated aliphatic solvents (e.g., tetrachloroethylene and trichloroethylene).
Biodegradation of these compounds under anaerobic conditions in the field
tends to produce lower substituted chlorinated hydrocarbons, and, as a final
by-product, either chloroethane or vinyl chloride. It is important to remem-
ber that these reactions in the field are natural under certain environmental
conditions. But such end products would be considered unacceptable in the
design of a bioremediation process which would result in their formation.
5.3 Cleanup Levels Achievable and Du-
ration of Treatment
Contaminant concentration often declines to a certain level during
bioremediation, after which little or no biodegradation occurs. A combina-
tion of contaminant and site-specific characteristics is responsible for this
asymptotic removal. Contaminant-specific characteristics that affect bio-
degradation include: 1) the biodegradability of the contaminant, 2) the pres-
ence of microorganisms adapted to degrade the contaminant, 3) whether the
contaminant serves as a growth source for the microorganisms, and 4) the
concentration of the contaminant.
Many environmental contaminants are biodegradable, at least in labora-
tory studies. In instances in which a biodegradable compound does not
biodegrade in the field, some environmental factor is limiting or the micro-
organisms are not adapted to degrade the contaminant.
In cases in which the contaminant does not serve as a growth substrate, a
primary substrate must be present for the microorganisms to grow on while
the contaminant (secondary substrate) is fortuitously degraded (see Section
3.2.1.3, Microbial Metabolism of Contaminants that are not Growth Sub-
strates). The rate and extent of contaminant biodegradation will be affected
by the concentration of the primary substrate. For some contaminants, such
as trichloroethylene, a high concentration of the primary substrate may
decrease contaminant biodegradation because both primary and secondary
5.3
-------
Process Evaluation
substrates compete for the same enzyme. For other compounds, high con-
centrations of a primary substrate may be necessary for microorganisms to
grow so that sufficient quantities of enzymes are available to effect desired
reactions.
The concentration of contaminant is also important. The greater the
mass of contamination, the greater the time required for cleanup. In addi-
tion, high concentrations of some compounds, such as benzene and trichlo-
roethylene, may be toxic to microorganisms and inhibit bioremediation
altogether.
Thresholds below which biodegradation of a contaminant does not oc-
cur, although it can serve as a growth source, probably exist in the field
(Alexander, 1985). Essentially, the substrate may be present at a concentra-
tion that can not support growth or induce the expression of enzymes neces-
sary to effect its degradation. Thresholds are probably organism-specific,
with high nutrient requiring (eutrophic) microorganisms having higher
thresholds than low nutrient requiring microorganisms (oligotrophs). In
addition, the presence of other growth substrates may decrease the threshold
concentration for contaminant biodegradation.
Site-specific characteristics also affect contaminant biodegradation. For
both in situ and ex situ treatment, sandy materials with low organic matter
content are more amenable to treatment than materials with high organic
matter and/or clay content. Organic contaminants may sorb to clay and
organic matter and are then rendered unavailable for biodegradation (see
Section 3.2.1.2.5, Contaminant Bioavailability; Section 3.3.3 Design Con-
siderations for In Situ Bioremediation of Aquifers). In the presence of high
concentrations of organic matter or clay, biodegradation may be 1 imited by
the rate of contaminant desorption.
Contaminant biodegradation may also be affected by the permeability of
the formation. Injection of electron acceptors and nutrients will be more
difficult in low versus high permeability materials, thus increasing the time
required for treatment and possibly decreasing the clean-up level achiev-
able.
In instances in which the results of bioremedial efforts do not meet regu-
latory standards, other remedial measures may be required. For in situ pro-
cesses, groundwater may be extracted and treated at the surface with a pol-
ishing step such as activated carbon or air stripping, if permitted. For in
5.4
-------
Chapter 5
situ and ex situ processes, surfactants may be added to render sorbed con-
taminants more bioavailable. Although adding surfactants to ex situ pro-
cesses is easily accomplished, transport of surfactants through the subsur-
face is more difficult. Surfactants added to the subsurface must be nontoxic
and biodegradable.
5.4 Cost
The cost of bioremediation varies depending on the type and quantity of
organic compounds present, site conditions, and the total volume of mate-
rial to be processed.
Biological processes can be applied directly to the aquifer and vadose
zone, or in reactors for aboveground systems. The main costs result from
movement of the liquid or soils to the reaction zone, oxygen supply in aero-
bic systems, and nutrient supply. There are too many variables in all of the
designs and in operational needs to give accurate ranges of cost. Each situ-
ation is unique and two designers looking at the same situation may develop
completely different methods for solving the practical problems of design-
ing and implementing the biological process.
Nevertheless, the following are generally true:
• biological treatment of groundwater or liquids from ponds is
more expensive than air stripping, but less expensive than air
stripping followed by air treatment or direct carbon adsorption
of the contaminants. But, this generalization does not apply
when very low concentrations in water (<1 mg/L organic con-
tent) are treated;
• solid-phase reactor costs relate more to the cost of materials
handling and, as a second consideration, the cost of mixing sup-
plying oxygen;
• biological treatment of soils costs less than incineration, but
about the same as thermal desorption or landfilling; and
• in situ bioremediation is usually more expensive than a soil
vacuum extraction system because it runs longer (usually used in
5.5
-------
Process Evaluation
conjunction with soil vacuum extraction instead of as a substi-
tute), and less than a pump-and-treat system.
The amount of soil moved is the key to comparing the relative costs of
ex situ soil bioremediation technologies. The major cost of in situ treatment
is that associated with the delivery of oxygen and nutrients to the bacteria,
and it is, therefore, less expensive than an aboveground soil reactor. Be-
cause monitoring can be a major expense for in situ designs, efforts should
be made to minimize the required monitoring.
5.6
-------
Chapter 6
6
LIMITATIONS
The key factor which controls bioremedial processes is the biodegrad-
ability of the organic waste. Provided that the waste is biodegradable, the
success of bioremediation will depend on a system that is designed to maxi-
mize the rate and extent of contaminant biodegradation. Such a system is
designed to deliver adequate concentrations of limiting nutrients and an
electron acceptor to the contaminant-degrading microorganisms. The limi-
tations of bioremediation are related to this delivery.
Application of bioremediation systems, like many other remediation
techniques, involves some risk because the system has to operate for the full
cycle time for a prognosis. Laboratory tests generally have not been accu-
rate predictors of field remediation rates or the extent of degradation under
field conditions.
6.1 General Limitations
All bioremediation processes are limited to biodegradable contaminants
or mixtures of biodegradable and nonbiodegradable contaminants where
bioremediation is combined with another technology. In general, soil
bioremediation processes are more difficult to apply to clayey and other
low-permeability soils. Sites containing biodegradable compounds might
not be suitable for bioremediation if the contamination levels are high and
the clean-up targets are low. This is particularly true with a contaminant or
mixture that is at least moderately recalcitrant. Bioremediation can also be
precluded when the contaminant is unavailable to the bacteria, such as when
polycyclic aromatic hydrocarbons are strongly adsorbed to the soils or when
oils are highly weathered or adsorbed by asphalt, ashes, or other material in
the soil.
6.1
-------
Limitations
Limitations of each bioremediation process were indicated in the last row
of table 4.2 (on page 4.4). As indicated, reliance on natural assimilative
capacity is limited by oxygen and possibly, nutrient availability. These
limitations are most severe when the contamination levels are high and
located deep below the surface, and when the permeability is low. The
practicality of natural bioremediation may also be limited by the cost of
demonstrating its effectiveness to obtain approval and meet potentially
stringent monitoring requirements by the controlling regulatory agency.
Land disposal restrictions (LDRs) require that RCRA wastes be pre-
treated if they are to be disposed in a landfill. EPA requires that wastes be
treated to a specific treatment standard based on the best demonstrated tech-
nology (BDAT), or by a specified technology. Bioremediation, along with
other technologies, has been designated as a technology applicable to treat-
ment of wastes prior to placement in a landfill. These rules are expected to
increase the use of bioremediation as well as other specified technologies.
6.2 Land Treatment Processes
The use of tillage methods is not appropriate for treating soils contami-
nated with volatile constituents that create air emissions of concern or
where the soils cannot be accessed or excavated. The use of tillage methods
to treat excavated soils may also be rendered impractical by the lack of
sufficient appropriate open space to apply the soils.
6.3 Bioventing
Bioventing is not applicable to surficial soils (less than 0.6 m (2 ft) from
the surface) or where the water table is less than approximately 3 m (10 ft)
unless the soils are capped. Clayey materials with low permeabilities usu-
ally are not amenable to bioventing; however, success has been achieved in
some clayey soils. The limitations of this process in different soil types are
not fully understood. Bioventing of nonvolatile, degradable compounds
may be impractical at sites where nutrient addition is needed.
6.2
-------
Chapter 6
6.4 Air Sparging
The use of air sparging for bioremediation is relatively new and there is
much to learn about its limitations. In most cases, air sparging requires
vapor and groundwater recovery systems or appropriate controls and pilot
testing to prevent loss of the contaminants from the treatment zone. In
heterogeneous soils with clay lenses, gravel stringers, etc., channeling of the
air bubbles through the permeable layers may direct the oxygen away from
the contaminated zone. The area of influence of air sparging systems is
related to the saturated interval or the depth below the water table at which
the air is introduced. When these distances are short, the area of influence
of the sparging system may be too small to be practical.
6.5 Liquid Delivery Processes
Bioremediation of aquifers using the liquid delivery process is most
commonly limited by the rate of groundwater recirculation relative to the
contamination level. These systems can also fail if soil pores become
clogged by precipitation of iron as result of the addition of the oxygen
source in high iron aquifers, by precipitation of nutrients that are incompat-
ible with the groundwater chemistry (hard waters), or through formation of
excess biomass.
6.6 Soil-Pile Treatment
Soil piles may not be applicable to clayey soils unless bulking agents are
added; however, the use of slow-release oxygen compounds, such as cal-
cium or magnesium hydrogen peroxide complexes, may overcome soil
permeability limitations in some cases. This technology also requires that
the contaminated soils are accessible and that sufficient land is available to
treat the soils.
6.3
-------
Limitations
6.7 Slurry Reactors
Slurry reactors require control of off-gases when volatile constituents are
present above regulated levels. This is difficult to accomplish when treating
a lagoon or other large system. While biodegradation rates can be quite
rapid in bioreactors, throughput is relatively small unless a lagoon or pond
is used. Each type of liquid bioreactor has its own limitations as shown in
table 4.2 (on page 4.4). In general, the selected system is likely to be lim-
ited by either the concentration of organic contaminants in the influent or
the requirements for constant supervision.
6.8 Air Bioreactors
Air bioreactors, which are used to reduce emissions from primary treat-
ment, can be limited by the size of the reactor required to treat the air flow
and the mass of contaminants in the vapor phase. In some cases, the re-
quired capacity of the system can be minimized by using an alternative
method, such as a catalytic converter, during the first few weeks of opera-
tion when the concentration of the recovered volatiles is highest. Alterna-
tively, the primary system can be made fully operational over several
weeks, although this practice would sacrifice the efficiency of the primary
system and its ability to control off-gas treatment costs.
6.4
-------
Chapter 7
TECHNOLOGY PROGNOSIS
7.1 Application of Bioremediation
Underground storage tank releases and other sources of environmental
contamination are still being found at a greater rate than sites are being
remediated. Increased environmental enforcement and the large number of
contractors and consultants promoting bioremediation are an indication that
bioremediation will be applied at more and more sites over the next several
years. Recent changes in regulations appear to remove some restrictions on
excavating, bioremediating, and reusing soils on site. However, other envi-
ronmental regulations may impede implementation of bioremediation pro-
cesses by restricting the use of nutrients during in situ treatment and by
stipulating clean-up standards that are overly conservative. Other factors
affecting the use of bioremediation include competition by other
remediation techniques, problems in bioremediating relatively recalcitrant
chemicals, and application where difficult geological conditions prevail.
Over the next several years, commercial bioremediation will continue to
focus on aerobic processes and will continue to be used widely to treat eas-
ily-biodegradable contaminants, such as low to moderate weight petroleum
hydrocarbons and many oxygenated hydrocarbons, in relatively permeable
soils. In situ bioremediation will continue to be attractive where the site
conditions restrict excavation. For excavated soils, disposal in landfills,
asphalt production, and low-temperature thermal methods will continue to
compete with on-site ex situ bioremediation. However, there soon will be
several soil recycling centers that use permanent soil-pile or land treatment
units to remediate soils for reuse or for clean fill at municipal landfills.
7.1
-------
Technology Prognosis
7.2 Process Improvements
Improvements in the engineering aspects of bioremediation, particularly
those relating to the delivery of nutrients and/or electron acceptors, will
expand the conditions under which bioremediation can be cost-effective and
compete favorably with alternatives. Since the late 1980s, bioventing and
air sparging techniques have found increasing use as a result of improved
techniques for using these methods. The U.S. Air Force has committed to
using bioventing at a large number of facilities. To date, air sparging has
been implemented mainly for remediation of source areas. Air sparging-
induced biodegradation will also be used in place of physical barriers and
pump-and-treat systems to prevent migration of biodegradable compounds
and as a means to introduce cometabolites for treating chlorinated solvents.
The large number of private companies and government agencies using
bioremediation and conducting pilot studies indicates that there will be
continued engineering improvements. The involvement of industry and
agencies, such as the U.S. Environmental Protection Agency, in promoting
the use of innovative technologies through testing, development, training
seminars, and symposia, will promote the rapid spread of improved
bioremediation processes.
Improved processes include treatment of contaminated vapors using in-
place soils, as well as porous bed ex situ reactors and aqueous phase
bioreactors. Natural attenuation (also known as passive or intrinsic
bioremediation) will be used more widely as the ability to predict and mea-
sure rates of biodegradation improves, and regulatory agencies recognize
the value of minimal disturbance to the environment and the cost-effective-
ness of such attenuation.
Since the 1980s, there have been widespread efforts to introduce new and
modified microbial approaches. Use of white rot fungus, cometabolic pro-
cesses, anaerobic processes, and methods for improving microbial. transport
through porous media are being pursued by several government and private
organizations. Many organizations are making rapid progress in developing
new strains of genetically-engineered microorganisms that have specialized
metabolic capabilities. The widespread use of such microorganisms, how-
ever, will require technical developments in transporting microorganisms
and improving survival traits. The use of genetically-engineered microor-
ganisms will also require acceptance by society and regulatory agencies.
7.2
-------
Chapter 7
Many of the improved processes have been shown to be effective in the
laboratory and, in some cases, pilot-scale demonstrations. Their commer-
cialization, however, will require additional effort. The small-scale results
will have to be applied to real world situations and the processes will have
to function routinely and continually. The question is not so much whether
a contaminant can be treated effectively, but when the technology will be
available. The issue of when may largely be controlled by the ability to
make the necessary engineering modifications to the processes used for site-
specific applications. To date, most applications of bioremedial processes
have involved the use of indigenous microorganisms to treat the readily
degradable constituents.
7.3 Site Characterization
Expanded use of bioremediation will depend on the ability to ask the
right questions and provide appropriate answers. Improvements are needed
in the methods of site investigations to provide more detailed delineation
and better real-time data. Site characterizations need to provide adequate
information to select and design a remedy and to follow remedial progress
addressing such issues as proximity of water supplies, water table, and
population density. However, generation of excessive amounts of data will
be costly and dilute the evaluation and interpretation of the essential data.
In addition to contamination identification and concentrations, improve-
ments are also needed in identifying microbial parameters including micro-
bial populations and metabolic capabilities, nutrient requirements and avail-
ability, and identification of soil heterogeneity.
7.4 Site Closure Criteria
The present criteria for site closure need re-evaluation. Many sites may
never reach closure because clean-up standards, particularly for total petro-
leum hydrocarbons, may be overly conservative and difficult, if not impos-
sible, to achieve. Instead of imposing unnecessary conservative clean-up
7.3
-------
Technology Prognosis
criteria, efforts should be focused on achieving remediation to a level that is
adequate and protective of human health and the environment. Using this
more realistic approach to remediation, the chemical of concern will be
detoxified, degraded, and immobilized.
As these issues are resolved, bioremediation will be used to treat an ever-
growing list of contaminant types under increasingly more demanding con-
ditions. Increased information and awareness will also lead to a more sci-
entific/technical approach to implementing and evaluating bioremediation.
As a result, implemented bioremediation projects will become more reli-
able, especially for the less recalcitrant contaminants.
7.4
-------
Appendix A
CASE STUDIES
BIOVENTING
There have been few well-studied examples of bioventing. The joint
U.S. Air Force/ U.S. EPA RREL bioventing demonstration project being
conducted at Hill Air Force Base near Ogden, Utah provides insight into the
bioventing process. A catastrophic release of 100,000 L (27,000 gal) of JP-
4 jet fuel contaminated the upper 20 m (60 ft) over an area approximately
4,000 m2 (1 a) in the delta outwash of the Weber River. The depth to water
was approximately 200 m (600 ft) with occasional clay stringers. Soil
moisture averaged less than 6%. Jet fuel concentrations were as high as
20,000 mg/kg with an average of approximately 400 mg/kg. Measurements
of total petroleum hydrocarbons (TPH) ranged to 20,000 mg/kg with an
average of 1,500 mg/kg (Dupont 1992b).
The vapor recovery system consisted of vertical wells at 12 m (40 ft)
intervals, which were screened from 3 to 18 m (10 to 60 ft) below the sur-
face. Off gases were treated by catalytic incineration. Initially, the total air
recovery rate was approximately 40 mVhr (26 acfm) or 0.04 pore volumes
per day. As the hydrocarbon levels in the vent gas decreased, the rate of
recovery was gradually increased to approximately 2,500 mVhr (1,500
acfm) or approximately 2.5 pore volumes per day. After one year of opera-
tion at the higher flow rate, the ventilation rates were reduced to between
500 and 1,000 mVhr (300 and 600 acfm) from wells located on the periph-
ery of the site to maximize the retention time within the contaminated zone
and to allow discontinuation of the off gas treatment while meeting the
prevailing air quality regulations.
A.I
-------
Case Studies
The process was monitored at several locations and depths for air pres-
sure, oxygen, and carbon dioxide. Carbon dioxide and oxygen levels were
compared to a background well that was located approximately 200 m (700
ft) from the site in the same geological unit. Oxygen consumption and
carbon dioxide production were observed throughout the treatment period.
The oxygen uptake rates appeared to be first-order.
Laboratory studies conducted at Battelle Memorial Institute in May,
1989, indicated that both moisture management and nutrient addition stimu-
lated biodegradation rates in the Hill Air Force Base soils. These findings
were incorporated into the field operation.
The upper 15 cm (6 in.) of soil was tilled to mix in the nutrients. Surface
spray irrigation was used to provide moisture and to distribute the nutrients
to depth. Soil samples at the end of the study detected both nitrogen and
phosphorus throughout the range of soil depths being treated. The results of
the field study showed that raising moisture levels to 30% to 50% of field
capacity statistically increased respiration rates. In contrast to the labora-
tory findings, there was little impact from nutrient addition on rates of bio-
degradation in the field. However, nutrients may not have been 1 imiting at
the time of the addition because TPH levels were already below 100 ppm.
Field respirometry tests conducted during the treatment period indicated
that oxygen uptake was a better indicator of biological activity than carbon
dioxide production. Oxygen uptake calculations were based on stoichio-
metric conversion of hexane to carbon dioxide. During the high How rate
period, 15% to 25% of the decrease in hydrocarbon concentration was at-
tributed to biodegradation. For the entire treatment period, approximately
45% of the decline in hydrocarbon concentration was due to biodegrada-
tion. However, because the hydrocarbons are assimilated into biomass, as
well as metabolized to CO2, it is likely that less than stoichiometric quanti-
ties of oxygen are required, and thus, these calculations probably underesti-
mate the percent removal due to microbial activities.
Over the first 10 months of operation, approximately 90,000 kg (200,000
Ib) of TPH was removed, reducing the soil TPH concentrations to approxi-
mately 80 mg/kg. The following 14 months of operation resulted in re-
moval of an additional 5,500 kg (12,000 Ib) of TPH, resulting in average
soil TPH levels of 8 mg/kg. The overall removal efficiency exceeded 99%.
A.2
-------
Appendix A
Another field bioventing study was conducted at Tyndall AFB in Florida,
where JP-4 jet fuel leaked from a tank farm. This test was conducted to
evaluate the effectiveness of low flow-rate bioventing as the sole treatment.
The unsaturated soils were largely sand with a depth to water of only 1.2
m (4 ft). Four plots were used in the test: two plots in the area of highest
contamination where levels ranged from 3,000 to 8,000 mg/kg (dry weight)
measured as TPH and two plots in a designated background location where
TPH levels ranged from 95 to 140 mg/kg. Each plot was constructed by
vertically installing plastic-wrapped plywood to a depth of 1.8 m (6 ft). The
test plots were 2 x 6 m (6 x 18 ft) while the background plots were 1.3 x 4
m (4 x 12 ft). A series of wells were used to dewater the area around the
test plots to a depth of 1.8 m (6ft). Plastic covers were placed over each
plot.
Each test plot contained ground monitoring wells and three banks of
multilayer soil gas sampling points. Air flows were maintained to provide
0.25 to 2.0 pore volumes of air per day. Moisture and nutrients were added
at the surface via drip irrigation. The nutrients consisted of fertilizer-grade
ammonium nitrate and triple superphosphate (Ca(H2PO4)2). The application
rate was at a C:N:P ratio of 100:10:1. Physical removal in the off gas was
measured as hexane equivalents while biodegradation was approximated
from oxygen depletion and from carbon dioxide production using a sto-
ichiometric conversion of 3.5 kg of oxygen per kg of hexane.
Nutrient addition did not significantly enhance biodegradation. Moisture
addition tests were inconclusive because climatic conditions maintained
moisture levels at 70 to 80% of field capacity. Temperature fluctuations
during the field test caused the soil temperature to vary from 17° to 25° C
(63° to 77°F). Respiration rates increased with temperature and approxi-
mated an Arrhenius relationship (see Section on Microbial Ecology and
Physiology). Variations in the air flow rate had the expected result on the
relative amounts of jet fuel that were removed physically versus biode-
graded. As the flow rate increased from 2 L per minute (0.5 pore volumes
per day) to 8 L per minute (two pore volumes per day), the percent removed
by biodegradation decreased from approximately 80 to 65%.
One of the background test plots was used as a vapor-phase bioreactor.
Hydrocarbon degradation rates averaged approximately 1.93 g of TPH per
day per m3 of bed at loading rates of 2.0 g TPH per day per m3 of bed vol-
ume. For these specific conditions, a 4:1 ratio of
A.3
-------
Case Studies
uncontaminated:contaminated soil was adequate for complete degradation
at soil vent gas flow rates of one to two pore volumes per day. These re-
sults are important because they provided some guidelines for designing
bioventing systems that could be operated with or without minimal off gas
treatment.
BIOREMEDIATION OF CONTAMINATED
SUBSURFACE MATERIALS USING LIQUID
DELIVERY: TRAVERSE CITY, Ml
(adapted from Ward et al. 1989; Fiorenza 1991; and Wilson,
Armstrong, and Rifai 1993)
Introduction
In March 1988, a quantitative demonstration of the liquid delivery
method for treating contaminated groundwater and subsurface materials
was initiated (Ward et al., 1989). Hydrogen peroxide, which decomposes to
form oxygen and water, was used as the source of oxygen for hydrocarbon
metabolism. The demonstration was conducted at a site where about 3,900
L (10,000 gal) of aviation fuel had leaked from an underground storage tank
at a Coast Guard Air Station in Traverse City, MI (Twenter, Cummings,
and Grannemann 1985). The spill occurred in 1969 but was not detected
until 1980. The aviation fuel spill as well as other sources contributed to
the resulting plume (figure A. Ion page A.5). Benzene, toluene,
ethylbenzene, and the xylene isomers were the major groundwater contami-
nants. In 1985, an interdiction field was installed to intercept the plume.
Materials and Methods
Site Characterization
A field demonstration was conducted in a 9 x 30 m (30 x 100 ft) plot
located upgradient and in the plume of the fuel spill. The overall
hydrogeology and contaminant regime at the site had been characterized
before the demonstration (Twenter, Cummings, and Grannemann 1985).
A.4
-------
Appendix A
Figure A. 1
Schematic Diagram of Fuel Spill and Resulting Plume
Wardetal 1989
To determine the concentration of hydrocarbons in the experimental plot,
core material was collected before implementing the liquid delivery system.
Samples were collected from boreholes located along transects, which were
parallel to groundwater flow, upgradient of the spill, and in the zone of
heavy contamination. Three contiguous 1 m (3 ft) long cores from each
borehole were collected to provide samples that extended:
• from the top of the capillary fringe to the lowest depth reached
by the water table during an annual cycle;
A.5
-------
Case Studies
• through the heavily contaminated zone; and
• from the slightly contaminated to the uncontaminated zone be-
neath the plume. The concentration of total extractable hydro-
carbons was determined using gas chromatography (Vandegrift
and Kampbell 1988). Using the data on total extractable hydro-
carbons, the computer model BIOPLUME II was used to predict
the amount of hydrogen peroxide that would be required to treat
the contamination in the treatment plot. BIOPLUME II is a two-
dimensional model for contaminant transport that is influenced
by oxygen-limited biodegradation (Rifai et al. 1988).
Treatment Schedule
The treatment schedule included (Ward et al. 1989):
1) February 25, 1988 - injection of water to equilibrate the system
hydraulically
2) March 1, 1988 - injection of water amended with pure oxygen
and inorganic nutrients to adapt the microflora to greater-than-
ambient concentrations of dissolved oxygen
3) June 1, 1988 - injection of water amended with 50 ppm hydro-
gen peroxide and inorganic nutrients
4) June 8, 1988 - injection of water amended with 100 ppm hydro-
gen peroxide and inorganic nutrients
5) June 15, 1988 - injection of water amended with 250 ppm hy-
drogen peroxide and inorganic nutrients
6) August 18, 1988 - injection of water amended with 500 ppm
hydrogen peroxide and inorganic nutrients
7) December 1988 - injection of water amended with 750 ppm
hydrogen peroxide and inorganic nutrients
Core and Groundwater Sampling
Core material and groundwater were collected periodically to determine
the progress of the demonstration. During the demonstration, dissolved
oxygen (DO), chloride (Cl), ammonia (NH3), phosphate (PO4), benzene,
toluene, ethylbenzene, the xylene isomers (BTEX), pH, conductivity, and
the water level were determined in samples of well water. Concentrations
A.6
-------
Appendix A
of DO, Cl, NH3, and PO4 were measured using standard methods (Standard
Methods for the Examination of Water and Wastewater 1985). Concentra-
tions of BTEX were measured using a modification of US EPA method 602
(Federal Register 1984) in which headspace analysis was used instead of
purge-and-trap.
Microbial Numbers in Samples of Groundwater and Sedi-
ment
Viable counts of microoorganisms in groundwater and sediments were
determined periodically during the demonstration (see previous section,
Core and Ground Water Sampling) using the spread plate technique (Ward
et al. 1989). Total heterotrophs were plated on Nutrient agar (Difco Indus-
tries, Detroit, MI) and hydrocarbon-degrading microorganisms were plated
on 1.5% Noble agar (Difco Industries), incubated in the presence of avia-
tion fuel vapors.
Treotability Study
The nutrient requirements of the microflora were determined by adding
different combinations of inorganic nutrients to a mixture of 50 mL of
groundwater, 2.0 ml of a 1:10 dilution of sediment, and 0.5 mL of aviation
fuel (Ward et al. 1989). Combinations of KH2PO4 + Na2HPO4, with
NH4C1, Na2CO3, CaCl2, MgSO4, and FeSO4 were tested. Two treatments
received only nitrogen or phosphorus. In addition, controls were prepared
which received 1) groundwater and inoculum, 2) groundwater, inoculum,
and gasoline, 3) groundwater and all inorganic nutrients, and 4) groundwa-
ter, all inorganic nutrients, and gasoline. The mixtures were incubated at
room temperature (~25°C [77°F]) for 2 weeks, during which time gasoline
was added every 2 to 3 days as needed. After incubation, the biomass was
harvested by centrifugation, dried, and weighed.
System Design Using BIOPLUME II
The design parameters for the demonstration were selected using
BIOPLUME II, and included the injection flow rate, number of injection
wells, concentration of injected oxygen, and time required for remediation.
System Layout. The well system is diagrammed in figure A.2 (on page
A.8). In December 1987, nine 10 cm (4 in) monitoring wells with 4 m
(12 ft) long well screens and 12 small cluster wells were installed. The
A.7
-------
Case Studies
Figure A.2
Schematic Diagram of Well System
Wells
$ Chemical Feed
© Injection
O Cluster (4)
• Cluster (6)
+ Monitoring
Well numbers =
ft. from injection
point
and
Sampling
Building
Injection &
Point ^ 9
t
N
o +109
->
0
Concrete Fire Lane
# 7A, B, C
+ 41
+ ^ ^e
0+3T +«50A,B,C *
+ 25 T49A,B,C Ir4
: ;"•
JHcuifiar Administration BuikhngF J.^hh'^,-
/'^^'•-'••.C^--^
Ward el al. 1968
cluster wells had four, five, or six screens at different depths depending on
their distance from the injection wells. The elevations of the screens were
based on historical maximum and minimum water levels in a monitoring
well (TP4) in the demonstration area. Cluster wells with four screens had
sampling intervals at 4.3, 5.5, 6.0, and 6.4 m (14.0, 18.0, 19.5, and 21.0 ft)
below the ground surface. Cluster wells with five and six screens had sam-
pling intervals at 4.3, 5.5, 6.0, 6.4, and 6.9 m (14.0, 16.5, 18.0, 19.5, 21.0
and 22.5 ft). The sampling depths were numbered from 1 to 6, ranging
from the shallowest to deepest screens. The cluster wells were constructed
of stainless steel and extended into the manifold and sampling building.
This building housed the manifold, sampling equipment, nutrients, and pure
oxygen or peroxide.
A.8
-------
Appendix A
Water was injected at a rate that would increase the water table by (0.3
m) (1 ft) to include the contaminated capillary fringe in the treatment zone
and deliver enough oxygen for biodegradation in the contaminated interval
(Ward et al. 1989). From a series of initial runs with BIOPLUMEII, a flow
rate of 150 L/min (40 gal/min) was selected. The injection water was
pumped from a well located 180 m (600 ft) southeast of the test site and
was split into two injection zones. Of the total flow, 110 L/min (29 gal/
min) was pumped into five injection wells screened below the fuel spill
between 7.7 to 8.8 m (25 to 29 ft) to raise the water table. The remaining
40 L/min (11 gal/min) was amended with nutrients and oxygen and pumped
into chemical feed wells screened in the contaminated interval between 4.3
and 5.8m (14 and 19 ft).
The highest concentration of extractable hydrocarbon detected, 5,590
mg/kg wet solids, was used to calculate the total amount of contamination
in the test plot; uniform distribution was assumed. The oxygen demand of
the contamination was estimated using a oxygen or hydrogen
peroxiderhydrocarbon ratio of 3.2:1 or 6.4:1, respectively. The model was
used to determine the amount of oxygen required to bioremediate the test
plot in 3 to 6 months using weekly, stepwise increases in oxygen concentra-
tion. The stepwise increase in oxygen concentration was used to allow the
subsurface microorganisms to adapt to higher-than-ambient levels of oxy-
gen.
The time needed for bioremediation of the test plot was estimated in a
worst case scenario using BIOPLUME II. Initial runs with the model var-
ied the retardation factor from 1 to 100. The retardation factor estimated
the rate at which the sorbed and/or entrapped contaminants leached into the
groundwater, with low numbers indicating little, and high numbers indicat-
ing a great deal of retardation. Although the use of different retardation
factors results in different concentrations of dissolved contaminants, the
total amount of contamination used in each scenario was the same. There-
fore, the time required for bioremediation is dependent on the retardation
factor used in the model. Using a retardation factor of 100 and an oxygen
supply of 15,060 kg (33,890 Ib), the worst-case scenario estimated that
remediation of the test plot would require six months. This amount of oxy-
gen was calculated assuming that hydrogen peroxide would be added at 8,
40, 100, 200, 400, and 800 mg/L during weeks 1, 2, 3, 4, 5, and 6, respec-
tively, and then at 2,000 mg/L for the remaining six weeks. However, this
design was not followed during the first 12 weeks of the demonstration.
A.9
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Case Studies
System Design Using Microbial Treatability Study
Table A. 1 (on page A. 11) shows the nutrient amendments that resulted in
more biomass than the control treatments, which received only nitrogen and
phosphorus (alpha = 0.05). Although several treatments yielded more bio-
mass than the nitrogen and phosphorus controls, the addition of NH4C1,
KH2PO4, and Na2HPO4 was chosen as the simplest, least expensive, but
effective amendment to enhance contaminant biodegradation.
System Operation
Water was injected into the test plot for 4 days to achieve hydraulic equi-
librium, after which oxygen and inorganic nutrients were added. The nutri-
ent amendment was prepared using food grade chemicals. Concentrated
solutions of NH4C1, KH2PO4, and Na2HPO4 were prepared to yield 56,000,
28,000, and 28,000 mg/L, respectively, and mixed in a 1,900 L (500 gal)
tank with injection water to produce groundwater feed concentrations of
381, 190, and 190 mg/L, respectively. The concentrations of chloride,
phosphorus, amd ammonia nitrogen were 250, 75, and 100 mg/L, respec-
tively. The injection water had a pH near neutrality and was 11" to 12°C
(52°to54°F). For each monitoring well, tracer tests were conducted to
determine the actual seepage velocity of a chloride tracer and the inorganic
nutrients along a flow path to that well (Wilson, Armstrong, and Rifai
(1993).
Oxygen or hydrogen peroxide was added after the groundwater feed left
the mixing tank. The average concentration of oxygen that was injected
was approximately 40 mg/L.
Oxygen was injected into the aquifer for the first three months of opera-
tion, after which it was replaced with hydrogen peroxide. The hydrogen
peroxide was added initially at 50 mg/L, and increased stepwise to a final
concentration of 750 mg/L (see Core and Ground Water Sampling section
for schedule). This schedule was different from that used by the
BIOPLUMEII model used to predict the time required for bioremediation
of the test plot.
A.10
-------
Appendix A
Table A.I
Mass of Cells Resulting From Different Inorganic Nutrient Amendments
KH,P04
Sample Na2flPO4
20 +
29 +
16 +
32 +
9 +
36 +
14 +
11 +
33 +
26 +
13 +
18 +
15 +
35 +
38 +
21 +
31 +
25 +
39 +
10 +
30 +
37 +
23 +
12 +
7 +
24 +
27 +
22 +
8 +
17 +
34 +
5 +
19 +
3 +
4
28 +
NH4C1 Na2CO3 CaCl2 MgSO4 MnSO4 FeSO4 Mass",g
+ -- + -- 0.0405
+ + + - + + 0.0405
+ - + +-- 0.0385
+ - + - + + 0.0380
+ -- + -- 00375
+ + - + - + 00350
+ + -- + - 0.0345
+ - - - + 0.0335
+ - + + + 00330
+ + + + + + 0.0330
+ + + 00330
+ + + 0.0325
+ + + 0.0325
+ + + + 00320
+ + + + 0.0320
+ + + 0.0320
+ + + + 0.0320
+ + + + + + 00315
+ + + + 0.0300
+ _ + 0.0290
+ + + + + 00285
•f + + + 00265
+ + + + 0.0255
+ + + 00250
+ + 0.0225
+ + + + + 00225
+ + + + + 0.0220
+ + + 00215
+ + 0.0215
+ + + 0.0210
+ + + + 0.0190
+ 00185b
+ + + 00180
00000
+ 0.0000
+ + + + + 00410
Fiorenza 1991
a Mean cell weight, a difference of 0 01593 g between treatments is significant
b Control treatment
A.ll
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Case Studies
Results and Discussion
Nutrient Transport in the Contaminated Interval
The seepage velocity of the injected water averaged 1.6 to 2.7 m (5 to 9
ft/day) (Wilson, Armstrong, and Rifai 1993). The results of tracer tests to
determine the actual seepage velocities of the chloride tracer and inorganic
nutrients are shown in table A.2. Transport of ammonium and arid phos-
phate ions was about half that of water, a finding which is different from
those in most other aquifers in which there is much stronger retardation of
these ions. Within 2 months after injection of the nutrients, greater than 10
mg/L of ammonia nitrogen and phosphate were distributed throughout the
test plot.
Oxygen Transport and BTEX Removal in Groundwater
The concentration of dissolved oxygen in groundwater collected from a
well (BD-7B) 2 m (7 ft) from the injection point was similar to that in the
chemical feed water (figures A.3 on page A. 13, and A.4 on page A. 14)
(Wilson, Armstrong, and Rifai 1993). Although the concentration of dis-
solved oxygen exceeded the equilibrium solubility of oxygen at ambient
Table A.2
Seepage Velocity of Fronts of Injected Chloride, Oxygen, Ammonia,
and Phosphate Between the Infiltration and Monitoring Wells
Distance from the
infiltration wells
(feet)
7B
(center well)
31
50B
(center well)
62
83B
(center well)
Level
2
3
2
3
2
3
2
3
2
3
Chloride
10.0
11.4
8.6
8.0
6.0
5.5
4.2
6.0
2.5
6.5
Oxygen
5.6
7.3
NET
4.4
NBT
NET
1.2
1.2
NBT
NBT
Ammonium
ion
6.5
8.0
3.9
39
2.1
3.3
26
3.1
1.9
3.3
Phosphate
5.6
7.1
3.6
3.7
2.1
2.')
1.7
2.8
0.7
2.2
Wilson, Armstrong, and Rifai 1993
NBT = No breakthrough during the tracer test
A.12
-------
Appendix A
Figure A.3
Schedule of Supply of Oxygen or Hydrogen
Peroxide to the Infiltration Wells
o
-50 0 50 100 150 200 250 300 350 400 450 500
Days Since Startup
Data prior to day 0 is the ambient concentration of oxygen in the aquifer moving into the spill under the natural gradient.
The concentration of hydrogen peroxide is expressed in terms of dioxygen equivalent
Wilson, Armstrong, and Rifai 1993
hydrostatic presssure and temperature for most of the demonstration, these
overpressures did not initiate bubble formation in the sand formation. The
oxygen did not escape because there was no air space in the injection sys-
tem.
Of interest were the patterns of oxygen and BTEX concentrations in
groundwater collected from a well (BD-31) 9 m (31 ft) from the injection
point (figure A.5 on page A. 15) (Wilson, Armstrong, and Rifai (1993). The
initial breakthrough of oxygen was at about the same concentration as that
of the injection water (25 mg/L), after which it declined to nondetectable
levels as the microflora acclimated to consume the oxygen. After 180 days
oxygen broke through again, which corresponded to an increase in hydro-
gen peroxide concentration in the injection water (compare figures A.3 and
A.5 on page A. 15). After oxygen reappeared in the groundwater, BTEX
soon disappeared.
A.13
-------
Case Studies
Figure A.4
Breakthrough of Oxygen and Depletion of Total Alkylbenzenes
in a Monitoring Well in the Most Contaminated Depth Interval
Seven Feet from the Infiltration Wells
BD-7B (Level 2)
~ 175 i
•I 125-
S
g 100-
5 75-
X
£ 501
m
•3 25-
-10 90 190 290 390 490 590
Days Since Startup
90 190 290 390 490 590
Days Since Startup
Wilson, Armstrong, and Rifai 1993
The dissolved oxygen concentrations in groundwater from wells located
15m (50 ft) from the injection point were never above 10 mg/L (figure A.6
on page A. 15) (Wilson, Armstrong, and Rifai 1993). Even though dis-
solved oxygen levels were low, BTEX was not detected after 300 days of
treatment. Oxygen breakthrough was not detected at wells located 19 and
25 m (62 and 83 ft) from the injection point (figures A.7 on page A. 16 and
A.8 on page A. 17). Although BTEX was still detectable in groundwater
from these wells at the end of the demonstration, the concentrations were
lower than initial levels.
Of interest was the apparent selective removal of benzene from the
groundwater; the xylenes were more recalcitrant (Wilson, Armstrong, and
Rifai 1993) (table A.3 on page A. 18). The pattern of benzene removal was
unusual. Benzene disappearance first occurred 25 m (83 ft) from the infil-
tration point below the contaminated zone; after 420 days of operation,
removal was uniform throughout the plot. Benzene was removed near the
infiltration wells using up to 80 pore volumes, whereas as little as eight pore
A.14
-------
Appendix A
Figure A.5
Breakthrough of Oxygen and Depletion of Total Alkylbenzenes
in a Monitoring Well in the Most Contaminated Depth Interval
31 Feet from the Infiltration Wells
BD-31 (Level 3)
90 190 290 390 490 590
Days Since Startup
~ 175 -i
=J
J|150-
| 125-
S
1 100-
a
a 75-
X
W 50-
co
3 25^
S -i
£ n1
M A
\ • A
1 \A\
I V \ \
* V V
-10 90 190 290 390 490 590
Days Since Startup
Wilson, Armstrong, and Rifai 1993
Figure A.6
Breakthrough of Oxygen and Depletion of Total Alkylbenzenes
in a Monitoring Well in the Most Contaminated Depth Interval
50 Feet from the Infiltration Wells
i
E
*0)
2
5
5
1
•3
5
80-
70-
-
60-
50-
40-
30-
20-
10-
0-
BD-50 (Level 2)
„ 1250;
|> 1125;
* 1000 :
1 875:
1 750 :
I 625"
o
U 500-
X
W 375;
CO 250 :
N A. 1 125^
rA
ff/ \ .
V \ A
i VV
* i i i i ?H-"t i"i'i*-f i
-10 90 190 290 390 490 590
Days Since Startup
-10 90 190 290 390 490 590
Days Since Startup
Wilson, Armstrong, and Rifai 1993
A.15
-------
Case Studies
volumes was required to remove the aromatic compound 33 m (108 ft) from
the infiltration wells.
Removal of BTEX and Petroleum Hydrocarbons from
Sediments
Initial concentration of total petroleum hydrocarbon ranged from 2,000
to 12,000 mg/kg of which 5 to 10% was alkylbenzenes (Wilson. Armstrong,
and Rifai 1993). Changes in concentrations of BTEX and total petroleum
hydrocarbons are shown in table A.4 (on page A. 19). Toluene and the xy-
lenes were removed in samples collected near the monitoring well 10 m (32
ft) from the infiltration point (see sample 50T3). After 8 months of opera-
tion when oxygen broke through and BTEX disappeared in monitoring well
BD-32, cores were collected and analyzed for BTEX and TPH. Concentra-
tions of BTEX were below detection; however, concentrations of TPH re-
mained at initial levels. After 12 months of operation, extensive: removal of
total petroleum hydrocarbons was detected (sample 50AQ3).
Figure A. 7
Breakthrough of Oxygen and Depletion of Total Alkylbenzenes
in a Monitoring Well in the Most Contaminated Depth Interval
62 Feet from the Infiltration Wells
BD-62 (Level 2)
80-
70-
60-
50
40-
30-
20-
10-
-10
90 190 290 390 490
Days Since Startup
590
1250;
1125;
iooo:
875-
750
625
500;
375;
250;
125-
0
-10 90 190 290 .390 490 590
Days Since Startup
Wilson. Armstrong, and Rifai 1993
A.16
-------
Appendix A
Figure A.8
Breakthrough of Oxygen and Depletion of Total Alkylbenzenes
in a Monitoring Well in the Most Contaminated Depth Interval
83 Feet from the Infiltration Wells
BD-83 (Level 2)
i
_§
"a>
_3
g
W)
1
1
1
Q
80 -i
70-
-
60-
50-
40-
30-
20-
10-
.
0-
-10 90 190 290 390 490
Days Since Startup
590
90 190 290 390 490 590
Days Since Startup
Wilson, Armstrong, and Rifai 1993
After 12 months of operation, toluene was the only alkylbenzene that had
been removed completely from samples collected just in front of the moni-
toring well 19m (62 ft) from the infiltration point; however, the concentra-
tion of TPH had been reduced from an initial concentration of about 6,500
to 3,100 mg/kg wet weight (table A.4 on page A.19) (Wilson, Armstrong,
and Rifai 1993). This removal occurred even though high levels of oxygen
had not broken through this section of the treatment zone (see figure A.7 on
page A. 16).
Concentrations of BTEX in samples of sediment at the end of the demon-
stration (522 days) were below 0.07 mg/kg in sediment samples collected 3,
10, and 16 m (10, 34, and 54 ft) from the infiltration point (table A.5) (Wil-
son, Armstrong, and Rifail 1993). In samples collected 19.5 m (64 ft) from
the infiltration point, concentrations of benzene, toluene, and o-xylene were
less than 0.07 mg/kg, whereas low concentrations of ethylbenzene and m-
p-xylene persisted. Even though BTEX was extensively removed through-
out the test plot, high concentrations of TPH remained.
A.17
-------
Case Studies
Table A.3
Depletion of Alkylbenzenes in Groundwater After 17 Months of
Infiltration of Nutrients and Oxygen
Distance from
infiltration wells
(meters feet) Benzene Toluene Ethylbenzene Total Xylenes
2.0(7) <0.1 <0.1 <01 <0.1
94(31) <0.
15 (50) <0.
19 (62) <0.
25 (83) <0.
33(108) <0.
<0.1 <0 1 08
<0.1 1.0 10.4
0.3 1.7 37.2
0.3 12 367
2 4 6.4 393
Wilson, Armstrong, and Rifai 1993
Microbial Patterns (adapted from Fiorenza, 1991)
The number of microorganisms in groundwater from monitoring wells
did not follow any consistent pattern. Most likely, changes in microbial
numbers were difficult to determine because these wells were screened over
a 3 m (12-ft) interval where there were vast differences in hydrocarbon
concentration. The number of heterotrophs or hydrocarbon-degrading or-
ganisms from a particular monitoring well approximated the average of all
levels from the corresponding cluster well (see figure A.2 (on page A.8) for
corresponding monitoring and cluster wells).
The data on cell numbers from cluster wells were more informative. The
most dramatic and consistent changes in microbial numbers were detected
in level 2 of the cluster wells, where the concentration of contamination was
the highest. In samples with little or no contamination, there was no obvi-
ous trend.
As the concentration of BTEX began to decrease during the demonstra-
tion, dissolved oxygen increased. As the concentration of BTEX decreased
in samples from the most contaminated interval (level 2 of the centerline
cluster wells), the number of heterotrophs and hydrocarbon-degrading mi-
croorganisms in groundwater usually increased (figures A.9 and A. 10 on
page A.20, A.ll on page A.21, A.12 on page A.22, A.13 and A.14 on page
A.23). This trend was not evident in wells located 25 and 33 m (83 and 108
A.18
-------
Appendix A
Table A.4
Changes in Concentration of Alkylbenzenes and Total Petroleum
Hydrocarbons in Core Material During Bipremediation of an Aquifer
Contaminated With Aviation Gasoline
Ethyl- o- m+p
Core* TPH Benzene Toluene Benzene Xylene Xylene
(nig/kg wet sample) - -
Near BD-31, collected June 1988 after 4 months of perfusion with mineral nutrients and oxygen
50T3 3,330 14 <1 7.3 23 <1
Near BD-31, collected after 8 months of perfusion with mineral nutrients and oxygen
50AE4 8,400 <0.3 <0.3 <0.3 <0 3 <0.3
50AE5 2,370 <0 3 <0 3 <0.3 <0 3 <0.3
Near BD-31, collected after 12 months of perfusion with mineral nutrients and oxygen
50AQ3 9 <03 <0 3 <0.3 <0.3 <0.3
Near BD-62, collected after 12 months of perfusion with mineral nutrients and oxygen
50AR4 3,100 1.5 <0.3 92 30 6.2
Wilson, Armstrong, and Rifai 1993
"Number designates distance from infiltration gallery, letters indicate sample
ft) away from the infiltration point where groundwater contamination still
existed at the end of the demonstration. Also of interest was the decrease in
both heterotrophs and hydrocarbon-degrading microorganisms to
nondetectable levels in samples from well BD-7B after the BTEX disap-
peared (figure A.9 on page A.20). This plummet in numbers coincided with
the detection of hydrogen peroxide in the groundwater from this well, and
suggests toxicity of the peroxide to the microorganisms. However, low
numbers of microorganisms (<100 cells/mL) were detected in samples from
this well later in the demonstration while peroxide concentrations were still
high, suggesting recolonization by peroxide-tolerant bacteria.
The number of catalase-postive microorganisms increased in samples of
groundwater from all levels of the cluster wells as the demonstration pro-
ceeded (data not shown). Catalase is the enzyme that catalyzes the break-
down of hydrogen peroxide to oxygen and water. A deviation from this
trend was seen in samples from well BD-7B in which the number of cata-
A.19
-------
Case Studies
Figure A.9
Effect of DO (•) and BTEX (x) on the Number of Heterotrophs (•) and
Hydrocarbon-Degraders (a) in Groundwater from BD-7B, Level 2
600
Fiorenza 1991
Figure A. 10
Effect of DO (• ) and BTEX (x) on the Number of Heterotrophs (•) and
Hydrocarbon-Degraders (a) in Groundwater from BD-31, Level 2
20-
600
Fiorenza 1991
A.20
-------
Appendix A
Figure A. 11
Effect of DO (• ) and BTEX (x) on fhe Number of Heterotrophs (•) and
Hydrocarbon-Degraders (n) in Groundwater from BD-50, Level 2
16
100
200 300 400
Julian Date
500 600
Fiorenza 1991
lase-positive organisms increased until the BTEX disappeared and peroxide
was detected (described earlier); then there were no detectable microorgan-
isms. At the end of the demonstration, the percentage of microoganisms
that were catalase-positive was greater than 90% in most groundwater
samples. These data suggest that catalase-positive organisms were selected
by exposure to peroxide, and those lacking catalase were unable to with-
stand the addition of peroxide.
Samples of core material from the shallow, contaminated and the deep,
uncontaminated zones also were analyzed for microbial numbers. The
change in number of hydrocarbon-degrading and heterotrophic microorgan-
isms was similar to that observed in groundwater. Counts were higher in
the shallow, contaminated samples than the deep, uncontaminated samples.
Both types of organisms increased in concentration in both shallow and
deep samples collected 9.4 and 19 m (31 and 62 ft) from the injection point
A.21
-------
Case Studies
and then declined (figures A. 15 on page A.24, A. 16 on page A.25). In
samples collected 33 m (108 ft) from the injection point, numbers of both
types of organisms initially were low and declined during the demonstration
(figure A. 17 on page A.25).
For samples from the 9.4 m (31-ft) zone, organisms were not detected in
the deep, uncontaminated samples and lower numbers than previously were
detected in the shallow, contaminated samples by Julian date 580; this oc-
curred about 400 days after BTEX was no longer detectable in the 9.4-m
(31-ft) well near this sampling point. Because BTEX was no longer detect-
able and therefore carbon was not available for cell metabolism and catalase
production, hydrogen peroxide may have sterilized the area.
The change in catalase activity in shallow and deep cores is shown in
figure A. 18 (on page A.27). Except for samples collected close to the infil-
tration point (2 and 9.4 m (7 and 31 ft)), catalase activity increased during
the demonstration, suggesting that the subsurface microflora adapted to the
peroxide addition by increasing catalase activity. The decline in activity in
Figure A. 12
Effect of DO (•) and BTEX (x) on the Number of Heterotrophs (•) and
Hydrocarbon-Degraders (n) in Groundwater from BD-62, Level 2
600
Fiorenza1991
A.22
-------
Appendix A
Figure A. 13
Effect of DO (•) and BTEX (x) on the Number of Heterotrophs (•) and
Hydrocarbon-Degraders (a) in Groundwater from BD-83B, Level 2
1400
600
Fiorenza 1991
Figure A. 14
Effect of DO (• ) and BTEX (x) on the Number of Heterotrophs (•) and
Hydrocarbon-Degraders (a) in Groundwater from BD-108, Level 2
rr450
600
Fiorenza 1991
A.23
-------
Case Studies
Figure A. 15
Heterotrophs (Squares) and Hydrocarbon-Degraders (Circles) in Shallow
(Closed Symbol) and Deep Subsurface (Open Symbol) Cores at 31 Feet
0 100 4oO 300 401
| Julian Date
BTEX not detected in Well BD3I-2 after this date
500
600
700
Fiorenza 1991
samples collected close to the infiltration point may have resulted from the
lack of a carbon source (BTEX) for microbial utilization during the latter
phase of the treatment and, therefore the inability of the microorganisms to
produce catalase.
Catalase activity was higher in the shallow, contaminated samples, which
contained BTEX that could be used as a carbon source, than in the deep,
uncontaminated samples (figure A. 18 on page A.26). The decreases in
activity in shallow samples observed at Julian day 279 occurred after a
decline in the water table elevation that prevented transport of nutrients and
oxygen through this zone; declines in heterotrophs were observed in shal-
low samples collected 9.4 m (31 ft) from the injection point (compare fig-
ures A. 15, A. 18 on page A.26). The decline in catalase activity in deep
samples collected 9.4 m (31 ft) from the injection point may have resulted
from toxicity of peroxide, which was detected at 167 mg/L at level 4 after
533 days of operation (compare table A.5 on page A.27 and figure A. 18 on
page A.26).
A.24
-------
Appendix A
Figure A. 16
Heterotrophs (Squares) and Hydrocarbon-Degraders (Circles) in Shallow
(Closed Symbol) and Deep Subsurface (Open Symbol) Cores at 62 Feet
600
Fiorenza 1991
Figure A. 17
Heterotrophs (Squares) and Hydrocarbon-Degraders (Circles) in Shallow
(Closed Symbol) and Deep Subsurface (Open Symbol) Cores at 108 Feet
Julian Date
Fiorenza 1991
A.25
-------
Case Studies
Figure A. 18
Catalase Activity in Shallow and Deep Subsurface Cores 7 (*), 31 (•),
62 (•), and 108 (A) ft from the Infiltration Wells
700
Fiorenza 1991
Nitrate and Alkylbenzene Biodegradation
Although high levels of oxygen were never detected in wells beyond 15
m (50 ft) from the infiltration point, significant amounts of alkylbenzenes
were removed in this region. Anaerobic biodegradation of these com-
A.26
-------
Appendix A
pounds was believed to be the result of using nitrate as an alternate electron
acceptor (Wilson, Armstrong, and Rifai 1993). The biodegradability of the
alkylbenzenes, except benzene, using nitrate as the terminal electron accep-
tor, has been demonstrated in the laboratory and field. The remaining
straight-chain fraction is recalcitrant under anaerobic conditions.
Although nitrate was not infiltrated into the test plot, nitrate accumulated
in zones that contained oxygen, little or no hydrogen peroxide (table A.6 on
page A.28), and no BTEX (figures A.4 to A.8 on pages A.14 to A.17); how-
ever, nitrate levels were lower in zones downgradient that still contained
BTEX. Calculations indicated that a significant fraction of the oxygen
added as hydrogen peroxide was converted to nitrate. These data suggest
that the nitrate produced from nitrification of the added ammonia served as
the terminal electron acceptor in BTEX biodegradation in zones that re-
ceived little or no oxygen.
Effectiveness of Treatment
After 17 months of treatment, removal of BTEX from groundwater was
extensive, whereas that for the TPH was low (Wilson, Armstrong, and Rifai
Table A.5
Concentrations of Alkylbenzenes and Total Petroleum Hydrocarbons in
Core Material From Most Contaminated Interval, After 17 Months of
Perfusion with Mineral Nutrients and Oxygen
Ethyl- o- m+p
Core TPH Benzene Toluene Benzene Xylene Xylene
(mg/kg wet sample) —
Near BD-7, 3 m (10 ft) from the infiltration wells
50AY3 922 <0 07 <007 <0 07 <0 07 <007
Near BD-31, 10 m (32 ft) from the infiltration wells
50BD2 2,310 <0.07 <0 07 <007 <0.07 <0.07
Near BD-50B, 16 m (53 ft) from the infiltration wells
50AW42 10,800 <0.07 <0.07 <0.07 <007 <007
Near BD-62, 19 5 (64 ft) from the infiltration wells
50BB7 1,280 <007 <007 3.2 <0 07 014
Wilson, Armstrong, and Rifai 1993
A.27
-------
Case Studies
Table A.6
Hydrogen Peroxide Decomposition, Nitrification, and
Potential Denitrification After Infiltration With Mineral
Nutrients and Hydrogen Peroxide
Distance
(feet)
Level °2
HA
NH4+
mg/Iiter-
Injection
7B
(center well)
31
SOB
(center well)
62
83B
(center well)
108
2
3
4
5
6
2
3
4
1
2
3
4
2
3
4
2
3
4
2
3
4
5
210
212
230
150
1.9
8.2
6.5
137
3.7
1.0
0.4
1.0
4.0
0.7
0.9
0.6
1.1
1.2
0.4
1.6
0.7
1.0
930
930
840
860
0.5
<0.1
<0.1
167
0.1
<0.1
<0.1
<0.1
0.2
<0.1
<0.1
<0.1
0.2
01
<0.1
<0.1
<0.1
<0.1
106
91
93
92
95
68
34
82
67
64
17
72
58
59
42
39
60
18
1.8
13
19
19
NO3-
NO2-
Dt nitrified
-mg N/liter
0.30
0.35
0.30
0.35
0.40
30
65
4.3
17
22
80
20
12
27
57
18
19
42
3.2
4.3
15
<0.05
<0.05
<005
<0.05
<0.05
1.9
24
1.9
10
2.8
1.9
2.4
12
1.4
3.2
22
7.0
4.4
<0.05
7.3
0.9
1.0
15
13
14
11
6
5
18
12
17
7
12
36
19
4
47
20
42
101
81
71
71
Wilson, Armstrong, and Rifai 1993
Hydrogen peroxide and oxygen data collected, 8/16/89 after 533 days of operation Ammonia, nitrate, and nitrite data
collected 7/13/89 after 499 days of operation.
1993). Total remediation of the test plot was not expected since title amount
of oxygen added was less than that required to achieve complete biodegra-
dation of the hydrocarbons.
At 34 m (110 ft) from the infiltration point, the concentration of benzene
was less than 1 mg/L and the concentrations of the other alkylben zenes
were below federal drinking water standards. The concentration of each
alkylbenzene in samples of core material collected up to 15 m (50 ft) from
the infiltration point was less than 0.07 mg/kg. However, the concentration
of TPH in the most contaminated interval ranged from 992 to 10,800 mg/
A.28
-------
Appendix A
kg. Based on the amounts of oxygen and nutrients added to the treatment
zone, the selective removal of the alkylbenzene fraction is difficult to ex-
plain. However, a significant amount of the oxygen supplied as hydrogen
peroxide was converted to nitrate. The overall selective removal of the
BTEX fraction probably resulted from the use of nitrate as the terminal
electron acceptor in regions that received little or no oxygen.
IN SITU BIOREMEDIATION OF DIESEL FUEL IN
SOIL Maureen E. Leavitt, IT Corporation,
Knoxville TN
The largest inland oil spill to date occurred in 1988 in western Pennsyl-
vania. A tank collapse at a petroleum terminal resulted in the release of a
million gallons of No. 2 fuel oil. After the emergency response was com-
pleted, an estimated 562,000 L (145,000 gal) of oil remained on the soils
surrounding the tank area and in a nearby wooded area. The IT Corporation
was contracted to investigate, design, and implement a bioremediation strat-
egy to reduce the TPH concentration in soil. The proposed approach was to
stimulate the indigenous hydrocarbon-degrading bacteria using inorganic
nutrients and regular tilling. Since the spill was recent, the contamination
was limited to the top two feet. Therefore, it was decided that all soils
would be treated in place.
This project was among the first bioremediation applications for this
region of the Pennsylvania Department of Environmental Resources, and
proof of the treatment was required prior to initiating the full-scale applica-
tion. Samples were collected within two months of the spill. A
bioassessment was completed to confirm that conditions were conducive to
bioremediation across the site. In addition, a biotreatability study was con-
ducted to demonstrate that indigenous organisms could reduce the TPH
content when supplied with nutrients, lime, and tilling.
Full-scale operations began within six months of the release. The soils
monitoring program documented a substantial increase in the bacterial
populations density, and maintenance of pH within the desirable range.
A.29
-------
Case Studies
Leachate quality improved during the bioremediation program, reaching
levels acceptable to the existing National Pollutant Discharge Elimination
System (NPDES) permit and the Consent Decree.
Although excessive rainfall often inhibited effective tilling, the 1,000
mg/kg target level was achieved in many areas after two growing seasons.
The third and final treatment season resulted in attaining the assigned treat-
ment target level in all areas, with the exception of one isolated area which
was identified as having non-spill related hydrocarbons. Extensive closure
sampling and analyses were completed and the soils treatment program was
considered a success.
Introduction
Land treatment (or land farming) for the reduction of petroleum hydro-
carbons in soil has been applied throughout the petrochemical industry for
many years. The conventional system utilizes a multi-acre field over which
the waste is applied. The field is regularly tilled to oxygenate the; soils, to
redistribute the waste throughout the soil medium, and to supplement the
soil with moisture and nutrients (Bartha and Bossert 1984). Successful land
treatments propagate an adapted hydrocarbon-degrading bacterial popula-
tion that is capable of mineralizing complex hydrocarbon loadings in excess
of 2% by weight.
The described adaptation of land treatment was used to remediate soils
that have been unintentionally contaminated with fuel oil. This method of
land treatment did not have the benefit of an even, calculated loading rate,
nor of an ideal soil matrix. Several physical, geographical, and regulatory
issues were addressed and resolved, resulting in successful remediation of
the subject soil.
Pertinent Site Factors
The subject fuel was No. 2 or diesel fuel, which is generally between 1%
to 10% volatiles (Reed and Associates, Inc. 1988). Although almost 16
million liters (4 MMgal) of fuel was stored in the tank, estimates stated that
approximately 4 million liters (1 MMgal) was released. After emergency
response efforts were completed, approximately 562,000 L (145,000 gal)
remained on the soil. Total petroleum hydrocarbon concentrations were as
A.30
-------
Appendix A
high as 100,000 mg/kg; however, most of the soil contained concentrations
in the 10,000 to 20,000 mg/kg range.
The area impacted by the fuel release was an operating petroleum termi-
nal (figure A. 19). The impacted area included several tank basins, an as-
phalt plant area, and a wooded/field area across a highway from the termi-
nal. The total volume of soil to be treated was approximately 11,500 m3
(15,000 yd3). The topography was generally flat and sloped toward the
river. The soils that received the fuel consisted of clays to clay/silts. The
ambient air temperature at the time of the release was -4°C (25°F).
Remedial Approach
Samples were collected from across the site to determine the potential for
successful bioremediation. A summary of the results is listed in table A.7
(on page A.32). A total of 27 samples were collected and analyzed for pH,
background nutrient concentration, and microbial population density. The
Figure A. 19
Site Configuration
Tanks
250'
A.31
-------
Case Studies
Table A.7
Initial Bioassessment of Site Soils
Area
Asphalt PlantA
Tank Basins8
Wooded Areac
A = Average of 7 samples
B - Average of 10 samples
C = Average of 10 samples
cfu = colony -forming unit
pH
6.9
5.2
5.5
Ammonia
(mg/kg)
8
8
12
Phosphate
(mg/kg)
0.5
0.5
47
Total
Heterotrophs
(cfu/g)
2.6x106
6.9 x 105
1.1 x 107
Diesel
Degraders
(cfu/j.)
2.5 x 106
7.8 x 105
1.5x 106
pH was near-neutral in the asphalt plant area; however, the tank basin and
wooded areas exhibited acidic pH values. The background nutrient concen-
trations were negligible. The microbial population density ranged from 104
to 107 colony-forming units per gram, for both heterotrophs and hydrocar-
bon-degrading bacteria. These data suggested that a significant microbial
population existed, although nutrient and lime augmentation would be re-
quired to stimulate rapid biodegradation of the diesel fuel.
To determine the specific requirements for the full-scale system, and to
establish that biodegradation could reduce the TPH content of the soil, a
biotreatability study was completed and has been described (Leavitt and
Jadlocki 1989). The study evaluated the benefit of lime addition, and four
different fertilizer loadings compared to moisture addition and tilling alone.
The specific treatments are described in table A.8 (on page A.33). The
treatments were maintained by maintaining the target moisture conlent (ap-
proximately 17% by weight) and thorough weekly mixing. The prescribed
nutrient dosage for each treatment was added in three aliquots, one every
two weeks. At two week intervals, duplicate samples from each treatment
were collected and analyzed for TPH.
The results of the TPH analyses are illustrated in figure A.20 (on page
A.34). There was no discernible difference between the nutrient dosages,
and the TPH in each nutrient treatment was considerably lower than the
untreated controls. In addition, the bacterial densities increased by as much
as two orders of magnitude in the nutrient-amended treatments. The great-
A.32
-------
Appendix A
est increase in bacterial density was observed in the manure treated soil.
Comparing the effect of lime addition, higher bacterial densities were ob-
served in soil samples after lime treatment. The results of the study sug-
gested that lime and nutrient addition increased the microbial population,
resulting in a lower TPH content compared to the control (moisture and
tilling only). These results were used to design a full-scale operating plan.
The initial proposal suggested treating soils using nutrient and lime addi-
tion and tilling to reach a 5,000 mg/kg target level over two growing sea-
sons. The final Consent Decree ordered treatment of the soils as described
to 1,000 ppm over three growing seasons.
Since the soils were to be treated in place, random sampling was not
recommended due to the potential for high variability in contaminant con-
centrations. Instead, specific areas were chosen to be monitored once every
three weeks to determine the nutrient and TPH content, as well as the mi-
crobial density and pH. Samples were collected at three distinct depths for
TPH analysis to identify any vertical migration of contamination and to
confirm that complete mixing was occurring.
In addition to soil, water that accumulated in several lysimeters placed
within the unsaturated zone were sampled to document any contaminant
migration through leachate resulting from the treatment program. Samples
Table A.8
Bench-Scale Biotreatability Study Treatment Scheme
Microcosm
Number
1
2
3
4
5
6
7
8
Water
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Lime
No
Yes
No
Yes
Yes
Yes
No
No
Fertilizer
Content
(N:P:K)
None
None
2:0.8:2"
20:4:10
10:6:4
20:4:10
20:4:10
10:6:4
Target Nutrient
Ratios (C:N:P)»
None
None
100:10:1
100:10:1
100:3:1
100:3:0.3
100:10:1
100:3:1
a If C:N:P ratios could not be met exactly, C:N was chosen as the priority ratio and nitrogen loading was used to
determine the quantity of fertilizer required.
b Fertilizer was processed cow manure.
A.33
-------
Case Studies
Figure A.20
Biotreatability Study Total Petroleum Hydrocarbon Analysis
--O--
-a
-A- -
246
Time (weeks)
Moisture Control •
Moisture + Lime A
Manure &-
100.10:1 CNP +Lime ---«-.
10
1003 1 C-N.P + Lime
100-3:3 CN:P +Lime
100.10-1 (C.N-P)
1003.1 (CN:P)
collected from these lysimeters, as well as soils, were tested for volatile and
semivolatile contaminants including benzene, naphthalene, 2-methylnaph-
thalene, toluene, xylenes, ethylbenzene and phenanthrene.
Field Progress
The first growing season began in July of 1988 and continued through
November of 1988 (table A.9 on page A.35). During this 105-day period,
there were 26 days of rain. The moisture accumulation resulting from the
rain limited the ability to mix soils and hampered oxygen diffusion into the
soil. Nutrient addition was applied in monthly doses to maintain a 100:10:1
C:N:P ratio (approximately 200 pounds per acre per application). The pH
was maintained between 6.0 and 7.6 in all areas using lime addition at a rate
of 4,000 pounds per acre.
A.34
-------
Appendix A
Table A.9
Bioremediation Program Operating Record
Growing Season
1988
1989
1990
Operating
Months
Jul - Nov
Apr - Oct
May - Oct
Total
Operating
Days
105
199
190
Percent Rain
Days Of Rain
26
82
73
Days
25
41
38
The microbial density of both heterotrophs and diesel-degrading bacteria
increased as much as three orders of magnitude, providing a sufficient mi-
crobial density for biodegradation (data not shown).
At the completion of the first season, an overall average TPH reduction
between 44 and 81 percent was observed (table A.10). Redidual TPH levels
ranged from 2,000 mg/kg to 8,000 mg/kg.
Analysis of leachate collected from lysimeters is shown for the begin-
ning and end of the first two treatment periods (table A.I la and b on page
A.36). Each value represents the average of between nine and fifteen
samples. In many samples, both volatile and semivolatile contaminants
were below the detection limits. Over the two treatment periods, an in-
Table A. 10
Summary of Soil Analytical Results — 1988 Season
Area
Basin A
Basin B
Basin C
Asphalt
Wooded
March
1988
— - TPH concentn
8,844
17,569
4,200
11,980
10,801
Autumn
1988
ition, mg/kg - —
4,588
7,823
2,340
6,737
2,026
Percent
Reduction
48
55
44
44
81
A.35
-------
Case Studies
Table A.I la
Analysis of Leachate Collected in Lysimeters Over Two Growing Seasons
Dale
July 1988
November 1988
April 1989
September 1989
Benzene
8
4
ND(5)
9
Ethylbenzene
28
ND(5)
ND(5)
12
Toluene
23
ND(5)
ND(5)
8
Xylenes
144
ND (5)
ND(5)
29
Table A.I lb
Analysis of Leachate Collected in Lysimeters Over Two Growing Seasons
Date
Methyl-Napthalene Naphthalene Phenanthrene
July 1988
November 1988
April 1989
September 1989
40
ND(10)
ND(10)
ND(10)
61
ND(10)
ND(10)
15
ND(10)
ND(10)
ND(10)
ND(10)
crease in contaminant concentrations was not detected. Therefore, it was
concluded that the treatment program did not cause a deterioration in the
leachate quality.
The second growing season began in April of 1989 and continued
through October of 1989. In this 199 day period, 82 days of rain were
documented (table A.9 on page A.35). The excessive moisture resulted in
severe tilling limitations. However, it was established that the stratified
sampling could be discontinued after statistical analysis proved the soils to
be vertically homogenized. The bacterial population densities were as high
as 1012 cfu/g, with an average between 109 and 1010 cfu/g.
A.36
-------
Appendix A
The concentrations of volatile and semivolatile compounds, in samples
collected from lysimeter continued to be below detection, therefore lysim-
eter sampling was discontinued. Average TPH values at the completion of
this season varied from 1,600 mg/kg to 6,500 mg/kg (table A. 12). Isolated
areas achieved the 1,000 mg/kg target level.
The third growing season began in May of 1990 and continued through
October, 1990. Severe rainfall hampered any activity prior to May. Ap-
proximately 73 days of rain occurred during the 190 day period. Also dur-
ing that period, the bacterial density declined from an average of 109 to an
average of 107 cfu/g. Since nutrient addition and tilling remained constant,
the decline was presumed to be attributable to lack of available organic
carbon for microbial maintenance and growth.
At the start of the third season, TPH values ranged between 2,000 and
7,000 mg/kg TPH. Initial analytical results indicated that most of the termi-
nal soils were nearing closure as defined in the Consent Decree. Therefore,
Table A. 12
Summary of Soil Analytical Results — 1989 Season
Sample
Location
Asphalt Area
Asphalt Area
Asphalt Area
Tank Basins
Tank Basins
Tank Basins
Tank Basins
Tank Basins
Tank Basins
Tank Basins
Tank Basins
Tank Basins
Wooded Area
Wooded Area
Wooded Area
Wooded Area
Wooded Area
Wooded Area
Average TPH
(mg/kg)
3,791
6,368
6,238
5,422
4,358
4,211
4,967
3,617
2,822
1,681
3,771
1,606
3,933
1,587
1,889
2,132
3,831
2,361
A.37
-------
Case Studies
while tilling and nutrient addition proceeded, the terminal was divided into
779 grids of approximately 250 square feet, as directed for site closure. As
the samples achieved a TPH value of less than 1,000 mg/kg, the corre-
sponding grid was closed, and no further tilling or nutrient addition oc-
curred. Most of the area achieved this criterion during the summer months.
Completion
A total of twenty of the 779 grids exhibited TPH values greater than
1,000 ppm at the end of 1990. Since these grids were treated exactly like
the rest of the terminal, and since the program was successful in removing
the diesel from the rest of the terminal, it was believed that the diesel had
been completely removed from these grids as well. Further investigation
into the nature of the remaining contaminants suggested that these areas
may have been impacted by the asphalt plant operation and could be attrib-
utable to tank bottom sludges.
The final TPH values across the site averaged 450 mg/kg (table A. 13).
Most of the area was found to have values within 50 percent of this value.
Two examples of actual sample location and variability are included in
figures A.21 and A.22 (on page A.39). The soil bioremediation program
was completed and considered successful in mitigating the petroleum hy-
drocarbon related to the January, 1988 spill.
Table A. 13
Summary of Soil Analytical Results — 1990 Season
Area
Basin A
Basin B
Basin C
Basin D
Wooded Area
Asphalt Area
Terminal Average
Number
Of Grids
73
197
233
100
119
57
779
Average TPH (mg/kg)
01/02/88
10,851
18,838
10,851
14,488
8,516
15,143
12,988
12/31/90
377
395
439
442
379
928
450
A.38
-------
Figure A.21
Closure Sampling Record
Appendix A
<50
610
400
<50
740
530
700
810
500
640
650
430
400
900
500
460
330
370
930
570
820
720
640
720
710
780
800
600
310
470
430
870
500
720
600
460
600
120
190
640
420
930
570
650
430
430
500
160
370
310
170
530
680
360
320
430
150
150
150
190
230
830
350
360
210
180
190
78
81
160
290
720
540
340
310
320
79
71
190
170
470 \
540 \
360
980
310
300
78
92
120
540
940
270
340
180
130 /
300/
180 /
Tank
TPH(
\
\
740 \
770/
HO/
/
Figure A.22
Closure Sampling Record
Drainage Swale
Asphalt Plant North Area
TPH (mg/kg)
A.39
-------
Appendix B
B
GLOSSARY OF TERMS
Abiotic Abiotic means nonbiological.
Aerobic Aerobic means in the presence of oxygen.
Anaerobic Anaerobic means in the absence of oxygen.
Anoxic Anoxic means in the absence of oxygen.
Aquifer An aquifer is an underground geologic formation that is filled with
water and which is permeable enough to transmit water to wells and
springs.
Arrhenius plot An equation relating the rate of a chemical reaction to tem-
perature can be expressed by an Arrhenius plot; in general, reaction rates
increase by a factor of two or three for each 10°C rise in temperature; de-
partures from the Arrhenius plot are indicative of certain processes, e.g.
reactions controlled by enzymes.
Bioaugmentation Bioaugmentation is the repeated addition of microbes
that degrade but do not grow on the contaminants in a bioremedial process.
Biogeochemistry Biogeochemistry is the involvement of living organisms
with the earth's chemistry. Biogeochemical processes control the global
cycling of the biologically important elements carbon, nitrogen, phospho-
rus, and sulfur, as well as the cycling of a variety of trace elements.
Bioreactor Bioreactors are ex situ biodegradation treatment facilities; many
types of bioreactors have been developed for various treatment conditions;
e.g., liquid-type reactors include suspended-growth, fixed-film, and sub-
merged fixed-film reactors. Solids can be treated using a slurry phase-type
reactor.
Bioremediation Bioremedition is the process by which organic or inor-
ganic waste is biologically degraded or transformed, usually to innocuous
B.I
-------
Glossary of Terms
materials; the process can occur naturally or can be enhanced by adding an
electron acceptor, nutrients, or other factors which would otherwise limit
the natural biodegradation rate.
Bioventing Bioventing is the process by which contaminants in the unsat-
urated zone are removed by volatilization and biodegradation as air or oxy-
gen is supplied by vacuum extraction and/or injection.
Bodenfilter A bodenfilter is a biofilter consisting of one or more beds of
biologically active material, primarily mixtures based on compost, peat or
soil through which gas is vented. Aerobic degradation of contaminants in
the gas occurs in the biofilm if microorganisms are present that can metabo-
lize them.
Cometabolism Cometabolism is the process by which an organic com-
pound is metabolized or transformed but not used for growth.
Composting Composting is a procedure for accelerating biodegradation of
contaminants at elevated temperatures by aerating and adding bulking
agents and possibly nutrients to waste in a compost pile.
Congener A chemical compound closely related to another in molecular
structure is a congener; e.g, PCB molecules may differ from each other by
the location or number of chlorine atoms attached to the carbon ring struc-
ture.
Constitutive mutant or strain A constitutive mutant or strain is a population
of organisms in which the enzyme controlling a process is functional with-
out having to be induced (see enzyme).
Darcy's Law Darcy's Law is expressed by an equation that can be used to
compute the quantity of water flowing through an aquifer; hydraulic con-
ductivity is a parameter in the equation.
Dehalogenation Dehalogenation is a chemical reaction in which a halogen
atom is replaced by a proton; reductive dehalogenation (dechlorination) of
trichloroethylene results in the removal of one chlorine and leaves as the
other product dichloroethylene.
Denitrification Denitrification is a microbial process that uses nitrate as a
terminal electron acceptor to oxidize an organic compound.
Enzyme An enzyme is a proteinaceous biochemical catalyst; a biological
reaction is controlled and accelerated by an enzyme; most enzymes are
B.2
-------
Appendix B
substrate-specific with a characteristic affinity for attaching to the substrate
to mediate the reaction. Induction of an enzyme is a process by which syn-
thesis of enzymes required to metabolize a substrate is initiated by the inac-
tivation of a represser substance.
Ex situ Ex situ refers to the execution of an environmental cleanup by re-
moving the contaminants from the existing location to another matrix, ei-
ther on- or off site for treatment, e.g., bioreactors.
Geochemistry Geochemistry is the science of the chemical composition of
the earth and the processes that cause the distribution of the elements.
Groundwater Groundwater is subsurface water, especially that water found
in the saturated zone below the water table, and is found in formations
known as aquifers.
Heterotrophic organisms Heterotrophic organisms derive their energy and
carbon required for growth from organic compounds, as opposed to au-
totrophic organisms which can grow and obtain energy from inorganic nu-
trients.
Hydrocarbon A hydrocarbon is a chemical that is composed of only carbon
and hydrogen, e.g., gasoline, benzene, methane.
Hydraulic conductivity Hydraulic conductivity is a coefficient of propor-
tionality describing the rate at which water can move through a permeable
medium (See Darcy's Law).
Inorganic substance An inorganic substance is a chemical that does not
contain carbon-to-hydrogen bonds, e.g. metals, nitrate, phosphate.
In situ In situ means "in place" in the environment.
Land treatment Land treatment of wastes and environmental contaminants
is a general term for processes that accelerate biodegradation of contami-
nants in situ or ex situ by aerating and possibly adding nutrients to contami-
nated surface soil or contaminated material that has been applied to soil.
Metabolism Metabolism is the chemical reaction in an organism resulting
in the breakdown of substances to produce energy and growth; products of
the metabolic process are called metabolites; catabolic metabolism is a
degradative process.
B.3
-------
Glossary of Terms
Methanogen Methanogens are strict anaerobic bacteria that produce meth-
ane while growing on CO2-type substrates (CO2, HCOOH, CO), methyl-
containing compounds (CH3OH, CH3NH3+, (CH3)2NH2+, (CH3)3NH+) or
acetic acid.
Methanotroph (methylotrophs) Methanotrophs are aerobic bacteria that
use methane and other one- and two-carbon compounds as a carbon source;
these organisms may also be capable of cometabolic processes that degrade
contaminants, such as chlorinated ethenes.
Michaelis-Menten kinetics Michaelis-Menten kinetics describe reaction
rate dynamics of enzyme-controlled chemical processes; the equation ex-
presses the relationship between reaction rate and substrate concentration
and is dependent on the relative affinity of an enzyme for a substrate.
Mineralization Mineralization is the process of decomposition of an or-
ganic compound to inorganic products.
Monod kinetics Monod kinetics is the application of the Michaelis-Menten-
type equation to describe the growth dynamics of a culture of microorgan-
isms, where growth rate is related to substrate concentration.
Mutant strain A mutant strain is a population of organisms genetically
different from their ancestors due to chromasomal changes.
Organic compound An organic compound contains carbon; the exception,
carbon dioxide, is generally considered inorganic.
Oxidation, oxidant In a chemical reaction, oxidation is the the loss of an
electron by a reactant.
Oxygenase enzyme An oxygenase enzyme incorporates one or both atoms
of molecular oxygen (O2) into an organic compound to yield hydroxyl
groups.
Permeability Permeability, or intrinsic permeability, is a property of a po-
rous medium independent of the nature of the liquid and pertains to the
relative ease with which a porous medium can transmit a liquid under a
hydraulic or potential gradient.
Plating technique Plating is a microbiological technique to enumerate
microorganisms which results in a "viable count;" the microorgani sms are
uniformly distributed in or on a growth medium which commonly contains
B.4
-------
Appendix B
agar in shallow containers called Petri dishes; the inoculated microorgan-
isms grow into colonies visible without magnification.
Porosity Porosity is a property of porous medium; it is the ratio of the vol-
ume of void spaces in a rock or sediment to the total volume.
Redox Redox is a chemical process, an oxidation-reduction reaction; the
loss (oxidation) and gain (reduction) of electrons among reactants affect the
charge of the medium and is expressed as electrical potential, Eh (volts) for
a particular hydrogen ion concentration, pH; most oxidizing environments
have a Eh level greater than +400 mV.
Reduction, reductant In a chemical reaction, reduction is the gain of an
electron by a substance. A reductant is a substance that induces the gain of
an electron by another substance.
Remedial Investigation (RI) The remedial investigation is part of the
Superfund remediation process of a hazardous waste site in which the site
conditions are characterized and appropriate treatability studies are per-
formed to evaluate potential cleanup technologies; risk assessment of expo-
sure and toxicity assessment are initiated during this time.
Reynold's number The Reynold's number is a dimensionless number used
to determine whether flow will be turbulent or laminar; in laminar flow the
particles follow paths that are smooth, straight, and parallel to the channel
walls.
Saturated zone The saturated zone is the subsurface region, an aquifer, in
which the voids in the rock or solids are filled with water at a pressure
greater than atmospheric. The water table is the top of the saturated zone in
an unconfined aquifer.
Seepage velocity The seepage velocity, also called average velocity, is the
actual rate of movement of fluid particles through porous media. The ve-
locity, taking into account the effective porosity of the medium, is calcu-
lated using Darcy's Law.
Soil Vapor Extraction (SVE) Test Soil vapor extraction test is a site charac-
terization test to pull vapor from the unsaturated zone of the subsurface to
analyze for volatile contaminants to evaluate the type and extent of con-
tamination.
B.5
-------
Glossary of Terms
Soil venting Soil venting is a type of remedial technology applied to the
unsaturated zone of the subsurface in which vapor is forced out of the for-
mation, or displaced with air and the volatile organic contaminants are re-
moved from the effluent by sorption.
Stoichiometry Stoichiometry is the method of calculation of the combining
amounts of reactants and products in a chemical reaction.
Surfactant Surfactant is a term formed from the contraction of "surface-
active agent;" surfactants are ionic or nonionic organic compounds with
amphipatic structure, i.e, the molecule is composed of groups of opposing
solubility tendencies, typically an oil-soluble (hydrophobic) and a water-
soluble, ionic or polar group (hydrophilic) section,
Transmissivity Transmissivity refers to the rate at which water is transmit-
ted through an aquifer under a particular hydraulic gradient due to the prop-
erties of the porous media and its thickness.
Unsaturated zone The unsaturated (vadose) zone is located between the
land surface and the water table. It includes the root zone (rhizosphere),
intermediate zone, and capillary fringe of an aquifer. The unsaturated zone
pore spaces contain water at less than atmospheric pressure, as well as air
and other gases.
Xenobiotic compound A xenobiotic compound is a synthetic compound.
B.6
-------
Appendix C
C
LIST OF ACRONYMS
BTEX Benzene, toluene, ethylxylene, xylenes
DNAPL Dense nonaqueous phase liquid
EPA Environmental Protection Agency
GC-FID/PID Gas chromatography - flame ionization detector/photoion-
ization detector
GC-FID/MS Gas chromatography - flame ionization detector/mass spec-
trophotometer
HOPE High density polyethylene
LNAPL Less dense nonaqueous phase liquid
MTBE Methyl tertiary butyl ether
NAPL Nonaqueous phase liquid
PAH Polynuclear or polycyclic aromatic hydrocarbon
PCE Perchloroethylene (Tetracholorethylene)
PCP Pentachlorophenol
PVC Polyvinylchloride
RCRA Resource Conservation and Recovery Act 1976
RI Remedial investigation
SVE Soil vapor extraction
TCE Trichloroethylene
TKN Total Kjeldahl nitrogen
TNT trinitrotoluene
TOC Total organic carbon
C.I
-------
List of Acronyms
TP Total phosphorus
TPH Total petroleum hydrocarbon
VES Vapor extraction system
ZOI Zone of Incorporation
C.2
-------
Appendix D
D
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American Petroleum Institute. 1983. Land treatment practices in the petroleum
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Apajalahti, J.H.A. and M.S. Salkinoja-Salonen. 1986. Dechlorination andpara-
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D.20
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THE WASTECH® MONOGRAPH SERIES ON
INNOVATIVE SITE REMEDIATION TECHNOLOGY
WASTECH® is a multiorganization effort which joins in partnership the Air and
Waste Management Association, the American Institute of Chemical Engineers, the
American Society of Civil Engineers, the American Society of Mechanical Engineers,
the Hazardous Waste Action Coalition, the Society for Industrial Microbiology, and the
Water Environment Federation, together with the American Academy of Environmental
Engineers, the U.S. Environmental Protection Agency, the U.S. Department of Defense
and the U.S. Department of Energy.
A Steering Committee composed of highly respected members of each participating
organization with expertise in remediation technology formulated and guided the
project with project management and support provided by the Academy. Each
monograph was prepared by a task group of five or more recognized experts. Their
initial manuscript was subjected to an extensive peer review prior to publication. This
1994 series includes:
Vol 1 - BIOREMEDIATION
The Principal Authors include- Calvin H. Ward,
Ph.D., Chair, Professor & Chair of Environmental
Science & Engineering, Rice University; Raymond C.
Loehr, Ph.D., P.E., DEE, Civil Engineering, University
of Texas; Robert Morris, Ph.D., Technical Director,
Eckenfelder, Inc.; Evan Nyer, Vice President, Techni-
cal Resources, Geraghty & Miller, Inc.; Michael
Piotrowski, Ph.D.; Jim Spain, Chief, Environmental
Biotechnology, AFESCA/RAVC, John Wilson, Ph.D.,
Process & Systems Research Division, U.S Environ-
mental Protection Agency.
Vol 2 - CHEMICAL TREATMENT
The Principal Authors include: Leo Weitzman,
Ph.D., Chair, President, LVW Associates, Inc.; Kim-
berly Gray, Ph.D., Assistant Professor of Civil Engi-
neering & Geological Sciences; Robert W. Peters,
Ph.D., P.E., DEE, Environmental Systems Engineer,
Argonne National Laboratory, Charles Rogers, Ph.D.,
Senior Research Scientist, USEPA Risk Reduction
Engineering Laboratory; John Verbicky, Ph.D.,
Chemfab Corporation
Vol 3 - SOIL FLUSHING/SOIL WASHING
The Principal Authors include: Michael J. Mann,
P.E., Chair, President, Alternative Remedial Tech-
nologies, Inc.; Donald Dahlstrom, Ph.D., Department
of Chemical Engineering, University of Utah, Patricia
Esposito, PAK/TEEM, Inc.; Lome Everett, Ph.D.,
Geraghty & Miller, Inc.; Greg Peterson, P.E., Director
of Technology Transfer, CH2M Hill, Inc.; Richard P.
Traver, P.E., General Manager, Bergmann USA.
Vol 4 - STABILIZATION/SOLIDIFICATION
The Principal Authors include: Peter Colombo,
Chair, Manager, Waste Management Research & De-
velopment, Brookhaven National Laboratory; Edward
Barth, P.E., Environmental Engineer, Office of Re-
search & Development, U.S. Environmental Protection
Agency; Paul L. Bishop, Ph.D., P.E., DEE, William
Thorns Professor, Department of Civil & Environmen-
tal Engineering, University of Cincinnati; Jim Buelt,
Staff Engineer, Battelle Pacific Northwest Laboratory;
Jesse R. Connor, Senior Research Scientist, Clemson
Technical Center, Inc.
Vol 5 - SOLVENT/CHEMICAL EXTRACTION
The Principal Authors include: James R. Donnelly,
Chair, Director of Environmental Services &
Technologies, Davy Environmental; Robert C. Ahlert,
Ph.D., P.E., DEE, Distinguished Professor, Rutgers
University; Richard J. Ayen, Ph.D., Director of
Chemical Processing, Chemical Waste Management,
Inc; Sharon R. Just, Environmental Engineer, Engineering-
Science, Inc.; Mark Meckes, Physical Scientist, USEPA
Risk Reduction Engineering Laboratory.
Vol 6 - THERMAL DESORPTION
The Principal Authors include: JoAnn Lighty, Ph.D..
Chair, Assistant Professor of Chemical and Fuel Engi-
neering, University of Utah; Martha Choroszy-
Marshall, Program Manager, Thermal Treatment,
CIB A-GEIG Y; Michael Cosmos, Project Director, Roy
F Weston, Inc.; Vic Cundy, Ph.D., Professor of Me-
chanical Engineering, Louisiana State University;and
Paul De Percin, Chemical Engineer, U.S. Environmen-
tal Protection Agency.
Vol 7 - THERMAL DESTRUCTION
The Principal Authors include: Richard S. Magee,
Sc.D., P.E., DEE, Chair, Executive Director, Hazard-
ous Substance Management Research Center, New Jer-
sey Institute of Technology; James Cudahy, President,
Focus Environmental, Inc.; Clyde R, Dempsey, P.E..
Chief, Thermal Destruction Branch, Office of Research
and Development, U.S. Environmental Protection
Agency, John R. Ehrenfeld, Ph.D., Senior Research
Associate, Center for Technology, Policy, & Industrial
Development. Program Coordinator, Hazardous Sub-
stances Management, Massachusetts Institute of Tech-
nology; Francis W. Holm, Ph.D., Senior Scientist &
Principal Deputy, Chemical Demiliterization Center,
SAIC; Dennis Miller, Ph.D., Science Advisor, U.S.
Department of Energy; Michael Model), Modell De-
velopment Corp.
Vol 8 - VACUUM VAPOR EXTRACTION
Paul Johnson, Ph.D., Chair, Research Engineer,
Shell Development; Arthur Baehr, Ph.D., U.S. Geo-
logical Survey, Water Resources Division; Richard A.
Brown, Ph.D., Vice President, Groundwater Technol-
ogy; Robert Hinchee, Ph.D., Research Leader, Battelle;
George Hoag, Ph.D., Director, University of Connecti-
cut, Environmental Research Institute.
The monographs can be purchased for $49.95 per volume or $349.65 for the entire series (plus shipping
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Annapolis, MD, 21401; phone 410-266-3311, FAX 410-266-7653. MC & VISA accepted.
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