(Soil Biology 1) Owen P. Ward, Ajay Singh (auth.), Dr. Ajay Singh, Dr. Owen P. Ward (eds.)-Applied Bioremediation and Phytoremediation- (2004) (1)

Series Editor: Ajit Varma
Springer-Verlag Berlin Heidelberg GmbH
Ajay Singh • Owen P. Ward (Eds.)
Applied Bioremediation
With 19 Figures and 27 Tables
Direetor R&D
Petrozyme Teehnologies Ine.
7496 Wellington Road 34, R.R.#
Guelph, Ontario NIH 6H9
e-mail: [email protected]
Professor of Mierobial Bioteehnology
Department of Biology
University of Waterloo
Waterloo, Ontario N2L 3G 1
e-mail: [email protected]
Adjunct Faculty Member
Department of Biology
University of Waterloo
Waterloo, Ontario N2L 3G 1
e-mail: [email protected]
ISBN 978-3-642-05908-7
ISBN 978-3-662-05794-0 (eBook)
DOI 10.1007/978-3-662-05794-0
Library of Congress Cataloging-in-Publication Data
Applied bioremediation and phytoremediation I Ajay Singh, Owen P. Ward (eds.).
p. cm. - (Soil biology; v. 1)
Includes bibliographical references and index.
1. Soil remediation. 2. Bioremediation. 3. Phytoremediation.
Ajay,1963- 11. Ward, Owen P., 1947- III. Series.
I. Singh,
TD878.A644 2004
628.5'5 - dc22
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Preface to the Series
Soil is a complex mixture of inorganic matter consisting of mineral particles, organic matter from decaying biomass, mieroorganisms, plants,
and animals. The uppermost soillayer supports the most life, including
plants that send their roots into the topsoil from whieh they derive
essential nutrients. In addition, soil is the horne of many insects, worms,
burrowing animals, invertebrates, and microorganisms. Soil mierobes
interact among themselves, with plants and animals, and provide a
perennial source of organie matter, whieh can be recycled for plant
nutrition. Soil provides the physieal support needed for the anchorage
of the root system and also serves as a reservoir for air, water, nutrients
and mieroorganisms, which are so essential for plant growth. Unfortunately, integrated information on different aspects of the soil biology,
chemistry, physies and topography is lacking. However, such data are
important in order to extend our knowledge of the soil sciences. The
Soil Biology Series, whieh is the first of its kind, deals with the study of
the nature, types and functioning of soil and the inhabiting organisms
(macro- and mieroorganisms).
Soil contains organie matter derived from mieroorganisms, animal
wastes, and plants, whieh are potential sources of various nutrients.
Much of this plant material is recycled by the biochemical activities of
mieroorganisms into usable nutrients, and humus. Humus is a darkcolored soil material that is composed largely of decay-resistant organie
matter. Soil influences a different group of microorganisms including
bacteria, actinomycetes, fungi, viruses, algae, protozoa, nematodes and
several other organisms. The fertility of soil depends not only on its
chemieal composition, but also on the qualitative and quantitative
nature of the microorganisms inhabiting it. Microorganisms carry out
a large number of metabolie activities of immense importance, namely,
nitrogen fixation and chemieal transformation of nitrogen, sulfur,
carbon, iron, phosphorus, etc. They also playa significant role in bioremediation, biodegradation and heavy metal accumulation.
Bioremediation, i.e. the application of biologieal methods, has gained
prominence as an option for soil remediation methods. Biologieal
processes are environmentally compatible and can be integrated with
Preface to the Series
non-biological processes for remediation of environmental pollutants.
Large-scale manufacturing, processing and handling activities in the
petroleum and chemical industries result in the production of a large
number of chemical compounds. Spills and unsafe disposal of these
xenobiotic compounds cause a serious deterioration in environmental
quality. These negative impacts have led to serious responses from the
government and non-governmental organizations. New regulations are
now being implemented requiring remedial action in support of environmental sustainability.
Bioremediation, the main topic ofVolume 1 of the Soil Biology Series,
Applied Bioremediation and Phytoremediation, has been used successfuHy to remediate contaminated sites. It is a rapidly advancing field and
new bio-based remedial technologies are continuing to emerge.
It is my pleasure to work as Editor-in-Chief of the Soil Biology Series,
primarily due to the stimulating co operation of the volume editors. 1
would like to thank Dr. Dieter Czeschlik, Editorial Director Life Sciences,
and Dr. Jutta Lindenborn, Springer-Verlag, for their help and friendly
cooperation during the preparation of this volume. I am grateful to Professor Ajay Singh and Professor Owen Ward for their efforts in compiling the volume in re cord time.
New Delhi, March 2004
Ajit Varma
The huge expansion of the chemical and petroleum industries in the
twentieth century has resulted in the production of a vast array of chemical compounds and materials that have transformed our lives. The
associated large-scale manufacturing, processing and handling activities have caused a serious deterioration in environmental quality and
created threats to human health. These negative impacts have led to
responses and regulations requiring remedial action in support of environmental sustainability.
Application of biotechnological methods through bioremediation,
has gained prominence as an option for soil remediation methods.
Bioremediation is a multidisciplinary approach where biologists, chemists, soil scientists and engineers work as team to develop and implement remediation processes. Bioremediation has now been used
successfully to remediate many petroleum-contaminated sites. However,
there are as yet no commercial technologies commonly used to remediate the most recalcitrant contaminants. Nevertheless, bioremediation
is a rapidly advancing field and new bio-based remedial technologies
are continuing to emerge.
Applied Bioremediation and Phytoremediation, Volume 1 of the series
Soil Biology, addresses a wide range of topics related to applied aspects
of microbial and plant-based technologies for treatment of environmental contaminants. Topics include bioremediation of petroleum
hydrocarbons, explosives, pesticides and metallic pollutants, bioremediation in extreme environments, natural attenuation and phytoremediation of persistent organic contaminants and metals. Chapters dealing
with innovative methods address current bioremediation technologies,
bio filtration and risk-based soil remediation approaches. Volume 2 of
the series Soil Biology, Biodegradation and Bioremediation will focus on
basic microbiological and biochemical pro ces ses in bioremediation.
This book contains contributions from authorities in the areas of
bioremediation and phytoremediation. The authors are from diverse
backgrounds, universities, government laboratories and industry, who
do basic, applied and industrial research. This book should prove to be
useful to under- and post-graduate students of biotechnology, micro-
biology, and soil and environmental sciences and engineering. We hope
that teachers, scientists and engineers, whether in academia, industry
or government, will find the contents, including its practical aspects,
We are grateful to all the authors for their excellent contributions.
Several of our colleagues provided encouragement and help during the
various stages of this editorial work. Continuous support and guidance
provided by Dr. Ajit Varma, Editor-in-Chief of the series Soil Biology,
and Dr. Jutta Lindenborn, Springer-Verlag, during the preparation of
this volume, are highly appreciated.
Guelph, Waterloo, Ontario, March 2004
Ajay Singh and Owen Ward
1 Soil Bioremediation and Phytoremediation -
An Overview ......................................
Owen Ward and Ajay Singh
1 Introduction ....................................
2 Major Environmental Contaminants . . . . . . . . . . . . . . . . . .
2.1 Chemical Contaminants ........................
2.2 Biological Wastes and Contaminants ..............
3 Microbial Transformation of Chemical Contaminants . . . .
4 Phytoremediation ................................
5 Criteria for Selecting Bioremediation as an Option and
for Selecting a Particular Bioremediation Configuration . .
6 Regulations and Public Attitudes ....................
7 Advantages of Bioremediation Approaches
to Environmental Sustainability .....................
8 Conclusions .....................................
References ......................................
2 Biodegradation and Bioremediation of Petroleum
Pollutants in Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Michael H. Huesemann
Introduction ....................................
Types of Petroleum Wastes and Their Composition .....
Soil Biotreatment Technologies . . . . . . . . . . . . . . . . . . . . . .
Loss Mechanisms Other Than Biodegradation . . . . . . . . . .
Optimizing Environmental Conditions . . . . . . . . . . . . . . . .
5.1 Moisture Content .............................
5.2 Oxygen .....................................
5.3 Fertilizers ...................................
5.4 pH .........................................
5.5 Temperature .................................
6 Addressing Other Potential Limitations ...............
6.1 High Contaminant Concentrations ...............
6.2 Presence of Inhibitors . . . . . . . . . . . . . . . . . . . . . . . . . .
6.3 Insufficient Number of Hydrocarbon-Degrading
Microorganisms ..............................
6.4 Lack of Cometabolism .........................
6.5 Inherent Recalcitrance .........................
6.6 Bioavailability Limitations ......................
7 Risk Assessment and Environmentally Acceptable
End Points ......................................
8 Conclusions .....................................
References ......................................
3 Bioremediation of Pesticide-Contaminated Soils
Ramesh C. Kuhad, Atul K. Johri, Ajay Singh, and
Owen P. Ward
1 Introduction ....................................
2 Biodegradation of Pesticides . . . . . . . . . . . . . . . . . . . . . . . .
3 Organochlorines .................................
3.1 Chlorophenoxy Acids ..........................
3.2 DDT........................................
3.3 y-Hexachlorocyclohexane ......................
3.4 Endosulfan ..................................
Organophosphates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Carbamates .....................................
s-Triazines ......................................
Other Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Conclusions and Future Prospects ...................
References ......................................
4 Biodegradation and Bioremediation of Explosives
Jian-Shen Zhao, Diane Fournier, Sonia Thiboutot,
GuyAmpleman, and Jalal Hawari
1 Introduction ....................................
2 Structural Properties and Effect on Biodegradation .....
3 Biodegradation of Cyclic Nitramine Explosives .........
3.1 Biodegradation of RDX and HMX by Anaerobic
Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.2 Biodegradation of RDX Under Aerobic Conditions . . .
3.3 Biodegradation of Polycyclic Nitramine Explosive
CL-20 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.4 Biodegradation of RDX and HMX in Sediments . . . . .
4 Biodegradation of TNT and Other Polynitroaromatic
Explosives ......................................
4.1 Biodegradation of TNT by Aerobic Bacteria ........
4.2 Biodegradation of TNT by Anaerobic Bacteria ......
4.3 Biodegradation of TNT by Fungi .......... " . . . . . .
5 Safety Procedures ................................
6 Conclusions .....................................
References ......................................
5 Biological Treatment of Metallic Pollutants . . . . . . . . . . . . . .
Brendlyn D. Faison
1 Introduction ....................................
1.1 Definition of Scope . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.2 Biologically Relevant Elements . . . . . . . . . . . . . . . . . . .
1.3 Iron Respiration as a Model for Dissimilatory
Metabolism ................................. .
2 Microorganisms as Remediation Tools for Suboxic
Environments ...................................
2.1 Problem Definition. . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.2 Relevant Biological Factors . . . . . . . . . . . . . . . . . . . . . .
2.3 Relationship to Abiotic Factors ..................
3. Practical Aspects of Bioremediation
3.1 Management and Operations ....................
3.2 Hydrological Data Needs .......................
3.3 Physicochemical Data Needs ....................
3.4 Spatial and Temporal Data Needs ................
3.5 Major Technology Needs .......................
4 Conclusions .....................................
References ......................................
6 Phytoremediation of Persistent Organic Contaminants
in the Environment ...........•......................
Saleema Saleh, Xiao-Dong Huang, Bruce M. Greenberg,
and Bernard R. Glick
Fundamentals of Phytoremediation ..................
Physical Remediation Strategies .....................
Phytoremediation Strategies ........................
Advantages and Disadvantages of Phytoremediation . . . . .
Phytoremediation of Organics ......................
5.1 Sources of Organic Contaminants
in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . ..
5.2 Factors That Affect the Uptake of Organic
Contaminants ................................
5.3 Contaminant Classes . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.4 Role of Microorganisms in Soil Bioremediation .....
5.5 Improved Mechanisms to Optimize
Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6 Conclusions ............................. . . . . . . . .
References ......................................
7 Phytoremediation of Metals and Inorganic Pollutants .....
Tomas Macek, Daniela Pavlikova, and Martina Mackova
1 Introduction ....................................
2 Phytoremediation and Rhizoremediation . . . . . . . . . . . . . .
2.1 The Role of the Rhizosphere ....................
2.2 Exudates and Enzymes Released .................
2.3 Phytoremediation Methods .....................
2.4 Artificial Wetlands ............................
2.5 Perspectives Regarding Plants Used for Detoxification
in Chemical Weapon Demilitarisation .............
3 In Vitro Plant Cultures in the Phytoremediation
of Metals .......................................
3.1 Callus and Cell Suspension Cultures ..............
3.2 Hairy Root Cultures ...........................
4 Breeding and Genetic Engineering . . . . . . . . . . . . . . . . . . .
4.1 Methods of Preparation of Transgenie Plants .......
4.2 Phytoremediation of a Mercury-Contaminated
Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..
4.3 Volatilisation of Selenium. . . . . . . . . . . . . . . . . . . . . . .
4.4 Increased Accumulation of Heavy Metals ..........
5 Other Approaches to Improve Phytoremediation
Processes .......................................
5.1 Symbiotic Bacteria ............................
5.2 Mycorrhizal Symbiosis .........................
5.3 Role of Secondary Plant Metabolites
in Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . .
6 Conclusions .....................................
References ......................................
8 Remediation of Organic Pollutants Through
Natural Attenuation ................................
Serge Delisie and Charles W. Greer
1 Definition of Natural Attenuation . . . . . . . . . . . . . . . . . . . .
1.1 Monitored Natural Attenuation (MNA) ............
1.2 Evaluating Natural Attenuation ..................
1.3 Abiotic Processes of Natural Attenuation . . . . . . . . . . .
2 Overview of Intrinsic Bioremediation ......... "......
2.1 Electron Acceptors ..................... "......
2.2 Intrinsic Bioremediation of Petroleum
Hydrocarbons ......................... "......
2.3 Intrinsic Bioremediation of Chlorinated Solvents ....
3 Case Studies of Monitored Natural Attenuation . . . . . . . . .
3.1 Intrinsic Bioremediation of Fuel Hydrocarbons at a
Former Canadian Forces Base, Chatham, New
Brunswick - Case 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.2 Intrinsic Bioremediation of Chlorinated Solvents
at Canadian Force Base, Trenton, Ontario - Case 2 ...
4 Conclusions ...................... . . . . . . . . . . . . . . .
References ......................................
9 Evaluation of Current SoH Bioremediation
Technologies ......................................
Owen P. Ward and Ajay Singh
1 Introduction ....................................
2 Factors Affecting Bioremediation and Choice
of Technology ...................................
2.1 Site Characteristics ............................
2.2 Soil Type ....................................
2.3 Moisture Content .............................
2.4 Nature of Contaminant . . . . . . . . . . . . . . . . . . . . . . . . .
2.5 Time Constraints .............................
2.6 Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.7 Biostimulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.8 Bioaugmentation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.9 Sampling and Monitoring . . . . . . . . . . . . . . . . . . . . . . .
3 Bioremediation Technologies and Their Evaluation . . . . . .
3.1 Natural Attenuation ...........................
3.2 In Situ Subsurface Bioremediation . . . . . . . . . . . . . . . .
3.3 Soil Pile and Composting Techniques .............
3.4 Marine Shoreline and Wetlands Remediation .......
3.5 Land Farming ................................
3.6 Slurry Bioreactors . . . . . . . . . . . . . . . . . . . . . . . . . . . ..
3.7 Phytoremediation .............. . . . . . . . . . . . . . ..
4 Conclusions .....................................
References ......................................
10 Bioremediation of Petroleum Hydrocarbon-Polluted Soils
in Extreme Temperature Environments. . . . . . . . . . . . . . . ..
Rosa Margesin
1 Introduction ....................................
2 Bioremediation in Cold Environments. . . . . . . . . . . . . . ..
2.1 Arctic Solls ..................................
2.2 Alpine Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..
2.3 Antarctic Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..
3 Bioremediation of Desert Soils After the Gulf War ......
3.1 Limiting Factors for Biodegradation ..............
3.2 Field Bioremediation ..........................
4 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References ......................................
11 Innovative Methods for the Biofiltration of Air
Contaminants .....................................
Zarook Shareefdeen
1 Introduction ....................................
2 Non-biological Methods ...........................
3 Biological Methods ...............................
3.1 Classical Biofilters . . . . . . . . . . . . . . . . . . . . . . . . . . . ..
3.2 Biotrickling Biofilters ..........................
3.3 Bioscrubbers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..
4 Biofilter Terminology and Parameters ................
4.1 Empty Bed Residence Time .....................
4.2 Elimination Capacity or Removal Rate ............
4.3 Removal Efficiency . . . . . . . . . . . . . . . . . . . . . . . . . . ..
4.4 Mass Loading ................................
4.5 Media Volume, Media Depth and Footprint . . . . . . . ..
5 Media Properties .................................
6 Feasibility of VOC and Odor Removal in Biofilters ......
7 Modeling of Biofiltration Processes ..................
8 Selected Industrial Applications .....................
9 Conclusions .....................................
References ......................................
12 Risk-Based Remediation of Contaminated Soil . . . . . . . . . ..
Tahir Husain
1 Introduction ....................................
2 Regulatory Standards .............................
3 Formulation ................................... "
3.1 Tier 1 Approach ..............................
3.2 Tier 2 Approach ..............................
3.3 Tier 3 Approach ..............................
4 Transport Models and Remediation Technology
Database .......................................
5 Conclusions .....................................
References ......................................
Subject Index ........................................
List of Contributors
Guy Ampleman
Defense Research Establishment Valcartier,
Department of National Defense, Quebec, Canada, G3J 1XS
Serge Delisie
Biotechnology Research Institute, National Research Council of Canada,
6100 Royalmount Avenue, Montreal, Quebec, Canada H4P 2R2
Brendlyn D. Faison
Department of Biological Sciences,
Hampton University, Hampton, VA 23668, USA
Diane Fournier
Biotechnology Research Institute, National Research Council of Canada,
6100 Royalmount Ave, Montreal, Quebec, Canada, H4P 2R2
Bernard R. Glick
Department of Biology, University of Waterloo,
Waterloo, Ontario N2L 3G1, Canada
Bruce M. Greenberg
Department of Biology, University of Waterloo,
Waterloo, Ontario, N2L 3G1, Canada
Charles W. Greer
Biotechnology Research Institute, National Research Council of Canada,
6100 Royalmount Avenue, Montreal, Quebec H4P 2R2, Canada
Jalal Hawari
Biotechnology Research Institute, National Research Council,
6100 Royalmount Ave, Montreal, Quebec H4P 2R2, Canada
Xiao-Dong Huang
Department of Biology, University of Waterloo, Waterloo, Ontario,
N2L 3G1, Canada
List of Contributors
Michael Huesemann
Pacifie Northwest National Laboratory, Marine Scienee Laboratory,
1529 West Sequim Bay Road, Sequim, Washington 98382, USA
Tahir Husain
Department of Environmental Engineering,
Memorial University of Newfoundland,
St. John's, Newfoundland AlB 3X5, Canada
Atul K. Johri
Channing Laboratory, Harvard Medieal Sehool,
181 Longwood Avenue, Boston, MA 02115, USA
Ramesh C. Kuhad
Department of Microbiology, University of Delhi, South Campus,
New Delhi-110 021, India
Tomas Maeek
Department of Natural Produets
Institute of Organie Chemistry and Bioehemistry
Aeademy of Scienees of the Czeeh Republie
Flemingovo n. 2,166 10 Prague 6, Czeeh Republie
Martina Maekova
Department of Bioehemistry and Mierobiology, Faeulty of Food and
Bioehemieal Teehnology, Institute of Chemieal Teehnology, Prague,
Teehnieka 3,16628 Prague 6, Czeeh Republie
Rosa Margesin
Institute of Microbiology, University of Innsbruek,
Teehnikerstrasse 25, 6020 Innsbruek, Austria
Daniela Pavlikova
Department of Agroehemistry and Soil Seienee, Faeulty of Agronomy,
Czeeh Teehnieal University, Kamyeka 129, 165 21 Prague 6, Czeeh
Saleema Saleh
Department of Biology, University of Waterloo,
Waterloo, Ontario, N2L 3Gl, Canada
Zarook Shareefdeen
Biorem Teehnologies Ine., 7496 Wellington Road 34,
R.R. # 3 Guelph, Ontario NIH 6H9, Canada
List of Contributors
Ajay Singh
Petrozyme Technologies Inc.,
7496 Wellington Road 34,
R.R. # 3, Guelph, Ontario NIH 6H9, Canada
Sonia Thiboutot
Defense Research Establishment Valcartier, Department of National
Defense, Quebec, Canada, G3J lXS
Owen P. Ward
Department of Biology, University of Waterloo,
Waterloo, Ontario N2L 3GI, Canada
Jian-Shen Zhao
Biotechnology Research Institute, National Research Council of Canada,
6100 Royalmount Ave, Montreal, Quebec, Canada, H4P 2R2
Soil Bioremediation and Phytoremediation An Overview
Owen P. Ward l and Ajay Singh2
In the 45-year period of 1930-1975, the global human population has
inereased by approximately 2 billion, rising to 4 billion. A further population inerease of 2 billion oeeurred in the 25-year interva11975-2000
and population is expeeted to reaeh 8 billion by 2020. Population growth
and the resultant development of large high -density urban populations,
together with parallel global industrialization, have plaeed major pressures on our environment, potentially threatening environmental sustainability. This has resulted in the buildup of ehemieal and biologieal
eontaminants throughout the biosphere, but most notably in soils and
Cases of uneontrolled contamination of soil and other media with
these toxie ehernieals emerged, drawing attention to the threats eaused
by these ehemicals to our environment and human health. For example,
the Love Canal area in Niagara Falls, New York State was the site of an
abandoned eanal that was used by the Hooker Chemieal Company as
a loeation for disposal of 22,000 tons of PCBs, dioxins, pesticides and
other ehemieal wastes du ring the 1940s and early 1950's. The site was
then eovered up and aequired by the City of Niagara Falls and subsequently used as a loeation for a sehool and housing. By 1978 toxie ehemicals had leaked from the soil into the basements of area hornes and high
rates of misearriages and birth abnormalities were reported among
There are many other signifieant historical examples, including the
dioxin erisis in Belgium, Italy and Bhopal. In addition, it was observed
Department of Biology, University of Waterloo, Waterloo, Ontario N2L 3G 1,
Canada, e-mail: [email protected], Tel: +1-519-8851211 Ext. 2427, Fax:
2 Petrozyme Technologies Inc., 7496 Wellington Road 34, R.R. #3, Guelph, Ontario N1H
6H9, Canada
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
O.P. Ward and A. Singh
that many of the toxic chemicals have a tendency to eventually transfer
from solid media to aqueous media and then to bioaccumulate in high
lipid-containing species such as fish (Bernard et al. 1999). A further
concern is the growing accumulation of pharmaceuticals in the environment, mostly excreted in urine (Kozak et al. 2001).
The historical foundation for modern environmental biotechnology
lies in the composting of organic wastes into soil fertilizers and conditioners. In its broadest sense bioremediation includes biological treatment of wastewater, sewage, food and agricultural wastes, contaminated
soils and groundwater. While its definition is the subject of debate,
Prince (1998) defines bioremediation as "the process of judiciously
exploiting biological processes to minimize an unwanted environmental impact; usually it is the removal of a contaminant from the biosphere". The term "intrinsic bioremediation" essentially involves taking
no action but rather monitoring a natural process of contaminant
reduction without intervention. Hence, intrinsic bioremediation can
hardly be termed as a technology, but it has met with some success as
a low cost approach.
Perhaps more than any other event, the Exxon Valdez oil spill, off the
coast of Alaska, demonstrated the potential for a large-scale application
of biological processes for cleanup of hydrocarbon-contaminated soil
(bioremediation; Prince et al. 1998). Since that event, the application
of bioremediation as a biotechnological method for soil remediation
has gained prominence as an alternative or used in combination with
physical or chemical treatment methods.
Soil bioremediation processes may be implemented using a variety
of different engineered configurations ranging from in situ subsurface
(unexcavated) processes to application of completely mixed soil slurry
reactor systems for treatment of excavated soils. The technology is
interdisciplinary, involving microbiology, engineering, geology, ecology,
chemistry and perhaps other disciplines. The choice of configuration
applied can shift the relative emphasis among the disciplines. The
common objective in the various processes is to create the necessary
environment to facilitate growth and contaminant degradation by the
appropriate biological organisms. Bioremediation has now been used
successfully to remediate many hydrocarbon-contaminated sites as weIl
as sites containing selected other contaminants (Singh et al. 1999; Van
Hamme et al. 2003).
It has been estimated that non-biological pro ces ses would cost
US$ 750 billion over the next 30 years to remediate all of the known hazardous waste sites in the United States (Pimentel et al. 1997). However,
the cost would be reduced to $75 billion with the use of bioremediation
over the same period of time. According to another estimate, worldwide
Soil Bioremediation and Phytoremediation - An Overview
Natural attenuation
Ce me nt kiln
Air stripping
Thermal desorption
Solvent extraction
Fig. 1. Soil remediation methods
use of bioremediation would cost $14 billion per year compared to the
use of current technologies at $135 billion per year (Hunter-Cevera
1998). Various biological and non-biological methods used in soil remediation are shown in Figure 1.
While there are many aspects of soil-related biodegradation and
bioremediation, this volume will primarily focus on remediation of
chemical contaminants in soil and will not address the issues of
biowaste recycling.
O.P. Ward and A. Singh
Major Environmental Contaminants
There are two major groups of environmental contaminants, namely
chemical and biological wastes that can accumulate in or be transmitted via soil as a result of the population expansion and industrial
intensification described above. The chemical contaminants can be
classified broadly into two groups, organic and inorganic contaminants.
Chemical Contaminants
The huge expansion of the chemical and petroleum industries in the
20th century has resulted in the production of a vast array of chemical
compounds and materials that have transformed our lives. For example,
annual volumes of individual bulk chemicals produced in the United
States range from 5-20 million metric tons for ethylene, propylene, vinyl
chloride, benzene and ethylbenzene and from 1-5 million tons for a
large number of other organic chemicals. Approximately 140 million
tons per annum of synthetic polymers/plastics are produced globally
(Shimao 2001). The global yeady volume of crude oil production is
approximately 72 million barrels per day (West 1996), while the total
wodd refining capacity is 74.4 million barrels per day (West 1996). The
latter huge numbers (ab out 25 billion barrels per year) indicate the scale
and volumes of refined fuels and other oil-based products produced and
used annually. Thus if one assumes that only 1% of these volumes enters
the environment through spills, waste disposal or volatilization, this
amounts to 266 million barrels per annum.
Industrial activities have also resulted in undesired contamination of
soil and other media with heavy metals, so often toxic to human and
animal health. Contamination of soil and solid wastes with high activity radionuclides, such as 235U, 99Tc and 241pU represents an additional
environmental hazard, with the potential for these metals to be radiotoxic to alilife forms (Lloyd and Macaskie 2000). Mention should also
be made here of excessive levels of inorganic fertilizer-related chemicals introduced into soil, such as ammonia, nitrates, phosphates, and
phosphonates, which accumulate, or lead to the contamination of our
water courses through run-off or of our air through volatilization.
Soil Bioremediation and Phytoremediation - An Overview
Biological Wastes and Contaminants
A second category of waste which threatens the environment is biological wastes, including raw and digested sewage (biosolids), raw and
digested animal manures, and vegetable wastes. While these biological
wastes have traditionaHy been recycled into soil for agricultural benefit
increasing urbanization and the continued expansion of cities requires
that these materials be transported over larger and larger distances for
application to farm land. Likewise the continuing shift to intensive livestock farming (factory farms), typicaHy located close to areas of high
population, means that tradition al land disposal practices of animal
wastes are often rendered uneconomic because of high transport costs.
Was te organic products of vegetable and food processing are also candidates for recycling back into soil, but again high volume intensive
factory operations require large areas of land for recycling. Alternative
waste reduction and beneficial use outlets or other economical disposal
approaches are being sought for these wastes.
An additional problem associated with biological wastes relates to the
risks of transmission of infectious diseases when infected materials are
applied to soil. This dimension has become more prominent through
the recent emergence of high profile infectious diseases, such as the
prion, BSE; viruses, such as foot and mouth or West Nile and intestinal
bacterial pathogens. Infected soils can facilitate disease spread through
direct or indirect contact with watercourses, plants, animals or humans.
The introduction to or disposal of recombinant organisms in soil is
a related concern, given the potential for recombinant strain proliferation, modification or recombinant gene transfer processes in the soil.
One smaH outcome arising from this concern is that the potential to use
recombinant organisms in bioremediation processes has been constrained. There is surely a need to understand better the nature of soils
as host media for retention, survival and propagation of these biological organisms/vectors.
Microbial Transformation of Chemical Contaminants
Processes for metabolism of organic contaminants may be differentiated depending on the organism's ability or otherwise to derive energy
or carbon from the transformation for growth. With primary substrates
the ceH gains energy and metabolites which can be used for ceH maintenance, division and growth. In the case of secondary substrates the
O.P. Ward and A. Singh
ceU derives energy from transformation of the contaminant but cannot
use the carbon for growth. Where the contaminant is transformed but
provides the ceU with neither energy nor carbon, the process is
described as cometabolism. Dissolved oxygen is the common electron
acceptor of aerobic bio degradation and bioremediation processes.
However, aU respiring microbes require a terminal electron acceptor
and in some cases an organic contaminant may act as an electron acceptor. N0 3-, S042-, Fe3+, CO 2 and some other substrates can act as terminal
electron acceptors under anaerobic conditions.
The organic contaminants for bioremediation inelude the alkanes,
monoaromatics, polycyelic aromatic compounds, chlorinated hydrocarbons, ineluding the polychlorinated biphenyls, nitroaromatics and
nitrogen heterocyeles, synthetic plastic and other man-made materials.
Often the organic contaminants are present as complex mixtures of
chemicals, as are present in petroleum. Processes for remediation of
petroleum poUutants, pesticides and explosives in soil are discussed in
Chaps. 2, 3 and 4.
Unlike organic chemicals, where there is potential for bioremediation
processes to convert the contaminants to harmless products such as CO 2
and H20, biological systems can at best only be exploited to concentrate
heavy metals rather than destroy them. Development ofbioremediation
processes for treatment of heavy met als and other inorganic compounds
have lagged behind use of bioremediation for degradation of organic
contaminants. Hence, where soils are contaminated with a mixture
of organic and inorganic contaminants, bioremediation cannot be used
as a one-step remediation system and typicaUy a treatment train is
required comprising separate steps for organic destruction (or recovery) and for removal/separation of inorganic compounds. However,
many microbes can transform met als from toxic to non-toxic species or
alter their solubility or availability (Harayama 2001). Some metal resistant organisms appear to have metal detoxification apparatus which may
have significant potential in future bioremediation processes. Potential
processes often exploit micro-organisms which have natural indigenous
roles in biogeochemical cyeling of metals (Lloyd and Lovley 2001). A
geneticaUy engineered heavy metal tolerant Ralstonia eutropha strain
decreased the toxic effects of Cd2+ on growth of tobacco plants (Valls et
al. 2000) and potentially useful strategies for metal detoxification are
being explored in E. coli as a result of recent genomic transcription
research (Brocklehurst and Morby 2000). Chapter 5 provides more
detailed discussion on the role of microorganisms in treatment of inorganic compounds.
Soil Bioremediation and Phytoremediation - An Overview
The phytoremediation method uses various plants to extract, contain,
immobilize or degrade contaminants from soil and water. Some plants
can remove contaminants from soil by direct uptake, followed by
subsequent transformation, transport and accumulation in a nonphytotoxic form. The diverse approaches in phytoremediation include
phytodegradation, phytoextraction, phytostabilization, phytovolatilization and rhizofiltration. Phytoremediation is still actively being
researched and plant-microbial associations seem to be the key to
enhancing removal of inorganic and organic pollutants. Genetic engineering methods are widely used for the improvement of different agricultural crops. A similar approach has been used for creating transgenic
plants with improved detoxification capabilities under field conditions.
Phytoextraction and rhizofiltration processes have shown pro mise for
commercialization. Phytoremediation technologies are most appropriate for large areas of low and moderately contaminated soils where the
application of conventional remediation technologies would be prohibitively expensive. Chapters 6 and 7 provide detailed discussion on
various phytoremediation strategies for treatment of organic and inorganic pollutants.
Criteria tor Selecting Bioremediation as an Option and
tor Selecting a Particular Bioremediation Configuration
Prince (1998) provided guidance for determining the suitability of
bioremediation as a clean-up option and questions to be addressed. If
the site has been contaminated for a considerable period of time, some
biodegradation of the easier-to-degrade compounds may have taken
place and the residual, perhaps more-difficult-to-degrade compounds
have to be degraded. Whether the microbial population at the site can
degrade the contaminants needs to be determined and the factors limiting growth of the population and degradation of the contaminants
need to be identified and the appropriate conditions applied. For inorganic contaminants, it needs to be determined that their toxicities can
be reduced by a change in redox or other physical or chemical state
and/or that microbes or plants can effectively extract, concentrate or
immobilize the contaminant from the soil matrix. For either organic or
inorganic contaminants, it is necessary to know that they are present at
O.P. Ward and A. Singh
non-tmQc concentrations that will also facilitate appropriate rates of
transformation and allow the clean-up criteria to be achieved. Very
often bioremediation of recalcitrant xenobiotic chemicals requires a
combination of chemical (or physical) and biological steps to increase
the efficacy of contaminant destruction (Paszczynski and Crawford
1995; Zhao et al. 2001).
Risk assessment is an em erging multi-disciplinary scientific practice
used to evaluate health and ecological risks posed by chemical contaminants, also known as "risk agents". Such evaluation helps in devising
risk-based management plans to achieve target risk reduction. However,
to develop a cost-effective remedial action plan, there is a need to introduce a systematic and scientifically sound methodology to assess the
associated risks at a site and identify appropriate remediation technologies. Chapter 12 deals with the risk-based remediation approach
for contaminated soils.
Hughes et al. (2000) evaluated some of the various bioremediation
configurations as options for treatment of different classes of chemicals.
Natural attenuation and electron donor delivery were considered as
options for treatment of chlorinated solvents and biostimulation was an
option for treatment of chlorinated solvents and phenols. Bioventing
was considered to be an option for remediation of PAHs. Land treatment or composting were options for nitroaromatics, phenols and PAHs
and bioslurry processes were considered suitable for treatment of all of
the above mentioned chemicals. All of the treatment methods, except
electron donor delivery, were options for monoaromatic hydrocarbon
bioremediation. Bioremediation was preferred for clean-up of contamination around leaking underground storage tanks in the United States
(Van Hamme et al. 2003). The specific bioremediation process applied
was (% of sites): natural attenuation, 28%; biopiles, 16%; landfarming,
7% and bioventing, 0.8%. Preferred non-bioremediation approaches
used in clean-up of leaking USTs were (% of sites): landfilling, 34%; soil
vapor extraction, 9%; thermal desorption, 3%; incineration, 2%; and soil
washing, 0.2%.
Remediation of organic contaminants through natural attenuation is
discussed in Chapter 8. The performance capability of the different
bioremediation configurations is further discussed in Chapter 9.
Chapter 10 describes bioremediation treatments of cold soils and desert
soils contaminated with petroleum hydrocarbons, as temperature plays
an important role in bioremediation processes.
In soi! bioremediation processes large quantities of VOCs may get
transferred to the atmosphere, thereby creating hazards to human
health. The effluent air streams from bioventing, biosparging and engineered soil pile systems are often heavily contaminated with VOCs. Air
Soil Bioremediation and Phytoremediation - An Overview
biofiltration, as discussed in Chapter 11, provides a cost-effective biological remedy to this problem and, as such, represents an important
process component in many soil bioremediation systems.
Regulations and Public Attitudes
The Love Canal environment al disasters, involving soil contamination, lead to the establishment of the Comprehensive Environmental
Response Compensation and Liability Act (CERCLA) (Cole 1994).
Together with the later Superfund initiatives, CERCLA provided a basis
for regulating the disposal of hazardous waste and the cleaning up of
sites contaminated by chemical pollutants in the United States. Other
industrial jurisdictions followed suit in establishing frameworks for
protecting and remediating the environment from hazardous waste
contamination. In addition to legislation, public pressure on chemical
and pharmaceutical companies in the Uni ted States appears to be promoting a greater commitment to environmental management. Zehnder
and Wulff (1999) provided data to indicate that investments in US
companies with better environmental management systems overall
performed better than similar investments in companies with poorer
environmental management approaches.
Advantages of Bioremediation Approaches
to Environmental Sustainability
It has been claimed that microbial biocatalytic processes and biomin-
eralization generally represent low cost remediation alternatives, for
example costing approximately one tenth the cost of incineration
(Grommen and Verstraete 2002). Additionally, bio degradation processes are argued to be flexible and adaptable to variable environmental conditions. Many synthetic chemicals, when first introduced to the
environment appeared recalcitrant. However, over time biological
organisms appear to have evolved which can degrade many of these
chemicals (Ralebitso et al. 2002). Another claimed advantage of bio remediation has more to do with public perception than with performance: biological processes are perceived as being environmentally
benign whereas incineration and more energy and equipment intensive
processes are perceived as being more environmentally polluting. Other
advantages often offered are that bioremediation processes can be
O.P. Ward and A. Singh
implemented on site, indeed sometimes in situ (without excavation) and
that the process is applicable to dilute or widely diffused contaminants
(Iwamoto and Nasu 2001).
The dis advantage most frequently cited for soil bioremediation
processes is that they are often very slow processes and frequently
desired end points may not be achieved. Failures occur more frequently
in more passive processes. With little intervention, removal of soil contaminants by pro ces ses of natural attenuation may occur and acceptable
endpoints may take many years to be reached, if indeed they are
reached. Natural attenuation processes in bioremediation are discussed
in Chapter 8. Intrinsic remediation is an acceptable low cost option
where the contaminated soil or associated groundwater appears to pose
little or no threat to the surrounding environment or to human or
animal health. There is also a concern that the expectation of gradual
reduction of contaminants through natural attenuation may be used as
a mechanism for postponing a more rigorous and perhaps more costly
clean-up approach. There can be many reasons for the slow bioremediation rates and failures principally that the environmental conditions
present are sub-optimal for selection and growth promotion of the
degrading strains. In addition, the kinetics of microbial growth and
bio degradation are such that as contaminant concentrations decline so
also do the rates of their further degradation.
A fundamental aspect of biodiversity is that organisms appear to have
evolved to utilize free energy, where ever it is generated as a substrate.
Soils and sediments represent a substantial system for support of microbial diversity. The intrinsic heterogeneity of these materials is the
essence of this diversity and calls for us to develop a greater understanding of the molecular and cellular interactions which occur in these
ecosystems. Indeed one proposed yardstick for measuring environmental quality of such ecosystems and to address handling of wastes is
to operate in a mann er which preserves or increases biodiversity. The
effectiveness of this approach was demonstrated at a petrochemical
complex in Brazil over more than 20 years (Verstraete 2002).
Rapid advances in molecular techniques should allow us to more
effectively understand the various dimensions related to microbial
biodegradation of chemical contaminants, microbial community structure in soil, the impacts of contaminants introduced to the environment
and the application ofbioremediation processes. Rence, it is certain that
Soil Bioremediation and Phytoremediation - An Overview
environmental biotechnology and bioremediation processes will continue to evolve and be applied as an effective mechanism for remediating, preserving and sustaining the soils that are such an integral part of
our natural world.
Bernard A, Hermans C, Broeckaert F, DePoorter G, DeCock A, Houins G (1999) Food
contamination by PCBs and dioxins. Nature 401:231-232
Brocklehurst KR, Morby AP (2000) Metal-ion tolerance in Eseheriehia eoli: analysis of
transcription profiles by gene-array technology. Microbiology 146:2277-2282
Cole GM (1994) Assessment and remediation of petroleum contaminated sites. Lewis,
Boca Raton, Florida, pp 287-312
Grommen R, Verstraete W (2002) Environmental biotechnology: the ongoing quest.
J BiotechnoI98:113-123
Harayama S (2001) Environmental biotechnology. Curr Opin BiotechnoI12:229-230
Hughes JB, Neale CN, Ward CH (2000) Bioremediation. In: Enclyclopedia of Microbiology, 2nd Edn, Academic Press, New York, pp 587-610
Hunter-Cevera JC (1998) The value of microbial diversity. Curr Opin Microbiol
Iwamoto T, Nasu M (2001) Current bioremediation practice and perspective. J Biosci
Bioeng 92:1-8
Kozak R, Dhaese I, Verstraete W (2001) Pharmaceuticals in the environment: focus on
17 alpha-ethinyloestradiol. In: Kummerer K (ed) Pharmaceuticals in the environment: sourees, fate, effects and risks. Springer Verlag, Berlin, pp 49-66.
Lloyd JR, Macaskie LE (2000) Bioremediation of radioactive metals. In: Lovley D (ed)
Environmental microbe metal interactions, ASM Press, Washington DC, pp 277-327
Lloyd JR, Lovley D (2001) Microbial detoxification of metals and radionuclides. Curr
Topics Biotechnol 12:248-253
Paszczynski A, Crawford RL (1995) Potential for bioremediation of xenobiotic compounds by the white-rot fungus Phaneroehaete ehrysosporium. Biotechnol Prog 11:
Pimentel D, Wilson C, McCullum C, Huang R, Dwen P, Flack J, Tran Q, Saltman T,
Cliff B (1997) Economic and environmental benefits of biodiversity. Bioscience
Prince RC (1998) Bioremediation. In: Kirk-Othmer Encyclopedia of chemical technology. Supplement to the 4th Edn, Wiley, New York, pp 48-89
Ralebitso TK, Senior E, HW van VerseveId (2002) Microbial aspects of arazine degradation in natural environments. Biodegradation l3: 11-19
Shimao M (2001) Biodegradation of plastics. Curr Opin Biotechnol12:242-247
Singh BK, Kuhad RC, Singh A, LaI R, Tripathi KK (1999) Biochemical and molecular
basis of pesticide degradation. Crit Rev BiotechnoI19:197-225
Valls M, Atrian S, deLorenzo V, Fernandez LA (2000) Engineering a mouse metallothionen on the cell surface of Ralstonia eutropha CH34 for immobilization of heavy
metals in soil. Nature BiotechnoI18:661-665
Van Hamme, JD, Singh A, Ward OP (2003) Recent advances in petroleum microbiology.
Microbiol Mol Biol Rev 7:503-549
Verstraete W (2002) Environmental biotechnology for sustainability. J Biotechnol 94:
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Ward OP, Singh A, Van Hamme J (2003) Accelerated biodegradation of petroleum
hydrocarbon waste. J Ind Microbiol Biotechnol 30:260-270
West J (1996) International Petroleum Encyclopedia, PennWell Publishing, Tulsa,
pp 1-15
Zehnder AJB, WulffH (1999) Business and the environment. Nature Biotechno117:25
Zhao J-S, Ward OP, Lubicki P, Cross JD, Huck P (2001) Process for degradation of
nitrobenzene: combining electron beam irradiation with biotransformation.
Biotechnol Bioeng 73:306-312
Biodegradation and Bioremediation
of Petroleum Pollutants in Soil
Michael H. Huesemann 1
Petroleum and its derived fuels, as a result of their widespread worldwide use, are probably the most ubiquitous organic pollutants found in
soil. Each year, approximately 40,000 barrels or equivalently 1,680,000
gallons of crude oil are spilled on land due to pipeline failures (Salanitro 2001). There are more than 200,000 underground storage tanks in
the US alone that have leaked gasoline and other fuels into vadose zone
soils (Bedient et al. 1994). A variety of physical and chemical treatment
technologies such as incineration, thermal desorption, soil washing, and
solvent extraction have been developed and tested for removing petroleum hydrocarbons from soils (Stegmann et al. 2001). Despite the fact
that these technologies are successful in cleaning contaminated soils,
they not only often destroy the soil structure or render the soil biologically impoverished, sometimes even completely sterile, but also are
usually cost-prohibitive. By contrast, biological soil treatment technologies such as land farming, composting or biopile treatment, slurry
bioremediation, and bioventing are not only cost-effective, but generally reduce hydrocarbon pollutant levels sufficiently to bring about a
significant improvement in soil quality as indicated by marked reductions in ecotoxicity. As a result, bioremediation is the most commonly
used treatment technology for petroleum polluted soils (Alexander
1994; Baker and Herson 1994; Cookson 1995).
During bioremediation, petroleum hydrocarbons are converted by
naturally occurring or indigenous soil microorganisms to carbon
dioxide, water, bacterial cells (biomass), and humus materials. Numerous factors are known to affect both the rate and the extent of
hydrocarbon bio degradation in contaminated soils. These include soil
Michael H. Huesemann, Pacific Northwest National Laboratory - Marine Sciences
Laboratory, 1529 West Sequim Bay Road, Sequim, Washington 98382, USA, e-mail:
[email protected], Tel: +1-360-6813618, Fax: + 1-360-6813699
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
M.H. Huesemann
properties such as moisture content, aeration, nutrient status, pH, and
temperature as weIl as waste characteristics such as the concentration
and molecular structure of hydrocarbon compounds or classes, the
presence of inhibitors and cometabolic substrates, and the degree of
contaminant sequestration which often leads to serious bioavailability
limitations, particularly in aged soils. It is the objective of this chapter
to outline a strategy for optimizing the hydrocarbon bioremediation
process by adjusting the various operational parameters so that none of
them become a limiting factor during treatment.
Types of Petroleum Wastes and Their Composition
Hydrocarbon-contaminated soils are found wherever crude oil or its
derivative fuels have been spilled during exploration, production,
refining, transport, or storage. During exploration and production,
crude oil is occasionally released onto soil as a result of pipeline failures
or via disposal of drilling muds or tank bottoms from storage facilities
(Salanitro 2001). Oily waste sludges that accumulate during refining
operations have conventionally been spread on nearby agricultural
fields for biotreatment by "land farming". Many refined fuels such as
gasoline, diesel, and jet fuel are often accidentally released during transportation and storage. For example, a fuel oil pipeline might rupture
and release large amounts of petroleum product into the surrounding
subsurface soil. Similarly, several hundred thousand underground
storage tanks, particularly those found at gas stations and military
installations, have leaked gasoline and other fuels into the vadose zone
and have often contaminated underlying aquifers (Bedient et al. 1994).
erude oil contains tens of thousands of different hydrocarbon compounds which for simplicity may be categorized as follows: (1) alkanes
(e.g., decane, hexadecane, etc.), (2) iso- or branched-chain alkanes (e.g.,
pristance, phytane, etc.), (3) cycloalkanes (e.g., cydohexane, decalin,
etc.), (4) mono-aromatics (e.g., benzene, toluene, ethylbenzene, xylenes
[BTEX], etc.), (5) polyaromatics or polynuclear aromatic hydrocarbons
(PARs) (e.g., naphthalene, phenanthrene, pyrene, benzo(a)pyrene, etc.)
and (6) heterocyclies or polars that contain nitrogen, sulfur, or oxygen
(e.g., resins, asphaltenes, etc.). Light crude oils (high API gravity) generally contain more mono-aromatics and fewer heterocyclics than heavy
viscous crude oils of low API gravity. During the refining process, crude
oil is separated into different fuel fractions that are characterized by
different hydrocarbon dass compositions. Gasoline consists primarily
of normal and iso-alkanes and mono-aromatics with carbon numbers
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
ranging from Cs to Cu, while diesel fuel eontains mainly normal and
eylco-alkanes (CS-C21 ) and polyaromaties (Salanitro 2001). Byeontrast,
Bunker C fuel (marine diesel #6) eontains large fraetions ofhigh moleeular weight PARs and heteroeydies (C9 -C60+) while motor/lubrieating
oil eonsists primarily of alkanes and eydoalkanes (C IO -C24 ) (Salanitro
2001). Considering that eompounds within a given hydroearbon dass
generally exhibit similar suseeptibilities toward bio degradation (see
Seet. 6.5 below), the overall biodegradability of a petroleum-eontaminated soil strongly depends on the type of erude oil or fuel that was
spilled. For example, a soil eontaminated with Bunker C, rich in recalcitrant five- and six-ring PAHs and eomplex heteroeydie eompounds,
is mueh more diffieult to bioremediate than a soil impaeted with
diesel fuel eontaining primarily easily degradable low moleeular weight
alkanes and PARs.
Soil Biotreatment Technologies
Rydroearbon-eontaminated soils may either be treated in plaee (in situ)
or exeavated, transported, and bioremediated elsewhere (ex situ). If only
surfaee soils are contaminated by erude oil or its fuel produets, either
aecidentally by spills and pipeline failures or intentionally via spreading of oily refinery wastes, the hydroearbons ean be treated in situ via
land farming. Land farming is generally eonfined to treating only the
top 30-em layer of the soil and involves the addition of fertilizers and
sometimes bulking agents and indudes periodie tilling and irrigation
to stimulate indigenous soil baeteria to break down the hydroearbon
eontaminants. Land farming is often the treatment method of ehoice
beeause of its eost-effeetiveness if it ean be assured that hydroearbons
do not leaeh into underlying aquifers and the volatilization of eontaminants does not pose a health risk to workers or nearby eommunities. If
subsurfaee or vadose zone soils are eontaminated by petroleum hydroearbons, either as a result of the downward movement ofhydroearbon
NAPLs from surfaee spills, storage tank leaks, or pipeline ruptures, the
hydroearbons may be treated in situ via bioventing whieh involves the
pumping of air through slotted vapor well easings into the eontaminated
vadose zone to stimulate aerobic biodegradation (Dupont 1993; Hinehee
and Alleman 1997; Tellez et al. 1998).
In eases where soils eannot be treated in situ either beeause the
hydroearbons pose a risk to groundwater resourees or volatilization
eauses unaeeeptable air pollution, or there are spaee limitations or other
eonstraints (e.g., property owners objeet to in situ treatment), the soils
M.H. Huesemann
must be excavated, transported, and bioremediated off-site in either
land treatment units (LTUs), compost piles, biopiles, or slurry bioreactors. Land treatment units are specifically dedicated areas where hydrocarbon contaminated soils can be treated by land farming where a
leachate collection system prevents the off-site migration of watersoluble hydrocarbons (Ryan et al. 1984; Pope and Matthews 1993;
Huesemann 1994). During composting, the contaminated soil is mixed
with organic materials such as sewage sludge, straw, or wood chips, and
placed in piles or windrows (Potter et al. 1999; Jorgensen et al. 2000;
Namkoong et al. 2002). The addition of organic bulking agents not only
improves soil aeration but also provides easily metabolizable substrates
whose rapid biodegradation increases the soil temperature which, in
turn, may speed mesophilic (30-45°C) or thermophilie (45-75°C)
hydrocarbon bio degradation (see also Sect. 5.5 below). A biopile is a
highly engineered composting system where aeration is controlled via
an embedded network of air distribution pipes and the runoff of watersoluble hydrocarbons is prevented via aleachate collection system (Cyr
and Spieles 1997; Von Fahnestock et al. 1998; Carrera et al. 2001). The
biopile is also well suited for the treatment of more volatile hydrocarbons (e.g., gasoline, jet fuels, etc.) whose escape into the surrounding
air is avoided by covering the pile with a plastic sheet and by applying
a slight negative pressure to the piped distribution system to pull air
through the pile and removing the volatile hydrocarbons from the
exhaust air prior to discharge. Compared to land-treatment units, the
biopile requires less space but its operation is more challenging and
costly (Long et al. 2001).
Even more involved is the treatment of contaminated soils in slurry
bioreactors that are continuously mixed and aerated to maximize the
transfer of oxygen into the aqueous phase (Pinelli et al. 1997; Villemur
et al. 2000; Launen et al. 2002; Castaldi 2003). However, as many environmental parameters such as pH, temperature, dissolved oxygen
concentrations, nutrient levels, and even microbial cell densities (via
inoculation with pure or mixed enrichment cultures) can be tightly controlled, small-scale slurry bioreactors are most often used in research
and development efforts to elucidate biodegradation mechanisms and
to optimize bioremediation performance. The intense mixing of contaminated soil particles with water prornotes desorption of hydrocarbons into the aqueous phase and thereby enhances the bioavailability
to microorganisms. As a result, hydrocarbon biodegradation rates are
generally higher in slurry bioreactors than in other "solid-phase" treatment systems, thus slurry treatment is often advantageous whenever
time is a limiting factor (see also Chap. 9 by Ward and Singh). Finally,
volatile emissions during treatment are easily captured and eliminated
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
from the bioreactor headspace and leaching is of no concern since the
slurry is always contained within a closed tank.
Loss Mechanisms Other Than Biodegradation
Before discussing in greater detail all the factors that affect hydrocarbon biodegradation in soil, it is important to recognize that there are
other processes such as volatilization, leaching, sorption and photooxidation that may cause the removal of certain hydrocarbon compounds or classes during the bioremediation treatment. At room
temperature (20°C), most hydrocarbons with carbon numbers up to CIS
or CI6 readily evaporate from soil if in free contact with air (Stronguilo
et al. 1994; Huesemann et al. 1995). Even heavier hydrocarbons (>C I6 )
including three- and four-ring PAHs are likely to volatilize if the soil is
heated by intense sunshine (Hawthorne and Grabanski 2000). Leaching
removes mostly highly water-soluble hydrocarbons such as BTEX and
naphthalenes from the soil during infiltration by rainwater (Bedient et
al. 1994). Highly lipophilic hydrocarbons, i.e., those compounds that are
characterized by high octanol-water partition coefficients (Kow ) such as
five- and six-ring PAHs, are known to strongly adsorb to soil organic
matter and thus may not only be effectively sequestered and no longer
bioavailable to microorganisms (see Sect. 6.6 below) but also less
extractable for chemical analysis (Hatzinger and Alexander 1995).
Finally, any hydrocarbons that are present in the thin soil surface layer
that is exposed to sunlight may be broken down by photo-oxidation
(Dutta and Harayama 2000).
Failure to consider these competing loss mechanisms during field or
laboratory bioremediation studies could lead to the erroneous conclusion that the soil was remediated by bacteria while in reality a large fraction of the contaminants was simply transferred to another medium
(i.e., air, water, or soil organic matter) or chemically broken down by UV
radiation. Thus, in order to ascertain that biodegradation is the primary
hydrocarbon removal mechanism, it is necessary to measure or estimate
all types of loss, either by calculating a complete mass balance or by carrying out microbially poisoned control experiments to assess the magnitude of all combined abiotic losses. Unfortunately, as noted in arecent
review by Salanitro (2001), most published field and laboratory bioremediation studies fail to account for losses due to evaporation and
weathering, and are therefore most likely overestimating the extent of
hydrocarbon bio degradation. For example, 15 to 60% of fuel hydrocarbons (diesel, jet fuel, and heating oil) was lost during soil bioremedia-
M.H. Huesemann
tion solely due to evaporation (Salanitro 2001), and even semi-volatile
four- and five-ring PAHs were volatilized to a significant degree from
soils during biotreatment (Wilson and Jones 1993; Hawthorne and
Grabanski 2000).
Optimizing Environmental Conditions
Moisture Content
Soil microorganisms can only biodegrade petroleum hydrocarbons
within a limited range of favorable soil moisture conditions. If the soil
is too dry, bacterial growth and metabolism will be greatly reduced or
even inhibited. Alternatively, if the soil is too wet or fiooded, soil aeration will be greatly impaired which, in turn, will result in anaerobic conditions that are not conducive to the bio degradation of most higher
molecular weight hydrocarbons. Although the soil is "fiooded" in a
slurry bioreactor, intensive mixing and aeration generally prevent
anaerobic conditions within the soil slurry.
The optimum moisture content for stimulating petroleum hydrocarbon bio degradation ranges from 50 to 80% of the moisture content at
field capacity for the contaminated soil (Bossert and Bartha 1984). Since
the moisture content at field capacity differs widely among soil types, it
is important to experimentally determine the field capacity for the soil
under investigation. It should also be noted that compared to clean soils,
the field capacity of oily soils is much reduced due to the water repelling
character of oiled soil particle surfaces (Dibble and Bartha 1979).
Because of this reduced water-holding capacity of petroleum contaminated soils, it is beneficial to add organic bulking agents (e.g., straw,
woodchips, sawdust, rice hulls, manure, sewage sludge, etc.) to the soil
in order to increase the water retention between irrigation events during
land farming and composting and thereby maintain the soil mo ist ure
content within a range that is optimal for hydrocarbon bio degradation.
Since the initial step in the breakdown of aliphatic, cyclic, and aromatic
petroleum hydrocarbons involves the enzymatic oxidation of these
compounds by oxygenases which require molecular oxygen for their
proper functioning (Leahy and Colwell1990), it is essential that aerobic
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
conditions be maintained in the contaminated soil throughout the
entire biotreatment period. It has been observed that hydrocarbon
biodegradation slows down considerably when the soil gas oxygen concentration drops below 2-5% (vIv) (Huesemann and Truex 1996; Hurst
et al. 1997; Leeson and Hinchee 1997; Hupe et al. 1999). In addition, most
petroleum hydrocarbons are not biodegraded to any significant degree
in the absence of oxygen.
The depth to which oxygen penetrates into the soil depends on the
balance between the rate of oxygen diffusion from the atmosphere into
the porous soil matrix and the rate of oxygen consumption by bacterial metabolism (Devinny and Islander 1989; Huesemann and Truex
1996). Consequently, the two primary mechanisms to maintain proper
soil aeration are (1) to create soil conditions that maximize oxygen diffusion rates and (2) to limit the thickness of the contaminated soillayer.
Oxygen diffusion in soil can be severely restricted if the soil is fiooded
with water, heavily compacted by vehicular traffic, or is rich in day
content. The best way to enhance oxygen diffusion into the contaminated soillayer is by adding bulking agents (i.e., during land farming
and composting) and by periodical tilling to break up large aggregates
and increase soil porosity.
In order to limit the demand of oxygen by soil bacteria, it is important not to overload the soil with too much high levels of petroleum
contamination. As outlined in Section 6.1, the optimum contaminant
loading level is ca. 5% (wt.) of petroleum hydrocarbons. In addition, the
depth of the soil layer during land farming should not exceed 30 cm
since oxygen transport by diffusion will most likely be severely limited
in deeper soil strata. If despite these precautions (i.e., the addition of
bulking agents, tilling, limited contaminant loading and depth) the soil
is still not properly aerated, it is possible to add solid peroxygen compounds to the soil (Davis et al. 1997; Heitkamp 1997). These peroxygen
additives slowly release oxygen into the soil and thereby enhance the
aerobic biodegradation of petroleum hydrocarbons.
As mentioned earlier, soil aeration in biopiles is provided by a
network of slotted pipes that either pushes (pressure) or pulls (vacuum)
air through the pile (Von Fahnestock et al. 1998; Koning et al. 1999).
Sufficiently high dissolved oxygen levels in slurry bioreactors are maintained by intensive mixing and aeration while low oxygen concentrations in deep vadose zone soils, where the rate of oxygen transport from
the atmosphere is diffusion limited, are effectively increased for stimulating hydrocarbon bio degradation by pumping air into the ground
during bioventing operations.
M.H. Huesemann
The rate of hydrocarbon biodegradation is severely limited if there are
insufficient amounts of nitrogen (N) and phosphorus (P) fertilizer in
the soil. A wide range of optimal C : N and C : P ratios have been reported
in the literature. While Frankenberger (1992) recommends a C: N : P
ratio of 100: 10 : 1, Dibble and Bartha (1979) found optimal petroleum
biodegradation with C: N and C: P ratios of 60: 1 and 800: 1, respectively. By contrast, Brown et al. (1983) reported that a C: N ratio of 9: 1
was optimal for refinery sludge biodegradation, and. Huddleston et al.
(1984) suggested that the C: N ratio should be maintained between 25
and 38. FinaHy, Morgan and Watkinson (1989) reviewed numerous
bio degradation studies and found that optimal C: N ratios between 9: 1
and 200: 1 had been reported for waste oil and sludges.
While the application of fertilizers according to a fixed C: N : P ratio
might be appropriate for relatively low petroleum contamination levels,
it can result in overfertilization for heavily contaminated soils. The
excessive addition of nitrogen fertilizers can create serious problems
such as the suppression of soil respiration, possibly due to ammonia or
nitrite toxicity (Bossert and Bartha 1984), the inhibition of hydrocarbon bio degradation (Morgan and Watkinson 1990; Genouw et al. 1994)
and groundwater contamination (Bedient et al. 1994). The toxic and
inhibitory effects of nitrogen fertilizers are expected to be particularly
severe in sandy soils that have limited water-holding capacity. This is
due to the fact that the application of fertilizer according to a fixed C:
N ratio in these sandy soils will result in extremely high (and inhibitory)
nitrogen concentrations in the relatively small amounts of soil water. As
a result, it has been recommended that nitrogen fertilizer be applied at
a concentration of ca. 2000 mg nitrogen per kg soil water to assure
optimal hydrocarbon biodegradation activity (Walworth et al. 1997). In
order to maintain these conditions, it is necessary to periodicaHy
monitor and adjust both the soil nitrogen levels as weH as the soil moisture content.
It is difficult if not impossible to add solid or liquid fertilizers to deep
hydrocarbon contaminated vadose zone soils during bioventing operations. In most published bioventing projects no nutrients were added,
because either sufficient soil organic matter-bound nitrogen was
present and maintained via biogeochemical cyeling or the addition of
liquid fertilizer solutions would have resulted in soil flooding with concomitant reduction in soil-air porosities which, in turn, would seriously
impair subsurface air flow during bioventing. In principle, however, it
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
is possible to add gaseous ammonia (NR3) to the injected air stream
and thus deliver nitrogen to subsurface soils. For example, MarshaH
(1995) reported that the introduction of NR3 to the subsurface
enhanced the bio degradation of diesel fuel contaminants as evidenced
by significant increases in evolved CO 2 and the rate of O2 utilization.
The optimum pR for hydrocarbon bioremediation in soil ranges from
5-7.8 (Dibble and Bartha 1979). Rowever, in order to avoid migration
of hazardous metals that may be present in the refinery oily wastes, it is
advisable to maintain the soil pR above 6 (Ruddleston et al. 1984). If
the soil is too acidic (pR< 6), lime or calcium carbonate may be added
to increase the pR to the required optimum range. If the soil is too alkaline (pR> 8), elemental sulfur, ammonium sulfate or aluminum sulfate
may be added to lower the pR. Detailed instructions regarding soil
pR adjustments may be obtained from a soil-testing laboratory
(Ruesemann 1994).
Bossert and Bartha (1984) found that the optimal temperature range for
the bio degradation of petroleum is 30-40°C although site-specific conditions may playa role in selecting a soil microbial population that can
function weH at a higher or lower temperature. Rydrocarbon biodegradation has been reported in soils at temperatures as low as -101°C
(Leahy and ColweHl990) but bio degradation rates are expected to slow
considerably below 15°C (Dibble and Bartha 1979). It is not known what
maximum soil temperature can be tolerated by hydrocarbon degrading
bacteria. Thermophilie bacteria appears to be able to degrade hydrocarbons at elevated temperatures (e.g.,60 0c) (Klug and Markovetz 1967;
Mateles et al. 1967; Kvasnikov et al. 1974; Castaldi et al. 1995; Ruesemann et al. 2002b). Thermophilie conditions may increase petroleum
hydrocarbon bio degradation rates and can be easily maintained in
bioreactors by heating the soil slurry or can be generated in compost or
biopiles by the addition of easily metabolizable organic materials.
Rowever, high soil temperatures might not be advantageous during land
farming since they promote rapid drying of the soil and the volatilization of hazardous hydrocarbons. In order to avoid excessive soil
warming by intense sunshine it is advisable to shade the soil surface
M.H. Huesemann
with porous mulch and irrigate the soil frequently to reduce the temperature by evaporative cooling.
Although numerous strains of hydrocarbon degraders were found to
biodegrade petroleum at 6°C (Belousasova et al. 2002), the rate of
bio degradation is very slow at these low temperatures and therefore
should be increased in active bioremediation projects by raising soil
temperatures. This can be done by covering compost piles with plastic
sheeting to retain heat (Turneo and Guinn 1997; Schoefs et a1. 1998) or
by circulating heated groundwater to warm overlying vadose zone soils
undergoing bioventing treatment (Leeson et a1. 1993).
Addressing Other Potential Limitations
High Contaminant Concentrations
Hydrocarbon bio degradation proceeds at optimal rates as long as contaminant levels in the soil are below 5%wt. Dibble and Bartha (1979)
reported that bio degradation slowed considerably at sludge loading
rates of 10 and 15%. If large amounts of oily waste have to be treated
on a limited land-farming area, it is best to apply multiple oilloadings
(s5% wt) several times throughout the growing season (Dibble and
Bartha 1979; Brown et a1. 1983; Bossert et a1. 1984; Ryan et a1. 1984; Sims
1984). The application of smaller, more frequent oily waste batches onto
the soil not only reduces the toxic and inhibitory effects observed at
high loadings but also enhances the co-metabolic bio degradation of
more recalcitrant compounds such as benzo(a)pyrene (see also Sect.
6.4). Finally, if as a result of an accidental oil spill the petroleum concentration in the soil happens to be significantly greater than 5% wt, it
is best to dilute the highly contaminated soil with clean soil or bulking
agents (via tilling) to bring the hydrocarbon levels down into the desired
range (i.e., ::;5% wt.) (Huesemann 1994).
Presence of Inhibitors
In addition to the inhibitory effects that are caused by high hydrocarbon concentrations and excessive nitrogen fertilization, the presence
of salts and heavy metals can also potentially inhibit petroleum
bio degradation.
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
Many refinery wastes contain salts (KCl and NaCl) that are the result of
desalting operations during the processing of crude oil. Frequent applications of these wastes onto land farms will result in the structural
deterioration of the soil and wi11lead to a reduction of soil permeability and a restriction of water availability (Pollard et al. 1994). This, in
turn, will have negative impacts on the activity of hydrocarbon-degrading soil microorganisms. For example, Ward and Brock (1978) found
that the rates of hydrocarbon metabolism decreased with increasing
levels of salinity. Similarly, Rhykerd et al. (1995) reported that bacterial
respiration during motor oil biodegradation decreased by 36% in soils
containing 6 g NaCl/kg compared to non-salty control soils. Excessive
salts can be removed from the impacted soil by water flushing or treatment with calcium (PoUard et al. 1994).
Refinery wastes often contain appreciable quantities of metals (Pb, Cd,
Hg, Zn, Cu, As, Cr, Ni, V) that are either derived from metaUophorphyrins or associated with unit operations that employ metal catalysts
or additives for enhancing product quality (Pollard et al. 1994). With
repeated applications of oily sludge to alandfarm operation, heavy
metals may accumulate to levels at which biodegradation may be
reduced (Frankenberger 1992). For example, Jensen (1977) studied the
effects of lead (Pb) on biodegradation of oily waste in soil and found
that a reduction of bacterial populations occurred at a Pb concentration
of SOOOppm. If high metal concentrations in the land-farm soil are a
problem, there are at least two different options to lessen their negative
impact on hydrocarbon bio degradation processes. First, the metalimpacted soil can be diluted with clean soil in order to reduce inhibitory
effects. Second, organic matter (e.g., woodchips, straw, rice huUs,
manure, etc.) can be added to the soil in order to promote the complexation of metals and make them less available to soil microorganisms (Frankenberger 1992).
Insufficient Number of Hydrocarbon-Degrading Microorganisms
Soils without previous exposure to hydrocarbon contamination should
have around 105-106 hydrocarbon degrading microorganisms per gram
M.H. Huesemann
of soil. Soils that have a long-term history of petroleum contamination
(e.g., land-farm soils) should have around 106 _10 8 hydrocarbon
degraders per gram of soil (Bossert and Bartha 1984). If the counts of
hydrocarbon degraders are significantly lower, it is likely that inhibitory
or toxic conditions are present in the soil. The first step is to address the
root causes of this problem by reducing potential inhibitory effects that
occur due to high soil contamination levels, excessive nitrogen fertilization, or high salt and metal concentrations. After the sources of
microbial inhibition have been reduced or eliminated, it is possible to
speed up the recovery of microbial populations by "inoculating" the
microbially impoverished contaminated soil with clean topsoil, which is
generally characterized by a rich microbial flora. Another option may
be to use sewage sludge (ideally from a refinery wastewater treatment
plant) to seed the contaminated soil with hydrocarbon degrading
If this fails, it may help to inoculate the soil with commercial bacterial preparations or enrichment cultures although the effectiveness of
this expensive approach is questionable due to the complex competitive
conditions that exist in the soil environment (Atlas 1977; Goldstein et
al. 1985; Leahy and Colwelll990; Forsyth et al. 1995). In fact, the majority of published inoculation studies have shown that the addition
of commercial cultures does not significantly enhance hydrocarbon
biodegradation in soils. Far example, inoculation with commercial bacterial preparations did not increase the removal of crude oi! (Venosa et
al. 1991; Huesemann and Moore 1994; Neralla and Weaver 1997), fuel
oil (Dott et al. 1989), or diesel and lubricating oil (Jorgensen et al. 2000).
Similarly, inoculation with a PAH degrading bacterial consortium did
not enhance PAH bio degradation in dredged sediments undergoing
slurry biotreatment (Launen et al. 2002). By contrast, Belloso (2001)
found that the addition of an enrichment culture that was started using
an inoculum taken from a land-farming site significantly increased
hydrocarbon rem oval efficiencies during the land treatment of refinery
oily waste (i.e., 84% in the inoculated soil versus 49% in the control soil).
Similarly, Morrison et al. (1997) reported that the addition of a commercial microbial culture reduced total petroleum hydrocarbon (TPH)
concentrations in a biopile by 71 % compared to only 44% without the
Lack of Cometabolism
It has been reported that five-ring PAHs are only biodegraded in the
presence of other hydrocarbons such as lower molecular weight PAHs
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
or complex hydrocarbon mixtures such as crude oil (Keck et al. 1989;
Kanaly et al. 1997; Kanaly and Bartha 1999). If these necessary cosubstrates are absent, the cometabolic bio degradation of higher molecular
weight PAHs cannot proceed. This situation can occur when three- and
four-ring PAHs and other lower molecular weight hydrocarbons have
biodegraded during the initial stages ofbioremediation treatment while
five- and six-ring PAHs are left behind without the necessary cosubstrates to stimulate their cometabolic transformation. This may be the
reason why multiple oily sludge application to aland-farm area actually
stimulates the biodegradation of complex aromatic compounds (Dibble
and Bartha 1979).
Inherent Recalcitrance
Susceptibility to biotransformation is a function of chemical structure,
the degree and nature of the substitution of the parent compound, and
molecular weight. The following generalized sequence of decreasing
susceptibility to biotransformation among chemical dasses has been
reported: n-alkanes > branehed chain alkanes > branched alkanes > low
molecular weight n-alkyl aromaties > monoaromatics > cydic alkanes,
polynudear aromaties »> asphaltenes (Pollard et al. 1994). There are
certain compounds and eompound dasses that are only very slowly, if
at all, biodegraded (Alexander 1973; Huesemann 1995, 1997; Salanitro
2001). For example, pentacydic saturated hydrocarbons such as
hopanes are so recalcitrant to biodegradation that they have been used
as eonservative biomarkers in bioremediation studies (Butler et al. 1991;
Prinee et al. 1994). Similarly, asphaltenes are so slowly biodegraded or
converted into soil humus that their concentration increases in land
farms that receive frequent waste applications (Bossert and Bartha
If indeed certain hydrocarbon compounds or dasses are essentially
reealcitrant to bio degradation, it should be in principle possible to
prediet the maximum extent of hydroearbon removal during bioremediation treatment as long as the initial composition of petroleum contaminants is known. Two different approaches to predict petroleum
biodegradation potential in soils have been reported. First, Huesemann
(1995) developed a predictive model for estimating the extent ofhydroearbon bio degradation, which is based on the experimental observation
that hundreds of specifie hydroearbon dasses exhibit similar biodegradation behavior even though they originate from different erude oils or
fuels. Thus, the hydroearbon dass eomposition of the original soil contaminants essentially determined the maximum extent of biodegrada-
M.H. Huesemann
tion. A second, simpler, predictive method has been reported by
McMillen et al. (2001) who found that the crude oil's API gravity, which
is a surrogate for crude oil composition because of its positive correlation to the amount of easily degradable hydrocarbons such as saturates
and n-alkanes and its negative correlation to the amount of more recalcitrant fractions such as asphaltenes and resins, could be used to predict
total petroleum hydrocarbon (TPH) and oil and grease (O&G) removal
from soil during bioremediation treatment.
If a contaminated soil contains a large fraction of hydrocarbons
or hydrocarbon classes that are inherently recalcitrant to microbial
bio degradation, it may be impossible to re ach specific cleanup targets
that are required by regulatory agencies. If this be comes a problem, it
may be possible to argue that these residual, recalcitrant waste components do not pose a significant risk to environmental receptors and
therefore might be safely left in place. In order to justify this decision,
it is necessary to perform a number of tests to assure that residual contaminants do not cause negative ecological impacts or groundwater pollution. These tests are discussed in more detail in Section 7.
Bioavailability Limitations
However, as both contaminant bio degradation and abiotic desorption
kinetics often exhibit similar biphasic behavior (i.e., long-term slow
rates that follow initial fast rates) (Linz and Nakles 1997; Cornellissen
et al. 1998), it has been postulated that hydrocarbon biodegradation is
likely to be mass-transfer rate or bioavailability limited, particularly
during the final stages ofbioremediation treatment. Bioavailability limitations are most commonly encountered in "aged" soils that have been
contaminated with petroleum hydrocarbons for many years. According
to the contaminant sequestration theory, lipophilic organic contaminants become with time increasingly sequestered in the soil matrix and,
as a result, less bioavailable to biodegrading microorganisms (Alexander 1995; Pignatello and Xing 1996; De Jonge et al. 1997; Huesemann
1997; Linz and Nakles 1997; Luthy et al. 1997; Alexander 2000). When
these aged hydrocarbon contaminated soils are subjected to bioremediation, any non-sequestered molecules are biodegraded very rapidly by
soil microorganisms, while contaminants that were sequestered during
the aging period are biodegraded only very slowly or not at all. The rate
of bio degradation of these aged, sequestered contaminants is believed
to be limited by the slow rate of desorption or dissolution from the soil
matrix. If the rates of contaminant release from the soil are exceedingly
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
slow, biodegradation is similarly very slow which, in turn, often results
in the incomplete removal of contaminants during time-limited bioremediation projects.
Although contaminant sequestration and lack of bioavailability are
often given as the main reasons - generally without further proof - for
incomplete bio degradation, several well-controlled studies found that
high molecular weight PAHs did not biodegrade despite their ready
bioavailability. For example, Cornelissen et al. (1998) measured the desorption kinetics of 15 PAHs from sediments before and after bioremediation and found that while some PAHs rapidly desorbed and
biodegraded, others did not degrade at all despite the fact that they were
bioavailable as indicated by the large rapidly desorbing fractions (up to
55%) at the end of biotreatment. Similarly, Huesemann et al. (2002a)
observed that after 90 weeks of slurry treatment, a number of high
molecular weight PAHs were recalcitrant to bio degradation although
they were released abiotically to a significant degree indicating that
microbial factors rather than bioavailability limited the bio degradation
of these PAHs. In another study, Huesemann et al. (2003) measured
both abiotic desorption rates and biodegradation rates for PAHs in six
differently aged soils at various time points during slurry biotreatment.
The biodegradation of PAHs was generally not mass-transfer limited
during the initial phase while it often became so at the end of the treatment period when bio degradation rates equaled abiotic desorption
rates. However, in all cases where PAH biodegradation was not observed
or PAH removal temporarily stalled, bioavailability limitations were not
considered responsible for this recalcitrance since these PAHs desorbed
rapidly from the soil into the aqueous phase.
If experimental tests indicate that bioavailability limitations are
indeed the cause for incomplete hydrocarbon bio degradation in aged
soils, it may be possible to increase contaminant bioavailability and thus
promote additional hydrocarbon biodegradation. In laboratory studies
involving aged soils, it has been found that the bioavailability of PAHs
can be increased by the addition of surfactants (Ghosh and Yeom 1997;
Tiehm et al. 1997; Boonchan et al. 1998) or by increasing the soil slurry
temperatures (Ten Hulscher and Cornelissen 1996; Cornelissen et al.
1997). While these techniques can be easily employed in slurry bioreactors, they are more difficult or impossible to implement in other soil
treatment technologies such as land farming or bioventing. Consequently, it has been argued that residual contaminants that remain after
bioremediation treatment are not bioavailable and therefore pose little
risk to environmental receptors. According to this argument, many
bioremediated soils have reached "environmentally acceptable endpoints" and can be left in place without further treatment.
M.H. Huesemann
Risk Assessment and Environmentally Acceptable End Points
Even if optimal bioremediation conditions are provided, the inherent
recalcitrance of certain hydrocarbons and the limited bioavailability of
sequestered and aged contaminants are two factors that not only cause
incomplete waste bio degradation but are also very difficult to overcome.
Consequently, it is advantageous to perform a number of tests to determine whether these residual hydrocarbons pose a significant threat to
ecological receptors and groundwater resources. If these tests indicate
that the hydrocarbons remaining after bioremediation treatment are
not toxie to soil microorganisms (e.g., Microtox), earth-dwelling organisms (e.g., earthworms, nematodes, etc.), and plants, and do not leach
into nearby aquifers used for drinking water, it may be concluded that
the residual hydrocarbons can be safely left in pi ace without further
A number of studies have shown that bioremediation treatment
reduces soil toxicity, contaminant leachability, and risk to environmental receptors (Alexander 1995,2000; Linz and Nakles 1997; Loehr and
Webster 1997; Chaineau et al. 2003). For example, a reduction in Microtox toxicity as a result of bioremediation treatment has been reported
by numerous investigators (Wang and Bartha 1990; Hund and Traunspurger 1994; Loehr and Webster 1996; Salanitro et al. 1997; Symons and
Sims 1988; Sayles et al. 1999; Haeserler et al. 1999; Huesemann et al.
2004). Soil bioremediation has also been shown to reduce the toxicity
of hydrocarbons to earthworms and nematodes, decrease genotoxicity
(i.e., Ames mutagenicity and chromosomal aberration), and improve
soil quality as measured by the ability of plants to germinate and grow
in the petroleum impacted soil (Wang and Bartha 1990; Loehr et al.
1992; Linz and Nakles 1997; Salanitro et al. 1997; Saterbak et al. 1999;
Sayles et al. 1999; Duncan et al. 2003). Finally, bioremediation treatment
also significantly reduces the leaching potential of hydrocarbons in soil
(Loehr and Webster 1996, 1997) and attenuates the mutagenicity of
land-farm leachates and runoff (Brown and Donnelly 1984; Barbee et
al. 1992; Haseseier et al. 1999). While in many cases bioremediation
reduces the toxicity of hydrocarbons to the background levels measured
in clean reference soils, under certain circumstances a residual negative
environmental impact may remain even after extensive biotreatment.
Biodegradation and Bioremediation of Petroleum Pollutants in Soil
The successful bioremediation of petroleum pollutants in soils depends
on numerous environmental parameters and operational factors, all
of which need to be optimized in order to achieve maximum contaminant treatment. Even under optimal conditions it is unlikely that all
hydrocarbon constituents will be removed from the soil. This incomplete bio degradation may be acceptable if the residual hydrocarbons
can be shown to have no significant impact on ecological receptors a
nd do not pose a risk to groundwater resources. Jf this can be demonstrated with a high degree of certainty, bioremediation is one of the
most cost-effective treatment technologies for petroleum contaminated
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Bioremediation of Pesticide-Contaminated
Ramesh C. Kuhad/ Atul K. JohrV Ajay Singh,3 and
Owen P. Ward4
About 4 million tonnes of pesticides are applied to agricultural crops
annually for pest control worldwide. It is estimated that less than 1% of
total applied pesticides generally gets to the target pests and most of the
pesticides remain unused and enter into the ecosystem. The ultimate
sink for excessive pesticides is soil and water. Despite their persistence
in the environment, with a tendency of residues to bioaccumulate and
be toxic to non-target organisms including humans, the use of chemical pesticides cannot be discontinued. Among various soil remediation
technologies available today for decontamination and detoxification of
pesticide-contaminated soils, bioremediation seems to be one of the
most environmentally safe and cost effective methods (Fogarty and
Tuovinen 1991; Häggblom 1992; Alexander 2000). Most of the pesticides generally fall under the major classes of chlorophenoxy acids,
organochlorines, organophosphates, carbamates and s-triazines. This
chapter focuses on microorganisms having the potential to degrade pesticides and on factors affecting pesticide bio degradation in contaminated soils.
1 Department of Microbiology, University of Delhi, South Campus, New Delhi-11 0021,
2Channing Laboratory, Harvard Medical School, 181 Longwood Avenue, Boston, Massachusetts 02115, USA
3 Petrozyme Technologies, 7496 Wellington Road 34, R.R. #3, Guelph, Ontario, N 1H 6H9,
Canada, e-mail: [email protected], Tel: +1-519-7672299, Fax: +1-519-7679435
4 Department of Biology, University ofWaterloo, Waterloo, Ontario, N2L 3G1, Canada
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and o. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
R.C. Kuhad et al.
Biodegradation of Pesticides
Mechanisms of bio degradation of certain pesticides are weIl understood under both aerobic and anaerobic conditions (Commandeur and
Parsons 1990; LaI et al. 1997). Recent advances in recombinant DNA
technology and increased knowledge of biochemical mechanisms have
revolutionized pesticide microbiology (Cork and Kruger 1991; Kumar
et al. 1996). Molecular probes have been used to identify and isolate
microbes with a degradation potential towards pesticides and related
compounds (Singh et al 1999). Studies with pure bacterial and fungal
cultures (Table 1) have provided detailed information on the catabolic
pathways of several pesticide compounds.
Microorganisms have evolved a variety of biochemical pathways to
degrade or detoxify pesticides. For example, removal of the halogen
substituent is a key step in the degradation of halogenated aromatic
pesticides, which occurs as an initial step via reductive, hydrolytic or
oxygenolytic mechanisms, or after the cleavage of the aromatic ring at
a later stage of metabolism. Oxidative dehalogenation, where halogen
is lost during oxygenation of the ring, occurs only under aerobic conditions. Hydrolytic dehalogenation, the replacement of halogen by a
hydroxyl group, can occur under both aerobic and denitrifying conditions. Reductive dehalogenation, the replacement of halogen by hydrogen, occurs only under methanogenic and sulfogenic conditions. In
addition, several biotransformation reactions such as methylation and
polymerization may take pI ace and produce more toxic or recalcitrant
Hydrolases and oxygenases are the two most important classes of
enzymes, which are responsible for catalyzing the pesticide biotransformation reactions. Hydrolases (helidohydrolases, esterases, amidases)
do not require co-factors and are stable at a wide range of pHs and temperatures. For example, halidohydrolases dehalogenate many halogenated aliphatic and aromatic compounds, whereas esterases, like
parathion hydrolase, attack phosphodiester bonds of organophosphates
and amidases degrade propanil etc. Oxygenases require molecular
oxygen as a substrate. They are generaIly less stable than hydrolases and
more complex enzymes.
Table 1. Pesticide-degrading microorganisms
Pesticide-degrading microorganisms
Alcaligenes eutrophus,
Alcaligenes xylosoxidans,
Flavobacterium sp.,
Pseudomonas putida,
Pseudomonas cepacia,
Comamonas sp.
Chaudhary and Huang
(1988); Gunulan and
Fournier (1993);
Bulinski and Nakatsu
Pseudomonas cepacia
herbicidivorans, Alcaligenes
Yadav and Reddy
Xun and Wagon
Yadav and Reddy
Zipper et al. (1998);
Tett et al. (1997)
Aerobacter aerogenes,
Alcaligenes eutrophus,
Agrobacterium tumefaciens,
Arthrobacter sp., Bacillus
cereus, Bacillus coagulans,
Bacillus megaterium, Bacillus
subtilis, Clostridium
pasteurianum, Clostridium
Hydrogenomonas sp.,
Klebsiella pneumoniae,
Nocardia sp., Serratia
marsescens, Streptomyces
annomoneus, Xanthomonas
chrysosporium, Trichoderma
Aerobacter aerogenes,
Bacillus cereus, Bacillus
megaterium, Bacillus
circulans, Bacillus brevis,
Citrobacter freundii,
Clostridium rectum,
Pseudomonas flourescens,
Pseudomonas putida,
Sphingomonas paucimobilis
chrysosporium, Trametes
versicolor, Phanerochaete
sordida, Cyathus bulleri
Anabaena sp., Nostoc
Nadeau et al. (1994);
You et al. (1996);
Fava et al. (1998);
Gray et al. (1999);
Singh et al. (1999)
Bidlan and
Manonmani (2002)
Shah et al. (1992);
Bumpus et al. (1993)
Sahu et al. (1993);
Johri et al. (1998);
Gupta et al. (2000);
Singh et al. (2000)
Singh and Kuhad
Kurtiz and Wolk
R.C. Kuhad et al.
Table 1. Continued
Pesticide-degrading microorganisms
Kim et al. (2001)
Arthrobacter sp.,
Pseudomonas cepacia,
Pseudomonas melophthora,
Pseudomonas aeruginosa
Chapalmadugu and
Chaudhry (1993);
Hayatsu et al. (1999)
Karns and Tomasek
(1991); Mohapatra
and Awasthi (1997);
Chaudhary et al.
Achromobacter sp.,
Arthrobacter sp.,
Flavobacterium sp.,
Pseudomonas cepacia,
Pseudomonas stutzeri,
Bacillus pumilis
Agrobacterium radiobacter,
Pseudomonas sp.; Klebsiella
pheumoniae; Rhodococeus
corallinus; Rhizobium sp.,
Nocardiodes sp.
Flavobacterium sp.,
Pseudomonas aeruginosa,
Pseudomonas diminuta,
Pseudomonas melophthara,
Pseudomonas stutzeri
Xu et al. (1996);
Walker and Keasling
(2002); Gilbert et al.
Bouguard et al.
(1997); Park et al.
(2003); Piutti et al.
Mougin et al. (1997)
Due to their persistence in the environment, the use of several
organochlorine pesticides is banned in many advanced countries, but
theyare still being applied in developing countries. Biodegradation and
bioremediation aspects of selected pesticides are discussed in this
Chlorophenoxy Acids
In agriculture, chlorophenoxy acids are extensively used as pesticides,
wood preservatives and herbicides and represent a major group of
Bioremediation of Pesticide-Contaminated Soils
recalcitrant environmental pollutants. 2,4-Dichlorophenoxyacetic acid
(2,4-D) and 2,4,5-trichlorophenoxyacetic (2,4,5-T) are two major representatives of phenoxyacetate herbicides.
Several 2,4-D-degrading strains belonging to genera Ralstonia,
Sphingomonas and Alcaligenes have been isolated from thc soil (Frantz
et al. 1987; Sandmann et al. 1988). The uptake ofherbicide 2,4-D by Ralstonia eutropha JMP134 was inducible and sensitive to metabolic
inhibitors (Leveau et al. 1998). Filer and Harker (1997) demonstrated
that dichloromuconate is the inducing agent in the 2,4-D pathway. The
first common metabolite identified in the degradation of 2,4-D by
Sphingomonas herbicidovorans MH was 2,4-dichlorophenol (Nickel et
al. 1997). In Alcaligenes eutrophus JMP134, genes responsible for 2,4-D
degradation are plasmid encoded (Don et al. 1985).
Biodegradation of 2,4-D and 3-chlorobenzoate was found to be a
common capability of microorganisms in the soils of undisturbed,
pristine ecosystems (Fulthorpe et al. 1996). The conjugal transfer of
catabolic genes can increase the rate of degradation of recalcitrant
pollutants in contaminated environments (Romine and Brockman
1996). Ralstonia eutropha JMP134 has been shown to be chemotactically attracted to 2,4-D (Hawkins and Harwood 2002). The chemotactic response was induced by growth with 2,4-D and the presence
of catabolic plasmid p JP4 which harbors the tfd genes for 2,4-D
During composting of 2,4-D, thermophilic microbes were dominant
with 46% of organic matter present lost at the end of composting.
About 47% of added 2,4-D was mineralized, about 23% was complexed
with humic acids and about 20% was non extractable (Michel et al.
Transfer of 2,4-D degradation plasmids pEMTl and pJP4 from an
introduced donor strain, Pseudomonas putida UWC3, to the indigenous
bacteria of two different horizons (A horizon, depth of 0 to 30 cm; B
horizon, depth of 30 to 60cm) of 2,4-D contaminated soil was investigated in a bio augmentation approach (Dejonghe et al. 2000). When soil
was amended with nutrients, plasmid transfer and enhanced degradation of 2,4-D were observed as compared to nonamended soil. The
choice of donor microorganisms might be a key factor to consider for
bio augmentation efforts in bioremediation of 2,4-D-contaminated soil
(Newby et al. 2000). The establishment of an array of stable indigenous
plasmid hosts at sites may be particularly useful. Laboratory and pilot
field studies have demonstrated that bio augmentation of both cells
and genes can be effective in enhancing degradation of 2,4-D and 3chlorobenzoates in metal (Cd) co-contaminated soils (Roane et al. 2001;
Pepper et al. 2002).
R.C. Kuhad et al.
2,4,5-T is poorly biodegradable as compared to 2,4-D. Pseudomonas
cepacia can transform 2,4,5-T into 2,4,5-trichlorophenol (2,4,5-TCP)
using 2,4,5-T oxygenase (Xun and Wagon 1995). 2,4,5-TCP is then converted to 5-chloro-2-hydroxy-1,4-benzoquinone (5-CHQ) in a two-step
hydroxylation catalysis (Xun 1996). When a 2,4-D degrading plasmid
pJP4 from Alcaligenes eutrophus JMP134 was transferred to a 2,4,5T-degrading strain, P. cepacia ACllOO, the transformed strain (RHJI)
completely metabolized and efficiently degraded both 2,4-D and
2,4,5-T (Haugland et al. 1990). Inhibitor studies have suggested that a
fiavin-containing enzyme is responsible for one of the two 2,4,5TCP-hydroxylation steps (Tomasi et al. 1995). The enzyme responsible,
chlorophenol-4-monooxygenase, consists of two subunits of 58 and 22
kDa polypeptides, encoded by the tftC and tftD genes (Xun 1996).
Mineralization of 2,4-D and 2,4,5-T by Phanerochaete chrysosporium,
and its peroxidase-negative mutant, in a high-nitrogen and malt extract
medium has been reported (Yadav and Reddy 1993). Higher rates were
observed when the compounds were tested as mixtures rather than
Phenoxyalkanoates, with side chains longer than two carbon atoms,
can be degraded by initial side-chain cleavage. Flavobacterium spp.
capable of degrading 4-(2,4-dichlorophenoxy)butyric acid and 2-(2,4dichlorophenoxy)propionic acid (Horvath et al. 1990) have been isolated from soil. Recently, Westendorf et al. (2003) have isolated and
characterized two alpha -ketoglutarate-dependent dioxygenases from
Delftia acidovorans MC1, capable of catalyzing the cleavage of the ether
bonds of various phenoxyalkanoate and phenoxyacetate herbicides,
including, 2,4-D.
When Pseudomonas pseudoalcaligenes POB310 and two modified
Pseudmonas sp. strains B13-D5 and B13-STl were introduced into soil
microcosms containing 3-phenoxybenzoic acid (POB), degradation of
POB by P. psuedoalkaligenes POB310 was incomplete, whereas strains
B13-D5 and B13-STl degraded PB (10-100ppm) to concentrations of
<50ppb with a concomitant increase in density from 106 to 108 CFU/g
dry wt. soil, suggesting the in situ bioremediation potential of the
modified strains (Halden et al. 1999).
DDT [l,l-bis( 4-chlorophenyl)-2,2,2-trichloroethane] bio degradation is
typically cometabolic and requires an alternate carbon source for
growth (Nadeau et al. 1994). The mechanisms of degradation include
dechlorination and ring-cleavage (Hay and Foght 2000). In aerobic
Bioremediation of Pesticide-Contaminated Soils
degradation by Alcaligenes eutrophus A5, oxidation on the phenyl ring
at adjacent ortho and meta positions occurs to form hydroxy-DDT,
which is further metabolized through meta cleavage to 2,3-dihydroxyDDT by dehydrogenase. Then, 2,3-dihydroxy-DDT is further metabolized through meta cleavage to form a yellow ring-fission product that
would then be catabolized to 4-CBA (Nadeau et al. 1994). The new
aerobic catabolic pathway of degradation of DDT, based on its oxidation, makes it possible to clone the catabolic genes responsible and to
understand the ability for concomitant degradation of PCBs, DDT and
related chloroaromatic compounds (LaI et al. 1995).
In anoxic soils, the dominant mechanism is the reductive dechlorination of DDT to 1,1-dichloro-2,2-bis-(4-chlorophenyl)ethane (DDD)
involving the substitution of an aliphatic chlorine for a hydrogen atom
(Aislabie et al. 1997). Under anaerobic conditions, DDD may be further
metabolized to 4,4' -dichlorobenzophenone (DBP). Under aerobic
conditions DDT is dehydrochlorinated to yield predominantly 1,1dichloro-2,2-bis( 4-chlorophenyl)ethane (DDE). Alternating anaerobic
and aerobic conditions can enhance DDT biotransformation to DBP
and subsequent aerobic aromatic ring cleavage (Foght et al. 2001).
Microbial transformation and volatilization are major routes for DDT
biodegradation in tropical soils, whereas in temperate soils, DDT may
persist for long period of times (half-life, 20-30 years) primarily as DDE
(Dimond and Owen 1996). Factors such as composition and activity of
soil microfiora, bioavailability of DDT, presence of organic matter, pH,
temperature and oxygen availability, significantly infiuence biodegradation of DDT in soil (Foght et al. 2001). Soils with high organic matter
have significantly lower concentrations of bioavailable DDT as compared to sandy soils. Low concentrations of DDT present in agricultural
soils along with poor bioavailability, significantly contributes to the persistence of DDT.
Successful bioremediation approaches must consider the reduction of DDT in soil as well as its metabolites. The addition of cometabolically simple (glycerol, hexadecane) and complex (alfalfa, green
manure, rice straw, cellulose, waste newspaper etc.) substrates, surfactants (Triton X-100, cyclodextrins), inducers (diphenylmethane, terpenes in orange peel), reducing agents (zero-valent iron, cystein, sodium
sulfide) and the use of bioaugmentation may accelarate the bioremediation of DDT-contaminated soils (You et al. 1996; Aislabie et al. 1997;
Fava et al. 1998; Sayler and Ripp 2000; Foght et al. 2001). In a full-scale
biopiling process with alternative aerobic and anaerobic stages, DDT
and DDE-contaminated soil, amended with chicken manure, newspaper, straw and woodchips, provided a 95% reduction in DDT levels
(Gray et al. 1999).
R.C. Kuhad et al.
The impacts of arsenic co-contamination on the natural breakdown
of 1,1,1- trichloro-2,2-bis(4-chlorophenyl)ethane (DDT) in soil has
recently been investigated (Van Zweiten et al. 2003). Microbial activity
was inhibited as residues of total DDTs and arsenic increased resulting
in an increased persistence of DDT.
The white rot fungus, Phanerochaete chryrosporium, is capable of
degrading several chlorinated organic compounds including DDT
(Morgan et al. 1991; Paszczynski and Crawford 1995). DDE can be mineralized to CO2 via DBP by P. chrysosporium (Bumpus et al. 1993). The
degradation pathway of DDT by white rot fungi differs from that of bacteria. Lignin peroxidases were thought to be involved in DDT catabolism by P. chrysosporium, but Kohler et al. (1988) argued that the
extracellular ligninase of P. chrysosporium has no role in the degradation ofDDT.
Technical preparations of y-hexachlorocyclohexane (HCH; lindane)
contain four isomers of HCH, a, ß, yand 8. Due to their high lipid solubility, bioaccumulation of HCH residues in aquatic and terrestrial
animals is high. Anaerobic degradation of y-HCH, occurs rapidly in
paddy soil. Clostridium spp. have been isolated and the pathway of yHCH degradation has been reported (Singh et al. 2000). Clostridium
degrades y-HCH via y-3,4,5,6-tetrachlorocyclohexane (y-TCCH), which
is further dechlorinated into chlorobenzene (Ohisa et al. 1980).
Pseudomonas putida and P. paucimobilis degrade y-HCH to y-PCCH,
y- TCCH and tetrachlorobenzene (Nagata et al. 1993). Sahu et al. (1993)
isolated a Pseudomonas sp. from the rhizosphere of an HCH-treated
sugarcane plant, which readily degraded a-, ß- and y-HCH as
sole source of carbon under aerobic conditions. Johri et al. (1996)
characterized the degradative pathway of y- HCH in Pseudomonas paucimobilis UT26 and found that a 500-bp fragment of DNA encoding yRCH dehydrochlorinase was responsible for dehydrochlorination of
y-HCH. Kumari et al. (2002) cloned and characterized lin genes responsible for the degradation of HCH isomers in Sphingomonas paucimobilis B90.
Biodegradation of a- and ß-HCH has been undertaken in soil slurries under different redox conditions such as aerobic, methanogenic,
denitrifying and sulfate-reducing and the aerobic conditions proved to
be the best for the microbial transformation of a-HCH (Bachmann et
al. 1988). Johri et al. (1998) reported the degradation of a-, ß-, y- and 8-
Bioremediation of Pesticide-Contaminated Soils
HCH by Sphingomonas paucimobilis. The bacterium degraded a-HCH
after 3 days; with ß- and y-, and with 8-HCH, respectively, 98 and 56%
degradation occurred after 12 and 8 days. y-Pentacholrocyclohexane (yPCCH) was the primary metabolite during the degradation of all the
HCH isomers.
van Eekert et al. (1998) studied an aerobic transformation of ß-HCH
by methanogenie granular sludge in a methanol-, volatile fatty acid- or
sucrose-fed upflow anaerobic sludge blanket (UASB). The sludge, not
previously exposed to HCH, transformed both ß- and a-HCH present
in contaminated soil to benzene and chlorobenzene upon incubation of
the soil und er anaerobie conditions. The biotransformation of four
isomers, a-, ß-, y-, and 8-HCH under methanogenie conditions was
investigated by Middeldorp et al. (1998). In a flow-through column
packed with polluted sediment, ß-HCH was completely eliminated.
Hence, there is a potential to use sequential anaerobie/aerobic conditions for the treatment of HCH-contaminated soils.
The white-rot fungus, Phanerochaete chrysosporium, cultured under
lignolytic conditions, partially mineralized lindane in liquid cuhure
(Kennedy et al. 1990; Singh and Kuhad 1999) and in corncob-amended
silt loam soil (Kennedy et al. 1990). Fungallignin-degrading systems
(LDS) mainly consist of lignin peroxidase (LIP), manganese peroxidase
(MnP) and laccase. Shah et al. (1992) suggested that the LDS could transform the pesticide. Singh and Kuhad (1999) studied the comparative
degradation of lindane and the formation of metabolites by four different species of fungi, Phanerochaete sordida, Coriolus hirsutus, P.
chrysosporium and Cyathus bulleri. Phlebia radiata, Coriolus versicolor
and C. bulleri were unique in their ability to produce all the three
enzymes (LiP, MnP and laccase) of LDS (Kuhad et al. 1997). Despite
having different sets of LDS, all studied white-rot fungi were found to
degrade lindane, suggesting the complete LDS may not be involved in
lindane degradation.
Both endosulfan and its oxidative metabolite endosulfan sulfate are
equally toxie and persistent in the environment. Enrichment of an endosulfan-degrading culture from soil with a history of endosulfan exposure, using the insecticide as a sole source of sulfur (Sutherland et al.
2000), produced a mixed culture able to both oxidize and hydrolyze the
compound; the degradation occurred concomitant with bacterial
R.C. Kuhad et al.
The addition of isolated bacterial cells to contaminated soil improved
the degradation of endosulfan isomers (Awasthi et al. 2000). The degradation was faster in wet soil than in flooded soils, and was inhibited by
sodium acetate and succinate. The degradation was optimal at pH 8.5
and did not occur at acidic pH. Phanerochaete chrysosposium transformed endosulfan under non-lignolytic conditions to endosulfan
sulfate and endosulfan diol (Kim et al. 2001).
Organophosphates (monocrotophos, chloropyriphos, parathion, methyl
parathion, diazinon, coumaphos), perhaps the most extensively used
agricultural insecticides, are comparatively less persistent, but possess
high mammalian toxicity. Microbial hydrolytic degradation of p-oalkyl and P-O-aryl bonds is considered the most important step in the
detoxification of organophosphorus compounds in soils. Organophosphate hydrolase (OPH) responsible for the hydrolysis has been isolated
from soil microorganisms and found effective in degrading a range of
organophosphate esters. Recombinant OPHs are equally effective in
hydrolyzing and detoxifying various organophosphorus pesticides (Xu
et al. 1996; RausheI2002).
Biodegradation of parathion by mixed microbial cultures produces
p-nitrophenol (PNP) as a major metabolite. PNP can be further utilized as a source of carbon and energy by other microorganisms.
Pseudomonas putida metabolized PNP to hydroquinone and 1,2,4benzene triol, which was cleaved by benzene triol oxygenase to maleyl
acetate (Rani and Lalitha Kumari 1994). A consortium of two engineered strains, E. coli SD2 (harboring two plasmids, one encoding for
parathion hydrolase and a second carrying green fluorescent protein
marker) and P. putida KT2440 pSB337 (containing genes for PNP mineralization) effectively hydrolyzed 146mg/L of parathion without the
accumulation of PNP (Gilbert et al. 2003).
Pseudomonas putida KT2442 was engineered to use the organophosphate pesticide parathion as a source of carbon and energy (Walker and
Keasling 2002), with the initial hydrolysis by organophosphate hydrolase (OPH) to PNP and diethyl thiophosphate, compounds that could
not be metabolized by P. putida KT2442. Transformation of P. putida
with the plasmids harboring opd and the PNP operons allowed the
organism to utilize 0.8 mM parathion and 0.5 mM PNP as sources of
carbon and energy.
Anchoring OPH onto the surface of E. coli using a Lpp-OmpA
(46-159)-fusion system, Richins et al. (1997) developed a unique
Bioremediation of Pesticide-Contaminated Soils
approach for organophosphorus pesticide detoxification. It was anticipated that immobilization of these live biocatalysts onto a solid support could provide an attractive and economical means for pesticide
detoxification as compared to immobilized enzymes or whole ceIls. The
recombinant strains effectively degraded parathion, but were relatively
unstable. To overcome this problem, a new plasmid was constructed to
express OPH onto the ceIl surface under the control of a tightly regulated tac promoter. Production of active OPH onto the ceIl surface was
highly host specific; a high rate of parathion degradation was observed
from strains JM105 and XLI-Blue (Kaneva et al. 1998).
Kim et al. (2002) reported a significantly enhanced efficiency of
degrading insecticide coumophos (CP) in cattle dip waste (CDW) using
a dense, non-growing ceIl population that functions without the addition of nutrients required for growing ceIl cultures. A recombinant
strain of Escherichia coli containing the opd gene for OPH, which is
capable of active hydrolysis of OP neurotoxins inc1uding CP, exhibited
significantly higher degradation rates with either suspended or immobilized OPH+ ceIls compared to rates with the microbial consortium naturaIly present in CDW. Of the two non-growing ceIl systems, the
detoxification rate with immobilized ceIls was approximately twice that
of freely suspended ceIls.
Carbamates, ester derivatives of N-substituted carbamic acid, are widely
used pesticides wh ich inc1ude carbofuran, carbaryl, aldicarb, propoxur
and methomyl. Several strains of Pseudomonas, Flavobacterium and
Arthrobacter degrade carbofuran (Karns and Tomasek 1991; Hayatsu
1999). Mohapatra and Awasthi (1997) identified two bacterial cultures,
P. stutzeri and Bacillus pumilis, causing a 98% loss of applied carbofuran within 30 days. A number of pathways for carbofuran biodegradation have been reported. In the major pathway, the methylcarbamate
linkage is hydrolyzed, yielding carbofuran-7-phenol (2,3-dihydro-2,2dimethyl-7-benzofuranol) and methyl amine). Growth-linked and cometabolic bio degradation appear to accelerate carbofuran degradation
(Charnay and Fournier 1994; Robertson and Alexander 1994).
The enzyme carbofuran hydrolase has been characterized from
Achromobacter sp. and the plasmid-borne gene that encodes this
enzyme has been c10ned (Karns and Tomasek 1991). Chaudhary et al.
(2002) purified a soluble carbamate hydrolase with a wide specificity
from Pseudomonas sp. 50432. Carbamate hydrolase c1eaved the
ester linkage of the N-methylcarbamates and hydrolysed several N-
R.C. Kuhad et al.
methylcarbamates and I-naphthyl acetate, but did not hydrolyse 0nitrophenyl dimethylcarbamate.
A Pseudomonas sp. was capable of growing on propoxur (2 g/l) as
a sole source of carbon and nitrogen with an accumulation of 2isopropoxyphenol as a metabolite in the culture medium (KamanavaUi
and Ninnekar 2000). Liquid ceU suspension of a carbofuran-degrading
culture strain C28 significantly improved pesticide removal from the
contaminated soil (Duquenne et al. 1996). Low and high concentrations
of glucose increased and reduced carbofuron degradation, respectively.
Among s-triazine pestieides, atrazine (6-chloro-N2 -ethyl- W-isopropyl1,3,5-triazine-2,4-diamine) is one of the most common selective herbieides used to contral an nu al grass and broadleaf weeds in a variety of
agricultural crops. Contamination of soil and groundwater and rivers
has caused the use of atrazine to be reduced in many countries and to
be banned in Germany (Dousset et al. 1997).
Degradation of atrazine in soil occurs through both biological and
non-biological mechanisms (De Souza et al. 1995; Ralebitso et al. 2002;
Wackett et al. 2002). The first step in atrazine degradation by bacteria
in the natural environment is dealkylation with the formation of diethylatrazine. The side chain of this herbieide is accessible for microbial
attack but mineralization of the s-triazine ring proceeds very slowly
(McMahon et al. 1992). However, in contrast to other herbieides, such
enhanced biodegradation has not been observed with atrazine,
cyanazine and metachlor (Rouchaud et al. 1997). Pseudomonas sp.
strain ADP, which can degrade atrazine (>1000mg/l) has been used to
study atrazine degradation under anoxie or denitrifying conditions
(Katz et al. 2000) and in atrazine-contaminated soil (Topp 2001).
White-rot fungi can transform atrazine, yielding hydroxylated and Ndealkylated metabolites (Mougin et al. 1997).
Shao et al. (1995) developed geneticaUy engineered Rhodococcus
strains capable of simultaneously degrading atrazine and simazine.
Piutti et al. (2003) described an atrazine-degrading strain of Nocardioides sp. SPI2, which contained a novel atrazine catabolic pathway
combining trzN with atzB and atze. The numbers of trzN and ITS
sequences of Nocardioides sp. SP12 were higher in a maize rhizosphere
than in bulk soil.
Park et al. (2003) studied the mineralization of sorbed atrazine in
soil and day slurries using Pseudomonas sp. strain ADP, Agrobac-
Bioremediation of Pesticide-Contaminated Soils
terium radiobacter J14a and Ralstonia sp. and developed a desorption-biodegradation-mineralization model. Characteristics of high
sorbed-phase concentration, chemotaxis and the attachment of cells to
soil particles appeared to contribute to the bioavailability of soil-sorbed
An increase in the population of atrazine-degrading microbes and
atrazine degradation was observed in soils with trisodium citrate (Katz
et al. 2000). On the other hand, sucrose addition did not significantly
improve atrazine degradation in contaminated soil by Agrobacterium
radiobacter J14a (Struthers et al. 1998). Glucose was found to reduce the
bioavailability of the soil-bound atrazine, whereas inorganic nitrogen
rather than organic nitrogen stimulated the catabolism (Ames and
Hoyle 1999). The addition of (bio) surfactants such as sodium dodecyl
sulfate and rhamnolipids was found to increase the atrazine biodegradation rates (Mata-Sandoval et al. 2000).
Strong et al. (2000) conducted a field study using recombinant E. coli
expressing the atrazine chlorohydrolase gene derived from Pseudomonas sp. ADP. Among different bio stimulation conditions such as pH,
carbon addition and phosphate addition, only phosphate increased the
atrazine bio degradation in soil plots. A combined approach of enzyme
and phosphate addition provided the best results, with a reduction in
atrazine concentration by 97 as compared to 84%, with only enzyme
addition and non-significant degradation in control plots.
The atrazine-mineralizing strain, Chelatobacter heintzii Citl, was
inoculated in two types of soils showing enhanced or low mineralization capacity (Rousseaux et al. 2003). In the case of soils adapted to
enhanced atrazine mineralization, the inoculation of C. heintzii did not
accelerate the rate of atrazine mineralization, which was essentially performed by the indigenous microflora. However, with soils that did not
mineralize atrazine, the introduction of 10 cfu g-l resulted in a three-fold
increase in atrazine mineralization capacity. The degradation patterns
of terbuthylazine (which is being used instead of atrazine in many
developed countries ) and atrazine in laboratory soil microcosms were
not significantly different (Dousset et al. 1997; van der Heyden et al.
Other Pesticides
Microbial degradation of chiral pesticides has been investigated to
understand the processes that govern enantioselectivity in biodegradation and the environmental fate of chiral herbicides. Sphingomonas her-
R.C. Kuhad et al.
bicidovorans MH, a pure bacterial strain isolated from soil, utilized both
enantiomers of a widely used chiral herbicide mecoprop [(RS)-2-(4chloro-2-methylphenoxy) propanoic acid], as the sole carbon and
energy source for growth (Zipper et al. 1998). Growth experiments
revealed a much faster disappearance of the (S) enantiomer than the (R)
enantiomer, indicating the involvement of specific catabolic enzymes in
the degradation of each enantiomer.
Two a-ketoglutarate-dependent dioxygenases have been characterized in S. herbicidovorans MH, one that is constitutively expressed and
specific for (S)-mecoprop and another specific for (R)-mecoprop which
is induced in the presence of (R) enantiomer (Nickel et al. 1997). Alcaligenes denitrificans exclusively degrades (R)-mecocarp (Tett et al. 1997).
In a laboratory soil experiment, a low concentration of mecocarp was
degraded by Alcaligenes denitrificans with extensive enantiomerization
(Buser and Müller 1997; Müller and Buser 1997). S. herbicidovorans MH
was able to completely degrade both enantiomers of the chiral herbicide dichloroprop [(RS)-2-(2,4-dichlorophenoxy) propanoic acid], with
preferential degradation of the (S) enantiomer over the (R) enantiomer
(Zipper et al. 1998).
Soil treated with linuron for more than 10 years showed a high
bio degradation activity toward methoxy-methyl urea herbicides (EIFantroussi 2000). The enrichment culture isolated from the treated soil
showed specific degradation activity towards methoxy-methyl urea
herbicides, such as linuran and metobromuran, while dimethyl urea
herbicides isoproturon, diuron and chlorotoluron were not degraded.
Mineralization of free and cell wall-bound isoproturon in soils was
found to be related to soil microbial biom ass, estimated by substrateinduced heat production (Lehr et al. 1996).
Conclusions and Future Prospects
Despite managed and controlled applications, pesticide usage will
continue to contaminate soil and groundwater. Bioremediation can
be a cost -effective method for the decontamination of pesticidecontaminated soils. The characterization of the microbial systems
involved in pesticide degradation is the important initial step in developing a bioremediation strategy. The potential for using genetically
engineered microbes for pesticide degradation has been confirmed in
the laboratory but desired results have yet to be achieved in the natural
environment. Several possible reasons for failures to date are: (1) pesticide concentrations in nature may be too low to support the growth of
microbes, (2) the microbes may be susceptibile to environmental stress,
Bioremediation of Pesticide-Contaminated Soils
and (3) more easily available organic compounds may be present.
Further study requires the identification and sequencing of gene(s)
encoding pesticide-degrading enzymes, a better und erst an ding of the
metabolie pathways which may be useful in recombinant strains, and
the construction of stress-resistant microbes. Until recently, research
into pesticide degradation by microorganisms has focused primarily on
bacteria, which grow faster and are amenable to genetic manipulation.
However, the application of white rot basidiomycetes, like Phanerochaete chrysosporium, and phytoremediation approaches hold great
pro mise for the remediation of pesticide-contaminated soil.
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Biodegradation and Bioremediation
of Explosives
Jian-Shen Zhao,I Diane Fournier,I Sonia Thiboutot,2
Guy Ampleman,2 and Jalal Hawar?
Explosives are highly energetic chemicals that release large amounts of
energy and gaseous products upon detonation in a short period of time.
The history of explosives dates back to the development ofblack powder
long before the industrial revolution started in Europe (Linder et al.
1980). Some of the most frequently manufactured and used secondary
explosives indude 2,4,6-trinitrotoluene (TNT), dinitrotoluenes
(DNT), 1,3,5-trinitrobenzene (TNB),N,2,4,6-tetranitro-N-methylaniline
(tetryl), trinitroglycerine (TNG), nitro guanidine (NQ), ethylene glycol
dinitrate (EGDN), nitrocellulose (NC), pentaerythritol tetranitrate
(PETN), glycidyl azide polymer (GAP), hexahydro-1,3,5-trinitrio-1,3,5triazine (RDX), octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX)
and 2,4,6,8,1O,12-hexanitrohexaazaisowurtzitane (CL-20) (Fig. 1).
Past and present practices with explosives such as manufacturing,
formulations, testing and training, demilitarization and open burningl
open detonation (OBIOD) have led to severe soil and groundwater contamination (Urbanski 1983; Haas and Schreiber 1990; Myler and Sisk
1991). It has been estimated that TNT alone is produced in amounts
dose to 2 million pounds a year (Harter 1985) and a single TNT manufacturing plant can generate over 1.8 x 103 m 3 of wastewater per day
(Yinon 1990). During the manufacture of RDX, up to 12mg/l may be
discharged to the environment in process wastewater (Jackson 1978).
Presently, soil and groundwater contamination by explosives is an envi1 Biotechnology Research Institute, National Research Council of Canada, 6100 Royalmount Ave, Montreal, Quebec, Canada, H4P 2R2
2 Defense Research Establishment Valcartier, Department of National Defense, Quebec,
Canada, G3J 1X5
3 Biotechnology Research Institute, National Research Council of Canada, 6100
Royalmount Ave, Montreal, Quebec, Canada, H4P 2R2, e-mail: [email protected]
Tel: +1 -514-4966267, Fax: +1-514-4966265
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
N0 2
CH 2"0-02 N
N0 2
\ON0 2 H/I
CH 2 0N0 2
, NH 2
NH =C 'NH -N0 2
,CH 2 0-02 N - 0
02 N - 0 - H 2 C '
02 N - 0 - H 2C ,
02 NN
1,3 -DNB
,N0 2
~NO' O'N~NO'
Fig. l. Structures and common names of most commonly used secondary explosives
CH 2 -{)-N0 2
CH 2 -O-N0 2
CH 2 -{)-N0 2
N0 2
°2 N
02 N 'Nr-N
CH-0-N0 2
2,4 -DNT
N0 2
CH 2 -O-N0 2
N0 2
H3 C,
Biodegradation and Bioremediation of Explosives
ronmental problem worldwide that started following intensive military
activities in World Wars land II, the Cold War and the subsequent
demilitarization activities (Pritsche et al. 2000).
Most energetic chemicals are toxic to aquatic and terrestrial species
(Taimage et al. 1999; Sunahara et al. 2001). Increased public awareness
of the toxicity and the risk associated with energetic compounds necessitate the development of cost-effective bioremediation technologies for
their removal. Little information is available on in-situ bioremediation
technologies for any explosive, but ex situ bioremediation technologies
such as composting (Bruns-Nage1 et al. 2000) and bioslurry technology
Oerger et al. 2000) are more advanced. The present review summarizes
past and current research undertaken to biodegrade these chemicals
.under both aerobic and anaerobic conditions.
Structural Properties and Effect on Biodegradation
Aromatic explosives such as TNT are characterized py having the highly
oxidized nitro (-NO z) functional groups directly attached to an aromatic Jr-system through a spz carbon atom. Therefore, the chemical, biochemical and microbial reactivity of TNT and related explosives are
determined by the presence of the -NO z groups and by the aromatic
character of the benzene ring. In the case of TNT, the Jr-e1ectrons and
the carbon atoms of TNT are shielded from external enzymatic (and
chemical) attack by the steric effects caused by the presence of the three
oxidized -NO z groups and one -CH 3 group (Hawari et al. 2000a). Consequently, TNT often undergoes sequential reduction under both
aerobic and an aerobic conditions to produce corresponding nitroso and
eventually monoamino (ADNT) and diamino (DANT) derivatives. Only
under strictly anaerobic conditions does ADNT reduce further to
produce 2,4,6-triaminotoluene (TAT) (McCormick et al. 1976; Preuss
and Rieger 1995; Rieger and Knackmuss 1995; Lewis et al. 1996; Hawari
et al. 1998) which is highly reactive, and can polymerize or irreversibly
bind to soi! (Rieger and Knackmuss 1995; Preuss and Rieger 1995; Lenke
et al. 2000).
Also, the reduction of the benzene ring has been reported to occur
in many aerobic bacteria and yeasts (Zaripov et al. 2002; Table 1)
through the addition of one or two hydride ions to the aromatic ring
with the subsequent formation of a Meisenheimer complex (Vorbeck et
al. 1994; Vorbeck et al. 1998). Apart from polymerization (Pak et al.
2000), the fate of the Meisenheimer complex is not c1ear.
The heterocyc1ic nitramine explosives RDX, HMX and CL-20 lack
the aromatic stability of TNT and a successful initial attack on these
sp. clone Ab,c
MAOl b,c
aeruginosa MX"
N0 2-
Transformation products a
Bacillus sp.
cloacae PB2
sp. HL 4NT-l
vaccae JOB-Sb
Table 1. Aerobic transformation of TNT by bacteria and their major products
VasiIyeva et al. (2000)
Kalafut et al. (1998)
Vanderberg et al.
Duque et al. (1993);
Ha"idour and Ramos
Alvarez et al. (1995)
Vorbeck et al. (1994)
Kalafut et al. (1998)
French et al. (1998)
Pasti-Grigsby et al.
Kalafut et al. (1998)
Montpas et al. (1997)
Vorbeck et al. (1998)
Pak et al. (2000)
Martin et al. (1997)
Fiorella and Spain
ND, no data available.
'Mineralization either not reported or <1%; +, produced; -, not produced; T, toluene; NT, nitrotoluene; DNT, dinitrotoluene; 2A4NT,
2-amino-4-nitrotoluene; ADNT, aminodinitrotoluene.
bThis strain produced diaminonitrotoluene.
'These strains produced azoxynitrotoluenes.
dThis strain produced aminodimethyltetranitrobiphenyl.
chromofuscus All b
Staphylococcus sp.
fluorescens I-C'·d
erythropolis HL
J.-S. Zhao et al.
molecules can lead to ring cleavage and decomposition. It has been
reported that the thermal cleavage of the N-N0 2 bond in RDX leads
to spontaneous decomposition because the inner C-N bonds in the
chemical become very weak «2kcallmol) (MeHus 1990). For instance,
thermal decomposition of both RDX and HMX pro duces formaldehyde
(HCHO), carbon dioxide (C0 2 ), nitrite (N0 2-), ammonia (NH 3 ), nitrous
oxide, and N2 (Linder 1980; Urbanski 1983). Recent experimental evidence has shown that the non-aromatic cyclic nitramines will decompose upon attack by enzymes and microbes under both aerobic and
anaerobic conditions. Evidence of RDX ring cleavage during degradation of RDX by Stenotrophomonas maltophilia PBl has been provided
by Binks et al. (1995). Subsequent work in our laboratory showed that
initial denitration of RDX by Klebsiella pneumoniae SCZ-l (Zhao et al.
2002), an isolate from an aerobic sludge, and Rhodococcus sp. DN22, a
soil isolate (Fournier et al. 2002), can lead to ring cleavage and spontaneous decomposition in water to produce HCHO, CO 2 , N20 and NH 3•
None of the above reactions described for RDX and HMX were observed
with the aromatic TNT, emphasizing the importance of molecular structure on the fate of explosives during biodegradation.
Biodegradation of Cyclic Nitramine Explosives
Biodegradation of RDX and HMX by Anaerobic Bacteria
Several research groups employed mixed an aerobic microbial cultures including methanogens (Adrian and Lower 1999; Adrain and
Chow 2001), acetogens (Beller 2002), nitrate reducers (Freedman and
Sutherland 1998), sulfate reducers from lake sediments (Boopathy et al.
1998) and individual isolates such as Clostridium bifermentans (Regan
and Crawford 1994), Providencia rettgeri, Citrobacter freundii, Morganella morganiii (Kitts et al. 1994), and Serratia marcescens (Young et
al. 1997) to transform RDX (Table 2), but in most cases no clear details
on ring cleavage products or mineralization were provided. Despite
these earlier efforts, there is still insufficient information regarding the
role of individual anaerobic isolates and their products during RDX
degradation. Without these details, particularly those concerning
specific initial enzymatic reactions and their subsequent products, it
will be difficult to understand degradation pathways and to optimize
degradation towards mineralization. Understanding the intermediary products and the underlying mechanisms of their formation is
Biodegradation and Bioremediation of Explosives
Table 2. Microbial degradation of RDX
CO 2
Anaerobic bacteria (facultative or obligate)
Eseheriehia eoli
Zhao et al.
Young et al.
Pudge et al.
freundii NS2 a
morganii B2b
rettgeri BI c
Aerobic bacteria
Rhodocoeeus sp.
Rhodoeoecus sp.
rhodoehrous 11 Y
Rhodoeoeeus sp. A
maltophilia PB1
sp. YH1
sp. 140
Young et al.
Young et al.
Kitts et al.
Kitts et al.
Kitts et al.
Regan and
Coleman et al.
Fournier et al.
Seth-Smith et al.
Jones et al.
Binks et al.
Brenner et al.
Yang et al.
Young et al.
Young et al.
Sheremata and
Hawari (2000)
}.-S. Zhao et al.
Table 2. Continued
CO z
Fernando and
Aust (1991)
Stahl et al.
ND, no data available.
a This strain degraded 10-15% of 50.uM of HMX.
b This strain degraded 85% of 50.uM of HMX.
C This strain degraded 10-15% of 50.uM of HMX.
essential to allowing prediction and enhancement of complete
bio degradation of the two explosives.
McCormick et al. (1981) studied bio degradation of RDX in anaerobic sludge and proposed a pathway based on the sequential reduction
of the -N02 groups to produce hexahydro-1-nitroso-3,S-dinitro-1,3,Striazine (MNX), hexahydro-1,3-dinitroso-S-nitro-1,3,S-triazine (DNX),
and hexahydro-1,3,S-trinitroso-1,3,S-triazine (TNX). The nitroso compounds were proposed to undergo further transformation to hydroxylamino derivatives, HONH-RDX, which subsequently cleave to produce
formaldehyde (HCHO), methanol (CH 30H), hydrazine (NH 2NH 2), and
dimethyl hydrazine [(H 3ChNNH2]. No microorganisms or enzymes
were identified in the study of McCormick et al. (1981). Recently,
Kitts et al. (2000) have reported the possible involvement of type I (2ereduction,oxygen-insensitive) nitroreductase in enterobacteria in the
degradation of RDX, but the authors did not identify any products.
We have found recently that anaerobic sludge successfully degraded
both RDX (Halasz et al. 2002) and HMX (Hawari et al. 2001b) to N20,
HCHO and CO 2. In contrast to the results of McCormick et al. (1981),
we were unable to detect hydrazine or dimethyl hydrazine. In addition to the above end products, we detected methylenedinitramine,
02NNHCH2NHN02 as a key RDX (and HMX) ring cleavage product.
Methylenedinitramine, was found to be unstable during incubation
and decomposed spontaneously to N20 and HCHO (Halasz et al. 2002).
Earlier studies suggested the presence of methylenedinitramine as an
RDX intermediate, often as abiphosphate salt, during the thermal and
photochemical degradation of the energetic chemical (Bose et al.
1998). Recently, Oh et al. (2001) also found the formation of methyl-
Biodegradation and Bioremediation of Explosives
enedinitramine during anaerobic RDX degradation. Interestingly,
methylenedinitramine was also found during photodenitration of RDX
at 350nm in water (Hawari et al. 2002), suggesting that initial denitration (enzymatic or chemical) of the chemicalleads to ring cleavage and
Zhao et al. (2002) used Klebsiella pneumoniae strain SCZ-1, a facultative anaerobic isolate from domestic anaerobic sludge, to successfully
degrade RDX and its mono-nitroso derivative MNX under an aerobic
conditions. Strain SCZ-1 degraded RDX to HCHO, CH 30H (12%), CO 2
(72%) and N2 0 (60%) through the intermediary formation of methylenedinitramine (Fig. 2). Likewise, MNX degraded to HCHO, CH30H and
N20 (16.5%) with a removal rate [0.39 fLlllolh-1 g-l (dry ceIl weight,
dcwt1] similar to that of RDX [OA1fLlllolh-lg-l (dcwt1; biomass, 0.91
mg (dcw) t 1]. These findings suggested the possible involvement of a
common initial reaction, possibly denitration, foIlowed by ring cleavage
and decomposition in water (Fig. 2). The trace amounts of MNX,
detected during RDX degradation, and DNX during MNX degradation,
suggested that another minor degradation pathway was also present
that reduced nitro groups to the corresponding nitroso groups.
Although the isolated strain improved our understanding of the pathways involved in the degradation of cyclic nitramines, its use for remediation might not be practical because of the slow rate of degradation.
However, the high mineralization yield observed during RDX degradation should encourage further research to optimize various physiological parameters to enhance rates of degradation of RDX.
Biodegradation of RDX Under Aerobic (onditions
Recent experimental evidence has shown that RDX and HMX can be
degraded under aerobic conditions (Hawari 2000). Several soil bacterial strains including Stenotrophomonas maltophilia (Bink et al. 1995),
Rhodococcus sp. strain A (Iones et al. 1995) and Rhodococcus sp. strain
D22 (Coleman et al. 1998), biodegraded RDX when used as a nitrogen
source for bacterial growth (Table 2). Most of these earlier studies
showed the initial formation of nitrite, but no other initial intermediates were reported. Binks et al. (1995) reported the cleavage ofRDX with
Stenotrophomonas maltophilia, and reported NMR data on the occurrence of ring cleavage. On the other hand, Sheremata et al. (2000)
employed a fungus, Phanerochaete chrysosporium, to mineralize RDX
using glycerol as a carbon source and detected the mononitroso derivative (MNX) as the sole RDX intermediate product, which did not accumulate. At the end of the experiment, nitrous oxide (60%) and carbon
02 N
H 2N- N02
N0 2-
'NO 2
/2 MN~X2
N2 0 , HCHO, MeOH
0 N/ '-...../N'NO
-1 ('I ': .
N 20 + H 20
" - N0 2
C02 + MeOH
N0 2 -
RD~X •
02 N /
02 N /
'-...../_ 'N0 2
N0 2
Nitroso route (path b)
Fig.2. Postulated pathways for an aerobic degradation of RDX with Klebsiella pneumoniae strain SCZ1 isolated from domestic an aerobic sludge. (Modified from Zhao et al. 2002)
Denitration route (path a)
Biodegradation and Bioremediation of Explosives
dioxide (55%) were obtained as end products. High mineralization (66.6
± 4.1%) of RDX (0.028mg!l) by Phanerochaete chrysosporium was
reported earlier by Fernando and Aust (1991), but with no description
of other metabolites.
Fournier et al. (2002) reported the biodegradation of RDX in
Rhodococcus sp. strain DN22 to nitrite (N0 2-) (30%), nitrous oxide
(N 20) (3.2%), ammonia (10%) and formaldehyde (HCHO) (27%), which
later converted to carbon dioxide. LC-MS (ES-) analyses indicated the
formation of a dead-end product with a deprotonated molecular mass
ion [M-H] at 118Da that was later identified as 4-nitro-2,4-diazabutanal by NMR. Based on the re action stoichiometry, Fournier et al.
(2002) proposed adegradation pathway for RDX that involves an initial
denitration step followed by ring cleavage to yield formaldehyde and
the dead-end product 4-nitro-2,4-diazabutanal (Fig. 3).
Subsequent work focused on the understanding of the enzymatic step
involved in the denitration of RDX. Bhushan et al. (2003) evaluated the
role of the cytochrome P450 enzyme in RDX biotransformation. The
authors used a cytochrome P450 2B4 (EC from rabbit liver as
a model system to provide insight into the initial reaction mechanism(s)
involved in RDX degradation in Rhodococcus sp. DN22. Both the
noN NH CH NH CHO autodecomposlllon
002 (30% in C)
Fig. 3. Postulated pathways for aerobic degradation of RDX with Rhodococcus sp.
strain DN22. (Modified from Fournier et al. 2002)
J.-S. Zhao et al.
cytochrome P450 2B4 and intact ceIls of Rhodococcus sp. DN22 catalyzed
the release of two nitrite ions from each reacted RDX molecule. The two
systems revealed substantial similarities in their product distribution
and their inhibition by cytochrome P450 inhibitors. Bhushan et al.
(2003) proposed that cytochrome P450 2B4 could catalyze two single
electron transfers to RDX, thereby causing double denitration and
leading to spontaneous hydrolytic ring cleavage and decomposition to
produce 4-nitro-2,4-diazabutanal (Fournier et al. 2002). The results
provide strong evidence that a cytochrome P450 is the key enzyme
responsible for RDX biotransformation by strain DN22 as suggested by
Coleman et al (2002). Seth-Smith et al. (2002) found that the gene
responsible for the degradation ofRDX in Rhodococcus rhodochrous 11 Y
was also a constitutively expressed cytochrome P450-like gene (2002).
Biodegradation of Polycyclic Nitramine Explosive CL -20
High density polynitropolyaza-caged compounds contain high energy
that attracts the military to use them as explosives and propeIlants. A
typical energetic chemical of this family is 2,4,6,8,10,12-hexanitro-,2,4,6,8,10,12-hexaazatetracyclo [,9.0 3,11] dodecane or hexanitrohexaazaisowurtzitane (commonly known as CL-20; Fig. 1), which has
recently been synthesized by Nielsen et al. (1998) and later adopted by
Thiokol for large-scale production. As mentioned ab ove , CL-20 is a
newly released energetic chemical, whose bio degradation is not weIl
known. However, the energetic chemical is a strained heterocyclic
nitramine which, like RDX and HMX, contains the characteristic NNO z functional groups that should determine the chemical and microbial properties of the explosive. In addition, both RDX and HMX are
cyclic oligomers of methylenenitramine, CHz-N-NO z, [(CH zNNO z)3
for RDX and (CH 2NN0 2 )4 for HMX], whereas CL-20 is a molecule with
a caged rigid structure that contains the repeating unit CH-NN0 2
(WardIe et al. 1996; Nielsen et al. 1998) and C-C bonds. Therefore CL20 should exhibit some differences in its enzymatic and microbial
degradation from those of RDX and HMX.
Trott et al. (2003) have recently found that when CL-20 was incubated
in various non-sterilized garden soils, the explosive was found to
degrade rapidly. After 2 days of incubation, about 80% of the initial CL20 disappeared. A CL-20 degrading bacterial strain, Agrobacterium sp.
JS71, was isolated from enrichment cultures with garden soil as inoculum, succinate as carbon source, and CL-20 as the nitrogen source.
Growth experiments revealed that the strain JS71 consumed 3 mol of
nitrogen per 1 mol of CL-20, which indicates that the molecule is
Biodegradation and Bioremediation of Explosives
degraded extensively. The authors concluded that CL-20 was biodegraded extensively in soil and not simply oxidized or reduced and therefore might be less persistent in the environment than RDX and HMX.
However, no information is available yet on the pathway or enzymes that
are involved in CL-20 degradation.
Biodegradation of RDX and HMX in Sediments
RDX and HMX in contaminated soil can find their way into coastal
water and estuarine locations and, in turn, can be transported further
to sediments. These toxie contaminants can subsequently migrate from
sediments and accumulate in other aquatie living organisms causing
adverse ecological effects. Marine sediments are mostly an aerobic with
low nutrients and high salt content (Bowman 2001; Karl and Dore 2001).
Also, marine sediments are subjected to high pressure under cold temperatures (ca. 4°C). Very little data are available on the degradation of
explosives in a very challenging environment such as sediments and
estuaries. Boopathy et al. (1998) reported the use of sulfate-reducing
bacteria to degrade RDX in fresh water sediments. Other groups have
isolated psychrophilic marine sulfate-reducing bacteria (Knoblauch et
al. 1999). Sulfate-reducing bacteria account for the degradation of more
than 50% of the organic matter in marine sediments (J0rgensen 1982).
Biodegradation of TNT and Other
Polynitroaromatic Explosives
Biodegradation of TNT by Aerobic Bacteria
Aerobic bacteria have been shown to use oxygenase or nitroreductase
as an initial enzyme to degrade mono- and di-nitroaromatic compounds (Spain 1995a, b; Nishino et al. 2000). Also, Zhao et al. (2000,
2001) found that a 3-nitrophenol-induced metabolie system in
Pseudomonas putida 2NP8 effectively converted a wide range of
nitroaromatic compounds, including dinitrotoluenes to ammonia. In
contrast, little information is available on the cleavage of the TNT ring
by aerobic bacteria.
However, many aerobic bacteria (Table 1) were reported to transform
TNT through either reduction of a nitro group or the benzene ring (Fig.
4). Transformation of TNT via the reduction of a nitro group pro duces
J.-S. Zhao et al.
H2 N
Stringent anaerobes
NH 2
N -formyJ ADNT
N -acetyl or
NH 2
GH 3
M eisenheimer I
M eisenheimer 11
Facultative anaerobes
N0 2 -
Facultative anaerobes
Biodegradation and Bioremediation of Explosives
mono amino dinitrotoluenes (ADNT) and diamino mononitrotoluenes
(DANT) (Spain 1995b; Fiorella and Spain 1997; Fuller and Manning
1997; Martin et al. 1997; Fritsche et al. 2000). The -NH z functional
groups in ADNT and DANT are reactive and thus can (bio )transform
further to several other products including azo, azoxy and acetyl derivatives with little or no mineralization (Lewis et al. 1996; Fritsche et al.
One possible route to degrade TNT is by either denitration or direct
attack on the benzene ring. The reduction of the benzene ring of polynitro-aromatic compounds, by a hydride ion, to Meisenheimer complexes,
was first reported by Lenke and Knackmuss (1992) in the degradation
of picric acid (2,4,6-trinitrophenol) by Rhodococcus erythropolis HL
24-2. The picrie acid Meisenheimer complex released nitrite and produced 2,4-dinitrophenol. Later, a similar Meisenheimer metabolite was
reported byVorbeck et al. (1994) in the degradation ofTNT in Mycobacterium sp. (Fig. 4). Vorbeck et al. (1994) found that the TNT Meisenheimer metabolite was unstable and either reverted back to TNT or
released nitrite, but no dinitrotoluene was detected. Vorbeck et al.
(1998) also found that the Meisenheimer adduct of TNT with one
hydride ion could be easily reduced further by a second hydride ion to
form a yellow adduct, Meisenheimer II (Fig. 4). The pentaerythritol
tetranitrate reductase in Enterobacter cloacae PB2 (French et al. 1998)
or the xenobiotie reductase from Pseudomonas fluorescens I-co (Pak et
al. 2000) were reported to reduce both the nitro group and the benzene
ring of TNT (Fig. 4). No dinitrotoluene was produced from denitration
of the TNT Meisenheimer complex in any of the above-described TNT
biotransformations. Pak et al. (2000) found that nitrite could be released
from Meisenheimer complex II when it condensed with hydroxylamino derivatives to amino-dimethyl-tetranitrobiphenyl (ADMTNBP)
through an oxidative abiotic reaction (Fig. 4). Fiorella and Spain (1997)
also suggested that nitrite might be released via oxidation of dihydroxylaminonitrotoluene. Furthermore, Duque et al. (1993) found that
Pseudomonas sp. ClSl, astrain capable of growth on TNT, dinitrotuluene and 2-nitrotoluene as nitrogen source, converted TNT to 2,4-
Fig. 4. Pathways summarizing biotransformation routes of TNT under both aerobic
and anaerobic conditions. NSDNT Nitrosodinitrotoluene; HADNT hydroxyldinitrotoluene; ADNT aminodinitrotoluene; ADMTNBP aminodimethyltetranitro-biphenyl;
AZT 4,4',6,6'-tetranitro-2,2'-azoxytoluene or 2,2'6,6' -tetranitro-4,4' -azoxytoluene; TAT
2,4,6-triaminotoluene. (Modified from Lewis et al 1997; Hawari et al. 1998,1999,2000;
Pak et al 2000; Esteve-NUfiez et al. 2001)
J.-S. Zhao et al.
dinitrotoluene, 2-nitrotoluene and toluene. Duque et al. (1993) and
Haidour and Ramos (1996) further proposed that the bacterium denitrated TNT via Meisenheimer complex, based on their reported observation of 2,4-dinitrotoluene and nitrite, however, this hypothesis was
not confirmed by the work ofVorbeck et al. (1998).
The release of nitrite and production of the denitrated product 2amino-4-nitrotoluene was also reported by Kalafut et al. (1998) in the
degradation of TNT by Bacillus sp., Pseudomonas aeruginosa, Staphylococcus sp. Also, Martin et al. (1997) observed the release of nitrite and
production of 2,4-dinitrotoluene from TNT by Pseudomonas savastanoi.
So far, the mechanism for the production of dinitrotoluene from TNT
is not clear. Recently, Zaripov et al. (2002) have reported that yeast
cells also transformed TNT via either the reduction of the benzene
ring to Meisenheimer complex or the reduction of the nitro groups to
Thus far, all literature reports indicate that TNT can be biotransformed under aerobic condition, but with poor mineralization (Table
1, Fig. 4). As the preceding discussion reveals, initial denitration of TNT
or its reduction to a Meisenheimer complex, followed by denitration,
can lead to mineralization. Thus far, arguments found in the literature
on the merits of these two important pathways are controversial and
require further investigation.
Biodegradation of TNT by Anaerobic Bacteria
Most reported studies on the an aerobic treatment of TNT have
been reviewed recently by Hawari et al. (2000b). Thus far, TNT biodegradation under anaerobic conditions leads to the initial production of
4-amino-2,6-dinitrotoluene (4-ADNT), 2-amino-4,6-dinitrotoluene (2ADNT), 2,4-diamino-6-nitrotoluene (2,4-DANT), and 2,6-diamino-4nitrotoluene (2,6-DANT) (Fig. 4). Naumov et al. (1999) observed that
two lactic acid bacteria, Lactobacillus fermentum BS3601 and Lactobacillus plantarum BS3604, transformed TNT mainly to hydroxylaminodinitrotoluene (95% of TNT removed). Reduction of the -N0 2 groups
in TNT is found to be regioselective and normally favors the reduction
of the -N0 2 group at the para position relative to the methyl group, to
produce the corresponding mono- and diamino derivatives 4-ADNT
and 2,4-DANT. 2,4,6-Triaminotoluene (TAT) is only observed under
strictly anaerobic conditions (McCormick et al. 1976; Preuss and Rieger
1995; Rieger and Knackmuss 1995; Lewis et al. 1996; Khan and Hughes
1997; Hawari et al. 1998; Nishino et al. 2000; Fig. 4). TAT was reported
to undergo deamination in sulfidogenic bacteria (Boopathy et al. 1993;
Biodegradation and Bioremediation of Explosives
Preuss and Rieger 1995), whereas Funk et al. (1993) reported the formation of p-cresol from TNT bio degradation. Later, Hawari et al. (1998)
showed that when [13 CH 3 ]TNT was treated with anaerobic sludge
neither [13 CH 3 ]toluene nor [pßCH 3 ]cresol was detected, indicating the
absence of denitration or deamination in the anaerobic sludge. We also
found that upon exposure to air at pH 2, TAT could abiotically convert
to 2,4,6-trihydroxyltoluene (Hawari et al. 1998). Once again, most reported studies on the degradation of TNT under anaerobic conditions
did not show any significant mineralization. Further research should
focus on microbes that lead to denitration and ring cleavage as prerequisite steps to enhance mineralization.
Biodegradation of TNT by Fungi
The widespread presence of fungi in nature and their ability to produce
extracellular enzymes including lignin peroxidase (LiP), manganesedependent peroxidase (MnP), and laccase, capable of breaking several
bonds, make them attractive candidates to degrade TNT. Close to 1.5
million different fungal species are reported to colonize a wide range of
habitats, of which 91 fungal strains were tested and most were found
capable of biodegrading TNT (Fritsche et al. 2000; Table 3). Phanerochaete chrysosporium is reported to mineralize TNT under ligninolytic conditions, astate identified by the presence of peroxidase
enzymes such as lignin peroxidase (LiP) and manganese-dependent
peroxidase (MnP). For example, Phanerochaete chrysosporium
(Hodgson et al. 2000) and Clitocybula dusenii TMb12 (Scheibner et al.
1997a) mineralized more than 30% of TNT in liquid cultures under
ligninolytic conditions. Also, several authors (Scheibner et al. 1997b;
Scheibner and Hofrichter 1998; Hofrichtetet al. 1999) reported that
MnP from basidiomycetous fungi was capable of converting TNT and
its reduction products (hydroxylamino- and amino-dinitrotoluenes) to
carbon dioxide in a relatively high yield.
The initial products from TNT biotransformation with
Phanerochaete chrysosporium were nitrosodinitrotoluene (NSDNT), 2hydroxylamino-4,6-dini trotoluene (2-HADNT), 4-hydroxylamino-2,6dinitrotoluene (4-HADNT) and mono- and diaminonitrotoluenes
(ADNT and DANT) (Barr and Aust 1994; Hawari et al. 1999; Fig. 4).
Under non-ligninolytic conditions (no LiP) the fungi transformed
initial TNT products further to azo, azoxy, phenolic, and acylated (acetylated and formylated) derivatives. Hawari et al. (1999) detected nine
different acetylated or formylated products. The azoxy compounds
(4,4',6,6' -tetranitro-2,2' -azoxytoluene, 2,2' ,6,6' -tetranitro-4,4' -azoxy-
J.-S. Zhao et al.
Table 3. Biodegradation of TNT by Fungi. (Adapted from Hawari et al. 2000b)
Fungal strain
CO 2
Wood-rotting/white rot basidiomycetes
Phanerochaete chrysosporium ATCC 1767
Phanerochaete chrysosporium BKM-F-1767
Phanerochaete chrysosporium BKM-F-1767
Phanerochaete chrysosporium BKM-F-1767
Phanerochaete chrysosporium
Fomes fomentarius MWFOl-4
Trametes versicolor TM5
Litter-decaying basidiomycetes
Agaricus eastivalis TMAest1
Agrocybe praecox TM70.3.1
Clitocybe odora TM3
Coprinus comatus TM6
Micromycetous fungi
Alternaria sp. TMRZ/WN2
Aspergillus terrus MWi458
Fusarium sp. TMS21
Mucor mucedo DSM810
Neurospora crassa TM
Penicillium frequentans ATCC96048
Rhizoctonia solani MWi5
toluene) were possibly the condensation products between HADNT
and NSDNT (Fig. 4). The azoxy compounds did not accumulate and
they reduced further to the azo derivatives (4,4',6,6'-tetranitro-2,2'azotoluene, 2,2',6,6' -tetranitro-4,4' -azotoluene) and the hydrazo-derivatives, 4,4',6,6' -tetranitro-2,2' -hydrazotoluene, 2,2',6,6' -tetranitro-4,4'hydrazotoluene. However, more of the intermediate(s) need to be
investigated further.
In general, biodegradation of TNT by ligninolytic fungi (Clitocybula
dusenii TMb12 and Phanerochaete chrysosporium) (Table 3) produced
high er mineralization amounts than those normally obtained by bacteria (Table 1). It has been suggested that TNT degradation by either LiP
or MnP (Barr and Aust 1994; Stahl and Aust 1995; Fritsche and
Hofrichter 2000; Fritsche et al. 2000) is mainly attributed to free radical
reactions, which can be verified using electron spin resonance spectroscopy (ESR). However, these free-radical reactions might be inhibited in soil because of its free-radical quenching effects.
Biodegradation and Bioremediation of Explosives
The current state of literature shows that TNT can biotransform quite
readily und er both aerobic and anaerobic conditions with little to no
mineralization. Therefore, the discovery of new microorganisms
capable of mineralizing TNT and other polynitroaromatic explosives,
such as tetryl, is needed. As an alternative solution to minimize the
risks associated with the presence of TNT in soil, several groups of
researchers promoted the formation of amine products from TNT and
forced them to undergo irreversible binding onto soil (immobilization)
(Lenke and Knackmuss 1992; Achtnich et al. 1999; Drzyzga et al. 1999;
Knicker et al. 1999). Such a strategy still requires development of an
understanding of the mechanisms of interactions of TNT and its transformed products with the soil matrix.
Safety Procedures
Due to their explosive characteristics, careless implementation of
sampling procedures can lead to detonation, deflagration or burning
under various stimuli such as friction and electrostatic discharge. Thus,
specific safety procedures must be followed while sampling at any explosive-contaminated site, particularly those that have unexploded ordnances (UXOs). It is expected that contaminated sites may contain more
than one explosive, and in the case of nitroglycerine (NG) an organic
vapor protective respiratory mask should be worn at all times during
sampling and sample preparation. Although this protective equipment
is not needed for other explosives, powder from other explosives might
be carried with soil dust. Under these circumstances a dust mask must
be worn during sampling and sample preparation. Also protective clothing, gloves and glasses should be worn at all times while working with
explosives-contaminated materials to avoid dermal contact and harmful
In general, explosives are labile molecules and, once in the environment,
they can (bio )transform, migrate through subsurface soil to reach
groundwater and/or be transported to plant tissues. Polynitroaromatic
explosives, such as TNT, biotransform quite readily under both aerobic
and anaerobic conditions to initially produce amines that can react
further to give the acyl-, azo- and azoxy-derivatives with little mineralization. Fungi, however, can mineralize TNT in liquid culture medium
J.-S. Zhao et al.
(Fritsche et al. 2000). No mineralization-based bioremediation technologies are known for TNT, or any other explosive, and therefore there
is a need for new natural or engineered enzymes and microorganisms
to promote mineralization of TNT and its biotransformed products. On
the other hand, the non-aromatic RDX and HMX explosives can be
biodegraded to produce nitrous oxide and formaldehyde. Preliminary
data with the newly synthesized polycyclic nitramine explosive CL-20
also showed that the compound can biodegrade via initial denitration.
As for sediments and estuaries, no rigorous data are available on the
degradation of explosives, particularly RDX and HMX. The enormous
marine military activities introduce undisclosed amounts of such energetic chemieals into marine aquatic systems.
Acknowledgements. The authors are grateful to the Defence Research
and Development Canada (DRDC), Valcartier (Quebec, Canada) of the
Canadian DoD for providing us with financial help and some of the
energetic chemicals while conducting parts of this research. We would
also like to thank the US DoD Strategie Environmental Research and
Development Program (SERDP) for financial funding of the project on
RDX and HMX (CU1213) and CL-20 (CP 1256). Finally, we would like
to thank the US Navy (ONR #N000140310269) for the financial support
on the bio degradation of RDX and HMX in sediments.
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Biological Treatment of Metallic Pollutants
Brendlyn D. Faison 1
This chapter addresses the use of biological technologies for the treatment of hazardous metals and radionuclides released into subsurface
media - soils and sediments. The subsurface is defined here as occurring at least 30 cm from the surface, and includes the vadose and saturated zones. The H20 solubility of metals is considered critical to their
human and environmental hazard assessment. Water is considered to
be present in arid, humid, and hydric environments. Water exists either
as a fluid filling all pores within sediments or saturated soils; as a vapor
in the unsaturated zone; or as films on particle surfaces and within the
interstices of all natural media. The latter kind of water includes "intrinsie" moisture, which is almost impossible to remove. The following text
will consider the (bio )chemistry that occurs within or adjacent to H20
in all its forms.
Definition of Scope
Modern periodic law, based on valence shell electronic configuration,
divides the known elements into four distinct blocks (Fig. 1). Environmental contaminants described here include the metallic f·block heavy
elements [including most radionuclides and artificial elements; so designated by the US Department of Energy (DOE)], the d-block metals
(including the radionuclide technetium), and the p-block elements
(metals, non-metals, and metalloids). Metals falling within the s-block,
including the inert elements, will not be treated here; nor will their
radioisotopes. Non-metals will be considered only as they affect metal
1 Department of Biological Sciences, Hampton University, Hampton, Virginia 23668,
USA, e-mail: [email protected], Tel: + 1-757-7275257,Fax:+ 1-757-7275961
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
.. ,
~ ----
Ar j
- - ,- -
Fig. 1. Periodic table of the elements, reflecting both the traditional chemistry of the elements and the current
understanding of their biology. The following classes of elements can be differentiated: hydrogen; alkali metals;
alkaline earth metals; heavy elements (US Department of Energy classification), including most radionuclides;
transition metals; "non-transition" metals; "other" metals; metalloids; non-metals; noble elements; known to be
bioderivatized (*)
Biological Treatment of Metallic Pollutants
(bio )chemistry. Elements that undergo oxidation/reduction (redox)
reactions (Text Box 1) and form ionic bonds and complexes are the
focus of this chapter. This text will examine bioinorganic chemistry as
it pertains to cell growth, activity and survival.
Biologically Relevant Elements
Microbiology and chemistry are usually thought to intersect only in the
lower periods - specifically at C, H, N, 0, P, and S. These are the elements
that define biomass, or cell structure. Cell function, however, exploits
the chemical energy contained within these and other elements. This
chapter focuses on prokaryotes - archae and bacteria - that are freeliving rather than associated with plant roots (see Chaps. 6 and 7 on
phytoremediation). The bioinorganic chemistry of eukaryotes has not
been studied as thoroughly in the context of bioremediation. However,
studies of these high er organisms have illuminated three key activities
of metals. These activities include the bridging of protein residues and
domains, e.g., Mg in chlorophyll; mediating protein-ligand interactions,
e.g., Zn in integrins; and catalyzing nucleophily or electron transfer, e.g.,
Fe or Mn in oxidoreductases. The last activity is the most relevant to
the biological treatment of metallic pollutants.
Microbial Transformation of Metals
Energetic metabolism is typically described as respiration, whereas
aerobic respiration describes the transfer of electrons to O2 , accompanied by the release of chemical energy. Anaerobic respiration targets
other elements as electron acceptors. Traditionally, these other elements
were considered only as their oxides: nitrate (NOn and sulfate (SO/-).
The discovery of the methanogenic archae revealed that one oxide of C,
carbon dioxide (C0 2 ), could also be considered an energetic substrate.
However, methylotrophic carboxydobacteria that use carbon monoxide
(CO) as an energy source have since been found. It is important here to
recognize that the use of these C oxides is being considered in the context of lithotrophy, not autotrophy. The relevant phenomenon common
to these anaerobic respirers is that electrons are transferred to C, N and
S rather than to 0; the atoms are electrochemically reduced. The microbial redox of low-atomic-number elements always involves transfers of
electrons in pairs. This activity is usually linked metabolically to the
oxidation of C, the main structural or nutrition al element. However,
B.D. Faison
Bioinorganic chemistry fundamentals
Atoms bond to achieve stability. Ionic bonds involve an apparently
complete transfer of electrons from one atom to another. The
exchange of ions between different dissolved species, or between
dissolved species and particulate matter that has a surface charge
(e.g., silicate rock or organic colloids), is common. Ion exchange is
reversible and usual in soils, where metal ions (once associated with
silicates or other geological materials) are present in high titers.
Ion-exchange reactions are affected by pH, as both H+and OH- may
participate in this activity. Covalent bonds involve a less drastic
transfer of electrons between or among elements. The bond results
from the sharing of electrons, such that each nuclide involved
appears to have a filled outermost shell. The distinction between
ionic and covalent bonding is not absolute; covalent bonds have a
partially ionic character.
Coordinate associations are a special sort of covalent bond that
were first described through the Lewis electron theory. Electrons
can be exchanged between substances via the formation of coordinate bonds - a form of covalent bond. Certain molecules or ions
can combine with others by forming a covalent bond with two electrons from a second molecule or ion. The first, or electron "acceptor", is called a Lewis acid or electrophile. The second substance,
which forms the covalent bond by donating a pair of electrons, is
called a Lewis base or nucleophile. Lewis acids and bases are thus
conceptually very similar to traditional, or Bronsted-Lowry, acids
and bases. However, it is specifically transition metals (those with
d- and f -shell electrons and thus including most radionuclides),
rather than other electropositive species, that tend to form coordinate complexes. All metal atoms or ions are Lewis acids. Most
anions are Lewis bases. Complexes can form between metals and
molecules of water, or between metals and ionic species such as
CO/ - or OH-.
Within aqueous systems - including soils and sediments - 0 may
participate in ionic, covalent, and/or coordinate bonding. Most
metals easily combine with 0 to form metal oxides, and many ores
consist of metal oxides. These compounds are, clearly, insoluble.
Thus, areaction with 0, or indeed with OH-, can effectively remove
metals from solution. Complexation of metals with oxyanions
(NO x - , SO/ -, PO/ -, CI04- , etc.) can also affect solubility.
Biological Treatment of Metallic Pollutants
Most metal ions are "hard" Lewis acids; they have low electronegativity and high charge density. However, hardness or softness is strongly affected by oxidation state: Fe(III) and Co(III) have
a high charge density and are hard; Fe(II) and Co(lI) have a low
charge density and are "borderline." Hard bases are electronegative
species: F- and most oxides, including OH-, N0 3- , PO/-, CO/ -, and
SO/ -. Soft bases are less electronegative, and include C, S, and P;
ct and Br- are borderline. Interactions between hard acids and
hard ligands are largely ionic in nature; those between soft acids
and soft ligands are largely covalent. Metals are easily coordinated
by N, 0, and S; P is a weak metal binder.
Proteins tend to complex metals more strongly than do nucleic
acids. However, neither macromolecule is typically involved in
metal-binding reactions in nature. In the subsurface, microorganisms employ biogenic organic acids (C/H/O) and siderophores
(C/H/O/N) to scavenge metals (Madigan et al. 1997). Low molecular weight, P-based chelators are in common industrial use and may
be co-disposed with metallic and particularly radioactive wastes.
B.D. Faison
partially-redueed C, N, and S - their moleeular oxides, rather than free
radicals - are reaetive. Although C and N ean undergo two-eleetron
redox to yield unstable intermediates, S ean also undergo one-eleetron
redox. When produeed as metabolie intermediates - or as the result of
eompeting or eomplementary abiotie reaetions - C/N/S oxides may
oxidize or reduee other elements in turn. Indeed, biotransformation in
vivo proeeeds via unstable, partially-redueed intermediates.
Metal Respiration
Non-nutritional, energetie metabolism of metals has been deseribed as
dissimilatory (rather than assimilatory) reduetion. The list of respirable
substrates includes various metals and non-metals as weIl as their
oxides (Table 1). Metal respiration may be a typieal microbial aetivity;
Table 1. Biologieally relevant respiratory transformations"
Oxygen tension or
metabolie mode
Element or
Mole of eleetrons
transferred per mole
of element redueed
Least energetie
2H+/H z
Anoxie; striet
2CO z/aeetate
S04 Z-/HSS40 /-/ 2SzO/-
Faeultative aerobes
AsO/-/AsO/UO/+/UO z
N0 3-/NOzNOz-/NH/
Se0 4 z- SeO/N0 3-IlI2Nz
Fe3+/Fe z+
MnOz/Mn z+
1/20 z/HzO
CI0 4-/CI-
Oxie; striet aerobes
Most energetie
"These values were measured under physiologieal eonditions of pH and temperature,
or ean be ealculated from measurements made at standard temperature and pressure.
The ranking of reduetion potentials (energy yields), however, will be unehanged.
Biological Treatment of Metallic Pollutants
this fact is uncertain as the range of potential respiratory substrates is
only now being studied. Even redox of Cl and its oxides has been
described; however, that topic is beyond the scope of this chapter.
Microorganisms capable of Fe redox appear to be ubiquitous (Sect. 1.3)
and Mn respiration is also common. The amount of energy released by
metal respiration can be either greater or less than that obtained from
CIN IS oxides. Some portion of the metabolized element may be diverted
to nutrition, but the amount of the element reduced is usually
significantly higher than that needed to produce biomass alone. Metals
that have multiple, reasonably stable oxidation states - namely, the dand f-block elements - facilitate electron transfer and energy transfer
in microbial systems. Moreover, biologicalligands can stabilize metals
in unusual oxidation states uncommon in nature. Metals shown
amenable to biotransformation then include, in order of increasing
atomic number, Si, Cr, Mn, Fe, Co, Cu, As, Mo, Tc, Hg, U and Pu; and in
some cases, biologically useful energy is also derived.
Metal Detoxification
Metal biotransformation is sometimes a survival function. Indeed,
derivatization, in the form of alkylation, of Si, S, and Se (which are
not classified as metals) or of Hg may support biomass formation or
detoxification. In the case of prokaryotes, biotransformation may aid
detoxification by removing the metal from the aqueous phase in which
most key metabolic processes occur, via volatilization or precipitation.
A third known approach to toxicant resistance, i.e., modifying the cellular target, will be considered briefly in Section 2.2.3. Metals can also
undergo abiotic redox reactions; the relative rates of the biological and
nonbiological reactions determine which will prevail under given conditions (Palmer and Fish 1997). Of the metals listed above, biological
relations with Fe are best understood.
Iron Respiration as a Model for Dissimilatory Metabolism
The microbial reduction of Fe is important to nutrition, energetics and
possibly survival, and this provides a conceptual basis for metal biotransformation in general (Text Box 2). The phenomenon mayaiso be
linked to both the origin of life on Earth and the shaping (and remediation?) of the terrestrial environment.
B.D. Faison
Iron respiration and contaminant removal
Iron in surface soils almost always occurs in the oxidized form,
because it reacts spontaneously with oxygen. Subsurface Fe may
also exist as Fe(III). Microorganisms that live in anoxic subsurface
media may carry out what is called a dissimilatory Fe reduction,
and are known as members of the dissimilatory metal-reducing
bacteria (DMRB; formerly, DIRB). Metal respiration sometimes
results in precipitation, depending on the element. Reduction
changes Fe(III) from a red, water-insoluble solid (hydr)oxide - rust
- to a greenish, water-soluble salto Incomplete reduction leads to
the appearance of a mixed Fe(II)/Fe(III) salt: "green rust". Green
rusts are layered, double hydroxysalt compounds (general formula:
[Fe 2+(l_x) Fe 3+.., (OH)2]x+ [xln A"- · m H20Y-, where x is the ratio Fe 3+1
Fetotad. These compounds are very reactive, and may contribute to
the abiotic oxidation of metal or radionuclide contaminants. Alternatively, biogenie green rusts may participate in such reactions,
yielding a coupled abiotic/biotic activity.
Some iron-respiring bacteria are able to reduce other metallic
elements, too. There is evidence that DM RB reduce U, Pu, Tc, and
possibly Cr. These metals, once reduced, are unable to dissolve in
and migrate with groundwater. Biotransformation (or chemical
reduction in general) of these specific contaminants decreases both
their toxicity and their mobility.
DM RB often use special binding compounds to simplify capture
and uptake of oxidized Fe; the compounds' ability to bind other
metals may be gratuitous. These compounds are siderophores, or
specific molecules that recognize metals' size, shape, and/or charge.
Siderophores chelate oxidized metal and increase its watersolubility; the metal-siderophore complex can then be absorbed
and/or reduced by the DMRB. Components of the cell surface,
including external slime layers, mayaIso be involved in the initial
recognition or subsequent capture and uptake. However, the details
of the binding, capture, absorption, and reduction of non-Fe metals
by DMRB are not weIl understood. Nor is it known whether the
reduction is a true respiration, i.e., whether the cells derive energy
for growth, survival, and/or reproduction.
Biological Treatment of Metallic Pollutants
Ubiquity of Fe on Earth
Iron comprises approximately 5% of the Earth's crust, by weight, and
is the fourth most prevalent element on Earth (surpassed only by 0, S,
and Al). Like most metals, Fe primarily exists as hydrous oxides, or
"(hydr)oxides" - a term including oxides (MOx), hydroxides [M(OH)x],
and oxyhydroxides [MOx(OH)y]. Several crystalline Fe(hydr)oxides
occur in soils: goethite (FeOOH), the most common Fe mineral;
hematite (Fe20 3), the most common Fe ore; magnetite (Fe304);
maghemite (Fe203); lepidocrocite (FeOOH); and ferrihydrite
(Fes(OH)g·4H20). All except magnetite contain reduced iron, or Fe(II),
only; magnetite is a mixture of Fe(III) and Fe(II). Moreover, magnetite,
unlike the other minerals, is a product of metamorphism as opposed to
pedogenesis. Fe(III) production may derive from abiotic reactions
between Fe(II) and O2in surface soils; but abiotic reactions may be supplemented by biotic activities. The role of biological processes in the
weathering of various substances has not been thoroughly examined
(see Chap. 8 on natural attenuation). Still, Fe biotransformation and
mineral formation may be linked within natural systems.
Significance of Fe (Bio)Energetics in Metal Remediation
Transformation of Fe(III) to Fe(II) is a feature of both assimilatory and
dissimilatory metabolism. Iron transformation may bridge abiotic and
biotic activity, in that Fe reduction involves the transfer of a lone electron in either system. The relatively high potential of the Fe(III)/Fe(II)
couple suggests this metal's ability to facilitate the redox of other
elements as a potential intermediary in coupled biological and nonbiological activities.
Microorganisms as Remediation lools
for Suboxic Environments
Problem Definition
The remediation of subsurface environments is complicated by the presence of buildings and geographical features; the cost of excavation
("muck, suck, and truck", or "pump and dump") or of site closure and
B.D. Faison
abandonment ("cap and run") mayaIso be formidable. Public concerns,
as in "not in my backyard;' cannot be ignored. Traditional wastedestruction technologies that are based on access to light or to molecular O2 are either not feasible or prohibitively expensive. Bioremediation
is therefore considered a relatively low-cost, socially acceptable
approach to cleanup. However, it is not clear that a one-time addition
of key microorganisms (bioaugmentation) or metabolie substrates
(biostimulation) would suffice over the course of decontamination. Nor
is it clear, based on the heterogeneity of the subsurface, that additions
made at one or several locations would yield a constant value of the
desired substance, over the entire site, over the entire treatment period.
Site restoration now focuses less on removal (leading to treatment,
storage, and/or disposal off site) and more on containment and stabilization. Therefore, some long-term stewardship, accompanied by monitoring, would be required. [Note: one frequendy stated goal is
contaminant immobilization in place, for long periods, that in the case
of radionuclides are defined as 10,000 years. Natural radioactive decay
was formerly considered an acceptable physical solution to contamination on ce ten half-lives had past. However, in the case of 238U, a main
component of depleted U preparations, that period covers 4.5 x 10 10
years. The longevity of all radionuclides in the original nuclide's decay
chain must also be considered. Site stewards hip is therefore truly a
lasting commitment.]
Relevant Biological Factors
Concerns surrounding metal contamination in the subsurface focus on
contaminant bioavailability to plants, animals, and fungi; bioavailable
metals are mainly those dissolved within, or adjacent to, groundwater
stores (Text Box 3). Bioremediation of metal-contaminated soils and
sediments requires the appreciation of hydrological (Text Box 4) and
geophysical (Text Box 5) phenomena as well as the chemical (Text Box
6). Inherent solubility of gases, produced biotically or abiotically, is also
relevant to the chemistry of metals in the environment (Table 2); gases
generated in situ by biotic or abiotic means can react with (trans form,
complex, or precipitate) dissolved contaminants or, if present as
bubbles, re du ce hydraulic conductivity. Meanwhile, matters of scale levels from atomic- to ecosystem-, or possibly planetary- - must be
considered (Text Box 7). Development of site-specific remediation
processes requires understanding of the site's geography, climate and
meteorology, surface fiows and subsurface hydrodynamics, and
geochemistry. Development of bioremediation processes requires
Biological Treatment of Metallic Pollutants
Contaminant plumes: migration of hazardous waste in the subsurface
Contaminant plumes are zones of pollution extending downstream
from sources of contamination (Riley et al. 1992). Contaminant
types vary in their rate of movement and distribution. If more than
one contaminant type has been released into the subsurface, multiple plumes can form with different spatial and temporal distributions, and with different relative coneentrations of eontaminants.
A eontamination source may be a single point such as aleaking
tank. Or, the plume may have spread out from the eontamination
of a large area. Point sources are frequently spills, treatment
lagoons, and disposal sites such as trenches, landfills, and
underground storage tanks. Once a eontaminant is released into the
environment, the plume can spread into soils, unconsolidated
sediments, rock formations, groundwater, and surface water.
Depending on the geology and hydrologie conditions at the si te
and the solubility of the contaminant, the plume may stay elose to
the source or be transported long distances by groundwater or rainwater infiltration events. In some cases, all of the contamination is
caused by a single spill or leak. In others, the souree of contamination may continue for decades - such as at an active was te disposal
site - or when surface water percolates down through the zone of
eontamination. In the groundwater, the shape of a plume will
depend on the migration rate, which is largely eontrolled by
groundwater ftow, the hydrogeologie setting, the physical and
chemical characteristics of the eontaminant, interactions between
the contaminant and other dissolved substanees, and the presence
of a continuing contamination source. If the inftow from the souree
has been stopped, the entire plume may migrate away from the original loeation, eventually becoming less eoneentrated through the
transport processes of advection, diffusion, and dispersion. Chemical and biologieal reactions mayaiso occur, further decreasing the
contaminant's concentration. In the case of radionuelides, the
physicochemieal behavior - transport and reaetivity - of each
radioactive or nonradioactive daughter produced by radiological
decay must also be eonsidered (Wilson et al. 2003).
Biological and ehemical reactions can affect the size and shape
of the plume by slowing or accelerating migration of the eontaminant. If the contaminants adsorb onto the geological materials, the
rate of plume movement will be retarded relative to water. Sometimes, however, eontaminants adsorb onto very small partieles,
B.D. Faison
called colloids that may themselves move with groundwater,
thereby transporting the contaminant. Both colloids and microbes
accumulate at air-water interfaces. In some cases, higher densities
of microbes and higher concentrations of contaminants are
observed at air-water interfaces, especially capillary fringe zones in
the vadose zone immediately above the water table. Water table
ßuxes can thereby cause unexpected concentration phenomena.
Chemical and biological interactions also can result in precipitation of the contaminant into a solid phase that is no longer mobile
and may cause plugging.
Glossary of hydrological terms
Groundwater flow and influences on contaminant transport.
Groundwater moves within aquifers to discharge points - springs,
streams, lakes, wetlands, or even the ocean. Flow rates can range
from <10- 2 to >10 1 m/day in specific areas; as a result groundwater
can remain in pi ace for hundreds of years. Groundwater is ultimately contiguous with surface water.
a) Direction: Route taken by water through soil, unconsolidated
material (sand and gravel), and bedrock. This vector is generally determined by hydraulic gradient (slope), but may be
inßuenced by permeability and porosity.
b) Velocity: Rate at which water moves through the subsurface.
This value is determined by the hydraulic gradient, permeability, and porosity of the material through which it moves.
Hydraulic gradient is related to the elevation difference, or pressure head.
c) Recharge rate: The rate at which groundwater is replaced by
rainwater or snowmelt infiltration, or by rivers, lakes, or the
d) Advection: The transport of dissolved solutes with the bulk
flow of water in the vadose zone or in groundwater. For highly
soluble contaminants that do not undergo chemical or biologi-
Biological Treatment of Metallic Pollutants
eal reaetions, it is the primary meehanism influencing the fate
and migration of the eontaminant.
e) Dispersion: The meehanieal mixing of solutes that oeeurs as the
solutes are adveeted through the groundwater system.
f) Presence 01 colloids: Particulate matter between 1 nm and 1 pm
in size.
Hydrogeologie setting. This concept seeks to integrate the behavior of water with loeal geology and, possibly, biology. Groundwater
oceurs below the root zone of plants, and lies beneath the vadose
zone. The depth to a given aquifer (water table) depends on loeal
geology, but may be as much as hundreds of meters.
a) Permeability: The capacity of a porous medium to transmit a
fluid. This attribute is also known as eonductivity; "hydraulic
conduetivity" refers speeifically to groundwater flow, and is a
measure of permeability.
b) Porosity: The volume of aquifer material that is not occupied by
solids. This concept of"void space" must indude the ratio of soil
minerals to soil organic matter, the crystalline strueture of solid
rock, and the potential presence of natural fissures deseribed
above. Permeability is eontrolled by porosity, pore eonnectivity,
and pressure; therefore, hydraulic conduetivity deereases with
depth, and dedines with microbial growth in pores.
e) Diffusion: Bulk movement of solutes resulting from thermally
driven moleeular motion of solutes. Through this random
molecular motion, contaminants move from areas of high eoneentration to areas of lower concentration. Diffusion is thought
to be partieularly important when a geologic formation has a
very low permeability or is very heterogeneous, such as a layered
sequence of sand and day.
Contaminant physieochemical characteristics. Physical attributes
of a given contaminant - its volatility, water solubility, density, etc.
- are eonstant whether measured on a laboratory bench top or in
the subsurface (Wilson et al. 2003). Its ehemical reactivity is similarly unchanged. However, eontaminants in groundwater must be
considered as existing in a heterogeneous solution, within a poorly
confined liquid, in avessei with reaetive walls.
B.D. Faison
a) Sorptiveness: The tendency of solutes to adhere to particulate
matter, including the walls of avesseI. This phenomenon
includes absorption and adsorption, and is used when the
precise mechanism of removal from solution is not known.
Absorption involves the penetration of one material into the
inner structure of another, such as occurs during the uptake of
water vapor by glycerol or other hygroscopic materials. Adsorption refers to adherence to surfaces. Finely divided or microporous materials presenting a large surface area (e.g., activated
carbon) are strong adsorbents. When at least two different
solutes are present, preferential adsorption may occur.
Chemisorption, or adherence to solids with high surface energies (e.g., Ni or Fe), involves bond formation and may result in
dissociation and catalysis.
b) Solubility: The tendency of one substance to blend uniforrnly
with another. Solubility is a complex phenomenon controlled by
electrolytic dissociation, diffusion, and thermodynamics. Materials that blend poody may form colloids. Solubility is a relative
term expressing degree to wh ich one material blends with
another in a closed system at equilibrium, which is not the case
in subsurface environments.
Contarninant interactions with other solutes. Competing chemical
reactions between solutes present within a mixture occur, and
render each solute's behavior in solution unpredictable - one
cannot extrapolate this behavior from that occurring in simple
aqueous solution. Carbonate, present either as either a component
of geological weathering or by the activity of microorganisms in
the subsurface, contributes to competing interactions (Madigan et
al. 1997). For complete explanations of interactions via ion
exchange, coordination, and chelation, see Text Box 1. Interactions
not described there, but important in subsurface chemistry, include
sequestration - the production of coordination complexes between
phosphates and metallic ions (usually, d- or f-block elements), rendering the metal soluble; and sorption - the common but poorlyunderstood physical adhesion of solutes to colloids.
Biological Treatment of Metallic Pollutants
Soils and sediments - a biochemical view
Soils and sediments are heterogeneous assemblages of solids,
liquids, and gases. Soils and sediments have identical origins in
fragmented materials produced by the weathering of rock; soils
differ from sediments in that the latter forms in layers beneath
liquids. Soil or sediment solids are a mixture of inorganic and
organic material, ranging from size from days (less than 0.002 mm
in partide diameter) through silt (0.002-0.05 mm) to sand
(0.0625-2 mm) and gravel (greater than 2 mm) or rock. The inorganic complement indudes quartz (SiO z) and feldspars (KAlSi30 s
or NaAlSi30 S), and day minerals such as kaolinite [Si6AI40 IO (OH)s]
and montmorillonite [Mx(AI,Fez+,Mg)4Sis0 2o(OH)4, where M is a
metal cation]. Soil or sediment minerals may have specific surface
areas, induding both extern al and internal surfaces, of 10 to
1000m2/g. The organic component is the total of organic compounds in soils or sediments exduding the undecayed tissues of
plants, animals, fungi, protists, and microbes and their partial
decomposition products. Soil- or sediment-associated natural
organic matter (NOM) and humus are synonymous. NOM content
ranges from less than 0.5 to almost 100% in agricultural soils and
sediments. The structure of NOM varies, but certainly is enriched
in aromatic hydroxyl groups (-OH) and carboxylic acids (-COOH).
These groups are acidic. Other functionalities (e.g., aliphatic -OH)
are neutral; other, nitrogen-containing groups (amines, amides) are
basic. The surface area of NOM and acidic groups influence the
behavior of metals in subsurface environments. At a lower pH (-pH
2), NOM is less important to cation exchange capacity than are
inorganic day minerals; at a higher pH (- pH 8), the contribution
of the two materials is approximately the same. Soils and sediments, unless extremely dayey and compacted, contain voids occupied by liquids, gases, and vapors. This chapter considers only the
activities that occur within or adjacent to these aqueous phases.
The pH of a soil or sediment is determined by the inorganic and
organic components; acidic groups contribute to a buffering capacity. The Eh' or redox potential, of a soil or sediment is determined
primarily by the matrix's degree of saturation with water. These
external influences are characteristic of the specific geology and
hydrology of the subsurface environment under study. The oxidation state of any element varies as a function of pH and Eh, and can
B.D. Faison
be described by reference to the element's Pourbaix diagram. An
element's speciation (molecular state) va ries as a function of pH,
and can be determined only by site-specific caIculation.
Chemical reactions within soils and sediments include ion
association, exchange or complexation, multivalent ion hydrolysis,
oxidation and reduction, crystallization, sorption, and solution/
dissolution. These activities da not proceed in a homogeneous
manner; for example, deposition of salutes may be even across a
soil surface, or may instead proceed at nucleation sites formed by
recently-sorbed salute. Reactions involve the transport of solutes
through bulk liquid, across a liquid film at the solid-liquid interface, and within a liquid-filled macroporej these activities precede
diffusion of sorbates across and throughout the solids.
Soil water activity (a w , which is determined in part by the presence and concentrations of solutes, and wh ich is different from
water content) determines the type and activity of resident
microflora. Soil and sediment organisms typically live at the soilfluid interface, often within very thin water films. The reactions
listed above may be biotic or abiotic. Or these reactions may reflect
a synergism between biological and nonbiological transformations:
excretion products of organisms - including CO 2 and acids from
normal metabolism, as weIl as other biogenic gases such as H2S or
H2 - released into subsurface water are available for reaction with
contaminants and natural substances without further microbial
Biological Treatment of Metallic Pollutants
Chemistry in the subsurface - general considerations
Chemistry became a true science in the 19th century when the periodic table was described, the first compendia of the properties and
reactions of compounds were published, and the first kinetic and
thermodynamic laws of chemical reactivity were stated. Subsequent investigators who wished to build on these findings, and
compare their results to those obtained by others, established standard physical conditions as a basis for calculations involving quantities - such as solubility - that vary with temperature and pressure.
S.T.P is defined, by convention, as a standard temperature of 273.15
degrees Kelvin (equivalent to 0° Celsius) for gases, or 298.15 K
(25°C) for liquidsi and a standard pressure of 101.325 pascals
(equivalent to 760.00 mmHg, or 1 atm). Data reported as "standard"
often also specify that measurements were taken at pH 7.0. It is
assumed that all work was done in the open air (approximately 21
wt% O2 and 0.04 wt% CO 2). Reactions are furthermore usually
studied in mixtures of known, homogeneous composition, and
including a limited number of possible reactants; equilibrium
concentrations of reaction products are generally reported.
Oxidation-reduction potentials listed in general electrochemical
tables, and relationships described in Pourbaix diagrams, were
obtained under S.T.P., unless otherwise indicated. Speciation diagrams, however, always refer to particular reagents and specify both
pH and Eh.
Deviations from ideal or standard conditions make bioremedial
investigations particularly challenging. Published data are useful in
showing a relative energy yield of reactions thought to occur in situ;
but it is important to remember that subsurface environments are
heterogeneous and rarely at equilibrium. Hydroxide ions are always
present in water, whatever its pH, and water molecules themselves
may form complexes with dissolved metal ions - as may CO/ -.
There is also the possibility of competing or interfering reactions
that involve co-contaminants, reaction intermediates, and/or components of the subsurface medium.
Data collected in experiments conducted at S.T.P., in well-defined
mixtures have practical qualitative value in that they suggest what
reactions are possible, and in some cases, probable. These data also
have quantitative value in establishing the relative probability of a
given reaction under specific conditions. Concentrations of a given
B.D. Faison
substance at various temperatures, pressures, or pH can be extrapolated by calculation from published data obtained at S.T.P. These
extrapolations may require the use of advanced computers and the
development of mathematical models.
Biotransformations in the subsurface - a matter of scale
Natural media contain atoms, ions, ion complexes, molecules, colloidal particles, cells, cell assemblages, soil or sediment particles,
microbial communities, soil aggregates or sediment blocks, and
entire microbial ecosystems. Sizes involved range from picometers
(l0- 12 m) for individual atoms, through nanometers (10- 9 m) for
minerals and micrometer (pm; 10-6 m) for microorganisms, to
thousands of square kilometers [(10+3 m )2] for sites. Soil or sediment minerals themselves may have specific surface areas, including both external and internal surfaces, of 10-1000 m 21g. A study of
hydrobiogeochemical phenomena at these greatly varying scales
would "bridge the gap"between single-atom and ecosystem studies.
Time, as weIl as physical dimensions of length, are important
parameters that affect bioremediation. Biotic reactions must
compete with simultaneous abiotic reactions that may increase or
decrease contaminant solubility. Since all reactions of interest here
occur within water, diffusion is critical. Many reactions in soil are
diffusion-limited, i.e., are controlled by the speed with which substrates move through pore water toward other dissolved reactants,
toward water-mineral interfaces, or toward microbial surfaces. The
subsurface transport of microorganisms and nonliving particles
less than 1 pm in diameter is similarly controlled by diffusion. The
kinetics of relevant metal-transforming or -transport phenomena
are therefore slow, and cannot be measured in optimized (wellmixed) systems.
The pH scale is also critical to subsurface scientific studies.
Studies of contaminant chemistry refer to values derived und er
standard conditions of temperature, pressure, and pH, in an air
Biological Treatment of Metallic Pollutants
atmosphere. However, subsurface environments are anoxie; and all
microorganisms produce CO 2 as a by-product of normal metabolism (Madigan et al. 1997). The pH of water in equilibrium with CO 2
at S.T.P. is 5.65, and the water solubility of CO 2 is over 25 times that
of O2• The partial pressure of CO 2 in soils is up to 400 tim es that of
air, yielding a typical pore-H 20 pH of 4.65- 5.15. This value would
be decreased further by microbial production of organic acids. The
background acidity of pore water affects subsurface reaction chemistry, requiring studies of microbial activity within soil microcosms
to confirm behavior in situ. The effects of pH extremes, as found
in some metallic wastes (ranging from 1- 12 in some DOE
effiuents), must also be considered.
An equally important scale describes Eh. Studies of abiotic
chemistry and bio transformation in the absence of oxygen must be
carried out either within sealed test tubes, or within anaerobic
chambers equipped either with stop-flow devices for nearinstantaneous measurement of ion speciation or with side ports for
continuous sampling. Laboratory experiments conducted in microcosms or mesocosms, under conditions analogous to subsurface
environments and with elose attention to kinetics, pH, and Eh, may
be extrapolated to the field. However, the inherent heterogeneity of
the subsurface may make such extrapolation, or computer simulations, difficult.
B.D. Faison
Table 2. Solubility of selected gases in water at 15-35°C
Solubility (mole fraction)
CO 2"
CH 4
"Includes all products from reaction with H20.
accommodation of these physicochemical realities plus microbiological
ones: nutrients, respiratory substrates, transport and trophic interactions (Faison et al. 2003). [Genomics may be useful only as a elue to
nutritional and energetic metabolism; see Sect. 3.4.] Estimates of microbial biomass in soil and sediment are high enough to suggest that all
sites support microbial populations and that information, coupled with
a current understanding of extremophilic lifestyles, would indicate that
indigenous, acelimated microbiota inelude organisms capable of useful
metal biotransformations. In short, bio augmentation - deliberate introduction of non-native microorganisms - may not be necessary. Environmental regulations may particularly preelude addition of genetically
modified organisms. The problem, then, is how to encourage native
microbiota to remove metals from groundwater. A possible solution lies
in linking the resident populations' energetic metabolism to their survival, with full understanding that element cyeling in nature requires
the participation of several ecotypes (Anonymous 1997).
Selection of Bioremediation Tool
A focus on one particular microorganism or genus is not necessarily
prudent. Axenic cultures rarely persist in nature. Indeed, at the ecosystem level, all natural systems are both redundant and degenerate.
Redundancy derives from the fact that cell types with identical biochemical function do not have identical genomes: case in point, the Ferespiring genera Geobacter and Shewanella. Degeneracy reflects the
finding that the same biochemical function can be carried out by dif-
Biological Treatment of Metallic Pollutants
ferent substances produced (excreted) by the same organism: metals can
be complexed in the subsurface by either siderophores (energetic facilitators) or by organic acids (nutritional waste products). Beyond the
behaviors of specific organisms, socio-microbiological issues must be
(ommunity Dynamics: Ecology
The activity of a given ecosystem depends on its complexity. Organisms
that precipitate dissolved contaminants may occupy the same niche as
do organisms that catalyze contaminant resolubilization. Symbiotic
issues of mutualism, commensalism, neutralism, amensalism, and
antagonism must all be recognized to avoid competitive interactions
that affect net contaminant immobilization rates. The worst forms of
competition - parasitism by phages and predation by eukaryotes - must
be considered. These latter activities will be controlled in part by
soil/sediment hydraulic conductivity, which affects microbial transport.
[The porous structure of these natural media minimizes the influence
of convective flow and hence bulk transport of different microbial populations. Although motile organisms can move independently through
fluids and across solid-water interfaces, this motility is usually
insignificant compared to groundwater flow on the mesoscale level. Diffusion is the parameter critical to subsurface movement of microbial
populations. ]
The possible impacts on the ecological succession of microbial invasion, site perturbation (by mining, sampling, or feeding) and periodic
selection (something caused by humans as part of site restoration activities) cannot be ignored. It is only by judicious site management that
ecosystem changes may be controlled. However, one type of biological
competition - known as evolution - is both unavoidable and
(ommunity Dynamics: Evolution
Current biological understanding describes evolution as the logical
result of replication plus random mutation (variation) plus competition
for limited resources (Text Box 8). Degeneracy leads to an evolutionary
selection of individuals with sufficient fitness, regardless of variation or
mutation. The lifespan of prokaryotes in situ has not been studied; nor
has the effect of long-term exposure to specific metals, including
radionuclides. However, there is some possibility that organisms,
B.D. Faison
Evolutionary microbiology and growth in the subsurface
Microorganisms inhabit every terrestrial habitat imaginable,
including those containing radionuclides, which may be considered
toxic. Microbial populations exposed to any toxic material suffer a
form of natural selection: only those individual members resistant
to the toxicant survive. Microorganisms generally develop resistance to toxic substances in three ways: by decreasing the effective
concentration via efftux; by altering the cellular site at which the
toxicity is expressed; or by attacking and modifying the toxicant.
Resistant subpopulations thus include those tolerant to relatively
high concentrations of the toxicant and those able to transform the
toxicant. Cultures capable of deriving energy or nutrition from this
detoxification activity have an advantage over those that cannot,
especially if such substrates are in limited supply. Natural selection
may thus be a common influence on cultures at contaminated sites.
Evolution can be defined as a genetic change, within a population, over time. Evolution within microbial populations in situ presumably occurs via the same mechanisms known to function in
vitro: horizontal gene flow via transformation, conjugation, and
transduction; successful competition within a population made
heterogeneous by random mutation; and a generalized genetic
drift. The profiles of subsurface microbial communities are known
to change in response to external influence, but the relative contribution of these three evolutionary mechanisms is not known.
Physiological adaptation is characteristic of individuals within a
population, occurs over the lifetime of the individual, and does not
represent any changes in an individuaI's genome. This activity
involves multiple layers of genetically encoded regulatory mechanisms, tuned to respond to changes in the state of internal or external variables. The change in structure or function can be ascribed
to alte red transcription, or post-transcriptional modification, of
gene products. In contrast, evolutionary adaptations occur in populations over generations, and represent a change in allele frequency over time. Both types of adaptation are controlled by
generation time, population size, recombination rates, strength of
selection, and various structural and/or physiological constraints.
However, neither individual cells' growth, behavior, or functioning
within complex ecosystems nor the activity of microbial populations has been studied in situ.
Biological Treatment of Metallic Pollutants
There is insufficient current knowledge to allow bioremediation
scientists to distinguish between physiological and evolutionary
adaptation. However, microorganisms are known to produce isoenzymes, or distinct forms of enzymes with a different structure but
similar function (Faison et al. 2003). This behavior suggests a physiological adaptation that allows microbes to utilize contaminants
with physical characteristics similar to those of natural substrates:
case in point, the binding of Pu(IV) by Fe(III) -targeted siderophores. Or the behavior may be an evolutionary adaptation
leading to the synthesis of completely new, different enzymes. The
probable mutagenizing effect of long-term exposure of radioactive
contaminants in situ does lend weight to the latter interpretation.
However, until more data are available, one can only speculate.
acclimated over years (a very lang time from a microorganisms' point
of view) to life in the presence of metal contaminants, would become
acclimated to, and preferentially utilize these materials as respiratory
substrates. Long-term biotransformation rates could not then be predicted, for the identity and nature of such acclimated cuhures is not
known. Lateral gene transfer, facilitated by cryptic growth, might
further complicate the issue.
Relationship to Abiotic Factors
There is doubtless synergism between the physicochemical. and biological factors explored above. Biotransformation of metals clearly can
occur, and may be wholly intracellular. However, biological activity can
also facilitate the following extracellular phenomena involving metals:
complexation (which can inhibit transformation); precipitation (by providing nucleation centers); mobilization (as colloids on the surface of
cells, within biopolymers, or as complexed species); corrosion (via reaction with biogenic H2S); and sorption to or desorption from natural
media. Microbial metabolites can solubilize Fe oxides, which can solubilize sorbed contaminants in turn. Biogenic H2ü 2, a reduced form of 0,
can react extracellularly with Fe(III) to form OH (hydroxyl radical), an
extremely unstable and reactive oxidizing species. However, greater
B.D. Faison
issues lurk beyond the nano- and microscale level problems described
Role of Nonliving Organic Matter
Dead microorganisms are a primary component of natural organic
material (NOM), which itself is the largest pool of C on the planet. NOM
includes cell metabolites and is characterized by complexing structures.
Natural media may contain high or low levels of NOM, depending on
their geographic location; NOM content ranges from less than 0.5% in
deserts to almost 100% in agricultural soils and sediments. Water that
infiltrates natural media may originate at some distance and bring NOM
from its source. The actual dissolved NOM content, then, of any aquifer
must be measured rather than estimated from local soil or sediment
characteristics. Dissolved NOM competes with microbiota for metals
and its presence can either facilitate or inhibit biotransformation rates
in the subsurface. NOM may be an essential cofactor (electron shuttle)
for U biotransformation; but NOM, when present at concentrations
above 100mg/l, inhibits U reduction completely. NOM is thus responsible for much of the abiotic activity with which the desired biotransformation activity competes. Lack of data concerning the lifespan of
microorganisms in situ prevents accurate prediction of the behaviors of
this dynamic system.
Role of (o-contaminants
The impact, at the molecular level, of co-contaminants on metal biotransformation is unknown. Issues surrounding multiple metals have
been largely covered in preceding paragraphs and the potential for
spontaneous electrochemical redox reactions is clear. However. metallic wastes frequently contain oxidizing acids, introduced according to
typical industrial practice. Uranium in particular is generally used as a
N0 3- salto Nitrate is a known microbial respiratory substrate, and its
reduction yields more energy than that of most metals. This ion will
thus be preferentially used by competent respiring an aerobic populations. Nitrogen is capable of one-electron redox, as is S. Sulfur oxides,
produced naturally during biotransformation or present as part of
the waste, also compete with metal respiration. Both N and S
(hydr)oxides can facilitate competing abiotic reactions that lead to
metal mobilization.
Biological Treatment of Metallic Pollutants
Practical Aspects of Bioremediation
Management and Operations
Bioremediation process development, as stated above, requires an intimate understanding of site characteristics - most notably, contaminant,
pH, Eh, aw , and past and planned usage - plus budget and time constraints. Field testing of metal bioremediation has been initiated (Text
Box 9), as have plans for site monitoring (Text Box 10). This work has
already identified numerous research areas for future scientists (Faison
et al. 1996).
Hydrological Data Needs
There is a lack of information regarding the bioavailability, to microorganisms, of contaminants sorbed to colloids. Moreover, the fate and
potential transport of contaminants sorbed to the microorganisms
themselves have not been studied, so the potential role of microorganisms in the physical (as opposed to chemical) remobilization of contaminants is unknown. Microbial access to fractures or to geologie
interfaces such as day layers, sediment interbeds, and mineral veins has
not been studied. This information is important to understand the
potential contribution of microbial activities to site restoration on the
meso- to macroseale levels. Perhaps more worrisome, however, is the
potential for microbial alteration of flow paths - via biomass proliferation, or by precipitation of metal (hydr)oxides, salts, or complexes. Biostimulation (amendment with nutrient or respiratory substrates) may
have a negative impact on site hydrology; this possibility is explored in
Text Box 11. However, colloid formation itself is not well understood;
nor particularly is the potential contribution of microbial activity to
that process. Supporting studies would be required to determine the
impact of increased particulates, formed by contaminant precipitation
by abiotic or biotic means, on groundwater flow.
B.D. Faison
Field testing of metal bioremediation processes
A current DOE study explores biostimulation for cleanup of Ucontaminated subsurface environments, via single-well, "push-pull"
tests that determine kinetics of mierobially mediated U reduction
in situ. The push-pull test methodology involves the injection of a
prepared aqueous test solution into the saturated zone, followed by
the extraction of the test solution/groundwater mixture from the
same loeation. The injected test solution contains various combinations of tracers, electron acceptors (including U), and/or electron
donors. By monitoring the composition of the injeeted test solution
through time, the kinetics of electron transfer may be quantified.
The project tests the following hypotheses: (1) that indigenous
microorganisms able to reduce U are present in subsurface environments at the site and are likely SO/- and FeH-reducing bacteria; (2) that the U reduction rate is limited by the availability of a
suitable eleetron donor, whieh can be eontrolled to promote U
reductionj and (3) that adding an electron donor to stimulate
reduction of N0 3- , Fe3+, and SO/- will enhance U reduction rates.
Field experiments that focused on the supply of eleetron donors acetate, ethanol, or glucose - to the subsurface confirmed tests performed in laboratory microcosms. In all cases, the addition of an
electron donor stimulated N0 3- reduction, resulting in transient
N0 2- aceumulation. However, in these tests, which were carried out
in U and/or Tc-contaminated soils at a former U-disposal pond site,
there was no evidence of radionuclide immobilization. Future electron donor additions will be followed by an additional series of
push-pull tests to measure the effect of repeated electron donor
additions on rates of N0 3- reduction and U/Tc immobilization.
Aseparate DOE project seeks to improve the understanding of
basic processes within heterogeneous subsurface environments speeifieally (1) to predict the rates and mechanisms controlling
microbial reduction of U in the field, and (2) to develop a system
capable of delivering electron donors, enabling spatially uniform
immobilization of dissolved U upon the passage of groundwater
through a subsurface biocurtain of introduced organisms. Microbial community dynamics are being characterized using molecular
biological approaches, coupled to more traditional hydrological
and geochemical measurement methods, to determine the spatial
distribution of biological activities. Signature tagged mutagenesis
(STM) is being used to identify bacterial genes that are neeessary
Biological Treatment of Metallic Pollutants
for survival in a particular ecosystem - here, the identification of
genes within SO/--reducing bacteria (SRB) are needed for survival
in anaerobic subsurface systems (Amann et al. 1995). SRBs may be
useful in an indirect immobilization of metal or radionuclide contaminants in situ as insoluble sulfides. Meanwhile, RNA arbitrarily
primed-polymerase chain reaction (RAP-PCR) is being used to
probe genes necessary for microorganisms to survive contaminant
exposure. Preliminary work has demonstrated that RAP-PCR is an
effective tool in assessing differential gene expression in environmental bacteria. RAP-PCR, along with STM, can now be applied to
questions of environmental significance - the identification of
genes that are upregulated in the presence of contaminants, such as
metals and radionuclides.
Subsurface monitoring
Environmental cleanup requires knowledge of the amount and
behavior of water underlying a contaminated site, its physicochemical status, and the types and concentrations of contaminants
dissolved within. Remediation plans therefore include a monitoring component that allows quantitative assessment of the physical
and chemical condition of subsurface environments, providing
useful clues about both contaminant speciation and indigenous
Groundwater monitoring is carried out above ground by analysis of water pumped to the surface from extraction wells. These
weHs are distributed throughout the site of concern, as an onsite
weH field. The placement of the weHs helps to define the aquifer.
Offsite wells, in nonimpacted areas, are also needed to establish
"background" levels for comparison. Creation of each extraction or
sampling weB involves the drilling of aborehole. Understanding of
the hydrogeologie setting must be obtained prior to drilling, via
separate geophysical measurements. Onsite monitoring weHs allow
measurements of changing groundwater levels. Data from offsite
weHs enable site workers to establish environmental quality base-
B.D. Faison
lines. WeIl-field design is therefore a eritical portion of monitoring
plans, as is sampling strategy (measurement Iocation and frequency). Influences on contaminant transport can often be
deduced by combining groundwater-monitoring data - specifically,
pH and Eh - with knowledge of the hydrogeologie setting and
physicochemical information about the contaminant(s) involved.
Analyses of metals and radionuclides are carried out through
specialized forms of spectroseopy (e.g., atomie absorption or
induetively eoupled plasma emission). Mieroorganisms themselves
can be observed via sophisticated imaging techniques such as eonfocal, fluoreseenee, or electron mieroseopy. Imaging techniques ean
be linked with X-ray methodologies to derive additional information about metal-mierobe interactions (Kirz et al. 1995).
Controlled biostimulation: a conceptual model
Field tests of biostimulation may involve addition of organie eleetron donors such as acetate to the subsurface. Increased contaminant reduetion levels have been reported in initial field trials, but
the sustainability of this approach has not yet been determined.
Potentially useful bioremediation cultures ean be distinguished
by their growth in defined liquid media containing a metallic contaminant (or its surrogate) plus an organie substrate under anaerobiosis (Faison et al. 2003). Growth in vitro is defined as an increase
in biomass, possibly achieved by ceU enlargement and aeeumulation of insoluble polymers, but more likely realized by ceIl reproduction. Although the growth medium may seleet for organisms
that possess a desired aetivity, neither growth nor reproduetion is
truly diagnostie of eontaminant reduetion. Biostimulation of
metal-reducing cultures by addition of organic electron donors in
vitra has led to "growth", measured as amount of biomass.
However, microbial overgrowth (biomass proliferation) in situ
would lead to deereased subsurfaee porosity. Such plugging would
Biological Treatment of Metallic Pollutants
constrain further delivery of argamc electron donors to the
Current biostimulation paradigms link energy metabolism
(reduction of inorganic contaminants) to nutrition al metabolism
(oxidation of exogenously supplied organic compounds). In this
sense, C within the nutrient is viewed as an electron donor.
However, a portion of that C may be diverted into anabolic metabolism and thus contribute to the amount of biomass present.
Rampant microbial growth in the subsurface may be problematic.
Clayey media may have pore sizes of less than 100 J1m in diameter.
Biofilms composed of microorganisms and/or their insoluble
excretion products may clog these pores, decreasing permeability.
Continued injection of nutritive electron donors into subsurface
media via groundwater ftow might then be restricted by microbial
growth. However, a controlled biostimulation may be achievable by
greatly diluting a nutritively rich organic electron donor, such as
acetate, with either an inorganic electron donor or a nutritively
poor organic supplement that will not support growth. Hydrogen
gas [H 2; H(O)] and formic acid [HCOOH; C(II)] are standards to
which other secondary electron donors may be compared.
Hydrogen gas costs US$ 1.00-2.00 per pound, depending on its
purity: that cost range is on the order of US$ 0.0044 per mole of
electrons ($/electron -mole) donated. (The cost would be lower if
generated on site or in situ.) Formic acid costs $35 per liter, or
$0.60/electron-mole. These values compares favorably to typical
organic electron donors: lactate ($63.0/electron-mole), acetate
($1.59/electron-mole), ethanol ($1.47/electron-mole), or glycerol
($0.68/electron-mole). Either a supplemental electron donar could
be used in great excess to a primary electron donor, diluting the
more expensive substrate and thus decreasing overall cost.
However, the minimum amount of organic substrate needed to
support anormal turnover of ceH components must be determined
Controlled biostimulation would provide an immediate cost
savings compared to the use of nutritive electron donors alone, and
would potentially provide future cost avoidance by forestalling any
plugging of the subsurface with superfl.uous biomass during longterm bioremediation actions.
B.D. Faison
Physicochemical Data Needs
The central roles of C oxides and pH in cell metabolism and metal
(im)mobilization were presented above. However, the two areas have
not been correlated. Studies are needed to establish the potential role of
biogenic CO 2 as a chelator and pH buffer. Perhaps more importantly the
metabolic origin of biogenic CO 2 - from nutrition versus respiration might be studied. The use of heavy-ion isotope analysis may be recommended; this technique may be equally useful to distinguish between
biotic and abiotically generated CO 2 , especially in the context of metal
carbonate formation. [Abiotic Ca carbonate production in situ may
affect site hydrology, and could be included in these studies.]
Hydrogen-specific isotope analysis may be equally useful to determine
the possible effect of biogenic H2 on site Eh/pH, and to distinguish
biologicaIly-produced H2 from abiotic H2 (which can be generated from
H20 by radiolytic decay, and has been lately implicated in subsurface
bioenergetics). A recalculation of electrochemical, speciation, and solubility data under conditions actually encountered in the field (such as
the presence of radiolytic heat, excess CO 2, or pH extremes) is recommended for these studies.
Possible changes in contaminant speciation or structure affect their
biological and "geological" decay rates, as weIl as their solubility and
sorptiveness. The effect of synergisms and antagonisms among contaminants in causing these initial changes is not known. Reaction rates
may further vary with contaminant concentration, yielding further
spatial heterogeneity at the site (i.e., adjacent to high-concentration
source areas, such as leaking tanks, vs. within dilute plumes).
Investigations might also focus on the activity of organic and
metallradionuclide contaminants within commingled plumes.
Spatial and Temporal Data Needs
Little is currently known about the distribution and heterogeneity of
microorganisms or ofNOM in the field. Ecological studies are called for,
as are investigations of the long-term sustainability of the desired
microbial populations. Traditional molecular approaches to mapping
community diversity, i.e., via genomic (DNA) or ribosomal RNA data,
may not be adequate; studies in real time of the microbial mRNA transcriptome, rather than of the proteome, may be more valuable. However,
research on community dynamics in situ would be necessary for model
Biological Treatment of Metallic Pollutants
validation. Key studies would include a foeus on the reciprocal impaet
of eeologieal sueeession on the ereation and interspecies transfer of substrates for nutritional or energy metabolism. The rate at whieh microbial populations are replaeed eould be established, as eould the relative
importanee of seleetive pressure and hydrologie transport in the
proeess. Related studies on the competitive ability of deliberatelyintrodueed microorganisms might also be useful to bio augmentation risk
Evolutionary forees applieable to bioremediation include both physiologieal and genetie adaptation. Physiologieal adaptation may oeeur
within the individual (i.e., non-reproducing mierobial populations).
Genetic adaptation may be limited to dividing eultures. However, the
line between parent and progeny is blurred in mieroorganisms. The relative frequeney of eaeh type of adaptation within growing, surviving or
reprodueing eultures is not known. Anormal turnover of ceU enzymes
during maintenanee metabolism may aUow either or both types of
adaptation to oeeur, as may produetion of isoenzymes with differing
specificities. Information on the physiologieal status of subsurfaee
mierobiota, including their long-term survivability in situ, is needed.
Knowledge of generation times within the subsurfaee would also be
useful. These studies of eryptie growth would define the fate of genetie
information released upon eeU death, thus potentiaUy eontributing to
eeologieal risk assessments for bio augmentation strategies.
Current studies on eontaminants within authentie natural media,
within meso- or mieroeosms, might be expanded to explore the hydrologieal behavior of partiaUy modified eontaminants, i.e., intermediates
in eontaminant degradation or transformation. These intermediates
would include eontaminants (or their daughters) that have interaeted
with key mierobial metabolites.
Major Technology Needs
Continuous methods for monitoring the progress of bioremediation
sehemes in real time, in'both spatial and temporal aspeets, are laeking.
Safe, remote sensing teehnologies (preferably operating on-line and in
situ) are needed. Early indieators of bioremediation sueeess or failure
are desirable, as is a method to define or prediet reasonable endpoints.
Computer-modeling teehnologies eapable of dealing with uneertainties
in kinetie rate eonstants, and in minimizing the propagation of
that uneertainty, through eomplex models, might be developed.
These models, ideaUy, would address synergism between eontaminants,
mierobial types, and both abiotie and biotie aetivities. A preliminary
B.D. Faison
conceptual model by which to integrate biologieal, chemieal, and chemical activities relevant to site remediation has been developed. The final,
sophisticated models would aHow numerical predictions of the cost, in
time or money, of action; the consequences of non-action; and the
ramifications of action, in the context of natural contaminant remobilization. Attention could then be paid to the life-cycle costs of various
treatments. It is generaHy understood that a given remediation technology must pay back its own capital costs within 1 year; plots of the
cumulative costs of various treatments as a function of time, when overlaid with biotreatment costs (treatment, feeding, monitoring), would
make biological approaches fiscally comprehensible. Scientists
informed by the modeling data would then be able to select among
and/or prioritize proposed remediation activities.
Metal remediation is a complex problem that is site-specific and involves
uniquely complex hydrologie and biogeochemical interactions. Site
properties are not fuHy understood, and process parameters - how clean
is clean enough? how soon is soon enough? - are poorly defined. An
understanding of microbial community dynamics over space and time
is minimal. Insertion of accumulating data, from hydrologieal, biogeochemical and socio-microbiological studies, into a modeling framework
might allow environmental professionals to understand the integrative
process that produces outcomes quite different from predictions based
on the sum of individual components. It is hoped that such a concerted
attempt to understand the environmental fate of nondestructible contaminants will weaken the somewhat artificial barriers among the
scientific disciplines, and between science, engineering, and mathematics. A complementary interdisciplinary approach that subsurnes various
specialties while acknowledging the central role of chemistry in biological and environmental processes is desired. This approach might
contribute to a systems-level analysis of natural phenomena and form
the foundation of future environmental policy decisions and adaptive
management strategies.
Acknowledgments. This research was funded by the Natural and Accelerated Bioremediation Research (NABIR) program, Office of Biological
and Environmental Research, Office of Science, US Department of
Energy (Award No. DE-OI-OIER63306). The author wishes to thank
investigators, administrators, and reviewers associated with NABIR, the
DOE Environmental Management Science Program, the DOE Office of
Biological Treatment of Metallic Pollutants
Environmental Management, the US Environmental Protection Agency,
the US Geological Survey, and the National Research Council for helpful
Amann RI, Ludwig W, Schleifer K-H (1995) Phylogenetic identification and in situ
detection of individual microbial cells without cultivation. Microbiol Rev
Anonymous (1997) Linking Legacies report DOE/EM-319. US Department of Energy,
Washington, DC
Faison BD, Hu MZ-C, Norman JM, Reeves ME (1996) Design and testing of a continuous metal biosorption system. Final report, 10 March 1994-9 June, 1995. report
number ORNL/M-5065
Faison BD, McCullough J, Hazen TC, Benson SM, Metting FB, Palmisano AC (2003)
Bioremediation of met als and radionuclides: What it is and how it works. A NABIR
primer. Prepared for the Natural and Accelerated Bioremediation Research
Program. 2003 revision of Lawrence Berkeley National Laboratory LBNL-42595. US
Department of Energy, Washington, DC
Kirz J, Jacobsen C, Howells M (1995) Soft X-ray microscopes and their biological applications. Q Rev Biophys 28:33-42
Madigan MT, Martinko JM, Parker J (1997) Brock biology of microorganisms, 8th edn,
Prentice-Hall, Upper Saddle River
Palmer CD, Fish W (1997) Chemically enhanced removal of metals from the subsurface. In: Ward CH, Cherry JA, Scalf MR (eds) Subsurface Restoration. Ann Arbor
Press, Ann Arbor
Riley RG, Zachara JM, Wobber FJ (1992) Chemical contaminants on DOE lands and
selection of contaminant mixtures for subsurface research. DOE/ER-0547T, US
Department of Energy, Washington, DC
Wilson S, Bernath PF, McWeeny R (2003) Handbook of Molecular Physics and Quantum
Chemistry, vol 3: Molecules in the physico-chemical environment - spectroscopy,
dynamics and bulk properties. Wiley, Hoboken
Phytoremediation of Persistent Organic
Contaminants in the Environment
Saleema Saleh,l Xiao-Dong Huang, l Bruce M. Greenberg, 1
and Bernard R. Glick2
Fundamentals of Phytoremediation
Phytoremediation is a technology that uses various plants to extract,
contain, immobilize or degrade contaminants from soil and water for
the purpose of the remediation of these contaminants from the environment. The remediation processes include all plant-influenced biological, chemieal, and physical processes that aid in the accumulation,
sequestration, degradation, and metabolism of contaminants, either by
the plants or by the free-living organisms that constitute a plant's rhizosphere (Cunningham et al. 1995, 1996; Macek et al. 2000). Phytoremediation can be used for in situ treatment of water, sediments, soils
and air. Regarding the removal of toxic organic compounds, it would be
ideal to mineralize them through a combination of plant and bacterial
processes. Phytoremediation is a long-term process that generally
requires several years in order to either remove the poHutant completely, or alter it so that it is no longer harmful to the environment
(Cunningham and Ow 1996). Along with an investment of time, there
is also a significant cost factor involved. Hence, the desire to develop
more cost effective and environmentally friendly remediation strategies
becomes apparent. In the past 10 years or so, green, plant-based
processes have received more attention due to their potential for cleaning up both organic and inorganic contaminants in soils (Cunningham
and Ow 1996; Alkorta and Garbisu 2001).
Plant-based remediation can occur either directly or indirectly. With
direct phytoremediation, plants function in detoxification by accumulating and processing considerable amounts of water-soluble contaminants and subsequently translocating them by means of the
Department of Biology, University of Waterloo, Waterloo, Ontario, N2L 3G 1, Canada
Department of Biology, University ofWaterloo, Waterloo, Ontario, N2L 3G1, Canada,
e-mail: [email protected], Tel: + 1-519-8851211 Ext. 2058, Fax: +1-519-7460614
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
S. Saleh et al.
transpiration stream. During root growth, roots can encounter, alter and
translocate elements and compounds against large chemical gradients.
As a result of transpiration and growth, plants bring the remediation
functions of both the plant - and root -associated microorganisms
together without incurring human engineering costs of soil manipulation or site destabilization (Stomp et al. 1994). Indirect phytoremediation, on the other hand, utilizes root-associated microorganisms to
accomplish detoxification. The root surfaces of plants support active
bacterial bio films and fungal extensions within the rhizosphere. Plant
roots not only supply organic nutrients and energy through exudates,
but also have a major impact on soil oxidation-reduction potential,
either directly by transporting oxygen via roots, or indirectly by changing soil porosity over time, thus providing an oxidation-reduction
microenvironment that is optimal for the growth of soil microorganisms. Vegetation also stabilizes soil, preventing erosion and pollutant
movement. In return, microorganisms greatly enlarge root surface area
and contaminant uploading capacity. They also partially regulate root
metabolie capacities and can alter most measurable soil physical and
chemie al parameters to enhance plant growth (Curl and Truelove 1986).
Increased soil microbial growth supported by root exudates has been
shown to translate into a greater metabolism of organic contaminants
in vegetated soils. Research has indicated that the disappearance of
PAHs (polycyelic aromatic hydrocarbons) from vegetated soil was
double that of unvegetated soil (Aprill and Sims 1990). In addition,
mycorrhizal fungus-plant associations have been found to increase
plant tolerance to heavy metals, metal uptake and translocation and
plant growth parameters, in a number of plant species (Wilkins 1991).
Phytoremediation is most applicable to pollutants that are elose to
the soil surface, relatively nonleachable and cover large surface areas.
Trees have special advantages because they support the growth of a large
diversity of soil microorganisms. Moreover, they not only have the most
massive root systems of all plants, which penetrate the soil for several
meters, but also the largest leaf surface area of all terrestrial plants.
Hence, microbial degradation rates and diversity of remediation activities could be enhanced to a great extent by using trees for bioremediation (Stomp et al. 1994).
There are two primary approaches to plant-based remediation, (1)
pollutant stabilization and containment, where soil conditions and vegetative cover are manipulated, and (2) decontamination where plants
and their associated microflora are used to eliminate and/or degrade
the contaminant. Decontamination reduces the amount of pollutants
within the soil by the rem oval of the pollutants. Pollutant stabilization
does not reduce the quantity of pollutant at a site, it merely alters soil
Phytoremediation of Persistent Organic Contaminants in the Environment
chemistry and either sequesters or absorbs the pollutants into the
matrix so as to reduce or eliminate environmental risks (Cunningham
et al. 1995).
Physical Remediation Strategies
Because of the hazardous nature and persistence in the environment of
many contaminants, it is absolutely essential to develop remediation
technologies to clean up these environments. Numerous physical and
biological techniques have been developed for this purpose.
Physical remediation strategies include air sparging, bioremediation,
the use of bioreactors, bio filters, bioventing, biosparging, capping, composting, flushing, in situ oxidation, extraction, land farming, natural
attenuation, the use of permeable reactive barriers, soil washing, solvent
extraction, thermal desorption and thermal enhancement. The remediation strategy used depends on the nature of the contaminant.
Several factors have to be considered prior to choosing and implementing a suitable technique for phytoremediation. One has to first consider what the contaminants are, the concentration of the contaminant
and the medium in which the contaminants are found, i.e. soil, sediment, groundwater or surface water, and then one has to consider the
cost and effectiveness of the technique for removing the targeted contaminants taking into account the environmental parameters of the contaminated site.
For instance, land farming has been used for in situ remediation. This
technique is effective in reducing concentrations of a variety of small,
volatile chemicals during the early stages of treatment, but degradation
rates severely decline thereafter, especially for recalcitrant compounds
such as PAHs. However, the presence of vegetation can enhance the
degradation of these larger and more complex compounds (Walton and
Anderson 1992). Schwab and Banks (1994) demonstrated that degradation of PAHs was greater in the presence of plants than in their absence.
In this case, enhanced microbial activity was found to be responsible for
increased dissipation of the target PAHs. To improve the effectiveness
of land farming, plant nutrient supplements such as nitrogen and phosphorus (i.e. fertilizer) have been applied to enhance natural microbial
degradation of contaminants. However, this approach is generally still
limited to the degradation of relatively small molecules such as those
found in TPHs (total petroleum hydrocarbons) or BTEX (benzene,
toluene, ethylbenzene and xylene). This technique is low cost, simple to
do, practical and feasible, it produces no secondary contamination and
S. Saleh et al.
has been proven to be effective for certain contaminants, mainly petroleum (US EPA 2000).
Bioremediation is the most popular technique used for the cleanup
of contaminated environments that contain organic compounds. Traditionally, bioremediation has been associated with degradation in a
bioreactor, however, more recently, it refers to in situ bioremediation,
which uses selected microorganisms to degrade contaminants in the
environments. Bioreactors are quite effective for the remediation of
TPHs, BTEX, PAHs and a few other groups of contaminants. However,
the contaminated soil or water has to be excavated and brought to the
In situ bioremediation processes are not as effective as a bioreactor,
mainly due to the in ability of microorganisms to generate a sufficient
biomass in the environment, to allow an acceptable rate of sequestration and degradation of contaminant molecules (Alexander 1995). An
additional complication is that few microorganisms can use high molecular weight contaminants as a sole carbon source; therefore, a readily
degradable organic carbon source must be supplied for co-metabolism
of high molecular weight compounds (Rock 1997; Alexander 1999).
Other methods like soil washing and solvent extraction are very
costly and disruptive to the environment. Very often, these methods
require secondary remediation processes for the extracted contaminants. In addition, physical methods, such as fractioning, thermal desorption, and thermal enhancement, have similar problems to chemical
methods. They are expensive to perform and are not complete processes
for remediation, which generates problems for a secondary remediation
process. Other physical methods for containment of contaminants such
as concrete capping for soil contamination and permeable reactive
barriers for groundwater contamination are widely used in field applications. These methods limit the risk by the containment of environmental contaminants thus reducing the potential for exposure.
Phytoremediation Strategies
Abrief discussion of various phytoremediation processes (Salt et al.
1998; Macek et al. 2000; US EPA 2000) is provided in this section.
Phytoextraction is the uptake of contaminants by plant roots and
translocation within the plants. Contaminants are generally removed by
harvesting the plants. This technology is used to concentrate contaminants from large areas into a much smaller mass to be disposed of. It
reduces or limits excavation and the disposal of large amounts of con-
Phytoremediation of Persistent Organic Contaminants in the Environment
taminated soils and water. This technology has primarily been applied
to metals.
Phyto(rhizo)filtration is the adsorption or precipitation onto plant
roots, or absorption into the roots, of contaminants that are in solution
surrounding the root zone. Phytofiltration results in containment in
which the contaminants are immobilized or accumulated on or within
the plants. Contaminants are then removed by physically removing
the plants. This technology is primarily used for extracted groundwater, surface water, and wastewater contaminated with metals and
Phyotstabilization is the immobilization of a contaminant in soil
through absorption and accumulation by roots, adsorption onto roots,
or precipitation within the root zone of plants, and/or the use of plants
and plant roots to prevent contaminant migration. The term phytolignification has been used to refer to a form of phytostabilization in
which organic contaminants are incorporated into plant lignin. Organic
contaminants can also be incorporated into humic material in soils in
a process related to phytostabilization. Phytostabilization is often used
to treat contaminated soils, sediments and sludges that contain metals.
Rhizodegradation is the breakdown of an organic contaminant in soil
through microbial activity that is enhanced by the presence of the root
zone. Rhizodegradation is also known as plant-assisted degradation,
plant-assisted bioremediation, plant-aided in situ bio degradation or
enhanced rhizosphere biodegradation. Root-zone bio degradation is
the mechanism for implementing rhizodegradation. Root exudates are
compounds produced by plants and released from plant roots. They
include sugars, amino acids, organic acids, fatty acids, sterols, growth
factors, nucleotides, flavanones, enzymes, and other compounds
(Alexander 1995). The microbial populations and activities in the rhizosphere are increased due to the presence of these exudates, and can
result in significantly increased organic contaminant biodegradation in
the soil. Additionally, the rhizosphere substantially increases the surface
area where active microbial degradation can be stimulated. Degradation of the exudates can lead to co-metabolism of contaminants in the
rhizosphere. Plant roots can also affect soil conditions by increasing
soil aeration and moderating soil moisture content, thereby creating
conditions more favorable for biodegradation by indigenous
Rhizodegradation is primarily used for the treatment of soils and
sediments contaminated with organic compounds, but groundwater and
surface water movement can also be treated by the transpiration of
plants bringing contaminants from the water body to the root zone to
be degraded. The contaminants that have been successfully remediated
S. Saleh et al.
through this technology include PAHs, BTEX, pesticides, chlorinated
solvents, PCPs (pentachlorophenols), PCBs (polychlorinated biphenyls),
and surfactants (Aprill and Sims 1990; Jordahl et al. 1997; Schwab 1998).
Phytodegradation or phytotransformation is the breakdown of contaminants taken up by plants through metabolie processes within the
plant, or (less commonly) the breakdown of contaminants external to
the plants through the effect of substances, such as enzymes, produced
by the plants. Any degradation caused by microorganisms associated
with or affected by the plant roots is considered to be rhizodegradation.
When phytodegradation occurs within the plant, contaminant uptake is
dependent on both, the properties of the contaminants including
hydrophobicity, solubility, and polarity, and the characteristics of plants.
Moderately hydrophobie organic compounds are readily taken up by,
and translocated within, plants. Very soluble compounds will not be
adsorbed onto roots or translocated within the plants. Hydrophobie
compounds can be bound to root surfaces or partitioned into the plants,
but cannot be further translocated within the plants. Nonpolar molecules with molecular weights <500 will adsorb onto the root surfaces,
whereas polar moleeules will enter the root and be translocated. Other
factors, like the age of the contaminant and the properties of the
soil, also affect the uptake of contaminants (Bell 1992; Schnoor et al.
1995). Phytodegradation is used in the treatment of soils, sediments,
sludges, groundwater and surface water contaminated with chlorinated
solvents, pesticides, explosives and phenols. Whether a plant can
degrade a particular contaminant depends on the enzymes that the
plant produces.
Phytocover, which is a vegetative cover system, refers to a long-term,
self-sustaining system of plants growing in and/or over material that
poses environmental risk; a vegetative cover reduces that risk to an
acceptable level and, in general, requires minimal maintenance (Macek
et al. 2000). There are two types of vegetative covers: an evapotranspiration cover and a phytoremediation cover. An evapotranspiration
cover is composed of soil and plants engineered to maximize the available storage capacity of soil, evaporation rates, and transpiration
processes of plants to minimize water infiltration. The evapotranspiration cover is a form of hydraulic control by plants.
A phytoremediation cover, on the other hand, consists of soil and
plants to minimize infiltration and to enhance bio degradation of underlying waste. The functions of a phytoremediation cover are to enhance
biodegradation of contaminants, prevent human and wildlife exposure,
and reduce leachate formation or movement. The mechanisms include
water uptake, root-zone microbial degradation and plant metabolism.
In addition to minimizing infiltration, isolating wastes, controlling
Phytoremediation of Persistent Organic Contaminants in the Environment
landfill gas and enhancing biodegradation, vegetative cover systems are
also used to restore forest ecosystems. These technologies are used for
landfill covers, groundwater containment, the uptake of infiltration
surface water, and in the treatment of contaminated soil, sludge and
Phytovolatilization is the uptake and transpiration of a contaminant
by a plant, with the release of the contaminant or a modified form(s) of
the contaminant into the atmosphere from the plant through contaminant uptake, plant metabolism and plant transpiration. Phytodegradation is a related phytoremediation process that may occur along with
phytovolatilization. Phytovolatilization has mainly been applied to
groundwater, but it can also be applied to soil, sediments and sludges.
The contaminants treated by the technology are chlorinated solvents
and metals.
Hydraulic control, also known as phytohydraulics or hydraulic
plumes control, refers to the use of plants to limit water movement
through uptake and consumption in order to contain or control the
migration of contaminants. This technology is used in the treatment of
groundwater, surface water and soil water, contaminated with water
soluble, leachable organics and inorganics at concentrations that are not
toxic to the plants. Hydraulic control requires plants with deep root
systems such as poplar, willow and cottonwood trees.
Riparian corridors or buffer strips are similar in concept to physical
and chemical permeable reactive barriers, in that they treat groundwater without extraction or containment. The mechanisms for remediation include water uptake, contaminant containment, contaminant
uptake and plant metabolism. Riparian corridors may incorporate
certain aspects of hydtaulic control, phytodegradation, rhizodegradation, phytovolatilization and phytoextraction and are generally applied
along streams and river banks to control and remediate surface runoff
and groundwater contamination moving into the river. They can also
be installed to prevent down gradient migration of a contaminated
groundwater plume and to degrade contaminants in the plume. This
strategy is used in the treatment of surface water and groundwater.
Phytoremediation technologies, in general, are considered as in situ
applications by the establishment of vegetation in areas of contaminated
soil or groundwater. However, contaminated soils can be excavated and
placed into a treatment unit where phytoremediation is applied. Similady, contaminated groundwater can be pumped into a place where it
will be treated with phytoremediation techniques. There are two main
classes of contaminated media that can be treated by phytoremediation
technology: solid media such as soil, sediment and sludge, and aqueous
media like groundwater, surface water and wastewater.
S. Saleh et al.
Phytoremediation technologies are most appropriate for large areas
of low and moderately contaminated soils that would be prohibitively
expensive to remediate using conventional technologies. Besides, the
depth and volume of contamination as weH as soil characteristics that
affect plant growth have to be considered for phytoremediation. Moreover, the contaminants in the soil should be within the root zone depth
of the selected plants (US EPA 2000).
Advantages and Disadvantages of Phytoremediation
Plants are solar-driven, pumping and filtering systems that have
significant uploading and degrading capacities. They are a more environmentaHy friendly alternative to excavation and treatment since they
preserve the natural structure and texture of soil. Plant roots can find,
alter or translocate elements and compounds against large chemical
gradients. Plants are also more cost-effective compared to other physical remediation systems. FinaHy, plants have the potential to be a rapid
cleanup strategy by providing large amounts of biomass. However, more
research needs to be done in order to optimize and realize this
Although using plants for remediation of persistent organic contaminants holds advantages over other methods, many limitations exist
for current application on a large scale (Rock 1996, 1997; McCutcheon
1996). For instance, when contaminant concentrations in the soil
are high, many plants will not grow sufficiently to provide adequate
biomass for successful remediation. In many cases, contaminated
soils are also poor in nutrients, which limit plant growth, thus slowing the remediation process. Furthermore, microbial populations in
contaminated soils are often depressed in both diversity and abundance.
Contaminated soils often do not contain the appropriate microorganisms for the efficient degradation of the contaminants, further limiting
the effectiveness of remediation. In addition, contaminants must be
within the root zones of plants. Therefore, phytoremediation processes
are, in general, slow and the time scale for complete remediation is
often unacceptably long (McCutcheon 1996; Rock 1997). To address
this problem, a multi-component phytoremediation system for the
removal of recalcitrant organic contaminants from soil was developed
(Huang et al., submitted). The remediation system includes physical
(volatilization), photochemical (photooxidation), bioremediation and
phytoremediation processes. The techniques applied to implement
these processes include land farming (aeration and light exposure),
Phytoremediation of Persistent Organic Contaminants in the Environment
and the introduction of contaminant degrading bacteria and enhanced
plant growth through the use of plant growth promoting bacteria. This
resulted in an overall remediation process that was very effective at
removing persistent, strongly bound PAHs from soil. This research indicates that a combination of techniques may be a viable solution for
remediating persistent organic contaminants from soils (Huang et al.,
Phytoremediation of Organics
Sources of Organic Contaminants in the Environment
Large amounts of hazardous waste have been released into aquatic and
terrestrial environments due to industrial and agricultural activities
and energy consumption (Safe 1984). Many of these organic compounds
are toxie, mutagenie and carcinogenie, and of great environmental
concern. They persist for a long time in the environment and pose a
significant hazard to ecosystems and human health (Neff 1979; Safe
1984; Piver and Lindstrom 1985). Not only are they present in high concentrations in the environment, but also the number and types of contaminant compounds are overwhelming.
One of the major types of environmental contaminants from agriculture is pesticides. For instance, more than 4.5 billion pounds of
chemieals are used annually as pesticides in the United States (US EPA
2000). Pesticide contaminants have a worldwide distribution, and can
be found in both urban and rural areas as weH as in aquatic and terrestrial environments. Other groups of persistent organic contaminants
that are distributed as widely as pestieides are TPHs and PAHs contaminants associated with fossil fuel processes and consumption.
TPH contaminants are mainly caused by oil spills during exploration,
drilling, transporting and refining.
Probably, the most degradation-resistant environmental contaminants are PCBs which were previously used as heat transfer fluids, diluents, plasticizers, lubrieant inks, fire retardants, paint additives, sealing
liquids, immersion oils, adhesives, de-dusting agents, laminating agents
and dielectric fluids for capacitors and transformers. From the 1930s to
the 1970s, millions of tons of PCBs were produced worldwide and more
than 30% of the chemicals was released into the environment and widely
distributed in waters, sediments, lands and biota.
S. Saleh et al.
Similarly, other chlorinated aromatics such as (PCP) and polychlorinated terphenyls (PCTs) were used as PCB replacements for a short
period of time in the 1970s. Other halogenated compounds like perchloroethylene (PCE), tetrachoroethene and chlorinated solvents are
produced by smelters, mines, battery-recyding plants, refineries, waste
incinerators and chemical factories, and are often used in dry deaning
solutions, adhesives and paint removers. They are major contaminants
in groundwater because they are in liquid form and highly soluble in
water. The other major groundwater contaminants are benzene, toluene,
ethylbenzene and xylene. They are components of gasoline released into
the environment by accidental spillage during filling and transporting
processes, and leakage from storage tanks. Explosives are another group
of contaminants and inc1ude compounds such as nitrotoluenes, nitrate
esters and nitramines. Not only are they persistent in the environment,
they are also toxic, dangerous and potentially explosive. They are
usually present in battlefields, military sites, and weapon and explosive
manufacturing sites.
Factors That Affect the Uptake of Organic Contaminants
The movement of an organic contaminant in soil depends on the chemical's water solubility, vapor pressure, molecular size and charge and
also on the presence of other organics in the soil. The ability of the soil
to absorb and sequester organic compounds is directly associated with
the organic matter in the soil, the type and amount of day present, soil
structure and pH, as well as with the age of the spill and the water Bux.
Since plants can alter any number of these parameters, the use of plants
is therefore site-, contaminant- and timing-specific. There is a strong
correlation between soil organic matter and bioavailability (Cunningharn and Ow 1996). The bioavailability of organic contaminants usually
decreases over time, hence older contaminants, as opposed to more
recent contaminants, are more difficult targets for phytoremediation.
There is also a direct relationship between the absorption of organic
compounds by plant roots and the relative lipophilicity of these compounds. Once absorbed, these compounds are bound into plant tissues
and subsequently made less biologically available, and furthermore,
may be unavailable for chemical extraction. Previously, McMullin
(1993) used carrots to absorb dichloro diphenyl trichloroethane (DDT).
The carrots were then harvested, solar dried and incinerated to destroy
the contaminant. Plants can also be used for the direct extraction of
organic contaminants from soil by root accumulation, xylem translocation and subsequent volatilization from leaf surfaces.
Phytoremediation of Persistent Organic Contaminants in the Environment
In contrast to inorganic pollutants, organic pollutants can be
degraded or even mineralized by plants or their associated microorganisms. The numerous metabolic activities that take place in the roots
and shoots of plants may be supplemented by the abilities of rhizosphere microorganisms to act upon these organic contaminants.
Research points toward a combination of both plants and microorganisms as being the best strategy in remediation of a number of organic
pollutants. Alternatively, transgenic plants may be engineered to express
microbial genes for biodegradation (Cunningham and Ow 1996).
(ontaminant (lasses
Traditionally, research on microbial transformations of organic compounds in the rhizosphere has mainly focused on agricultural chemicals such as insecticides and herbicides. Researchers have shown an
increase in pesticide degradation in the rhizosphere of a variety of plant
species. This increase in degrading capacity correlates with increased
numbers of pesticide-degrading microorganisms. It has also been suggested that microbial consortia are generally responsible for degradation, rather than a single member of the microbial community.
Sandmann and Loos (1984) found, following herbicide treatment,
an increase in the number of 2,4-dichlorophenoxyacetate (2,4-D)degrading bacteria in the rhizosphere of previously untreated sugarcane
soils. The degradation of nonagricultural organic chemicals is also
accelerated in the root zone. Aprill and Sims (1990) evaluated the persistence of a variety of PAHs in the root zone of eight prairie grasses
and found that the PAH disappearance was consistently greater in vegetated than unvegetated controls, probably due to the microbial degradation of PAHs. In addition, these researchers speculated that the
humification of PAHs (transformation of PAHs into secondary by-products in the soil) may have accounted for their increased disappearance
in the presence of plants. Plant uptake and the metabolism of organic
compounds mayaiso have contributed to the enhanced degradation of
this material (Bell 1992). The uptake of organic compounds from soil
solution through plant roots depends on the physical and chemical
properties of the compounds, as weIl as environmental conditions and
plant characteristics (Walton and Edwards 1986). Furthermore, research
has also shown the ability of ectomycorhizal fungi to degrade certain
congeners of PCBs in vitro (Donnelly and fletcher 1992; Katayama and
Matsumura 1993). It was postulated that the prolonged exposure of soil
microorganisms to the toxicant may speed up degradation through the
selective enrichment of biodegradative species in the microbial envi-
S. Saleh et al.
ronment. Overall, plant uptake is favored for small and low molecular
weight polar compounds, whereas large, high molecular weight compounds tend to be excluded from the root.
The remediation of chlorinated solvents is one of the most successful examples of the use of phytoremediation. This is due to the
hydrophobie nature of chlorinated solvents that allows plants to readily
uptake and translocate them within the plant tissues, where they can be
degraded and/or volatilized. However, contaminated sites often contain
a mixture of pollutants that must be tolerated and then degraded. When
a variety of different forage grasses were combined with several bacterial inoculants and tested for the ability to degrade various concentrations of mono- and di-chlorinated benzoie acids, it was found that the
plant -bacteria associations better tolerated and degraded mixtures of
contaminants rather than individual compounds (Siciliano and
Germida 1997,1998).
PCBs are a group of halogenated polycyclic aromatic hydrocarbons
of toxic environmental pollutants. The lower chlorinated species are
more soluble, less adsorbed, more volatile and tend to be lost, leaving
residues of the more highly chlorinated PCBs (O'Connor et al. 1990).
Furthermore, PCB-contaminated sites contain mixtures of chlorinated
benzoie acids from PCB degradation. This may pose a problem because
one toxic component of the mixture may be inhibitory and limit the
degradation of other components. Siciliano and Germida (1997) also
determined that inoculated plants reduced chlorobenzoic acid levels in
soil, however, the inoculants themselves were ineffective in enhancing
degradation in the absence of a plant. This suggests that plant-bacteria
associations are imperative in reducing chlorobenzoic acid levels, and
are formed during phytoremediation.
Another example of the successful application of phytoremediation
involves hydrocarbon contaminants such as TPHs and BTEX, both of
which are quite widespread contaminants of soils and waters. Many different plant species (wheat, badey, corn, alfalfa, fescue) have been used
in the phytoremediation of petroleum and related contaminants. These
contaminants are readily taken up and translocated within plants. The
functional processes of phytoremediation are phytoextraction, phytodegradation, rhizodegradation and phytovolatilization. There are
many field applications for the phytoremediation of petroleum and its
products in contaminated soils and waters (Anders on et al. 1997;
Alexander 1999).
The bioavailability of organic environmental contaminants depends
on the relative lipophilicity of the compound (Bell 1992 ), on the soil type
(organic matter, pH, and clay content) and on the age of the contaminant (Hatzinger and Alexander 1995). Contaminants that are tightly
adsorbed to soil particles and resist uptake by microbes or plants are
Phytoremediation of Persistent Organic Contaminants in the Environment
poor targets for phytodegradation. When contaminants are not readily
mobile, or fail to interact with other organisms, phytostabilization
should be considered.
Both organie and inorganic contaminants can form chemical and
biologieal associations of varying intensities in biologically active soils.
These associations can decrease contaminant availability for bioremediation, but at the same time reduce the likelihood of leaching. Such
sequestration processes indude the incorporation of organics into
lignin or soil humus. Phytostabilization techniques exploit these processes to decrease bioavailability further, with the aim of significantly
reducing or eliminating the toxic effect of the contaminant in the environment (Cunningham et al. 1996).
Once inside the plant, compounds are either stored unchanged,
bound to plant structural constituents, metabolized, or passed through
the plant and volatilized (Hathaway 1989). Plants have significant metabolie activities in the root and shoot that may be exploited for phytoremediation (Sandermann 1992; Schnoor et al. 1995). Moreover,
not all plant-associated degradative capacities need to be internal to
the plant. Plant oxidoreductases have been proposed for the decontamination of phenolics and anilines in water-treatment systems (Dec
and Bollag 1994). Research on freshwater sediments has been found
to contain degradative enzymes of plant origin (Schnoor et al.
1995). This suggests that plant enzymes released after plant death
may have significant catalytic effects and may be useful for
Plant-associated mieroflora are being investigated for their role in
bioremediation. It has been determined that accelerated rates of degradation for many pesticides, as weIl as trichloroethylene and petroleum
hydrocarbons, occur in the rhizosphere (Aprill and Sims 1990; Anderson et al. 1993). This suggests that microorganisms in dose proximity
to plants playavital role in enhancing the phytoremediation of organic
Role of Microorganisms in Soil Bioremediation
A variety of naturally occurring toxic and recalcitrant organic compounds exist on Earth, many of which are naturally mineralized. Soil
organic matter is recyded by a diverse array of microorganitsms that are
responsible for mineralizing most organic matter to carbon dioxide.
Residual organic matter that is not readily mineralized is then incorporated into humus where it is no longer available to exert its toxie effect
in the environment. One can apply these naturally occurring systems in
the deanup of man-made pollutants.
S. Saleh et al.
Microbial communities offer a potentially important treatment strategy for in situ biologie al remediation of chemically-contaminated soils.
Moreover, research has shown that vegetation enhances microbial
degradation rates of organic chemical residues in soils. These findings
are important because vegetation may provide a low-cost alternative to
expensive capital investment technologies for soil cleanup (Anders on
et al. 1993).
The rhizosphere is an area of increased microbial activity that
enhances transformation and degradation of pollutants (Bollag et al.
1994). The overall effect of plant-microbe interactions is an increase in
mierobial biomass by an order of magnitude or more, compared with
the microbial population in bulk soils. This increase in microbial
growth and activity in the rhizosphere may be responsible for the
increased metabolie degradation rate of certain xenobiotic compounds
in the rhizosphere. The actual composition of the microbial community
is dependent on root type, plant species, plant age, and soH type, as weIl
as the exposure history of the plant roots to xenobiotics (Sandmann and
Loos 1984; Campbell 1985; Atlas and Bartha 1992; Bolton et al. 1993).
Degradation rates in the rhizosphere can be stimulated by supplying
inorganic nutrients and oxygen, degradative microbial inocula and
enzymes. Rhizosphere soil also has higher concentrations of carbon
dioxide. In addition, it is usually at least 1-2 pH units different from
bulk soil. Oxygen concentrations, osmotic and redox potentials and
moisture content are other parameters influenced by vegetation (Curl
and Truelove 1986).
Several suggestions have been given as reasons behind the formation
of plant-bacteria associations in contaminated soils. pfender (1996)
postulated that bacterial inoculants might protect plants from toxicants
in soil. Crowley et al. (1996) suggested that plants may provide a niche
for bacteria to maintain their degradative plasmids. Siciliano and
Germida (1997) found that inoculants did not decrease chlorobenzoic
acid phytotoxicity per se, and many inoculants caused plant death in
contaminated soil, perhaps due to a toxie intermediate being formed in
the degradative pathway, as noted by Barriault and Sylvestre (1993).
Other researchers have hypothesized that bacterial toxin production
may be the cause of inoculant death in soil mierocosms (Havel and
Reineke 1992). Conversely, plants may provide a safe niehe for bacterial
inoculants and also inoculants may require specific plants to reduce
contaminant levels in soil. Some plants have intrinsic bioremediation
activity, whereas others require bacterial inoculants to reduce contaminant levels in soH. When studies were done with bean (Phaseolus vulgaris), ryegrass (Lolium perenne) and crested wheatgrass (Agropyron
cristatum), it was shown that these species degrade contaminants in soil
Phytoremediation of Persistent Organic Contaminants in the Environment
in the absence of bacterial inoculants (Crowley et al. 1996; Gunther et
al. 1996; Pfender 1996). On the other hand, millet (Panicum miliaceum
1.) had no effect on PCP levels in soils until it was inoculated with PCPdegrading bacteria (pfender 1996).
Improved Mechanisms to Optimize Phytoremediation
The first step in optimizing the phytoremediation of organic contaminants is the selection of species of plants that have significant capacity
to take up these compounds and are able to establish microbial associations that facilitate their degradation. The selected plant species should
possess characteristics that enable them to grow on contaminated sites
and detoxify the contaminant. They should also be integrated in field
deanup efforts with relative ease and adhere to any regulatory concerns.
A variety of plant species have been investigated and successfully used
for remediation of organic contaminants. These indude trees such as
poplars, cottonwoods and willows, as well as shrubs like elderberry,
mulberry and hackberry (US EPA 2000).
The low permeability of day soils poses achallenge for in situ
biotreatment of PAH contaminants in soils. The low fiux of nutrients
and electron acceptors mayaiso restrict microbial activities. Furthermore, bacteria and fungi cannot use high molecular weight PAHs as a
sole carbon source. Laboratory studies have demonstrated a significant
increase in PAH removal through the use of a grass-enhanced system.
PAHs partition in day-soil organie matter complexes, and oil globules
get trapped in soil pores, becoming unavailable for bio degradation. The
carbohydrates and amino acids exuded by roots in the rhizosphere can
sustain a dense microbial community, which in turn may enhance
bio degradation rates of organic contaminants. In addition, organic exudates from roots may induce mierobial co-metabolism of high molecular weight PAHs.
Examples of grasses that have been used in phytoremediation studies
successfully indude tall fescue, dover, ryegrass, prairie grass and turf
grass. Other plants that are capable of taking up toxie contaminants
indude alfalfa, sorghum, wheat, barley and corn as well as aquatic plants
like phragmites (i.e. reeds and cattails) (US EPA 2000).
One of the key elements for successful phytoremediation is to use
plant species that have the ability to proliferate in the presence of high
levels of contaminants. The degree of tolerance to contaminating chemicals varies among plant species. Certain physiologieal characteristics
of plants make given species less susceptible to the negative effects of
some compounds. The identification of these characteristies will allow
S. Saleh et al.
for the establishment of selection criteria for plants that are more effective in phytoremediation. Huang et al. (submitted) demonstrated that
these properties included increased water content in tissue, accelerated
root growth, and maintenance of both chlorophyll content and the
chlorophyll alb ratio. This study also showed that PGPR can accelerate
plant growth, especially roots, in heavily contaminated soils, thereby
reducing the toxic effects of various contaminants to plants. Moreover,
since much of the phytodegradation takes place in or around plant
roots, the increased root biomass in contaminated soilleads to a greater
remediation potential.
Work by Fletcher and Hedge (1995) has shown that a number of
plants such as Morus rubra 1. (mulberry) have sufficient levels of phenolic compounds to support PCB degradation. These researchers postulated that the phenolic compounds released by plants can support the
growth of PCB-degrading bacteria and suggested that phenol provision
via root exudates may allow the development and maintenance of PCBdegrading communities in the rhizosphere. Siciliano and Germida
(1997, 1998) postulated that plants and bacteria might form beneficial
associations that would degrade toxicants in soil, and developed a
screening methodology based on the germination of test plants in contaminated soils and the inoculation of grass es that are known to thrive
on contaminated soils, with PGPR, or bacteria capable of contaminant
Other strategies for improving the phytoremediation of organic
compounds include screening bacteria for degradative capabilities,
as well as the application of growth promoting bacteria to enhance
degradation and plant growth. Recently, Johnsen et al. (2002) have
developed a novel microtiter plate assay to i.solate PAH mineralizing
strains. Numerous bacterial strains able to grow on three- and four-ring
PAHs have been isolated (Kastner et al. 1994; Mueller et al. 1997;
Bastiaens et al. 2000; Ho et al. 2000). Growth on various PAHs as sole
sources of carbon and energy cannot easily be determined by standard
growth assays, due to the low water solubility and bioavailability of most
PAHs, which leads to slow bacterial growth and low cell yields.
However, this novel screening technique overcomes these challenges.
Furthermore, optimal media to isolate phenanthrene-degrading strains
have been developed recently (Andersen et al. 2001). Hybrid techniques
or multi-component phytoremediation systems using degradative
bacteria and/or plant growth-promoting bacteria with plants have
also been found to optimize phytoremediation (Huang et al., submitted). It is quite possible that plant growth-promoting bacteria
function to lower the toxicity of organic pollutants by lowering
ethylene levels in plants, through the production of ACC deaminase,
an enzyme that is known to function under stress conditions such as
Phytoremediation of Persistent Organic Contaminants in the Environment
flooding, pathogens as weH as sah and drought stress (Wang et al. 2000;
Grichko and Glick 2001a-c; Mayak et al., submitted). Land farming, the
modification of soil and water pH and addition of soil supplements such
as chelating agents, nutrients and organic matter prior to planting can
also enhance microbial degradation rates and plant toxicant uptake
Phytoremediation is still being actively researched, and novel strategies
as weH as the optimization of plant-microbe associations seem to be the
key to enhancing the degradation of organic poHutants. Additional
investigations need to be conducted to determine the influence of the
size and structure of plant roots on toxicant degradation as weH as the
potential for roots to release surfactant compounds that may solubilize
xenobiotics. In addition, the isolation of contaminant-degrading strains
using superior selection criterion will aHow the integration of these
species into contaminant sites, and subsequently enhance degradation
rates at these sites.
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Phytoremediation of Metals
and Inorganic Pollutants
Tomas Macek, l Daniela Pavlikova,2 and Martina Mackova3
Within the last two centuries of industrial production, the mmmg
industry and different urban activities caused environmental contamination on a large scale. Agricultural practices, especially those of the
twentieth century, are also responsible for the widespread contamination of soil, sediments and water, not to mention wars and different
conflicts responsible for production, use and storage of explosives and
chemical warfare agents (Macek et al. 1998; French et al. 1999). Considering the rapidly growing world population and the detrimental
impact of agricultural systems on the environment (Zechendorf 1999),
the cleaning of large contaminated sites is important for ensuring sustainable development, because it might reduce pressure to expand into
the wilderness, rainforests and marginal lands, thus supporting biodiversity and preservation of vital ecosystems. Air contamination by
volatile toxic compounds is also an important problem.
One approach to the removal of contaminants from soil involves the
use of physico-chemical methods. These techniques are unfortunately
very costly and often further destroy the environment. For this reason,
in the last decade, much of the research was oriented towards the use of
biological remediation methods. One option is the use of plants for
transfer, accumulation and the removal of pollutants from the environment, or at least to reduce their spread (Cunningham et al. 1995;
1 Department of Natural Products, Institute of Organic Chemistry and Biochemistry,
Academy of Sciences of the Czech Republic, Flemingovo n. 2, 166 10 Prague 6, Czech
Republic, e-mail: [email protected], Tel +420-220-183340, Fax +420-224-310177
2 Department of Agrochemistry and Soil Science, Faculty of Agronomy, Czech Technical University, Kamycka 129, 165 21 Prague 6, Czech Republic
3 Department of Biochemistry and Microbiology, Faculty of Food and Biochemical
Technology, Institute of Chemical Technology, Prague, Technicka 3, 16628 Prague 6,
Czech Republic
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
T. Macek et al.
Macek et al. 2000). This approach can be used for the removal of both
inorganic and organic xenobiotics and pollutants present in the soil,
water and air (Sah et al. 1995; Lasat 2000; Ryslava et al. 2003). The type
of contaminants ranges from inorganic fertilisers to pesticides, from
heavy metals and trace or radioactive elements to recalcitrant organic
compounds like PCBs or PAHs.
It has been known since prospectors started looking for metalore
that some plants are resistant to high concentrations of heavy metals.
Finding or developing plants that acquire high levels of metal contaminants in harvestable tissue was thought impossible until the (re)discovery of a small group of remarkable plants called hyper-accumulators.
These plants are relatively uncommon, but are present throughout the
plant kingdom, and can accumulate surprisingly large amounts of some
met als (Brown et al. 1995), probably by mechanisms similar to the
uptake of met als essential for their enzymatic activities. These so-called
hyper-accumulators have a selective advantage when growing on high
metal-containing soil, i.e. they are being protected, e.g. against herbivores. Approximately 400 plant species have been classified as hyperaccumulators of heavy metals, most of them accumulate nickel (Baker
et al. 1994). Hyper-accumulators are found usually at the surface
in places where ores have a high metal content (e.g. Congo, New
Caledonia). These locations are unfortunately characterised by a
relatively low biomass formation and are therefore not weIl suited for
phytoremediation purposes. The mechanisms involved in hyper-accumulation are not yet fully understood, but already attempts have been
made to increase the accumulation of heavy metals in plants producing
high amounts of biomass (Krämer and Chardonay 2001; Clemens et al.
2002). Morel et al. (1999) discussed some important aspects of plantsoil interaction with respect to hyper-accumulation. Data are gradually
being collected on the role of peptides and proteins in metal accumulation (Kotrba et al. 1999; Mejare and Bülow 2001).
One of the main aims of phytoremediation is to prevent the migration of pollutants to a site of ac tu al danger to human health (Salt et al.
1995; van der Lelie et al. 2001). The establishment of vegetation on a
contaminated site also reduces soil erosion by wind and water, which
helps to prevent the spread of contaminants and reduces the exposure
of humans and animals. Olson and Fletcher (2000) described the ecological recovery of vegetation at a former industrial sludge basin,
together with its implications for phytoremediation. Most of the highly
productive plants only accumulate amounts of heavy metals too low to
be economically feasible for the cleaning of contaminated soil (Sah et
al. 1995; Macek et al. 1999; Bock et al. 2002). Metal concentrations in
shoots of some known hyper-accumulators can re ach the high levels
indicated in Table 1.
Phytoremediation of Metals and Inorganic Pollutants
Table 1. Heavy metal accumulation in some hyperaccumulating plants. (Adapted from Cunningham and
Ow 1996)
Level (mgkg- 1
dry wt.)
Macadamia neutrophylla
Thlaspi caerulescens
Psychotria douarrei
Ipomoea alpina
Haumaniastrum robertii
Thlaspi rotundifolium
Thlaspi caerulescens
One of the main tasks of basic research in the field of phytoremediation can be defined as the selection of plant species most suitable for
phytoremediation purposes (Macek et al. 2002a). Not all plant species
are equally well suited to metabolise or accumulate pollutants. Plants
removing heavy metals must grow fast in the contaminated environment, to be resistant to the metals, to be able to accumulate toxic metals
and transfer cations or oxyanions into the harvestable (aboveground)
parts, or transform them into less-toxic forms (Krämer and Chardonay
2001). Plants which are able to remove more than one pollutant are
especially useful, because contamination is usually caused by a mixture
of more toxic compounds. Many elemental pollutants enter plants by
the basic transport systems designated for nutrient uptake. A number
of xenobiotics are then stored in vacuoles as protection against their
toxic effects.
Plants have developed their own systems for the binding of heavy
metals based largely on the synthesis of phytochelatins (Grill et al.
1989). Heavy metal binding in plants is normally achieved by different
peptides, phytochelatins and phytosiderophores (reviewed by Kotrba et
al. 1999). The latter are known to be ubiquitous in plants since Rudolph
et al. (1985) compared the distribution of nicotianamine in the Plant
Kingdom. There have been various attempts to improve the capacity of
high biomass-yielding plants to accumulate heavy metal. The other
option is to use traditional breeding to produce plants that grow faster
artd taller (e.g. alpine pennycress).
Examples of the application of phytoremediation in the field include
the use of hydroponically grown sunflower in artificial lagoons on
a river near Chernobyl and at a uranium-enrichment plant in Ohio
(where the culture was absorbing radioactive metals), or phytoextrac-
T. Macek et al.
tion of heavy metals and radionuclides from soil by the Indian mustard
plant in Trenton, New Jersey (Salt et al. 1995).
Especially during the accumulation of heavy metals, but also during
the removal of organic contaminants, the main limiting factor in largescale exploitation of plants is the long duration required for soil decontamination. Kayser and Felix (1998) evaluated the efficiency of the use
of metal-accumulating plants based on long-term (5-year) field trials at
three heavy metal-contaminated sites in Switzerland. The long time
required for significant reduction of the toxic metal content in soil was
considered the limiting factor for the wider exploitation of plants for
this purpose. Nevertheless, phytoextraction of metals presents large
economic opportunities because of the size and scope of environmental problems associated with metal-contaminated soils and the competitive advantages offered by a plant-based technology (Chaney et al.
1997). In the case of radioactive compounds, showing high biological
effect even at very low concentrations, complete decontamination by
plants seems to be practically impossible (Soudek et al. 2003).
Much effort has been devoted to the preparation of plants, specifically
tailored for phytoremediation purposes, by breeding or genetic manipulations. Recently, more information has appeared in the literature
about possibilities to increase the expression of the genes already
present, or to introduce bacterial or mammalian genes into plants, in
order to increase the natural ability of plants to cope with xenobiotics
(Macek et al. 2002a). Particularly in the case of metal uptake, basic
research studies are quite extensive. Complex interactions of transport
and chelating activities control the rates of metal uptake and storage. In
recent years, several key steps have been identified at the molecular
level, enabling us to initiate transgenic approaches to engineer the transition metal content of plants. Clemens et al. (2002), in an excellent
review, summarised the determinants of metal accumulation, mobilisation, uptake, sequestration and xylem transport, and discussed the role
of different families of transporter proteins. Many problems are still
unresolved, the transport and storage forms of transition metals are
largely unknown, and many systems have yet to be analysed. Despite
the unsolved questions, many attempts were made to prepare transgenic
plants with improved properties for phytoremediation purposes. Transgenic plant technology exhibits many obvious advantages over conventional plant breeding approaches for crop improvement.
The most important parameter for the selection of suitable plants is
not the tolerance of the plant to heavy metals, but its effectiveness in the
accumulation of heavy metals. In addition to accumulation capacity,
biomass production must be considered in order to determine the total
metal uptake. To achieve better economics of the whole process, energy
Phytoremediation of Metals and Inorganic Pollutants
plants or some technologically useful plants have to be considered as
candidates for use in phytoremediation, like hemp or flax, as suggested
by Griga et al. (2003). The tolerance to some heavy metals has recently
been studied by Bringezu et al. (1999). In the case of contaminated
water, a wider range of organisms seems to be promising, from water
plants to microalgae, from root filters to immobilised bacteria. Was te
products, such as sludge, ash, wastewater, mine drainage water and
urban refuse city dumps, have high nutrient values but often also high
heavy metal content. An overview of the decontamination of heavy
metal-contaminated sludge was published by Gerth et al. (2000).
Rhizofiltration, a process in which plant roots grown in water absorb,
concentrate and precipitate toxic metals from polluted effluents, has
also been tested.
The contaminant must be in a biologically accessible form, and root
absorption must take pi ace. Translocation of the contaminant from root
to shoot makes tissue harvesting easier and lessens exposure to the contaminant. Decontamination of a site requires a plant with both a high
biomass yield and a metal accumulation capacity of 1 to 3% of the metal
on a plant dry matter basis. Currently, there is little or no information
available about the agronomics, genetics, breeding potential and disease
spectrum of most of the phytoremediation candidates considered to be
suitable based on the above criteria. This fact favours the use of known
agricultural crops.
The response of high er plants to cadmium in their environment is
summarised in much detail by Sanita di Toppi and Gabbrielli (1999).
The authors discussed many factors, like sequestration, cell wall immobilisation, plasma membrane exclusion, stress responses ete., explaining
models of tolerance and detoxification mechanisms. The wide range of
metals and metalloids, studied from the phytoremediation perspective,
may be illustrated by the studies of rapid accumulation of technetium
in plants and proteins that bind it (Krijger et al. 1999).
After harvesting, a processing step is necessary. Alternatively, the harvested biomass could be reduced in volume and/or weight by composting, an aerobic digestion, low temperature incineration and leaching.
This step would further decrease the costs of handling and processing,
as discussed by Salt et al. (1995). After stabilisation, subsequent landfilling can be considered. With some metals, after leaching or smelting
in kilns (e.g. Ni, Zn, and Cu), the value of the reclaimed metal may
provide an additional incentive for remediation. This step reduces the
generation of hazardous waste and generates recycling revenues.
The compounds which plants use for the binding of heavy metals
have recently been reviewed by Kotrba et al. (1999). In general, there is
a significant lack of knowledge of heavy metal transport, vacuolar
T. Macek et al.
uptake, ABC transporters and other factors important for a really effective exploitation of plants in heavy metal remediation. Recently, many
laboratories have star ted addressing these aspects of phytoremediation
using transgenic plants.
Phytoremediation and Rhizoremediation
Plants remove compounds from soil by the direct uptake of the contaminants, followed by their subsequent transformation, transport and
accumulation in non-phytotoxic forms, which are not necessarily nontoxic for humans (Schnoor et al. 1995). If the biotransformation reactions lead to a decrease in the toxicity of a given xenobiotic compound,
the reaction is called detoxification. In some cases, the biotransformation reactions can lead also to an increase in the toxicity of the
compounds. This is mostly the case with organic compounds, but may
change a recalcitrant compound into a more easily degradable one.
Thus, a more toxic or more soluble compound generated in an early
metabolic step can be rendered harmless in further steps. In some cases,
non-mutagenic compounds (prornutagens) can undergo metabolic activation into mutagenic compounds, a topic which is discussed in detail
in the review by Plewa and Wagner (1992).
The Role of the Rhizosphere
Rhizoremediation is a term inseparable from phytoremediation. The
living plant roots exert strong changes in the physical, chemical and biological properties of the soil. The complexities of roots and their rhizosphere have been discussed recently by McCully (1999). The soil-root
interface, the rhizoplane, and the narrow volume surrounding roots
(a few millimetres), called the rhizosphere, is characterised by several
processes such as the exudation of organic compounds, root respiration
(absorption of oxygen and release of carbon dioxide), the release of
protons and other mineral ions, and the uptake of water and solutes.
The influence of plant roots on the soil environment has been discussed
by Morel et al. (1999).
Various groups of microorganisms are present in the soil (bacteria,
actinomycetes, fungi, algae and viruses). In the rhizosphere, their
numbers are high er than in the bulk soil. Qualitative changes in the
microbial population were also observed (Fletcher et al. 1995; Morel
et al. 1999). The composition of the microbial population is controlled
Phytoremediation of Metals and Inorganic Pollutants
by both soil and plant factors, including compounds with allelopathic
effects. The intensive microbial activity in the rhizosphere is due to the
presence of high amounts of available carbon released as exudates by
roots (Donnelly et al. 1994). Plants respond to the presence of microorganisms by modification in growth, e.g. symbiosis or interactions
with free-living organisms. Rhizobium-legumes or endo- and ectomycorrhizas can serve as examples of direct and positive actions
(Batkhuugyin et al. 2000), while pathogenicity is a negative effect. Donnelly and Fletcher (1994) studied the potential use of mycorrhizal fungi
as bioremediation agents. The use of arbuscular mycorrhizal (AM)
fungi in soil remediation has been recently discussed byVosatka (2001).
The AM symbiosis affects many aspects of plant physiology, rooting,
nutrient cycles, nutrient acquisition and plant protection.
Indirect actions involve growth-promoting substances, antibiotics or
siderophores released by both plants and bacteria. Positive effects are
expected from the inoculation of roots by selected microorganisms
(suppliers of growth-promoting substances, nitrogen fixation, acquisition of phosphorus etc.).
Exudates and Enzymes Released
Root exudation is the major process associated with the rhizosphere. In
addition to the uptake and direct phytodegradation, plants support the
bioremediation by the release of exudate and enzymes that stimulate
both microbial and biochemical activity in the surrounding soil. The
composition of exudates, sites of exudation and various factors affecting the root exudation are summarised by Morel et al. (1999). These
include nutritional stress, excess of metals or the presence of some
microorganisms. Exudates include compounds of low molecular weight,
like phenolics or sugars and amino acids and other compounds, secretions, lysates, and also compounds of high molecular weight, i.e.
Phytoremediation Methods
The diverse approaches include phytostabilisation, phytoextraction
(Morel et al. 1999; Nedelkoska and Doran 2002), direct phytodegradation (Chroma et al. 2003), rhizofiltration (US EPA 1998, Meagher 2000),
phytovolatilisation (Terry et al. 2000), phytoremoval of air pollutants,
hydraulic control (US EPA 1998), formation of artificial wetlands and
T. Macek et al.
lagoon systems, co-operation with microorganisms in the process of
rhizoremediation, and development of plants tailored for specific phytoremediation needs using genetic engineering (Cunningham et al.
1995; Macek et al. 2002a). A detailed account ofvarious phytoremediation methods is provided in Chapter 15 (Sect. 3).
Artificial Wetlands
The use of wetlands for the treatment of contaminated water is well
established and cost-effective. The most commonly used plant species
in treating wetlands are the monocotyledonous species common reed
(Phragmites australis), cattail (Scirpus lacustris) and sweetgrass (Glyceria fluitans). Many such surface flow wetland systems exist around the
world, in most cases treating wastewater for nitrogen, phosphate and
biological oxygen demand (BOD) originating from municipal works,
farming and small industries. The effect of wetland plants on the biogeochemistry of metals beyond the rhizosphere has been addressed by
Wright and Otte (1999). The majority of treated wetlands rely on a combination of moderately reducing substrate conditions and the oxidising
conditions surrounding plant roots (Hammer 1989; Vymazal et al.
1998). For several types of wastewater, wetlands with strongly reducing
substrates facilitate the reduction in sulfate.
O'Sullivan et al. (1999) discussed the use of wetlands for the rehabilitation of metal-containing mine wastes. However, knowledge is lacking
about the fundamental processes in such systems and especially metal
fixation rates are not available. An interesting technological approach is
that of ponds with a floating plant cover. The plant cover reduces wave
formation and, as a consequence, oxygen input into the water body is
restricted. Furthermore, the plants, as in surface flow systems, play an
important role as a carbon source (decaying plant matter and root exudates) for microbial, anaerobic processes.
Perspectives Regarding Plants Used for Detoxification
in Chemical Weapon Demilitarisation
There are possibilities of the exploitation of plants in chemical weapon
demilitarisation (CWD). Macek et al. (1998) proposed that plants can
be used for the detoxification of some metal-containing chemical
warfare compounds, and also for the detoxification of some products
formed during the breakdown or neutralisation of chemical warfare
agents in demilitarisation processes. Problems are arising from incom-
Phytoremediation of Metals and Inorganic Pollutants
plete reactions, from products themselves and also from contaminated
soil caused by the leaking of corroded munitions, which presents a
specific problem. Plants are often expected to be used only for the final
polishing of sites decontaminated in other ways, but it was shown that
plants and their enzymes might play an active role in the degradation
of many toxic compounds directly as part of the technological setup
(Macek et al. 1998, 1999). Some chemical war agents like Adamsite
(diphenylaminechloroarsine) and Clark (Clark 1, diphenylchloroarsine,
and Clark 2, diphenylcyanoarsine) contain a high percentage of arsenic,
which requires special treatment. It is known that plants are able to take
up and concentrate a variety of metals including arsenic (Tlustos et al.
1997, 2002). The possibility to remove products of pyrite breakdown
from soil has been studied recently by Zakharova et al. (2000). The
sulfur-containing products were effectively taken up from contaminated
soil by plants.
In Vitro Plant Cultures in the Phytoremediation of Metals
Most of the experiments used to establish phytoremediation techniques
were carried out with normal soil-grown plants or hydroponics. The
plant cultures used were homogeneous, non-differentiated callus and
suspension cultures, differentiated embryogenic cultures, shooty
teratomas, and hairy root cultures obtained by genetic transformation
with Ri plasmid from Agrobacterium rhizogenes or with Ti (tumourinducing) plasmid from A. tumefaciens. The cultures can be grown
under standard laboratory conditions. Major advantages of in vitro
plant cultures are their rapid growth characteristics and the need for
less analytical input and expense as compared to the normal soil-grown
(allus and (eil Suspension (ultures
The concept of using cells as a model system for phytoremediation is
not new; in vitro plant-cultivated cells have been used in studies of pesticide resistance and metabolism (Wimmer et al. 1987; Rarms and
Kottutz 1990) for many years. Callus and cell suspension cultures have
proven to be very useful tools, but the results obtained under in vitro
conditions need to be carefully evaluated, especially in the case of nondifferentiated callus and suspension cultures, because the variability
within cultures of the same species may be high (Macek 1989; Mackova
et al. 1997a). Many callus cultures have been used in work with heavy
T. Macek et al.
metals, e.g. for selection and biochemical characterisation of zinc- and
manganese-adapted lines in Brassica sp. (Rout et al. 1999). Metal (Hg,
Cu, Cd, Zn) stress response and tolerance in connection with heat stress
proteins were studied using ceH cultures of Silene vulgaris and Lycopersicon peruvianum by WoHgiehn and Neumann (1999). The transformation activity depends to a great extent on the level of morphological
and biochemical differentiation of the strain or clone tested. A comparison of various cultures shows the trend to predict the usefulness of
a given species and, independently, to evaluate the "negative" clones
where some enzyme with the capability to detoxify or transform may
be missing.
Hairy Root Cultures
The so-caHed hairy root cultures are formed by the genetic transformation of a single plant ceH by the soil bacterium Agrobacterium rhizogenes. Due to their fast growth, unlimited propagation in culture
media, genetic and biochemical stability and growth in hormone-free
media, these tissues proved to be very good model systems for plant
metabolism and physiology (Doran 1997; Mackova et al. 1997b; Gleba
et al. 1999; Pletsch et al. 1999; Shanks and Morgan 1999). A. rhizogenestransformed plant roots exhibit nearly aH features of normal plant roots
and grow rapidly under defined conditions in vitro, thus aHowing us
to distinguish between plant metabolism and the effect of the complex interaction between plants and microbial communities in the
rhizosphere during the phytoremediation of metals and xenobiotic
Interesting results for metal phytoremediation have been obtained
using hairy root cultures derived from hyper-accumulator plants. A
range of Agrobacterium rhizogenes-transformed hairy roots have
been tested and compared with hairy roots of closely related non-hyperaccumulator plant species for metal accumulation. The exploitation of
hairy root cultures in phytoremediation studies has been reported for
cadmium uptake by Macek et al. (1994, 1997a, b). Soudek et al. (1998)
compared the uptake of different metals by using horseradish hairy
roots. Nedelkoska and Doran (2002) used long-term hairy root cultures
to demonstrate that in Alyssum species, hyper-accumulation does not
necessarily depend on the presence of shoots or root -shoot transfer.
A choice of available unique hairy root cultures of different hyperaccumulating species, and their non-hyper-accumulator relatives, allows
the important comparison and selection of traits responsible for high
metal accumulation.
Phytoremediation of Metals and Inorganic Pollutants
Breeding and Genetic Engineering
Enhancement of metabolie abilities of plants can be achieved by traditional breeding, protoplast fusion and direct insertion of novel genes.
The methods of genetic engineering are widely used for the improvement of different crop plants. A similar approach is expected to largely
improve the abilities of plants in the field of environmental detoxification (Macek et al. 1996,2000).
There have been many attempts to breed willow, poplar and other
plants with properties useful for phytoremediation. The aim is the formation of plants combining the high ability to accumulate, detoxify or
degrade xenobiotics and pollutants, with the resistance towards the
toxic compounds present and with suitable agronomie characteristics
(Macek et al. 2002a). Transgenic plant technology showed many obvious
advantages over conventional plant breeding approaches for crop
improvement. Interesting papers on genetically engineered plants, suitable for metal accumulation, have recently appeared, indicating the
growing interest in the application of biotechnology to the phytoremediation of inorganic compounds (Clemens et al. 2002; Pilon-Smith and
Pilon 2002).
It is possible to modify and improve normal plant mechanisms, thus
improving overall plant yields applicable to the remediation process.
Improved bioavailability of metals, caused by changes in the exudate,
increased excretion of organie acids, or co operation with rhizospheric
microorganisms, might be itself a target of genetie engineering. Alternatively, the genetic engineering approach may be to increase the other
aspects of performance of transgenie plants. Our very early screening
of different plant species for the production of phytosiderophores
(Rudolph et al. 1985), as shown by the example of nicotianamine, indicated that their presence is ubiquitous in high er plants. Their formation
in plants might be increased or, alternatively, such compounds can be
produced by bacteria.
Methods of Preparation of Transgenic Plants
The main vectors used for routine plant genetic transformations are different plasmids derived from agrobacterial plasmids Ti or Ri. Protoplasts are useful tools for gene manipulations, because under specific
conditions fusions can be achieved between cells of closely related
species, as well as between very different plant species, allowing the for-
T. Macek et al.
mation of somatic hybrids. Other approaches inc1ude ballistic methods,
allowing the transfer of genes into cells and intact plants. These
methods are based on the use of accelerated heavy metal partic1es (gold
or wolfram) covered by genetic material (Maliga 1990). The relatively
new methods are used for species where the use of Agrobacterium failed.
Such approaches can be used with success also in other applications,
inc1uding the direct transformation of organelle genome or the fast
introduction of genetic information into cells without tissue damage,
e.g. into growing meristems (Old and Primrose 1994).
Phytoremediation of a Mercury-(ontaminated Environment
Concerning the treatment of heavy metal-contaminated fields, an interesting approach is the phytoremediation of mercury or methylmercury
by genetically modified plants. The bacterial genes merA, which
encodes bacterial mercuric reductase (reducing mercury ions into
volatile metaI), and merB, a bacterial organomercuriallyase, have been
used for this purpose. Separately or together, both genes were c10ned
into Arabidopsis or tobacco by Heaton et al. (1998) and proved to function well in soil. In this way, a mercury-resistant transgenic plant was
produced that volatilizes mercury into the atmosphere. Rugh et al.
(1998) described the transformation of poplar proembryogenic masses
by microprojectile bombardment with modified merA constructs,
obtaining plantlets releasing elemental mercury at high rates. The question arises, whether volatilisation is an acceptable solution in this case,
but the combined use of both enzymes solves the problem. During the
preparation of such plants, the expression of merA from bacterial transposon Tn21 in plants was the first step (Bizily et al. 2000). Rugh et al.
(1996) constructed a mutant merA with a modified coding sequence
(merA9) and introduced it into Arabidopsis thaliana plants. The seeds
germinated and seedlings were able to grow in media containing up to
100,lLM of mercury (HgC12 ). Transgenic plants released between two to
three times more elemental mercury HgO than the control plants. The
plants proved to be resistant to a toxic concentration of Au3+.
To test the efficacy of using transgenic plants for the phytoremediation of Hg-contaminated soil, the yellow poplar was chosen (Liriodendron tulipifera) because of its promising biological and structural
characteristics (Rugh et al. 1998). It was proved that transgenic poplar
released ten times more elemental HgO in comparison with the control
plant. Although this system was not tested under field conditions, the
results c1early showed that genetic engineering is able to increase the
ability of plants to remove undesired heavy metal compounds from the
Phytoremediation of Metals and Inorganic Pollutants
soil. However, as the release of mercury into air through volatilisation
is for ecological reasons undesirable, further plants were prepared that
are able to accumulate the mercury instead of releasing it (Bizily et al.
Volatilisation of Selenium
Volatilisation of selenium was the aim of research in California; this
time with an expected positive effect of evaporated selenium (Berken et
al. 2002). This is only one of the aspects discussed by Terry et al. (2000)
in their review on selenium in higher plants. It was also found that rhizosphere bacteria enhanced the accumulation of Se and Hg in wetland
plants. Montes-Bayon et al. (2002) studied Se speciation in wild-type
and genetically modified Se-accumulating plants with HPLC separation.
Increased Accumulation of Heavy Metals
Among other natural mechanisms that allow plants to grow in heavy
metal-contaminated environments, many studies have dealt recently
with the phytochelatins. It is a general feature within plants to complex
heavy met als with phytochelatins. These peptides are formed from glutathione by phytochelatin synthase, when plants are challenged with Cd
or other heavy metals. Much effort has been devoted to describing the
biochemical route to their formation (Chasaigne et al. 2001; Oven et al.
Transgenic plants, prepared with the increased formation of glutathione synthase or phytochelatin synthase, exhibited improved
Cd accumulation (Kärenlampi et al. 2000). Transgenic plants bearing
foreign genes for proteins transporting metal across membran es have
also been prepared, as summarised in arecent review (Krämer and
Chardonay 2001). Although plants introduced with genes co ding different types of metallothionein (mammalian, yeast, insect and human)
showed increased resistance to some heavy metals (Macek et al. 1996,
1997b; Kärenlampi et al. 2000), increased accumulation could not be
achieved (Liu et al. 2000). The process of cadmium partitioning in
transgenic tobacco plants, expressing a mammalian metallothionein
gene, is described in detail by Dorlhac de Borne et al. (1998).
In order to increase the metal-binding capacity, the possibility of
introducing an additional high affinity metal-binding domain into the
target protein was studied (Macek et al. 1996). In this way, by combin-
T. Macek et al.
ing the gene CUP 1 for yeast metallothionein from Saccharomyces cerevisiae and the gene for the histidine anchor with high affinity to heavy
metals from the commercial plasmid pTrcHis (Invitrogene), a construct
was prepared and introduced by A. tumefaciens into tobacco (Macek
et al. 1997b, 2002b). The comparison of genetically modified Hnes of
tobacco with control plants showed an increase in resistance and good
growth in cadmium-contaminated soil, together with increased accumulation of cadmium during cultivation in hydropony, in sand with
cadmium and in real contaminated soil (190% of the control; Macek et
al. 2002b). Before transformation of the plants, the different constructs
were tested in E. coli in aseries of growth curves and exhibited improved
resistance and a substantial increase in Cd accumulation (Macek et al.
Other Approaches to Improve Phytoremediation Processes
Symbiotic Bacteria
The finding of a plant growth-promoting bacterium that decreases
nickel toxicity in seedlings represents an interesting example. Burd
et al. (1998) described Kluyvera ascorbata SUD165, which produced
a siderophore and displayed 1-aminocyclopropane-1-carboxylic acid
deaminase activity. The presence of the bacterium did not re du ce the
nickel uptake by seedlings in contaminated soil and probably promoted
plant growth by lowering the level of the plant stress hormone ethylene
that is induced by nickel.
Genetically modified symbiotic bacteria represent a further development of mutual co-operation between plants and bacteria. Here, we
should mention the construction of a novel system, based on symbiosis between leguminous plants and genetically engineered rhizobia
(Sriprang et al. 2002). The authors introduced the gene for human metallothionein MTL4 into a Rhizobium strain forming nodules on Astragalus plant roots. The symbionts increased Cd accumulation in nodules
up to two-fold, but no significant increase in Cd accumulation was noted
in the aboveground part.
Studies with genetically modified bacteria illustrate other approaches
to the genetic manipulation of the properties of phytoremediation
systems, e.g. growth and the promotion of better biomass formation.
Glick and coworkers (Saleh and Glick 2001; Ma et al. 2002) manipulated
the plant growth-promoting bacterium Enterobacter cloacae, using
Phytoremediation of Metals and Inorganic Pollutants
genes from Pseudomonas fluorescens, in order to facilitate plant growth.
Factors tested included indole acetic acid production, antibiotic production, 1-aminocyclopropane-1-carboxylic acid deaminase activity
(decreases root growth inhibiting ethylene production) and siderophore
production. Overproduction of the genes in bacteria, inoculated into
the rhizosphere of canola, resulted in significant increases in root and
shoot elongation, which is beneficial in phytoremediation. Grichko et
al. (2000) described the increased ability of transgenic plants expressing the bacterial enzyme ACC deaminase to accumulate Cd, Co, Cu, Ni,
Pb and Zn.
Mycorrhizal Symbiosis
Arbuscular mycorrhizal fungi (AMF) represent another important
biosystem influencing the behaviour of plants in contaminated soil.
Over 80% of vascular plants can establish symbiotic associations with
AMF that have an important effect on nutrient uptake by plants (e.g.
phosphorus), especially during starvation stress (Jolicoeur et al. 2002).
Establishing symbiosis of plants and AMF in contaminated soil is
expected to support plant growth and survival rate. However, it is still
an open question as to whether root colonisation will lead to increased
accumulation of trace elements.
One of the approaches to evaluate the effect of mycorrhizal fungi on
the phytoremediation process, in the case of cadmium accumulation, is
to inoculate transgenic tobacco, characterised by increased Cd uptake
capability, with mycorrhizal fungus (Vosatka and Janouskova, pers.
comm.). Such pot experiments were performed in sand with nutrients and Cd, and in real Cd-contaminated soil sterilised by gammairradiation, using control tobacco and transgenic tobacco, with and
without colonisation by Glomus intraradices. While phosphorus uptake
was much increased in the presence of root colonisation by G.
intraradices, both in transgenic and non-transgenic tobacco, Cd uptake
and accumulation in aboveground parts were suppressed by the presence of AMF G. intraradices in this model system. The potential uses of
mycorrhizal fungi, and the stimulatory effects of plant and root compounds on degrading bacteria, have been discussed (Donnelly et al.
1994; Fletcher et al. 1995; Gleba et al. 1999).
T. Macek et al.
Role of Secondary Plant Metabolites in Phytoremediation
The contents of root exudate (lactate, acetate, oxalate, succinate,
fumarate, malate, citrate, iso citrate, aconitate, sugars, amino acids) and
compounds released from dying and decaying roots represent a vast
array of molecules that might be involved directly in the metal uptake
and stimulation of bacterial growth (Fletcher et al. 1995; Singer et al.
2003). Biosynthesis of such compounds could also be a suitable target
for genetic manipulation to benefit phytoremediation of metals (Gleba
et al. 1999). Kos and Lestan (2003) have discussed the induced phytoextraction/soil washing of lead using biodegradable chelate and permeable barriers.
Not all waste types and site conditions are comparable and each site,
therefore, must be individually investigated and tested. Engineering and
scientific judgement must be used to determine whether a technology
is appropriate for a site, e.g. United States Environment Protection
Agency states (US EPA 1998). Researchers are finding that the use of
trees rather than smaller plants allows them to treat deeper contamination because tree roots penetrate more deeply into the ground. Very
deep-Iying contaminated groundwater may be treated by pumping the
water out of the ground and using plants to treat the recovered contaminated water.
Specifically, two subsets of phytoremediation are nearing commercialisation. The first one is phytoextraction, in which high biomass
metal-accumulating plants and appropriate soil amendments are used
to transport and concentrate metals from the soil into the harvestable
part of roots and aboveground shoots, which are harvested with conventional agricultural methods. The second one is rhizofiltration, in
which plant roots grown in water absorb, concentrate and precipitate
toxic metals and organics from polluted effiuents. In both cases, the cost
of the remediation of organic contaminants and heavy metals can be
expected to be at the lower and high er end, respectively, of the expected
price ranges (Boyajian and Carreira 1997).
The main advantages of phytoremediation in comparison with classical remediation methods are (Salt et al. 1995; Schnoor et al. 1995):
Phytoremediation of Metals and Inorganic Pollutants
less disruptive for the environment
less need for soil disposal sites
better public acceptance
avoids excavation and transportation
allows a much larger scale cleanup
potential versatility to treat a diverse range of hazardous materials
Considering these factors and the much lower cost expected for phytoremediation, it appears that it will facilitate cleanup operations at
much larger scales than are possible by other methods. The process is
relatively inexpensive, using the same equipment and supplies that are
used in agriculture.
Phytoremediation also has a number of inherent technical limitations. The contaminant must be within, or drawn toward, the root zoneß
of plants, which implies that water content, depth and nutrient content,
exposure to the atmospheric, physical and chemical conditions may all
represent limitations. The site must be large enough to make farming
techniques appropriate. There mayaiso be a considerable delay in time
needed for obtaining satisfactory cleanup results between phytoremediation and "dig and dump" techniques. In addition, formation of
vegetation may be limited by extremes of environmental toxicity. Contaminants collected in leaves can be released again to the environment
during litter fall, or accumulated in energy crops. In the case of the
planned use of proper incineration with metal removal from exhaust,
this approach can be considered as beneficial. In some cases, the solubility of contaminants may be increased, resulting in greater environmental damage and/or pollutant migration.
The growing knowledge about the factors that are important in phytoremediation will provide a basis for the genetic modification of plants
directed to improved performance. We trust that molecular biology will
facilitate the tailoring of plants for particular applications (Raskin 1996;
Macek et al. 2000, 2002a, b). These changes will include transforming
plants to produce specific proteins or peptides for binding and transporting contaminants, increasing the quantity and activity of enzymes
(Francova et al. 2001,2003), including those that are transferred to the
rhizosphere and the surrounding soil to improve the performance of
soil bacteria. Plant-fungal interactions also appear to be important for
exploitation in this area, particularly mycorrhizal associations.
Manipulating plant genes that regulate toxic metal uptake or transformation represents cutting-edge research. These studies have focused
so far on basic research, to explain the yet unknown facts, as a longterm investment to improve plants. The likelihood that this approach
will gain public acceptance, and the fact that permits for field testing of
T. Macek et al.
genetically engineered plants now vastly outnumber permits for genetically altered microbes, are causes for moderate optimism concerning
the use of phytoremediation.
Other discussed approaches might increase the efficiency of the
process, and probably the concerted action of several methods will allow
the real potential of transgenic plants to be exploited, e.g. by increasing
the bioavailability of lead using biodegradable chelates (Kos and Lestan
2003). Phytoremediation is a potential method used to reduce environmental contamination that can prevent the toxic compounds entering
the food chain. It can help maintain human health and conserve biological diversity (Macek et al. 2002a) and can also contribute to sustainable development. The pioneering work on phytomining, using
hyper-accumulating plants with good biomass yield, is an especially
promising approach (Brooks et al. 1998).
Acknowledgements. The authors are grateful for the support of their
research by grant 526/02/0293 of the Grant Agency of the Czech Republic, research projects JI9/98:2232500003 and Z4-055-905
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Remediation of Organic Pollutants Through
Natural Attenuation
Serge Delislei and Charles W. Greer 2
Definition of Natural Attenuation
The basic concept of "natural attenuation" is to allow naturally occurring processes to reduce the mass, volume, or concentration of contaminants in the environment, while addressing the potential threat to
human health and the environment.
Over the last decade, remediation by natural attenuation has become
increasingly accepted as an alternative for managing the risks posed
by organic compounds present in soil and groundwater. Although the
biodegradation of some organic contaminants in the subsurface environment, such as gasoline (referred to as BTEX - benzene, toluene,
ethylbenzene and xylenes) and polycyclic aromatic hydrocarbons
(PARs) composed of two rings or more, is well documented, further
research is required in several areas to investigate the bio degradation
pathways of recalcitrant organic compounds, the effect of mixed contaminants on bio degradation potential, and to assess microbial interactions during the biodegradation process.
Natural attenuation is not a "no action" alternative. Application of
natural attenuation requires analytical data showing that site conditions
(geochemical, hydrogeological, geological, ete.) are suitable, historical
trends showing plume evolution over time, and microbiological data
which show evidence of bio degradation processes. The use of natural
attenuation as aremedial approach is often referred to as monitored
natural attenuation. The US Environmental Protection Agency (EPA
1997) defines natural attenuation and monitored natural attenuation as
1 Biotechnology Research Institute, National Research Council of Canada, 6100 Royalmount Avenue, Montreal, Quebec, Canada H4P 2R2
2 Biotechnology Research Institute, National Research Council of Canada, 6100 Royalmount Avenue, Montreal, Quebec, Canada H4P 2R2, e-mail: [email protected], Tel: +1-514-4966182, Fax: +1-514-4966265
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
S. DelisIe et al.
1. Natural attenuation: Naturally occurring processes in soil and
groundwater environments that act without human intervention to
reduce the mass, toxicity, mobility, volume or concentration of contaminants in these media. These in situ processes include biodegradation, dispersion, dilution, sorption, volatilization, and chemical
or biological stabilization, transformation, or destruction of
2. Monitored natural attenuation: The reliance on natural attenuation
processes (within the context of a carefully controlled and monitored
site cleanup approach) to achieve site-specific remedial objectives
within a time frame that is reasonable compared to that offered by
other more active methods.
The processes involved in natural attenuation can be divided into two
major categories; non-destructive and destructive. The non-destructive
mechanisms result only in the reduction of the contaminant concentration and include hydrodynamic dispersion (mechanical dispersion
and diffusion), sorption, dilution, volatilization and chemical stabilization. The destructive mechanisms result in the reduction of the
contaminant mass and volume. Destructive mechanisms include
bio degradation, and abiotic oxidation and hydrolysis.
The term natural attenuation refers to all processes involved in the
reduction of pollutant mass, volume and concentration. The biodegradation process is referred to as "intrinsic bioremediation".
Principal advantages of monitored natural attenuation include:
• Contaminants can ultimately be mineralized to form carbon dioxide
and water or be transformed into less toxic compounds (ethene,
• Relatively less intrusive;
• Lower overall remediation costs than active re medial technologies;
• Potential for application to all, or part, of a given site;
• Utilizes inherent natural processes;
• Used in conjunction with or as a follow up to, other active remedial
• May reduce the use of energy, and creation of emissions from more
active technologies.
Potential limitations to the use of monitored natural attenuation
• Longer time frame may be required to achieve remediation objectives;
• Site characterization will normally be more complex and costly;
• Toxicity of transformation products (e.g. vinyl chloride) may increase
the risk;
Remediation of Organic Pollutants Through Natural Attenuation
• Potential for contaminant migration;
• Long term monitoring plan is usuaHy necessary;
• Possible renewed mobility of previously stabilized contaminants (e.g.
Monitored Natural Attenuation (MNA)
The design of a groundwater monitoring program involves installing
monitoring weHs upgradient of the potential receptors and determining the frequency and duration of the subsequent monitoring. Methodology should also be developed to obtain sufficient representative
sampies to analyze data on a statistical basis. A long-term monitoring
plan should consider downgradient receptors and toxic breakdown
products resulting from degradation (Wiedemeier et al. 1995). The
objectives of an MNA program will be to determine the plume status
(shrinking, stable, expanding), to detect new release of contaminants
and identify changes in geochemical conditions that could infiuence
natural attenuation (Wiedemeier and Haas 1999; UK Environmental
Agency 2000).
Evaluating Natural Attenuation
The natural attenuation of organic compounds dissolved in groundwater can be assessed by using several lines of converging evidence
(Wiedemeier et al. 1999):
1. Historical trends of data demonstrating plume stability and con-
taminant loss over time;
2. Geochemical data indicating that active biodegradation is occurring
(consumption of electron donors and acceptors, decreased concentration of parent compound, and production of metabolites or end
3. Microbiological data supporting biodegradation.
Abiotic Processes of Natural Attenuation
The abiotic mechanisms that affect the transport of organic contaminants dissolved in groundwater can be divided into two groups:
Fluid mixing due to groundwater
movement and aquifer
heterogen ei ti es
Spreading and dilution of
contaminant due to molecular
Reaction between aquifer matrix
and solute whereby contaminants
become sorbed to organic carbon
or day minerals
Movement of water across the
water table into the saturated zone
Volatilization of contaminants
dissolved in groundwater into
the vapour phase (soil gas)
Chemical transformations that
degrade contaminants without
microbial facilitation, such as
hydro lysis and hydrogenolysis
Dependent on contaminant properties
and groundwater geochemistry
Dependent on compound vapour
pressure and Henry's law constant
Dependent on aquifer matrix properties
(organic carbon and day mineral content,
bulk density, specific surface area and
porosity) and contaminant properties
(solubility, hydrophobicity, octanolwater partitioning coefficient)
Dependent on aquifer matrix properties,
depth to groundwater, surface water
interactions, and dimate
Dependent on aquifer properties,
mainly hydraulic conductivity, effective
porosity and hydraulic gradient.
Independent of contaminant properties
Dependent on aquifer properties and
scale of observation. Independent of
contaminant properties
Dependent on contaminant properties
and concentration gradients. Described
by Fick's laws
Movement of solute by bulk
groundwater movement
Can result in partial or complete
degradation of contaminants. Rate
typically much slower than for
bio degradation
Causes dilution of the contaminant plume
and may replenish electron acceptor
concentrations, especially dissolved
Removes contaminants from groundwater
and transfers them to the soi! gas phase
Causes longitudinal, transverse and
vertical spreading of the plume. Reduces
solute concentration
Diffusion of contaminant from areas of
relatively high concentration to areas of
relatively low concentration. Generally
unimportant relative to dispersion at
most groundwater fiow velocities
Tends to reduce apparent solute
transport velocity and remove solutes
from the groundwater via sorption to the
aquifer matrix
Main mechanism driving contaminant
movement in the subsurface
Table 1. Summary of important abiotic processes in natural attenuation. (Adapted from Wie dem eier et al. 1999)
Remediation of Organic Pollutants Through Natural Attenuation
physical (dispersion, dilution, diffusion) and geochemical processes
(sorption, volatilization, hydrolysis and chemical stabilization). Table 1
summarizes the main processes affecting contaminant transport.
These processes are weIl documented (Domenico and Schwartz 1990;
Wiedemeier et al. 1999).
Overview of Intrinsic Bioremediation
Microorganisms obtain energy for growth and activity by oxidizing,
reducing or fermenting substrates, which can be primary substrates
(common organic compounds) or organic pollutants. Oxidation and
reduction reactions involve the transfer of electrons from electron
donors to electron acceptors. In fermentation reactions the contaminant
or carbon source acts as both an electron donor and electron acceptor;
this generally occurs when external electron acceptors (oxygen, nitrate,
etc.) have been depleted. Cometabolism, which can be a significant
process in the degradation of recalcitrant organics, is the transformation of a substrate that does not result in any benefit to the microorganism: the substrate is neither a carbon nor an energy source.
The energy produced by the transfer of electrons from an electron
donor to an electron acceptor is termed Gibbs free energy C1G), which
represents the energy difference between the initial and final state of the
reaction under conditions of standard temperature and pressure. A negative value for L1G indicates that the reaction is exothermic, or energy
producing, whereas a positive value for L1G indicates an endothermic,
or energy requiring reaction.
As electron acceptors are consumed, groundwater be comes more
reduced and the oxidation-reduction, or redox potential (ORP) of the
water decreases, favoring the activity of facultative and strictly anaerobic bacteria. Under an aerobic conditions, alternate electron acceptors
are used in the following order: nitrate, manganese (Mg4+), ferric iron
(Fe3+), sulfate and finally carbon dioxide. This sequence of events is
depicted in a contaminated groundwater plume in Fig. 1. Each successive reaction in the process yields less energy than the previous one
(Table 2). Facultative bacteria, such as the denitrifiers, will consume
oxygen first, before switching to nitrate as an electron acceptor. It is
important to note that an aquifer is a dynamic, open system, and that
various oxidation-reduction reactions occur simultaneously and not
in a strictly ordered sequence. Groundwater geochemical data from
the site is presented in the form of an isopleth map, which shows the
concentration contours of the different electron acceptors in the
S. Delisle et al.
ground surface
vadose zone
saturated zone
Fig. 1. Oxidation - reduction zones identified in a contaminant plume
Table 2. Energy yielding reactions catalyzed by bacteria (presented in order of
decreasing free energy yield)
Metabolie process
Carbon source
Electron acceptor
LlGo (kcal/mol)
Nitrate reduction
Aerobic respiration
Manganese reduction
Ferric iron reduction
Sulfate reduction
Organic matter
Organic matter
Organic matter
Organic matter
Organic matter
Organic matter
N0 3O2
Mn H
groundwater plume. These data are used to quantify any changes in concentrations of electron acceptors or their products, spatially and temporally, to identify whether natural attenuation is occurring at the site.
Electron Acceptors
The use of different electron acceptors by bacteria during the degradation of petroleum hydrocarbons is summarized below.
Dissolved oxygen: This is the most thermodynamically favorable electron acceptor for the biodegradation of aliphatic, aromatic and polycydic aromatic hydrocarbons. The dissolved oxygen concentration will
be at its maximum at the periphery of a dissolved hydrocarbon plume
and absent dose to the source of the contamination. Under normal,
non-contaminated conditions, dissolved oxygen concentrations in
groundwater can range between 0 and 12 mg/I.
Remediation of Organic Pollutants Through Natural Attenuation
Nitrate: Denitrification results in the transformation of nitrate into
dinitrogen in a multi-step process that produces nitrite as one of the
intermediates. The nitrate concentration is maximal at the periphery of
the contaminant plume, and this represents the average background
concentration in the aquifer. In the active denitrification zone, the
nitrate concentration is minimal, and the difference between the background nitrate concentration and the concentration near the source of
the contamination provides an indication of the extent of contaminant
degradation linked to dentrification. The nitrate concentration in noncontaminated groundwater is typically less than 5 mg/l.
Manganese: The degradation of petroleum hydrocarbons associated
with manganese reduction under anaerobic conditions pro duces Mn 2+
concentrations that are measurably high er than the aquifer background
levels. The Mn2+ concentrations are based on measurements and not on
the available Mn02 concentration in the aquifer. It is therefore possible
that the bio degradation capacity associated with manganese reduction
is greater than measured values. Manganese concentrations in noncontaminated groundwater are typically less than 1 mg/l.
lron: The degradation of organic compounds linked to iron reduction
results in the transformation of insoluble ferric iron (Fe 3+) into soluble
ferrous iron (Fe2+). To calculate the iron-reducing capacity of the
aquifer, the concentration of ferrous iron is measured, since ferric iron
is asolid and its total availability is difficult to determine. The concentration of ferrous iron in the hydrocarbon contaminant plume is compared to the background concentration, which in non-contaminated
groundwater varies between 1 and 10 mg/l. Under strongly reducing
conditions, where sulfate reduction and methanogenesis are present,
ferrous iron can combine with hydrogen sulfide or with carbon dioxide
to precipitate as pyrite (FeS2) or siderite (FeC0 3 ). Because of these
processes, it is possible that bio degradation under iron-reducing conditions is underestimated in the contaminant plume.
Sulfate: Sulfate reduction coupled to the an aerobic degradation of
petroleum hydrocarbons results in the transformation of sulfate into
hydrogen sulfide (H 2S). Sulfate-reducing bacteria are sensitive to the
physico-chemical conditions in the aquifer, and the pH optimum for
the reduction of sulfate is 7. The concentration of sulfate in noncontaminated groundwater is typically less than 500 mg/l. Natural
sources of sulfate in groundwater are pyrite, gypsum and anhydrite, and
considerable quantities of sulfate are produced by combustion of, for
example, coal.
Carbon dioxide: Under very low reducing conditions, the consumption of organic matter (contaminants or nutrients) by anaerobic bacteria produces methane. The biodegradation capacity of the aquifer due
S. DelisIe et al.
to methanogenesis is based on the concentration of methane in relation
to its background concentration.
Intrinsic Bioremediation of Petroleum Hydrocarbons
The biochemical pathways for the aerobic biodegradation of various
petroleum hydrocarbons, including PAHs composed of up to fOUf rings
is well documented, and is covered elsewhere in this volume. Evidence
is now accumulating that many petroleum hydrocarbons can be biodegraded using a variety of electron acceptors other than molecular
oxygen (Spormann and Widdel 2000) and it is evidence of the consumption of these alternate electron acceptors that is required to
demonstrate the intrinsic bioremediation of petroleum hydrocarbons
and other environmental pollutants.
Intrinsic Bioremediation of Chlorinated Solvents
Chlorinated solvents (i.e. chloroethenes and chloroethanes) are widespread groundwater pollutants (Vogel 1994), which can be biodegraded
under both aerobic and anaerobic conditions (McCarty and Semprini
1994). Anaerobically, chlorinated ethenes and ethanes are transformed
by reductive dehalogenation, or dehalorespiration using hydrogen as an
electron donor (McCarty 1997). During reductive dehalogenation, perchloroethene (PCE) is transformed to trichloroethene (TCE) then to
cis-dichloroethene (cDCE), which is subsequently transformed into
vinyl chloride (VC). Dehalococcoides ethenogenes is the only characterized bacterium that is able to dehalogenate VC to ethene (Maymo-Gatell
et al. 1997). Chloroethanes, such as 1,1,1-trichloroethane (TCA), can be
degraded under a variety of anaerobic conditions including nitrate- and
sulfate-reducing and methanogenic conditions (McCarty and Semprini
1994). It has been shown that dechlorination is faster for highly chlorinated compounds than for compounds that are less chlorinated (Vogel
and McCarty 1987; Bouwer 1994).
Remediation of Organie Pollutants Through Natural Attenuation
(ase Studies of Monitored Natural Attenuation
Intrinsic Bioremediation of Fuel Hydrocarbons at a Former (anadian
Forces Base, (hatham, New Brunswick - (ase 1
The former Canadian Forces Base (CFB) at Chatham is located in northeast New Brunswick, south of the city of Miramichi. The former CFB
is situated on a plateau sloping east-southeast to the Napon River and
north toward the Miramichi River. Buildings in the study area, in the
northeast of the base, were used as a maintenance and fuel service
building (building 66) and a hangar (hangar 1) (Fig. 2). A former underground storage tank (UST) storing gasoline, and its associated pump
island, were removed at some unknown time prior to site characterization. A production weH (B90-BW) is located less than 80m northeast of
hangar 1. This weH was used to supply water needs for the base, and
since 1994, it has been used as a potable water source for the city of
Former Structures
Fig. 2. Site map and loeation of the produetion weil
.s m
S. m
iIt __IIIt:==j
s. Delisle et al.
A site characterization study performed in 1996 revealed a plume of
dissolved fuel hydrocarbons in the groundwater. The plurne, which consists largely of BTEX (benzene, toluene, ethylbenzenes and xylenes)
extended southeast towards hangar 1.
The objectives of this study were to (1) define the nature and extent
of groundwater contamination at the study site with respect to the New
Brunswick Department of the Environment (NBDOE) guidelines; (2)
demonstrate the occurrence of intrinsic remediation in the shallow
aquifer by collecting geochemical data; (3) perform laboratory studies
to characterize the indigenous microftora and its capacity to degrade
hydrocarbons and (4) evaluate the possibility that the impacted shallow
aquifer was in contact with the intermediate aquifer in the vicinity of
the production weIl, B90-BW.
Geology and Hydrogeology
The overburden on the base is characterized by a thin mantle of glacial
till, which can be up to 3 m thick, and consists of sandy, glacially
reworked, marine sediments. The bedrock below the overburden is
Pennsylvanian age (320 to 280 million years B.P.) sandstone with
interbedded layers of shale, and lesser amounts of siltstone and
There are also minor occurrences of coal, conglomerate and day
units. The bedrock is known to be relatively transmissive and is the local
aquifer. The mechanism of deposition resulted in a geologic unit that is
characterized by complex layered vertical stratigraphy and horizontal
facies changes. Aquifers within the unit te nd to be confined or semiconfined due to the adjacent siltstone.
The hydrogeological setting at the base is characterized by three
aquifers: a shaIlow, unconfined sandstone aquifer to a depth of 27 m, a
lower, semi-confined aquifer and a deep confined aquifer. The water
level in the shallow aquifer is 3 m below ground surface and the direction of groundwater ftow is topographically controlled. The relative
permeability of the shallow aquifer is controlled by weathering and vertical fracturing. The first aquidude consists mainly of red and gray shale
with an approximate thickness of 15 m. The intermediate aquifer is
located at a depth ranging between 42 and 121 m. This aquifer lies below
the shallow aquifer and is characterized by many layers of sandstone,
ranging in thickness from 5-15 m, separated by shale layers varying
between 3 and 9 m in thickness. The intermediate aquifer is semiconfined to confined, and the water level ranges from 10-33 m, depending on weIl pumping. The screen of the pumping weIl, B90-BW, is located
in the intermediate aquifer. The deep confined aquifer is below 121 m.
Remediation of Organic Pollutants Through Natural Attenuation
Nature and Extent of Contamination
Chemical analysis of the groundwater indicated the presence of three
dissolved hydrocarbon (BTEX and TPH) plumes in the shallow aquifer.
The larger plume probably originated from the former UST and its associated pump island located south of building 66, whereas the other two
plumes, around MW53 and BRI-MW62, probably originated from a spill
or from fuel manipulation in this area. The vertical extent of the dissolved contamination of the largest plume was between 10 and 15 m
below ground surface. The vertical extent of the dissolved contamination in the two smaller plumes was between 5 and 10 m below grade.
Figure 3 is an isopleth map showing the approximate extent of BTEX
contamination in the groundwater at 5, 10 and 15m below ground
surface in 1999. The isopleth map for TPH contamination was very
similar to the BTEX isopleth map. The highest BTEX and TPH concentrations observed in the groundwater during this study were 29,400 and
20,538 jig/l, respectively.
Hydrogeological Test
A hydrogeological test was performed to examine the possibility of a
hydraulic connection between the shallow aquifer and the intermediate
aquifer in the immediate vicinity of production well B90-BW. Following
a 24-h shutdown, the production pump was turned on at the normal
pumping rate used for B90-BW. Water levels in all the surrounding monitoring wells were measured at different times during the shut down
period and 7 h following the restart of the production well pump. The
water level ranged between 2.0 and 3.4 m below ground surface in the
shallow aquifer and 20 m below ground in the intermediate aquifer.
Results of the hydrogeological test are presented in Fig. 4 where the
time-drawdown curves for MW54-L (intermediate aquifer) and MW54
(shallowaquifer) are shown. The time-drawdown curve for MW54-L
indicates a strong response with the production well, B90-BW. The water
level increased from 20.0 to 15.8m below ground surface in 24h when
the pumping well was turned off. The drawdown after the pump was
turned on was 2.5 m. The time-drawdown curve at each of the shallow
aquifer monitoring wells showed no significant fiuctuation (Fig. 4B).
The fiuctuations that were observed are probably due to atmospheric
pressure, a dry period or a previous rainfall event. Based on these
results, there is no evidence of hydraulic connection between the
shallow aquifer and the intermediate aquifer in the vicinity of production weIl B90-BW. The risk of introducing contaminated shallow
S. Delisie et al.
BTEX Isopleth Map for Groundwater in 1999 . 5 m below surface
__ ;r
M on itoring Wells
Dissolved Benzene
Groundwate: Plume
, "
50 ~Q/L App:ox imale Contou: - J
Level 11 NBDOE
Olssolved Benzene
G:oundwale: Plume
2_5 ~ g/L Approximate Conlou
BTEX Isopleth Map for Groundwater in 1999 ·10 m below surface
- +'
SCALE .... /'
Dissolved Benzene
G:oundwate: Plume
,- ,
50 pg/L App:oximate ConlOu:
Level 11 NBOO E
\- ,
Monitoring We lls
Oissolved Benzene
Groundwaler Plume
2.5 pg/L Approximate Coolou
Fig. 3. BTEX isopleth map for groundwater in 1999 at 5, 10, and 15 m below surface
Remediation of Organic Pollutants Through Natural Attenuation
BTEX Isopleth Map for Groundwater in 1999 -15 m below surface
Dissolved Benzene
Groundwaler Plume
50 pg/L Approx imate Conlour , J
Lavalll NBDOE
M oniloring Wells
Dissolved Benzene
Groundwater Plume
2.5 pg/L Approximale Conlou
Fig. 3. Continued
groundwater into the intermediate aquifer by pumping B90-BW is
therefore negligible.
Geochemicallndicators of Biodegradation
The interpretation of the groundwater chemistry data was performed
using the methodology ofWiedemeier et aI. 1995, and the capacity of the
aquifer to dissimilate BTEX was calculated for each biological process.
Aerobic respiration: Dissolved oxygen concentrations in non-impacted groundwater at the site are in the order of 8.9 mg/I. A comparison of the dissolved plume with the isopleth map of dissolved oxygen
concentrations (Fig. 5) shows that the plume has depleted the dissolved
oxygen. This is a strong indication that aerobic bio degradation of
organic contaminants is occurring at the site. In the absence of microbial cell production, the shallow aquifer has the capacity to assimilate
2.8 mgll of total BTEX.
Denitrification: Nitrate concentrations in the non-impacted groundwater at the site are approximately 2.4 mg/I. The highest dissolved BTEX
concentrations are associated with low concentrations of nitrate, which
S. Delisie et al.
A) MW-54-L
time (min)
O,OO r---,---,--,--r---,---,----,---------,
Fig. 4. Time-drawdown plot of
A) MW54·L (intermediate
aquifer) and B) MW54 (shallow
---- .•.. ... . . . . .
4,50 ' - - - - ' - - - ' - - - ' - -- ' - - - ' - - - - ' - - - ' - - - - - '
B) MW-54
time (min)
·0,1 0
~ -{I,OS
:t ·0,04
..... ..........
- - - - - - - . - . - - .j . ...
O,OO ~
0,02 ' - - - ' - - - - ' - - - ' - - - ' - - - - - ' - - ' - - - ' - - - - '
provides strong evidence that processes of denitrification of the contaminants is occurring at the site (Fig. 5). In the absence of microbial
cell production, the shallow aquifer has the capacity to assimilate 0.50
mgll of total BTEX during denitrification.
Manganese reduction: Manganese (IV) concentrations in the nonimpacted groundwater at the site are less than 1 mg/I. The highest manganese (IV) concentration (25.1 mg/I) observed at the site is located in
Fig. 5. Isopleth
map showing
electron acceptor
-, '
Llne 01 Equal Dos.olved
O'ygen Concentr.tron (ppm)
-+-, - ,05 mgil
-----t5 mg ll
10;';911 - - - --
Moniforin g We lls
llne 01 Equa l Nitrale
Coneenlra"on Ippml
Mangan ... (IV)
.- -+-
Monltorjng We lts
L:ne 01 Equa l Mangan ... (l VI
Coneenllal ion Ippm I
s. Delisie et a1.
Fig. 5. Continued
Mon iloring Walls
lme 01 Equa l Ferrous
Iron Concentralion (ppm I
line 01 EQual Sulfate
Concentration (Pilm
., .
the portion of the dissolved plume where the BTEX concentration is at
a maximum. The relationship between the concentration of dissolved
hydrocarbons and the manganese concentration in the plume (Fig. 5)
provides strong evidence that manganese reduction is occurring. In the
absence of microbial ceU production, the shaUow aquifer has the capacity to assimilate 4.4 mg/l of total BTEX during manganese reduction.
Iron reduction: The ferrous iron concentration in the non-impacted
groundwater at the site was be10w the detection limit of 10 mg/I. The
Remediation of Organic Pollutants Through Natural Attenuation
highest measured ferrous iron concentration was 30 mg/l in a weUlocated down gradient of the source. The isopleth map of ferrous iron
concentrations (Fig. 5) shows a good correlation with the dissolved
BTEX plurne. This is an indication that ferric iron is being reduced to
ferrous iron during the degradation of hydrocarbons. In the absence
of microbial ceU production, the aquifer has the capacity to assimilate
1.4 mg/l of BTEX during iron reduction. This calculation is based on the
observed ferrous iron concentration and not on the amount of ferric
hydroxide available in the shallow aquifer.
Sulfate reduction: The highest sulfate concentration in the nonimpacted groundwater was 34.5 mg/I. The isopleth map of sulfate has a
good corre1ation with the extent of the dissolved hydrocarbon plurne.
Sulfate concentrations are below the detection limit where the dissolved
BTEX concentrations are maximal. This is a strong indication that
sulfate reduction is occurring at the site. The shaUow aquifer has the
capacity to assimilate 7.5 mg/l of BTEX compounds during sulfate
Methanogenesis: No c1ear methane concentration trends were apparent and it appears that methane is not being produced at the site.
Groundwater geochemistry: The total alkalinity at the site varied
between 17 and 317 mg/I, which is not sufficient to buffer potential
changes in pH caused by biological activities. The pH ranged from
5.6-7.2, which is within the optimal range for BTEX bio degradation.
The groundwater temperature varied between 10 and 15 oe.
Expressed assimilative capacity: The groundwater data suggest that
natural attenuation of BTEX and TPH is occurring at the site, and
involves the processes of aerobic respiration, denitrification, iron
reduction, manganese reduction and sulfate reduction. Based on the
calculations of these biological processes, the expressed BTEX assimilative capacity of the shaUow aquifer is at least 16,600 pg/1 (Table 3). This
estimate is conservative, since it does not consider microbial biomass
Table 3. Expressed assimilative capacity of the CFB-Chatham aquifer
Biological process (electron acceptor)
Expressed BTEX assimilative
capacity (.ug/l)
Aerobic respiration (oxygen)
Denitrification (nitrate)
Iron reduction (ferric iron)
Manganese reduction (Mn IV)
Sulfate reduction (sulfate)
Expressed assimilative capacity
Highest observed total BTEX concentration
S. Delisie et al.
production, and it does not provide an accurate estimate of the total
available iron. The highest dissolved-phase total BTEX concentration
observed at the site was 29,400 pg/l, therefore the shallow aquifer does
not have enough assimilative capacity to degrade all the dissolved-phase
Laboratory Treatability Study
A laboratory treatability study is frequently used to assess the presence
of hydrocarbon degrading microorganisms in the contaminated ground
water or geologie material, and to determine their potential capacity
to degrade the contaminants under different treatment conditions.
Biodegradation activity can be assessed in microcosms using representative 14C-labeled compounds as the substrates. An advantage of the
microcosm assay is that a variety of different potential treatment conditions can be evaluated, inc1uding the use of supplementary substrates,
or alternate electron acceptors under different atmospheres. In the
present study, laboratory microcosm studies demonstrated a low indigenous bacterial population density, and very poor mineralization rates
for dodecane and benzene in the absence of supplementary carbon
sources and additional electron acceptors. When biostimulation conditions were applied (addition of carbon sources and electron acceptors)
mineralization rates were increased dramatically. The treatability study
demonstrated that hydrocarbon-degrading indigenous microorganisms
were present in low numbers, and the conditions in the aquifer were not
favorable for their activity. The results also demonstrated that their
hydrocarbon degradation activity could be enhanced by biostimulation.
Intrinsic Bioremediation of (hlorinated Solvents at (anadian Force
Base, Trenton, Ontario - (ase 2
The Canadian Force Base (CFB) at Trenton occupies approximately
7 km 2 and is bounded by a railway corridor to the north, White's Road
to the east, the Bay of Quinte to the south and the city of Trenton to the
west. The study site, adjacent to building 151, is relatively flat, and areas
surrounding the building are paved. Building 151 is an active mechanical shop and consists of a slab on grade and steel structure. One 2400-
Remediation of Organic Pollutants Through Natural Attenuation
Conlaur 01 Water
T.b .. Aqu.ter
EIe". on 01 Waler
Table AQuiler
tlow erlltCllOn
B ILDI G 151
--%~._---72 m
Fig. 6. Site map and groundwater elevation contours
I underground storage tank (UST), used to store waste solvents from the
machine shop, located at the northeast corner of the building was
removed in March, 1995. A site map is presented in Fig. 6.
Soil sampies were secured from the walls of the excavation of the UST
in the spring of 1995. Sampies were submitted for BTEX, TPH, PAH
and metal analyses. None of the soil sampies exceeded the Canadian
Council of Ministers of the Environment (CCME) guidelines for commerciallindustrialland. In November 1995, one borehole instrurnented
as a monitoring well (MW13) was installed. Analytical results revealed
the presence of volatile organic compounds (VOC) in the soil and
groundwater at concentrations exceeding the CCME and the Ontario
Ministry of the Environment and Energy (OMEE) criteria. Fourteen
additional boreholes were drilled and instrurnented as monitoring wells
in the vicinity ofbuilding 151 in 1996. The results delineated a dissolved
plume of 1,I,l-trichloroethane (TCA) and 1,l-dichloethane (DCA) in
the groundwater. All soil sampies satisfied applicable remedial criteria.
In 1998, a groundwater collection trench and a groundwater treatment system were installed. The collection trench intersected the dissolved plume perpendicularly, downgradient of the source. The bottom
of the collection trench was located at the junction of the overburden
and bedrock. The groundwater treatment system is comprised of a coa-
S. Delisie et al.
lescing product/water separator, a flow totalizer and an air stripper.
Volatile compounds are stripped from the water and discharged
through an exhaust stack.
The objectives of this study were to:
1. Define the nature and extent of soil and groundwater contamination
at the study sites with respect to the OMEE criteria;
2. Detail the physical flow regime of groundwater within the overburden and bedrock aquifers;
3. Collect geochemical data (i.e. dissolved oxygen, pH, redox potential,
ferrous iron, mangane se, methane) to demonstrate the occurrence of
intrinsic remediation in the aquifers;
4. Evaluate the potential of natural attenuation in groundwater at the
site using field results.
Geology and Hydrogeology
Generally, the overburden consisted of sand fill ranging in thickness
from 0.75-0.9 m, which extended to as much as 2 m below ground dose
to the former UST. Below this filllayer the natural overburden consisted
of a silt deposit with traces of sand having interbedded layers of fine
sand. Under the silt deposit is a gray sandy till with traces of gravel
extending from a depth of approximately 3.4 m to the bedrock, located
between 4.8 and 5.3 m. The bedrock is mainly altered limestone with
embedded layers of shale. Fresh bedrock was located at 7 m below grade.
Bedrock consists of a fossiliferous limestone of the Ottawa Group with
thin shale beds.
The confined aquifer consists of a till material and the upper part of
the bedrock (fractured). The aquifer has a relatively high permeability
as a result of the till material (gravelly sand), weathering, and the
fractured bedrock. Based on water level measurements and hydraulic
tests, the groundwater located inside the gravelly sand and bedrock is
assumed to be part of the same aquifer. The aquifer is overlayed by silt,
which represents an aquidude with low transmissivity.
Static water levels indicated that the depth of the water table varies
between 1.0 and 2.5 m from the ground surface. Water level information
also shows that the groundwater flows downward from the overburden
into the bedrock. Groundwater gradients across the overburdenl
bedrock interface are relatively strong, up to 0.2 mim. As indicated in
Fig. 6, the overall groundwater flow direction was determined to be in
a south-southeast direction.
Remediation of Organic Pollutants Through Natural Attenuation
Table 4. Dissolved concentrations of contaminants and metabolites in the till and
bedrock aquifers
Till material (pg/l)
Bedrock (pg/l)
Biological pathway
Abiotic pathway
Vinyl chloride
Nature and Extent of Contamination
Based on the operational history and remedial activities, the potential
sources of contamination for the site were (1) the former UST used to
store waste solvent; (2) activities inside the building associated with a
degreasing shop, and (3) spills on each side of the door located south of
the former UST.
Soil and groundwater data were compared to the industriallevels of
the OMEE Guidelines for Use at Contaminated Sites in Ontario (1996).
Dense non-aqueous phase liquid (DNAPL) was not detected in the monitoring wells, but the high dissolved concentration of TCA in the till
(215,OOOppb) and bedrock (47,OOOppb) aquifers, suggested the presence
of DNAPL. All soil sampie results showed VOC and TPH concentrations
either below Tables Band D of the OMEE Guidelines (1996) or below
the detection limit of the analytical apparatus. Dissolved TCA, DCA,
vinyl chloride (VC) and chlorobenzene were detected in excess of the
OMEE Guidelines (Table B, 1996) in the till and the upper part of the
bedrock (fractured).
Table 4 summarizes the dissolved concentrations of TCA and
metabolites observed in both aquifers. Figure 7 shows an isopleth map
delineating the distribution of chlorinated ethanes in the till aquifer.
The distribution of chlorinated ethanes in the bedrock followed a
similar profile to the till material.
The literature reports that the biodegradation of TCA occurs under
an aerobic conditions by reductive dehalogenation (de Best et al. 1997;
l,l ,l-TRICHlOROETHANE (119/L)
C) ~~l:~Y::=~:-C;=~:~'fu'
.u01'l40hllO W . . .
100 rn
l ,l-0ICHlOAOETHAN E (119/l)
t.l01I1I0rln9 W,
11. 01 G I
1 ______ _
Fig.7. Distribution of
dissolved chlorinated
ethanes in the till
Remediation of Organic Pollutants Through Natural Attenuation
Fig_ 7. Continued
C) ~~~~~~:~~~o:, 7::~~o;t
u on.olW'l~ w.n.
0 :
___ I
- - - F,,.,e
Chen et al. 1999), and that aerobie co-metabolie degradation of TCA
by methanotrophs was incomplete and limited (Oldenhuis et al. 1989).
TCA is transformed biologieally under anaerobic conditions into 1,1dichloroethane (DCA), which can be further reduced to chloroethane
(CA). CA can be transformed biotically into ethane, however, it is biologically relatively stable, and may be transformed abiotically into
ethanol and chloride.
The abiotic transformation of TCA to either 1,I-dichloroethene
(DCE) or acetic acid has also been reported (Vogel and McCarty 1987;
Gälli and McCarty 1989; Chen et al. 1999). Abiotic transformation of
S. Delisie et al.
TCA occurred in the absence of live cells due to the presence of either
iron complexes or humic-acid-like substances, or of naturally occurring
materials such as vitamin Bl2 and cytochrome F430 , which are released
by dead cells into the environment (de Best et al. 1999). DCE can be
degraded biologically to vinyl chloride (VC) and to ethene under anaerobic conditions. The anaerobic transformation rate of TCA has been
estimated to be 34 to 150 times faster than abiotic transformation
(Klecka et al. 1990). Reductive dechlorination and not abiotic transformation is therefore expected to be the predominant phenomenon under
an aerobic conditions.
One line of evidence that demonstrates the occurrence of biodegradation is that concentrations of DCA detected in groundwater for both
aquifers are about 10 to 20 times the DCE concentration. The elevated
ratio of DCA to DCE strongly suggests that TCA is being reduced by
biological processes. Another line of evidence to determine whether
bio degradation is occurring on the site was to use the site geochemical
data in the bioattenuation screening protocol developed by Wiedemeier
et al. (1996).
Geochemical data were used from at least six monitoring wells
located in the till material and the bedrock dose to the source, in the
dissolved plurne, and from the non-impacted groundwater (upgradient,
down gradient and laterally). Table 5 presents the geochemical screening criteria adapted for TCA as the released contaminant.
Table 6 presents the analytical parameters and weighting for the preliminary screening. Some analytes (volatile fatty acids, ethane, DOC and
alkalinity) were not measured, so no values were assigned. Table 7 summarizes the range of possible scores and gives an interpretation for each
score. The score for both units at building 151 is 16, indicating that adequate evidence exists to support the reductive dechlorination of TCA.
Case 1: Three dissolved plumes of hydrocarbons were identified at the
study site. These plumes were identified as containing primarily gasoline. No contamination higher than level II (NBDOE Guidelines) was
observed at 15 m below ground. Monitoring wells installed between the
dissolved plumes and the production well showed no evidence of dissolved BTEX and TPH in the groundwater. Moreover, a hydrogeological
test showed no evidence that the pumping of the production well B90BW at normal pumping rates infiuenced groundwater in the shallow
Table 5. Analytical parameters and weighting for preliminary site screening. (Adapted
from Wiedemei~r et al. 1996)
Concentration in
contaminated zone
Oxygen a
<0.5 mg/l
0.5 mg!l < DO < 1 mg/l
>1 mg/l
<1 mg/l
Ferrous iron"
>1 mg/l
<20 mg/l
>0.5 mg/l
Redox potential"
>20 mg/l
>2x background
5< pH < 9
5> pH > 9
>2x background
Volatile fatty
>0.1 mg!l
>0.1 mg/l
Vinyl chloride"
"Required analyses.
>0.01 mg/l
>0.1 mg/l
Suppresses the reductive
pathway at higher
May compete with reductive
pathway at higher
Reductive pathway possible
May compete with reductive
pathway at high er
Ultimate reductive daughter
Reductive pathway possible
Reductive pathway likely
Carbon and energy source;
can be natural or
Results from interaction of
carbon dioxide with aquifer
Optimal range
Outside the optimal range
Biochemical process is
Daughter product or organic
Intermediates resulting
from bio degradation of
aromatic compounds
Carbon and energy source,
drives dechlorination
Material release
Material release
Daughter product of 1,1,12
Material release
Daughter product of 1,1,1TCA
Material release
Daughter product of 1,1DCA
Daughter product of
Material release
Daughter product of DCE
S. Delisie et al.
Table 6. Evaluation of the potential for natural attenuation at CFB-Trenton
Redox potential'
Volatile fatty
1,1dichloroethane P
Concentration in
contaminated zone
Concentration in
3.77 mg/l
4.84 mg/l
3.55 mg/l
9.76 mg/l
Not analyzed
Not analyzed
32,000 mg/l
>2x Background
Not analyzed
1.85 mg/l
Not analyzed
Not analyzed
42,000 mg/l
>2x Background
Not analyzed
36.128 mg/l
4.930 mg/l
47 mg/l
2.280 mg/l
2.380 mg/l
0.574 mg/l
0.329 mg/l
Not analyzed
0.020 mg/l
Not analyzed
a Superscript
s denotes source, p denotes plurne.
Table 7. Interpretation of points awarded
o to 5
Inadequate evidence for bio degradation of chlorinated organics
Limited evidence for bio degradation of chlorinated organics
Adequate evidence for bio degradation of chlorinated organics
Strong evidence for bio degradation of chlorinated organics
6 to 15
16 to 21
Remediation of Organic Pollutants Through Natural Attenuation
The indigenous microbial population did not demonstrate significant
mineralization of dodecane or benzene under a variety of tested conditions without bio stimulation. Moreover, the field geochemical indicators showed that the shallow aquifer did not have sufficient assimilative
capacity to degrade the dissolved-phase BTEX. These results suggested
a low potential for intrinsic bioremediation of the site. Based on the laboratory treatability study, biostimulation seemed to be a good remediation strategy for this site.
Case 2: Groundwater at the study site showed TCA, DCE and VC
concentrations above Table B of the OMEE guidelines. The impacted
groundwater fiow seems to be restricted in the source area, probably due
to the low hydraulic conductivity of the silt material. The groundwater
also contains the TCA bio degradation products, DCA, and CA.
The chemical and geochemical data provide good evidence for reductive dechlorination ofTCA. The breakdown products ofTCA are present
in both (till and bedrock) aquifers. Reduced nitrate, elevated ferrous
iron and elevated methane concentrations in the groundwater indicated
that conditions are sufficiently reducing to support the localized reductive dechlorination of TCA.
Bouwer EJ (1994) Bioremediation of chlorinated solvents using alternate electron
acceptors, In: Norris RD, Hinchee RE, Brown R, McCarty PL, Semprini L, Wilson JT,
Kampbell DH, Reinhard M, Bouwer EJ, Borden RC, Vogel TM, Thomas JM, Ward CH
(eds) Handbook ofbioremediation, Lewis Publishers, Boca Raton, pp 149-175
Chen C, Ballapragada, BS, Puhakka JA, Strand SE, Ferguson JF (1999) Anaerobic transformation of 1,1,1-trichloroethane by municipal digester sludge. Biodegradation 10:
de Best JH, Jongema H, Weijing A, Doddema HJ, Janssen DB, Harder W (1997)
Transformation of 1,1,I-trichloroethane in an an aerobic packed-bed reactor at
various concentrations of 1,1,I-trichloroethane, acetate and sulfate. Appl Microbiol
Biotechnol 45:417-423
de Best JH, Hage A, Doddema HJ, Janssen DB, Harder W (1999) Complete transformation of 1,1,1-trichloroethane to chloroethane by a methanogenic mixed population.
Appl Microbiol Biotechnol 51:277-283
Domenico PA, Schwartz FW (1990) Physical and chemical hydrogeology.Wiley, New
Gälli R, McCarty PL (1989) Biotransformation of 1,1,1-trichloroethane, trichloromethane, and tetrachloromethane by a Clostridium sp. Appl Environ Microbiol55:837844
Klecka GM, Gonsior SJ, Markharn, DA (1990) Biological transformations of 1,1,1trichloroethane in subsurface soils and groundwater. Environ Toxicol Chem 9:
Maymo-Gatell X, Chien Y, Gossett JM, Zinder SH (1997) Isolation of a bacterium that
reductively dechlorinates tetrachloroethene to ethane. Science 276:1568-1571
S. Delisle et al.
McCarty PL (1997) Breathing with chlorinated solvents. Science 276:1521-1522
McCarty PL, Semprini L (1994) Groundwater treatment for chlorinated solvents,
In: Norris RD (ed), Handbook of Bioremediation. Lewis Publishers, Boca Raton,
pp 87-116
Oldenhuis R, Vink RLJM, Janssen DB, Witholt B (1989) Degradation of chlorinated
aliphatic hydrocarbons by Methylosinus trichosporium OB3b expressing soluble
methane monooxygenase. Appl Environ Microbiol55:2819-2826
Spormann AM, Widdel F (2000) Metabolism of alkylbenzenes, alkanes, and other
hydrocarbons in anaerobic bacteria. Biodegradation 11 :85-1 05
UK Environment Agency (2000) Guidance on the Assessment and Monitoring of
Natural Attenuation of Contaminants in Groundwater. R&D Publication 95, Environment Agency, Bristol, UK
US EPA (1997) Use of Monitored Natural Attenuation at Superfund, RCRA Corrective
Action and Underground Storage Tank Sites, Draft Interim Final, US EPA Office of
Solid Waste and Emergency Response Directive 9200.4-17, United States Environmental Protection Agency, Washington, DC
Vogel TM (1994) Natural bioremediation of chlorinated solvents. In: Norris RD (ed)
Handbook of Bioremediation. Lewis Publishers, Boca Raton, pp 201-225
Vogel TM, McCarty PL (1987) Abiotic and biotic transformations of 1,1,1trichloroethane under methanogenic conditions. Environ Sei Technol21:1208-1213
Wiedemeier TH, Haas PE (1999) Designing monitoring programs to effectively evaluate the performance of natural attenuation, US Air Force Center for Environmental
Excellence, San Antonio, TX
Wiedemeier TH, Wilson JT, Kampbell DH, Miller RN, Hansen JE (1995) Technical protocol for implementing intrinsic remediation with long-term monitoring for natural
attenuation of fuel contamination dissolved in groundwater, US Air Force Center for
Environmental Excellence, San Antonio, TX
Wiedemeier TH, Swanson MA, Moutoux DE, Gordon EK, Wilson JT, Wilson BH,
Kampbell DH, Hansen JE, Haas P, Chapelle FH (1996) Technical protocol for evaluating natural attenuation of chlorinated solvents in groundwater, Draft revision l.
US Air Force Center for Environmental Excellence, San Antonio, TX
Wiedemeier TH, Rifai S, Newell CI, Wilson JT (1999) Natural attenuation of fuels and
chlorinated solvents in the subsurface. Wiley, New York
Evaluation of (urrent Soil Bioremediation
Owen P. Ward l and Ajay Singh 2
Bioremediation competes with other methods as an approach to an
environmental cleanup of contaminated soil. Physicochemical methods
available for treatment of petroleum-contaminated soils are generally
costly and capitallenergy-intensive, and provide destructive or separation approaches, including thermal desorption, incineration, and
solvent extraction. Destructive methods often emit large quantities of
greenhouse gases. Separation methods are difficult to manage because
of the variable nature of the contaminants and soils, and residuals may
still require appropriate environmental disposal. Comparisons between
methods indicate that biological, thermal and scrubbing methods may
be used to remediate soil of contaminants such as petroleum hydrocarbons, polycyclic aromatic hydrocarbons (PAR) (except some high
molecular weight PARs), phenols, nitroaromatic and halogenated compounds. Thermal or scrubbing methods have generally been used to
degrade polychlorinated biphenyls (PCB), polychlorinated dibenzodioxins (PCDD) and polychlorinated dibenzofurans (PCDF). Where
bioremediation is effective it is typically the least expensive technology
option. This chapter is divided into two main parts. The first section
evaluates the component elements or factors which influence technology choice. The second section examines the major technology types
and their applicability, supported by performance-based case study
[Department of Biology, University of Waterloo, Waterloo, Ontario N2L 3G1,
Canada, e-mail: [email protected], Tel: +1-519-8851211 Ext. 2427, Fax:
2Petrozyme Technologies, Guelph, Ontario N1H 6H9, Canada
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
O.P. Ward and A. Singh
Factors Affecting Bioremediation and Choice of Technology
Site Characteristics
Choice of technology will often be determined by contaminated site
characteristics and, in particular, by accessibility to the contaminated
material, its mobility on the site and other information which would be
generated in a site assessment report (Cole 1994). Important elements
in this will indude: site description - location of site, buildings, roads,
structures, pipes; subsurface description - depths of saturated and
unsaturated zones and seasonal groundwater fluctuations and flow
direction. Selby (1991) has provided a detailed description of site
assessment methods.
Soil Type
Contaminated soil type may range from days, silts and sediments to
sands and gravels. In bioremediation, soil porosity (the ratio of pore
space to total material volume) and soil permeability (its capacity to
transmit fluids) will influence remediation effectiveness and hence
remediation approach. Partide size as measured by grain surface area
is an important property related to contaminant sorption - the greater
the surface area, the greater the capacity for contaminant sorption
which can reduce bioavailability. Contaminants can become associated
with soil particles by entrapment, sequestration and sorption and contaminants may diffuse into micro-regions to which microbes have no
access, especially in day, thereby limiting bio degradation (Guerin and
Boyd 1992; Rarms 1996). Rydrophobicity of the soil mayaiso playa role
in reducing contaminant bioavailability (Nam and Alexander 1998).
Furthermore, contaminant-soil tight associations, as for example those
occurring with PARs and TNT, often result in residues that are no longer
extractable, and hence not properly analyzed, a phenomenon described
as humification (Rieger and Knackmuss 1995; Lenke et al. 1997).
Moisture Content
Soil moisture content, measured as degree of water saturation, will vary
with soil depth in a typical soil column, from the upper unsaturated
Evaluation of Current Soil Bioremediation Technologies
vadose zone (20-60% saturation) to the capillary zone (60-80% saturation) to a water table fluctuation zone (60-100% saturation) and the saturated groundwater zone (100% saturation).
Nature of Contaminant
While indigenous microorganisms have developed the capability to
degrade many organic compounds, many molecular species have structures or physico-chemical properties which are not readily transformed
or may be transformed but do not support the growth of the transforming organisms, and these moleeules can persist in soil for many
years. They are described as being recalcitrant. Some of these molecules, for example certain polymers like plastics, are persistent but nontoxie. However, many recalcitrant moleeules are toxic. Examples of
persistent chemieals, which have been shown to survive in soil for 15
or more years (Alexander 1999), include highly chlorinated PCB congeners, chlordane, DDT, heptachlor, dieldrin, lindane and simazene.
Other moleeules may be biodegraded but are not easily accessed by
microorganisms because of their low water solubility and their tendency to bind to soil. Examples of the aqueous solubilities of contaminants benzene, o-xylene, naphthalene, anthracene and benzo[a]pyrene
are: 1780,175,58,30,0.07 and 0.004 mg/I, respectively (Mahro 2000). The
more soluble compounds are generally more biodegradable.
A generalized sequence of petroleum components in order of
decreasing biodegradability is represented as follows: n-alkanes >
branched-chain alkanes > branched alkenes > low molecular weight
n-alkyl aromatics > monoaromatics > cyclic alkanes > polynuclear
aromatics »» asphaltenes. Predictive models have been designed
for estimating the extent of petroleum hydrocarbon biodegradation
(Huesemann 1995). Some compounds or their non-degradable metabolites may be toxic to the microbial populations with a possible negative
impact on bioremediation potential. Such considerations may contribute to the decision as to whether bioremediation may be a cleanup
option or regarding a choice of bioremediation approach.
Time Constraints
There are many cases where short-term real estate development plans
require accelerated remediation of contaminated sites and where the
long time scales required for conventional bioremediation and natural
attenuation processes are not acceptable. Accelerated bioremediation
O.P. Ward and A. Singh
typically exploits increased levels of microbial expertise and there are
opportunities for advances in our knowledge of the microbial biochemistry, physiology, genetics and recombinant technology to be used
to optimize rates and extents of bio degradation. Generally, it has been
shown that increased microbial intervention typically comes with a
decrease in process cyde time.
Many organic chemicals that are environmental contaminants are
volatile. A major disadvantage of some bioremediation processes relates
to the large quantities of volatile organic carbons, hazardous to human
health, present in the contaminated soil, which get released into the
atmosphere. These wastes also cause tropospheric ozone production
(Field et al. 1992; Moseley and Meyer 1992; Coppock and Monstrom
1995). In the Exxon Valdez spill in the relatively cold Alaskan dimate,
15-20% of the oil was reported to be lost to the atmosphere by
volatilization (Field et al. 1992; Cole 1994) and other reports indicate
that up to 60% of remediation may be due to volatilization (Preslo 1988;
Salanitro 2001).
Untermann et al. (2000) described a hierarchy of technology choices for
bioremediation. Where the degrading population exists in the contaminated zone but where nutrients or other conditions are not sufficient
to promote microbial activity (oxygen is often the most limiting substrate, which may be supplied by the introduction of air, O2 , H2 0 2 or
manganese peroxide), the process may be accelerated by the addition of
other nutrients, typically sources of nitrogen and phosphorus, for remediation of carbon contaminants. Added co-substrates may promote
growth and may be especially important where the contaminant is
degraded by co-metabolism.
Oxygen/Electron Acceptor
In surface bioremediation processes, aeration is achieved by the direct
contact of the soil medium with the atmosphere. Where the contaminated soil is in an unsaturated zone in situ or indeed in an engineered
Evaluation of Current Soil Bioremediation Technologies
soil pile, oxygen is typieally supplied in the form of air. However, where
soil porosity is low or where it is desirable to keep gas flow rates in
the soillow because of contaminant volatilization problems, compressed
oxygen may be used in place of air. In the latter case, and where the contamination is in a water-saturated zone, hydrogen peroxide may be used
as an alternative source of oxygen, given the low solubility of oxygen in
water (McCarthy 1998). HzO z is highly soluble in water and can slowly
breakdown in the aquifer to free oxygen. However, when injected into
the ground problems ofbiomass plugging dose to the point of injection
have been experienced. Slow release of inorganic peroxide formulations, such as magnesium peroxide, have met with success (Prince 1998).
In environments depleted of oxygen, aerobic microbes may be displaced
by other species capable of using organic materials as electron acceptors or inorganic electron acceptors such as nitrate, sulfate or CO z• Ferric
iron mayaiso serve as an electron acceptor enabling some strains to
anaerobieally oxidize some simple aromaties, in du ding benzoate and
phenol (Lovely 2000). In general, rates of these anaerobie transformations are slow and very long periods are required to acdimate the
microbial population.
Co-metabolie Substrate
Transformation of organic compounds by microbes, which are unable
to use the substrate as a source of energy or for significant nutrition al
benefit is termed co-metabolism (Alexander 1999). The substrate is
typically metabolized in the presence of a second substrate that may
support growth, but in certain circumstances metabolie reactions may
occur even in the absence of the growth-supporting co-substrate. Cometabolie conversions involve a single enzyme reaction or aseries of
reactions. The failure of the resulting "dead-end" product to participate
in further metabolism may be due to the substrate specificities of the
downstream metabolizing enzymes, or the presence of a substituent
which prevents a forward re action occurring or the reaction product may re ar range or otherwise chemically react to form a nonmetabolizable end product. In a mixed contaminant environment, for
example in hydrocarbon oil bioremediation, smaUer and often more
volatile hydrocarbons are generally more biodegradable and more
supportive of microbial growth and ceU energy provision than more
complex components which may rely on co-metabolie pro ces ses. Consequently, in prolonged hydrocarbon bio degradation processes where
volatile materials are lost to the atmosphere, the development of a
microbial population on these substrates may be prevented which might
O.P. Ward and A. Singh
contain the potential early stage catabolic enzymes with relaxed substrate specificities to initiate transformation of the more recalcitrant
compounds. Kanaly and coworkers (1999,2000) showed that diesel fuel
or other hydrocarbons stimulated co-metabolie mineralization of
benzo[a]pyrene in culture and in soil. Anaerobic biostimulation of
chlorinated solvent degradation in situ was achieved by the addition
of selected organic acid and alcohol substrates (Sewell et al. 1998).
The hydrophobie nature of many organic contaminants often limits the
capacity of microorganisms, which generally exist in aqueous environments, to assimilate and degrade these compounds. Hydrocarbondegrading bacteria produce a variety ofbiosurfactants, either associated
with the cell surface or secreted into the extracellular medium (Fiechter
1992; Barathi and Vasudevan 2001; Makkar and Cameotra 2002).
However, production of biosurfactants by fermentation, as a possible
amendment for bioremediation, appears to be more costly, so the use of
commercially-produced biosurfactant products in bioremediation is
arguably uneconomic. Hence chemical surfactants, which are much
cheaper to produce, have a potential role. Properly chosen, chemical surfactants may enhance bio degradation (Rouse et al. 1994; Nelson et al.
1996; Bruheim et al. 1997; Bruheim and Eimhjellen 1998; Van Hamme
and Ward 1999). However, both enhancements and the inhibition of
bio degradation of hydrocarbons have been observed (Mulligan et al.
Biodegradation of certain poorly soluble petroleum hydrocarbons
may be inhibited by surfactants as a result of (1) toxicity by a high concentration of surfactant or soluble hydrocarbon; (2) preferential metabolism of the surfactant itself; (3) interference with the membrane uptake
process; or (4) reduced bioavailability of miceller hydrocarbons
(Efroymson and Alexander 1991; Rouse et al. 1994). There is always the
concern that the surfactant may get used preferentially as a carbon
source instead of the contaminant. Hence, there is a need to provide a
perspective as to when or how surfactants may be exploited in petroleum hydrocarbon degradation processes, to improve rates and extents
of degradation. Properties of chemical surfactants that influence their
efficacy include charge (nonionic, anionic or cationic), hydrophiliclipophilic balance (HLB, a measure of surfactant lipophilicity) and critical micellar concentration (cmc, the concentration at which surface
tension reaches a minimum and surfactant monomers aggregate into
micelles). Perhaps the largest problem in attempting to evaluate the
Evaluation of Current Soil Bioremediation Technologies
utility of surfaetants in bioremediation is due often to the eomplexity
of the systems tested, with a tendeney to generalize on eonclusions
drawn from very partieular sets of cireumstanees.
The following observations may be relevant to this diseussion: In
general, surfaetant eoneentrations required to solubilize and mobilize
hydrophobie eontaminants from soil are typieally an order of magnitude high er than requirements in a predominantly aqueous solution.
Typieal surfaetant eoneentrations required to wash eontaminants out of
soil are 1-2%, whereas the same eontaminants may be solubilized in an
aqueous solution at a surfaetant eoneentration of 0.1-0.2%. It appears
that mueh of the surfaetant added to soil is ineffeetive as it beeomes
sorbed to soil particles. In eases where mieellarization of the eontaminant (at or above the surfaetant eme) prevents aeeess by the mieroorganism, the higher surfaetant eoneentration required for soil washing is
likely to retard mierobial degradation of that eontaminant in the washings. In some eases, diluting the washings to get the surfaetant eoneentration below its eme ean facilitate mierobial aeeession and eontaminant
degradation. This has been observed with PCB bio degradation
(Billingsley et al. 1999,2002). Nonylphenol ethoxylate ehemieal surfaetants enhaneed the bio degradation of water-insoluble substrates by a
eo-eulture through the disruption of hydrophobie eeIl-substrate interactions of one of the strains (Van Hamme and Ward 2001). This observation illustrated that the range of stimulatory and inhibitory effeets
reported in the literature may not be contradictory but simply deseribe
unique eases based on surfaetant properties and the physiology of the
organisms involved (Hommel 1990). Understanding how baeteria
respond differentially to surfaetant amendments in aeeessing insoluble
hydrophobie substrates, and how these responses vary in the presenee
of other species, is useful when developing bio degradation protoeols
involving mixed eultures (Van Hamme and Ward 2001).
Solubilizing Agents
Lower moleeular weight hydroearbons ean assist in the solubilization of
the more reealcitrant and more hydrophobie moleeules, making them
more bioavailable to the mierobial population. Jimenez and Bartha
(1996) attributed the ability of solvents such as paraffin oil to promote
mineralization of pyrene to their solubilizing and mass transfer action.
Addition of vegetable oil to soil facilitated mobilization of the PAHs
present and aeeeierated their degradation (Pannu et al. 2003).
O.P. Ward and A. Singh
Bulking Agents
Soils with low porosity, especially days and silts, are difficult to aerate
because of their low permeability and this is particularly problematic
in static processes such as in situ or engineered soil pile processes and
even in some mixed solid phase processes. Bulking agents, ranging from
sand to plant material, especially woody materials with large partide
sizes, may be used to increase the porosity of the soil. The major drawback of adding bulking agents is that they increase the mass of the contaminated medium by up to 100% at the start of the process.
Several laboratory and field investigations have indicated that inoculation with selected and acdimated microbes (bio augmentation)
(Venkateswaran and Harayama 1995; Venkateswaran et al. 1995; Chatre
et al. 1996; Mohandass et al. 1997) did not significantly enhance rates of
oil bio degradation over that achieved by nutrient enrichment of the
natural microbial population (biostimulation) (Fayad et al. 1992;
Pritchard et al. 1992; Venosa et al. 1992). Successes observed under controlled laboratory conditions may have been due to reduced competition from indigenous microflora. The comprehensive studies on the
bioremediation effects related to the Exxon Valdez oil spill established
that augmentation with hydrocarbon-degrading inocula had no
significant impact on the process over and above the application of
nitrogen-containing fertilizers (Glaser 1993). This experience has been
viewed by many as a general rule that bio augmentation is ineffective in
petroleum and other biodegradation processes.
Hence, there is sufficient data to indicate that possible advantages of
bio augmentation are not easily predicted and variations in the strains
selected, the environment to be treated and in critical physico-chemical
parameters influencing bioremediation, are not always understood
(Vogel 1996). At the very least, the data raise valid questions as to
whether a role exists for inocula in petroleum degradation processes
and under what circumstances. A further interesting question relates to
whether there is any potential to exploit recombinant organisms in the
practice of environmental bioremediation and was te treatment. Bioaugmentation has been effectively used in agriculture and wastewater treatment (Jasper 1994; Rittmann and Whiteman 1994). In bioremediation
there are examples of successes where the introduction of competent
Evaluation of Current Soil Bioremediation Technologies
microbes appeared to enhance the bioremediation rate and where the
indigenous population appeared non-competent (Briglia et al. 1994;
Lamar et al. 1994; Shin and Crawford 1995; Dave et al. 1995; Riggle
1995). Bioaugmentation has been beneficial in treating recalcitrant
compounds such as 2,4,6-TNT, carbon tetrachloride and PCP (Dybas et
al. 1995; Edgehill 1995; Holroyd and Caunt 1995; Shin and Crawford
1995; Witt et al. 1995). The importance of using microbes that have
demonstrated long survival times in a particular soil environment may
be an important dimension in this approach. In modeling experiments,
Pseudomonas species had much longer survival times than Escherichia
coli (Vandepitte et al. 1995). The use of very high microbial numbers in
bioaugmentation (>10 7/g soil) has been shown to be effective where
bioavailability was not the limiting factor (Briglia et al. 1994; Pearce et
al. 1995).
Sampling and Monitoring
The implementation of proper sampling and monitoring protocols is a
big issue and an important requirement in remediation processes in
general. It represents a large cost component particularly in the case of
in situ bioremediation processes, which may be characterized by variations in the soil type, water saturation level and concentrations of contaminants, nutrients, oxygen/electron acceptors, horizontaHy and
verticaHy across the "plume". Monitoring weHs need to be installed to
monitor progress of contaminant removal in a three-dimensional grid.
Sampling and monitoring systems are somewhat simpler in excavated
above ground engineered soil piles, where some of the variations may
be eliminated through proper pile construction. Sampling accessibility
problems are almost completely eliminated in surface remediations
such as shoreline treatments and land farming. In mixed systems such
as land-farming systems, the greater the degree of mixing, the more
homogeneous the system and the lower the number of samples required
for process characterization. This reaches a limit in weIl-mixed slurry
systems, where homogeneity is approached or achieved, meaning only
a single sampie location is required. Sampling protocols and methods
for soil bioremediation are described by Keith (1988) and Roemer
(2000) and analytical methods are provide in EPA publications (EPA
O.P. Ward and A. Singh
Bioremediation Technologies and Their Evaluation
Bioremediation methods for the treatment of contaminated soils, using
indigenous or augmented microorganisms in land farming, biopiling/composting, bioventing and bioreactor configurations, have been
well documented (Ward 1991; Prince 1993; Atlas and Cerniglia 1995;
Korda et al. 1997). In reality, strategies for bioremediation manifest
themselves in processes having different degrees of complexity and
technological requirements. At the extremes, bioremediation of contaminants in soil by natural attenuation requires no human intervention, whereas the implementation of accelerated and controlled
bioreactor-based processes may be directed to exploiting microbial
technology and bioprocess engineering to optimize rates and extents of
contaminant degradation. In systems which require little or no microbiological contribution for bioremediation effectiveness, process-limiting factors often relate to nutrient or oxygen availability or the lack
of relatively homogeneous conditions throughout the contaminated
medium. These conditions result in microbial growth and degradation
processes operating under conditions which are both variable and suboptimal and which, at best, lead to prolonged degradation cycles.
Processes are often unpredictable and unreliable with respect to performance and required contaminant degradation endpoints are often
not achieved though the whole medium or hot spots of contamination
remain. These processes ignore the realities of enzyme and cell substrate saturation kinetics which result in a tendency far rates of degradation to slow down as contaminant concentrations fall and lead to
associated reductions in the viable microbial populations. In cases
where contaminants are degraded by co-metabolism early elimination
of the co-substrates necessary for the degradation of these contaminants can halt the degradation processes. This often occurs with the
biodegradation of easily metabolizable co-substrates and frequently the
latter substrates mayaiso be depleted through volatilization. The nonhomogeneous and unpredictable nature of these processes makes them
intensive in terms of sampling, and analytical activities as patterns of
contaminant removal have to be monitored at different locations
throughout a three-dimensional grid. The latter disadvantages represent a major justification for the implementation of more controlled and
optimized biodegradation processes, which ensure contaminants are
efficiently biodegraded to defined criteria.
In this section, the range of petroleum bio degradation processes will
be reviewed starting with the processes requiring the least microbial
Evaluation of Current Soil Bioremediation Technologies
expertise and moving on to processes with increasing levels of microbial technological complexity.
Natural Attenuation
Natural attenuation or intrinsic remediation is the least invasive
approach to bioremediation, requiring no intervention other than to
demonstrate that an indigenous population exists that can degrade the
contaminants and to monitor the progress of the degradation. Natural
attenuation is more frequently applied to groundwater remediation
than to soil and its efficacy in meeting remedial objectives remains controversial (Hughes et al. 2000). The limiting nutrient in the subsurface
in the presence of carbon contaminant is typically oxygen. Soil becomes
increasingly anaerobic with increased soil depth. The rate of remediation is a function of nutrient availability and size and the composition
of any indigenous bacterial population. Best bioremediation will occur
in the top organic soillayer. Deeper contamination can last for years. In
general, this approach is effective with the lightest refined oil products
such as gasoline, diesel and jet fuel. However, abiotic losses due to evaporation, leaching, dispersion and photo-oxidation also playa major role
in the decontamination of an oil-spilled environment. In natural attenuation anaerobic bio degradation may occur if electron acceptor compounds are present (Hinchee and OlfenbutteI1991). A classical example
of natural attenuation was the extensive dechlorination observed with
PCB contaminants in the Hudson riverbed (Table 1). Comprehensive
details of natural attenuation processes are discussed in Chapter 8.
In Situ Subsurface Bioremediation
In the case of in situ subsurface bioremediation processes, arguably the
greatest challenge relates to the physical properties and makeup of the
subsurface and the challenge of engineering that environment in order
that microbial activity can thrive there and effectively degrade the contaminants present. Oxygen, other electron acceptors and/or nutrients
are supplied to the subsurface with minimum onsite operations and at
a relatively low cost. Processes work best with loose porous soils. Injection wells are required to provide nutrients. However, it may be necessary to contain the contaminant plume boundary to prevent subsurface
spreading. Surface containment may be needed to prevent volatilization.
For the deeper unsaturated zone and the lower water-saturated zone,
O.P. Ward and A. Singh
Table 1. Natural attenuation and in situ subsurfaee proeesses
Example 1.1. Intrinsie bioremediation (Abramowiez 1990; Bedard and May 1996)
• Using intrinsie bioremediation approaehes, potential for a bioremediation solution
to the PCB eontamination of the Hudson riverbed was investigated in detail.
• Performance: Extensive deehlorination of highly ehlorinated PCBs was observed
under anaerobie eonditions, especially in a methanogt~nie environment. SampIes of
long-term PCB-eontaminated sediments showed as much as a 45% deerease in the
hexa- and hepta-chlorinated biphenyls (CBs) relative to Arochlor-60 with an
increase in the tri-, tetra- and penta-CBs.
Example 1.2. Anaerobic decontamination of chlorinated solvents (SeweIl et al. 1998)
• Decontamination under anoxie conditions was also achieved using electron
acceptors. Anaerobie bio stimulation of chlorinated solvent degradation in situ was
achieved by the addition of benzoate, lactate and methanol to an aquifer.
• Performancelcycle time: All eontaminants were signifieantly redueed in 5 months.
Example 1.3. Benzene-eontaminated soil (Chiang et al. 1989)
• In a minimal intervention in situ system soil eontaminated with benzene 10-25 ft.
below the surfaee was remediated by aeration.
• Performancelcycle time: 90% of the eontaminant was redueed in 21 months. Only
5% of the reduetion was attributed to volatilization.
Example 1.4. TCE-contaminated aquifer (McCarthy et al. 1998)
• Oxygen, hydrogen peroxide and toluene were injected into a TCE-eontaminated
aquifer to stimulate growth of toluene oxidizing baeteria to eo-metabolically
degrade TCE.
• Performance: When groundwater, containing 1,000 j.Lg/I TCE, was subsequently
pumped through the 480-m 2 treatment zone, 97-98% of TCE removal was
Example 1.5. Petroleum hydrocarbon-eontaminated soil and groundwater (Nelson
et al. 1994)
An integrated approach to in situ bioremediation to treat petroleum hydroearboneontaminated soil and groundwater under a formerly used oil sump area.
Biostimulation was promoted by the injeetion of nutrients, hydrogen peroxide and
air upstream of the direetion of groundwater fiow through the plume with soil
vapor extraetion wells downstream.
• Performance: 16,144kg of contaminant mass was removed in a 30-month period,
94% through bio degradation and the remainder through the recovery of phase
separated hydrocarbons or by vapor extraetion.
bioventing and biosparging are employed as the most effective systems
for oxygen delivery to treat the subsurface soil (Klein 2000). Biosparging involves the injection of atmospheric air into the aquifer, generally
into the saturated zone and typically below the contaminant plume.
Bioventing typically treats the unsaturated soil zone and may involve
active or passive air injection. Bioventing normally includes enhanced
soil vapor extraction (Held and Dorr 2000). Nutrients are supplied as
Evaluation of Current Soil Bioremediation Technologies
required, perhaps additional substrates if needed for co-metabolism
and moisture is controlled. Extraction wells may be used to draw oxygen
through the vadose zone, with gas mobility facilitated by the porous soil
(Cole 1994) with surface entry of oxygen or H20 2 and nutrients (spraying). The inflow of air and its passage between the sources (surface or
injection wells) through the contaminated zone to the extraction weHs
is driven by pressure differences (Brown et al. 1994; Reddy et al. 1995).
Air-flow rates are typicaHy maintained low to favour bio degradation
over volatilization. The approach is best suited for soils with high gas
permeability with gravel or sand consistencies giving better performance than day, and is effective for example for treating petroleum or
chlorinated hydrocarbon -contaminated soils. The extraction weHs or
independent monitoring weHs located outside the plume may be used
to measure the decrease in oxygen and increase in CO 2, reflecting contaminant bio degradation (Hinchee and Olfenbuttel1991).
With substantiallevels ofboth volatile and non-volatile hydrocarbons
in the subsurface, with the latter present as a non-aqueous phase liquid,
one remediation approach, which has met with considerable success, is
termed "bioslurping". Remediation involves liquid and vapor extraction
and bioremediation of the vapor. As the liquid layer is pumped out, air,
often together with nutrients, is drawn in or injected via aeration weHs,
which stimulates aerobic bio degradation (Keet 1995).
Soil Pile and Composting Techniques
Soil piles are typically engineered to promote the transfer of oxygen
to excavated soil. Aseries of pipes in the pile promotes aeration by
distributing air to the pile either by positive pressure or preferably
by drawing a vacuum in the pipes so that the direction of airflow is
from the surface of the pile through the soil into the pipes. The latter
approach allows any containment of a volatile nature released from the
soil to be directed to a VOC treatment system, for example activated
carbon. The excavated soil may be mixed to achieve a homogeneous
distribution of contaminant and perhaps to add bulking agent and
Static composting of soil is implemented in a variation of the soil pile
approach, perhaps the major difference being that bulking agents, are
added typically at a ratio of approximately 1 : 1 w/w to soil. The more
porous nature of the mixture aHows for simpler pipe work arrangements for aeration, located at the center of the pile. Bulking agents
include biodegradable material, induding wood and plant materials, that enhance composting and heat generation (Ro et al. 1998;
O.P. Ward and A. Singh
Table 2. Soil pile and composting processes
Example 2.1. Composting of munitions contaminated soil (Williams et al. 1989)
• Thermophilie composting was used to remediate soil contaminated with TNT,
RDX, HMX, tetryl and nitrocellulose. Bulking agents, such as alfalfa, straw and
woodchips, together with nutrients were applied. Volume to mass ratio was 1.65: 1
m 3 /metric ton.
• Aeration/mixing: A radial blade blower pulled air through the compost pile, linked
to a thermistor to maintain temperature at 55°C.
• Performancelcycle time: Contaminants were reduced from 17,870 to 74mg/kg in
153 days.
Example 2.2. Composting of chlorophenol-contaminated soil (Valo and SalkinojaSaltonen 1986)
• Soil was combined 50: 50 by weight with plant bulking materials and N, P and K
fertilizer added to two 50-m3 windrows and remediated by composting with weekly
irrigation. Temperature of the compost was I-15°C above ambient and ranged
from 15-32°C depending on the season.
• Inoculum: While no inoculum was used the contaminated soil had developed a
strong chlorophenol-degrading population of 4.5 x 104 colonies/g, which rose to 5
x 106 colonies/g during composting.
• Performancelcycle time: Contaminants were reduced from 200-300 to 15mg/kg
over 17 months. All monochlorophenols were degraded except polychlorinated
Example 2.3. Anaerobiclaerobic composting process for explosives (Lenke et al. 1997)
In this process, organic matter was added to the soil, the degradation of which
created strictly anaerobic conditions and raised the soil temperature to 60 oe.
Under these anaerobic conditions, TNT was converted to triaminotoluene (TAT).
The soil was then turned to create aerobic conditions which promoted TAT bin ding
to the soil.
• Performancelcycle time: In aperiod of 35 days the concentration of explosives
contaminants was reduced from 325 to <1 ppm.
Alexander 1999). Other eomposting eonfigurations may be employed to
aehieve aeration, especially the mixing/turning the eomposting material
in windrows. Contaminants are degraded/transformed or stabilized by
mesophilie (15-45°C) or thermophilie baeteria (50-70°C), although a
higher thermophilie end may not be optimal beeause it reduees mierobial and henee metabolie diversity. The optimal range is 50-65°C
(Cookson 1995). Examples of some soil pile and eomposting proeesses
are shown in Table 2.
Marine Shoreline and Wetlands Remediation
Bioremediation is an effeetive teehnology for the treatment of oil
pollution sinee the majority of moleeules in the erude oil and refined
Evaluation of Current Soil Bioremediation Technologies
products are biodegradable and oil-degrading microorganisms are
ubiquitous (Fayad and Overton 1995; Aislabie et al. 1998; Chaineau et
al. 2000). Abiotic losses due to evaporation, dispersion and photooxidation also playa major role in decontamination of oil-spilled environment (Garrett et al. 1998; Salanitro 2001). Nutrient availability,
especially N and P, appears to be the most common limiting factor
(Pritchard et al. 1992; Rosenberg 1992) and applications of inorganic
and organic fertilizers, surfactants and bulking agents have shown
success (Lee et al. 1993, 1995; Venosa et al. 1996; Boufadel et al. 1999).
Accidental marine oil spills have generated a major worldwide environmental concern and earlier reviews have described bioremediation
efforts (Pritchard et al. 1992; Lee et al. 1993; Fingas 1996; Swannell et al.
1996; Prince 1998; Cohen et al. 2001). In open-water environments successful strategies have involved the application of nutrients in oleophilic
forms which are retained at the oil-water interface and are not washed
away by the tide (Swannell et al. 1996). A micro-emulsion mixture
of urea in brine, encapsulated in oleic acid as an extern al phase
with lauryl-ether phosphate as a surfactant, and a slow release N, Pformulation with a modified urea-formaldehyde polymer were effective
for the treatment of beach surface or subsurface, respectively, accelerating the rate of oil removal from two- to five-fold (Pritchard and Costa
1991; Button et al. 1992; Swannell et al. 1996; Lee and DeMora 1999; Van
Ramme et al. 2001). Bioremediations involving shorelines, present
specific challenges in that applied nutrients and developing microbial
populations may be washed away by tidal waters. Added inoculants did
not have a beneficial effect on bioremediation emphasizing the importance of relying on the indigenous developing microbial population.
Presumably, the surviving and selected indigenous organisms, like the
oleophilic fertilizers, resisted tidal washing by their abilities to associate with oily surfaces. Treatment with oleophilic fertilizers, of the
oil-contaminated beaches and cobbled surfaces, effected significant
degradation of the oil, as compared with no nutrient controls, in a 6week period (Glaser, 1993).
Given the observed regenerative powers of brackish marshes, there
has been some interest in using constructed wetlands as bioremediation systems. Often biodegradation is accompanied by other contaminant removal and physico-chemical destruction mechanisms
(Untermann et al. 2000). Rowever, degradation processes were found to
be predominantly aerobic and the acute toxicity of contaminants or
metabolites was often a problem, and the availability of fertilizer and
oxygen was often rate limiting (Rarris et al. 1999; Lin et al. 1999; Mueller
et al. 1999; Ruddleston et al. 2000; Shin et al. 2000). Wetland processes
that deserve more attention include volatilization, partitioning, photodegradation, bio degradation, and plant and animal uptake processes.
O.P. Ward and A. Singh
Land Farming
Land-farming processes can be used for in situ treatment of shallow
contaminants in the upper unsaturated zone. While land farming of
refinery and wellhead oily sludges is no longer considered environmentally acceptable, it is still used as an oily sludge treatment and disposal method in many parts of the world (Atlas 1981; Bartha 1986;
Huesemann 1995). Large tracts of land are deliberately contaminated,
followed by the subsequent bioremediation of the less recalcitrant oil
fractions. These land-farming operations can tie up large areas of land,
which later will have to be decommissioned. Maximum contaminant
degradation occurs in the tilled surface, typically amounting to a depth
of 1O-20cm , although deeper aeration and mixing (up to 50cm, using
ploughing and rotavating equipment) have also been effectively implemented. Bulking agents are sometimes added to the soil to promote
greater oxygen diffusion. Sprinklers are used to control moisture. Soil
moisture content is maintained at 40-60% of saturation level. Some
petroleum hydrocarbon land-farming examples are given in Table 3 to
provide insight into degradation performance. Examples 1-3 refiect
degradation rates of 1% petroleum hydrocarbons per month with adequate aeration. Example 4, simulating a process land farm in a petrochemical plant refiects an approximate TPH degradation rate of 1% per
2-months but shows a high accumulation of TPHs in the soil and very
poor degradation of high molecular weight PAHs.
While oily sludges have traditionally been processed by landfarming, bioremediation of these practices are banned in the United
States and are being phased out in many other jurisdictions. Oil companies have therefore been forced to seek other disposal solutions.
Most of the rate-limiting and variability factors observed in land
farming may be eliminated by employing the more homogeneous conditions in a bioreactor. Bioreactors can accommodate solids concentrations of 5-50% w/v. Bioreactor systems, including waste lagoons, unless
maintained as simple configurations, may be costly to operate as compared to the traditionalland-farming method (Cookson 1995; Hughes
et al. 2000). Bioslurry processes help break up solid aggregates and disperse insoluble substrates, exposing increased particle surface areas and
maximizing hydrocarbon contact with the aqueous phase, increasing
Evaluation of Current Soil Bioremediation Technologies
Table 3. Land farming of petroleum hydrocarbons
Example 3.1. Petroleum-contaminated soil in refinery (EHis 1994)
o A refinery soil contaminated with 1.3% oil was treated with nutrients and
surfactants. Air temperatures were maintained around 25°C.
o Inoculum: Microbial inoculants were applied to the soil.
• Mixing/aeration: The soil was regularly mixed and aerated with deep tilling
• Performance!cycle time: TPRs were reduced by about 90% in 34 days.
Example 3.2. Land farming of number 6 fuel oil-contaminated soil (FogeI1994)
o Application of land farming with nutrient application and control of moisture level
to remediate soil containing 6% of number 6 fuel oil.
• Aeration/mixing: Aeration provided by both ploughing and rotavating.
• Performance!cycle time: 80-90% reduction in TPRs was observed in 6 months.
Example 3.3. Land farming of kerosene-contaminated soil (Dibbles and Bartha 1979)
• Following a pipeline break in a wheat field, land farming of kerosene-contaminated
soil up to depth of 45 cm using lime and nutrients.
• Aeration/mixing: Frequent tilling.
• Performance!cycle time: Contaminant concentration was reduced from 8,700 to
30-3,OOOppm (depending on soil depth) in 24 months. The variation with depth
suggested oxygen limitation. Seed germination and yield data showed the field
returned to normal wheat production capability after 1 year.
Example 3.4. Process land-farm simulation in a petrochemical plant (Bosert et al.
o The fate of petroleum waste hydrocarbons during a laboratory study of oily sludge
application to soil was monitored. The hydrocarbon application period was 751
days with the active land-farming experiment lasting 920 days.
• Aeration/mixing: Provided to simulate an active land-farming bioremediation.
• Performance!cycle time: A gradual accumulation of petroleum hydrocarbons up to
13.8% w/w occurred in the soil over time. Of the total PARs applied to the soil in
the waste, the percentages remaining at the end of treatment were 1.4,47.4,78.5
and 78.3 for the 3-,4-, 5- and 6-ringed PARs, respectively. Residual concentrations
for pyrene and benzo[a]pyrene were 245 and 28ppm, representing extents of
degradation of 14.4 and 44.4%, respectively. At the end of the treatment period,
53% (155 mg hydrocarbons per g soil) of the applied hydrocarbons was removed
from the soil, representing a degradation rate of 1% (w/w) per 2 months.
desorption and accelerating biodegradation. Slurry reactors can
provide a finer control of key factors, such as pR, temperature, moisture, mixing, thereby increasing the bioavailability of nutrients including the rates and extents of microbial growth and oil transformation,
and can thus facilitate optimal system performance. Much higher rates
and extents of degradation than are observed in land-farming systems
due to the minimization of mass-transfer limitations and increased
desorption of contaminants by continuous mixing are achieved
O.P. Ward and A. Singh
(Christodoulatos and Koutsospyros 1998). Bioreactor systems indude
batch and continuous stirred tank reactors and air-agitated systems,
induding fiuidized beds (Kleijntjens and Luyben 1987). Bioreactor
cascades or sequential batch reactors have been described to achieve
semicontinuous modes of operation. Liquid/solid treatment (LST) by
bioremediation may be used to treat soils contaminated with petroleum,
phenols, chlorinated hydrocarbons, pesticides, phthalate esters (Ward et
al 1993, 2003; Cookson 1995; Alexander 1999; Juneson et al. 2001; Ward
and Singh 2001) and is recognized as a technology applicable to the
degradation of petroleum refinery sludges (Stroo 1989).
Containment of the sludge bio degradation processes in bioreactors
allows for the management of off gas ses, for example through the use
of surfactants or sorbents in the medium, to reduce volatilization. By
creating culture conditions, which accelerate the process of bioremediation ofVOCs, the bio degradation process rather than volatilization can
become the dominant VOC removal mechanism. The bioslurry reactor
examples shown in Table 4 illustrate a trend towards the use of more
controlled systems with more effective inocula. These processes have to
be accelerated because of their more capital- and energy-intensive
nature (Kleijntjens and Luyben 1987). The first three pro ces ses, having
reactor cyde durations of 1-4 months, were implemented in the late
1980s and, early 1990s (Oolman et al. 1992; Coover et al. 1993). ExampIes 4 and 5 have cyde times of 10-14 days.
Plants and their rhizospheric microorganisms have been considered as
a candidate technology for hydrocarbon remediation (Radwan et al.
1998; Salt et al. 1998; Meagher 2000; Dietz and Schnoor 2001; Mejare and
Bulow 2001). Examples of some phytoremediation process are shown in
Table 5. Plant root exudates contain organic compounds which supply
carbon and nitrogen sources for microbial growth (Pletsch et al. 1999).
Densities of rhizospheric bacteria can be two to four orders of magnitude greater than the population in the surrounding soil (Cunningham
et al. 1995; Salt et al. 1998; Alkorta and Garbisu 2001) and enzymes may
be produced that degrade organic contaminants (Schnoor et al. 1995;
Boyajian and Carreira 1997; Macek et al. 2000). Rhizosphere environments of the wild desert plants, crop plants, alfalfa and various grass es,
and various legurnes have been investigated for hydrocarbon degradation potential (Dzantor et al. 2000; Magot et al. 2000; Suominenen et al.
2000; Yateem et al. 2000; Hou et al. 2001). Progress in the development
of this technology is described in Chapters 6 and 7.
Evaluation of Current Soil Bioremediation Technologies
Table 4. Bioremediation in slurry bioreactors
Example 4.1. French Limited Superfund Site at Crosby, Texas, USA (ENSR 1998)
• Perhaps the highest profile reactor-based study at a former waste disposal facility
containing 70 million gallons of petroleum wastes.
• Inoculum: The indigenous microfiora were used to promote hydrocarbon
• Aeration/mixing: A novel "MixFlo system" used pure oxygen.
• Performance!cycle time: In an ll-month treatment time 300,000 tons of tar-like
material and associated subsoil was remediated to criteria, with 85% destroyed in
122 days.
Example 4.2. Gulf Coast Refinery (Coover et al. 1993)
• A I-million gallon bioreactor-treated petroleum impounded sludges (l0% solids).
• Inoculum: Hydrocarbon-degrading culture from refinery wastewater-activated
• Aeration/mixing: Aeration achieved through fioat mounted mixer/aerators.
• Other operating parameters: Average temperature, 22.6 °C.
• Performance!cycle time: Overall PAHs removal of 90%, with a 50% reduction in oil
and grease in 80-90 days.
Example 4.3. Amoco Oil Company, Sugar Creek, MI, USA (AMOCO 1989)
• LST combined with land treatment to treat refinery sludges in a 5-million gallon
• Inoculum: Activated sludge and prepared hydrocarbon cultures.
• Aeration/mixing: Aeration supplied by a fioat-mounted aeration and mixing
• Performance!cycle time: The bioreactor was operated to reduce oil and grease by
66% «60 to 90 days) land application of solids to reduce residual PAHs to below
Example 4.4. The Petrozyme Process (Ward and Singh 2000; Singh et al. 2001)
• The process is in operation at 1.2 million liter scale for treatment of sludges
produced from about 75% ofVenezuela's refining capacity for the past 6 years.
Inoculum: For the first batch, the inoculum was a mixed culture, acdimated on
crude oil. Each subsequent batch was inoculated by culture carryover from the
previous batch.
Aeration/mixing: Sparged air lift aeration system with no mechanical mixing.
Other operating parameters: Nutrients and a surfactant addition, optimized to
maximize hydrocarbon accession, microbial growth rates and hydrocarbon
degradation; temperature control.
Performance!cycle time: By using a well-acdimated inoculum, optimized nutrients
and operating conditions, the process operates on a much shorter cyde time of
10-12 days with average biodegradation rates of 1% TPHs per day (initial TPH,
about 10%).
Example 4.5. PCP-contaminated soils using bioslurry reactors (Compeau et al. 1991)
Application in treatment of PCP-contaminated soil from a wood treatment
contaminated site.
• Inoculum: PCP-degrading organisms; little or no degradation without inoculation.
• Aeration/mixing: Achieved in sparged 2 x 25,000-1 stirred tank reactor.
• Other operating conditions: Soil solids content up to 40%.
• Performance!cycle time: 370mg/kg PCP was degraded to <O.5mg/kg in 14 days.
O.P. Ward and A. Singh
Table 5. Phytoremediation processes
Example 5.1. Phytoremediation of PAHs with grasses (Pradhan et al. 1998)
• Phytoremediation of PAH-contaminated soil using alfalfa (Medicago sativa), switch
grass (Panicum virgatum) and litde bluestem grass (Schizachyrium scoparium).
• Performance!cycle time: Reduction in total PAH concentration by 57% after 6
months of treatment.
Example 5.2. Phytoremediation of petroleum hydrocarbons (Nedunuri et al. 2000)
• Evaluation of phytoremediation for field-scale degradation of petroleum
hydrocarbon with rye grass and St. Augustine grass.
• Performance!cycle time: TPH reduction of 42 and 50% in 21 months by rye grass
and St. Augustine grass, respectively.
Example 5.3. Phytoremediation of PCBs with hairy root culture (Mackova et al. 1997)
• Ability of hairy root cultures of Armoracia rusticana, Solanum aciculare, Atropa
belldona and Solanum nigrum to degrade PCBs demonstrated.
• Performance!cycle time: Reduction of PCB concentration from 100 to 40ppm was
achieved in 30 days; efficient degradation of PCBs was associated with increased
peroxidase activity in the root system.
Example 5.4. Phytoextraction of heavy met als (McGrath et al. 1993)
• Field-based experiment on accumulation of heavy metals by Thlaspi caerulescens.
• Performance!cycle time: Accumulation of 2000-8000,LLgZng-1 dry weight in shoots;
total Zn uptake of 40kgha- 1 in a single growing season.
Phytoremediation is not a suitable method for the remediation of
high volume oily wastes. VOCs can be taken up by plants and transpired
to the atmosphere without transformation in a process known as phytovolatilization, which is not an acceptable environmental solution.
There is limited plant uptake of more hydrophobie and larger petroleum
components. Petroleum hydrocarbons, especially the smaller molecular
weight compounds, can be degraded in the rhizosphere under aerobic
conditions by rhizosphere bacteria, and promoted by leakage of exudates from the plants (Schnoor 2000).
Bioremediation has been successfully used to treat a wide range of soil
organic contaminants. In general, there is a need to make adjustments
to the contaminated medium, by supply of electron acceptor and/or
nutrients or co-substrates, to promote growth of the appropriate biodegrading microbial population. Greater control and optimization of the
microbial process require a greater level of intervention and usually
higher capital and perhaps material costs. This can result in the achieve-
Evaluation of Current SoH Bioremediation Technologies
ment of much higher degradation rates, greater process dependability
and facilitates a bioremediation option in situations where short remediation timelines are required. Sampling and analytical costs can be
high in in situ processes where process intervention is minimized.
Where the contaminated material is excavated and mixed during
biotreatment sampling and analytical costs are greatly reduced, compared to in situ treatments. This component is minimized in reactorbased systems where homogeneous mixing is the goal. The more
accelerated approach must to some extent offset the high er capital
outlays such that treatment costs per unit weight or volume of contaminated material are competitive.
While there is disagreement on the role of bio augmentation, this
approach may have merit in accelerated processes, especially where contaminants have a higher degree of recalcitrance and can benefit from
carefully selected and acclimated microbial species, especially when
inoculated at high count levels. While surfactants often inhibit
bio degradation and can be strongly sorbed to soil, reducing their effectiveness or generating a need for higher dose rates, they have been effective in accelerating bio degradation in bioreactor systems for oily sludge
treatment. The beneficial effects of stimulating bio degradation of recalcitrant compounds by the addition of solubilizing agents, such as more
refined hydrocarbons or vegetable oil, have been demonstrated with the
resulting promotion of contaminant accession and/or co-metabolism.
Advances in molecular biology and the impacts of genomics and proteomics will provide better insights into how more recalcitrant compounds may be biodegraded. This new knowledge may lead to the
creation of new strains expressing enzymes capable of high rates and
extents of catabolism of these molecules.
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Bioremediation of Petroleum
Hydrocarbon-Polluted Soils in Extreme
Temperature Environments
Rosa Margesin 1
Extreme environments are colonized by extremophilic organisms which
have specifically adapted to survive, grow and multiply under environmental extremes, such as extreme of temperature, pH, salinity, nutrients,
etc. Temperature plays a significant role in bioremediation processes' .
This chapter describes bioremediation treatments of cold soils and
desert soils contaminated with petroleum hydrocarbons.
Bioremediation in Cold Environments
A large part of the Earth's surface is characterized as having low temperatures. Vast land areas of the Arctic and Antarctic are permanently
frozen or are unfrozen for only a few weeks in summer. These environments are colonized by cold-adapted microorganisms able to grow at
O°C and even below. Psychrophiles (cold-Ioving) are defined as
organisms with a maximum growth temperature of 20°C or lower and
prevail in permanently cold habitats. Environments that are subject to
periodic, diurnal or seasonal temperature fluctuations are favourable to
cold-tolerant (facultatively psychrophilic) microorganisms; they grow
over a wide temperature range and restart their metabolic activity
rapidly after thawing. Cold-adapted microorganisms have evolved a
complex range of adaptations for all their cellular constituents. Adaptation strategies with regard to growth, enzyme production and enzyme
activity enable them to compensate for the negative effects of low temperatures on biochemical reactions (Margesin andSchinner 1999a;
Margesin et al. 2002).
1 Institute ofMicrobiology, UniversityofInnsbruck, Technikerstrasse 25,6020 Innsbruck,
Austria, e-mail: [email protected], Tel: +43-512-5076021, Fax: +43-512-5072929
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
R. Margesin
A wide range of metabolic activities have been described in cold
ecosystems, induding biodegradation of organic compounds such as
petroleum hydrocarbons. Special challenges to microorganisms in
hydrocarbon-contaminated cold soils indude reduced enzymatic reaction rates, increased viscosity of liquid hydrocarbons, reduced volatility
of toxic compounds, limited bioavailability of nutrients and contaminants, extremes in pR and salinity. Depending on the local conditions,
water activity mayaiso be limiting.
Until recently, frozen soils have been considered to be a practically
impermeable barrier to pollutants. Meanwhile, studies have confirmed
that hydrocarbons can penetrate into frozen soils, which may lead to
changes in the physical, mechanical and engineering soil properties.
Even ice-saturated soils are not an absolutely impermeable barrier for
oil penetration. The ability of oil to spread on the surface and to penetrate into the frozen soils depends on the characteristics of the soil and
the oil, and on temperature (Chuvilin et al. 2001). The time of the year
atwhich an oil spill occurs determines the extent of oil spreading on the
ground surface. A winter spill may cover a larger surface area (oil
spreads mainly on the surface of frozen soil and/or snow cover) than a
summer spill (the lateral distribution is reduced due to detention by the
vegetation cover). Freezing of oil-saturated soils causes aredistribution
of the oil components. In sandy soils, the oil concentrates in a thawed
zone in front of the freezing front, while in day soils the oil can accumulate in the frozen zone.
Arctic Soils
The term "arctic" implies a high latitude zone with a sparse and
specialized fauna and flora, a cold dimate with ice and frozen ground.
The subarctic zone covers the forest-tundra ecotone at high latitudes
(Löve 1970).
Limiting Factors tor Biodegradation
Organic decomposition is slow in arctic soils. Limiting factors for
microbial activity are low temperature and permafrost conditions, low
annual precipitation, low soil moisture, and low contents of available
nutrients (organic carbon, nitrogen, phosphorus) (Braddock et al. 1997;
Mohn and Stewart 2000). Permafrost in some arctic regions is continuous and extends from near the surface to depths of up to 300 m. There
is a short an nu al thaw season of about 1-3 months in summer, which
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
limits microbial processes like contaminant transport. Maximum soil
temperatures of 7 and 14 oe were reported at arctic and subarctic sites,
respectively (Walworth et al. 2001).
Hydrocarbon pollution is common in the Arctic, especially on
military sites. A number of studies have demonstrated the presence
of cold-adapted indigenous microorganisms able to degrade petroleum
hydrocarbons at low temperatures in arctic and subarctic soils (Turneo
and Guinn 1997; Whyte et al. 1999,2001; Soloway et al. 2001). Generally,
mineralization potentials and the population of hydrocarbon degraders
were higher in soils from contaminated sites than from reference sites,
and were stimulated by nutrient addition in laboratory treatments
(Braddock et al. 1997; Whyte et al. 1999,2001). Nitrogen is often the
major limiting factor for effective bioremediation but stimulation of
microbial numbers and activity was maximal in the presence of both
nitrogen and phosphorus, so attention has to be paid to the nutrient
concentration. Recently, the use of chitin as an inexpensive new agent
to accelerate oil bio degradation in the field and to reduce toxicity has
been proposed (Richmond et al. 2001).
Walworth et al. (2001) demonstrated the interaction of temperature
and nutrients using contaminated arctic (typical of gravel pads used for
construction on permafrost soils in the Arctic) and subarctic (typical of
soils from alluvial deposits in subarctic Alaska) soils. Both soils were
sandy and nutrient-deficient and contained very little organic carbon
(<1%). In both soils, a temperature increase from 1-5 to 20 oe and a
simultaneous nitrogen addition (50-100mg/kg) had a positive effect on
bio degradation and respiration, higher temperatures had a negative
effect. This effect of soil warming was even more pronounced in the
presence of nitrogen plus phosphorus, although the interaction between
temperature and phosphorus was lower than the interaction of temperature and nitrogen.
Biodegradation rates of crude oil in artificially contaminated (20 g/kg
soil) sandy soil were similar at 5 and 21 oe after an acclimation period
of 3 months (Gibb et al. 2001). Freeze-thaw cycles (alternating 24-h
periods at 7 and -5 Oe) had a stimulatory effect on hydrocarbon
biodegradation in fertilized diesel-fuel-contaminated Arctic tundra soil
in laboratory microcosms (Eriksson et al. 2001). The positive effect
of freeze-thaw cycles was explained by an increased hydrocarbon
bioavailability due to a change in the physical soil properties.
Bioaugmentation with non-indigenous microorganisms is allowed in
the Arctic, but not in Antarctica. Field and laboratory bio augmentation
using enrichment cultures from the contaminated site accelerated
bio degradation in unfertilized Arctic soils but not in fertilized soils
(Whyte et al. 1999; Thomassin-Lacroix et al. 2002), regardless of the
R. Margesin
initial density of the inoculum. The initially stimulating effect of
bio augmentation of soil treated in field biopiles was smaller than the
effect of fertilization (Mohn et al. 2001). The opposite result was
described after in situ bio augmentation of an artificially oilcontaminated tundra soil (Koronelli et al. 1997).
Field Bioremediation
In Situ Bioremediation
Most sites are remote, so the excavation for treatment in more temperate regions is too expensive. Furthermore, excavation is disruptive to
the immediate soils, permafrost and the surrounding habitat. Therefore,
in situ bioremediation is often preferred. Various treatment options
have been considered to enhance the low in situ bio degradation rates
in arctic soils (Table 1).
The Exxon Valdez oil spill (about 35,500 m 3 of Prudhoe Bay crude oil)
in Prince Williams Sound, Alaska, in 1989, resulted in the most expansive use of bioremedation yet undertaken. The fate of the oil has been
determined from the most complete and accurate mass balance of any
oil spill. The most significant term in the mass balance was about 50%
of the oil that biodegraded either in situ on beaches or in the water
column (Wolfe et al. 1994). Biodegradation rates were influenced by
environmental factors, such as the intensity of physical mixing, the
availability of nitrogen, phosphorus and alternative carbon sources, the
concentration of nitrogen in sediment pore waters, the oilloading and
the extent to which natural biodegradation had already taken place
(Bragg et al. 1994; Sugai et al. 1997). About 30% of hydrocarbons was
lost in a short time by physical weathering such as vaporization and dissolution (Bragg et al. 1994). About 2 months after the spill, the number
of oil-degrading microorganisms had significantly increased (byabout
four orders of magnitude) relative to uncontaminated sites, and hydrocarbon mineralization potentials were elevated (Braddock et al. 1996).
The application of an oleophilic nutrient-containing fertilizer to
surface-exposed oil enhanced bio degradation significantly, as measured
by both changes in oil composition and oil residue weights.
In situ treatment of a zone in the Canadian Arctic, contaminated for
more than 20 years with diesel fuel from a leaking undergound storage
tank, started in 1999. The area is located slightly above the permafrost
zone, with average temperatures of -23 to -14°C in winter and 8-20°C
in summer. The treatment included the use of biodegradable surfac-
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
Table 1. Field bioremediation treatments of oil-contaminated Arctic, Alpine and
Antarctic soils
Arctie soils
(mg TPH/kg soil)
1 Summer
2 Summers
65 days
Biopiles + TIS b 2 Summers
9 months
Land farming
Land farming
4 weeks
2 Summers
1 Summer
Antaretic soils
In situ
1 year
In situ
6 months
Land farming
5 weeks
( artificial
1 Summer
2 Summers
3 Summers
Mohn et al.
ThomassinLacroix et al.
Filler et al. (2001)
(250-11 ,000)
Land farming
Alpine soils
TPH a loss (%)
Tumeo and
Gawde (1997)
Piotrowski and
Aaserude (1992)
(0-10cm) Zytner et al.
(1O-30cm) (2001)
95 (0-3cm)
83 (3-6cm)
Kerry (1993)
DeHlle et al.
Tumeo and
Guinn (1997)
Margesin and
Schinner (2001a)
"Total petroleum hydrocarbons.
bThermal insulation system.
tants, micro- and macro-nutrients for biostimulation. These agents were
distributed, via biosparging, to promote bio degradation by increasing
dissolved oxygen and redox potential. After 1 year of treatment, oxygen
delivery had been identified as the major limiting factor. There was no
significant hydrocarbon removal, probably due to the increased partitioning of hydrocarbons to the aqueous phase as a result of surfactants
(Soloway et al. 2001).
R. Margesin
On-Site Bioremediation (Biopiles, Engineered Biopiles and Land Farming)
On-site bioremediation requires excavation, hut avoids the expensive
transport of material off site. Field experiments have indicated the
feasibility of on-site bioremediation of Arctic soils contaminated with
weathered diesel fuel, using small-scale biopiles (0.25-0.5m3; Table 1).
Biopiles in an Arctic desert (Canadian Forces Alert, Nunavut), with low
annual precipitation (mean 155 mm) were amended with phosphorus,
urea, surfactant and lyophilized inoculum enriched from the treated
soil. The contamination (2.9 g/kg soil) was reduced by 83% after 65 days
at an average daily temperature of 10-14 °C (Thomassin -Lacroix et al.
2002). Treatments of soil from two military radar sites on the Canadian
Arctic tundra ineluded N-P fertilization, addition of pe at for soil
bulking, inoculation with lyophilized hydrocarbon-degrading enrichment cultures (from other soils than those treated), and covering of
the biopiles with a elear plastic sheet which resulted in increased soil
temperatures (Mohn et al. 2001). At the end of the first summer season,
hydrocarbon loss was 72 and 85% in all amended biopiles. After 1 year,
hydrocarbons had completely, or almost completely (92%), disappeared
in amended biopiles compared to a loss of 64-66% in non-amended
controls, and nutrients had the greatest stimulatory effect.
According to laboratory studies, the rate of hydrocarbon biodegradation in Arctic soils can be increased by raising the temperature in
the range of 1O-20°C, without negatively affecting the indigenous soil
population. The bioavailability of contaminants that are less soluble
at low temperatures can be increased. The availability of excess energy
capacity is required for such a treatment. Thermal insulation systems
have recently been developed for innovative bioremediation efforts in
cold regions (Filler and Carlson 2000). A thermally enhanced biopile,
based on the combination of bioventing with active warming, fertilization and power cyeling, was installed at a former oil-industry working
camp in Prudhoe Bay, Alaska, in the summer of 1999 (Filler et al. 2001).
Contaminated tundra soil mixed with gravel was used to form a biopile
with an average contamination of 2 g diesellkg soil. The combination of
thermal insulation, an engineered heat mat and cost-effective power
utility optimization extended the period of annual treatment. Temperatures of 1.5-6.5 and 0.5-7.8 °C were realized during the first and second
season of bioremediation, respectively. Numbers of heterotrophie and
oil-degrading microorganisms increased in situ and were active during
the warming period. Successful bioremediation, to below Alaska
eleanup standards for gravel pads in the Arctic (500 mg/kg) , was
obtained after two seasons with a residual diesel contamination of
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
39-397 mg/kg and significantly reduced contents of aromatic
The application of land farming was studied in arctic Alaska where
freezing temperatures occur from late September through early June.
Tumeo and Gawde (1997) used multiple plastic-lined cens filled with
sandy soil and covered with plastic. After about 9 months, only 34% of
the contamination (1.3 g diesellkg soil) was lost in nutrient-amended
soil, whereas 26% was lost in the untreated control. The effect of surfactants was higher in the presence (49-81% loss) than in the absence
(46-54% loss) of nutrients. Most of the hydrocarbon removal had
occurred during the first 6-8 weeks. Little or no loss was recorded in
the winter. Piotrowski and Aaserude (1992) excavated soil fill, contaminated (11.5 g TPH/kg soil) from a 5-year-old diesel fuel spill without
damaging the underlying permafrost. Treatment induded fertilization
and tilling. TPH loss was 44% after 4 weeks due to biodegradation,
volatilization and leaching. An average reduction in diesel fuel of 75%
was measured after the second summer period. The land-farming treatment of contaminated silty day loam (18g TPH/kg soil) induded the
addition of peat, lime and a slow release urea-based nitrogen fertilizer,
and tilling to mix the additives and provide aeration. After three
summer months, the largest loss (58%) occurred in the top lOcm, while
onlya 10% loss was found in lower soil depths (10-20 and 20-30cm)
due to reduced oxygen levels. Biodegradation was the dominating
process, hydrocarbon volatilization was low, and leaching was negligible (Zytner et al. 2001).
Alpine Soils
The term "alpine" is generally accepted as a term for a high-altitude belt
(for example, about 1800-2500m above sea level in Europe) above continuous forests on mountains in Europe, USA, South America and Asia.
The subalpine belt has been defined as covering the forest-tundra
ecotone at high altitudes (Löve 1970).
Limiting Factors tor Biodegradation
Alpine soils are subjected to large temperature fiuctuations (air
temperatures vary from -5 to more than 20 Oe), high precipitation
(2000-3000 mm/year) , and regular freeze-thaw events. The alpine
microbial communities may differ from those in Arctic and Antarctic
habitats because strong valley winds from boreal and Mediterranean
R. Margesin
landscapes continuously transport microorganisms to the Alps. Soil
microbial biomass in alpine soils was unaffected by freeze-thaw events
(Lipson et al. 2000).
The characterization of 12 oil-contaminated soils (0.4-30 g TPH/kg
soil) near diesel storage tanks, petrol stations or garages, and 8 corresponding pristine soils sampled from various alpine sites in Tyrol,
Austria, showed that most of the soils had a low organic matter content
«2%), while pH values ranged from 4.8 to 9.2. Significant microbial heterotrophic and oil-degrading, cold-adapted populations were present in
all of the soils, independent of the contamination level (Margesin et al.
2003). The enrichment of cold-adapted oil degraders from 29 habitats,
mainly soils and glaciers at high altitudes, demonstrated that both
oil-contaminated samples and samples from uncontaminated sites
contained efficient oil degraders (Margesin and Schinner 1998).
A number of studies, on the potential of bioremediation for the treatment of diesel oil-contaminated European alpine soils, demonstrated
the important role of indigenous soil microorganisms. Enrichments of
oil degraders occurred soon after contamination. The diesel decontamination of the investigated soils at 10 oe was comparable to that reported
at 25-30 oe with soils from a temperate climate. eomparable decontamination results could be obtained with experimentally and chronically
contaminated alpine soils in short-term laboratory studies (for a review,
see Margesin 2000). Long-term studies, however, showed big differences
depending on the age of the contamination. While only 10% of the
initial contamination of two artificially contaminated alpine soils was
detected after 150 days at lOoe, oil-polluted soil sampies from a tank
farm showed only areduction of 20-49 and 23-66%, without and with
fertilization, respectively, under the same time and temperature conditions. Weathered contamination may be characterized by possible
threshold concentrations for bio degradation, possibly due to the presence of recalcitrant compounds and very limited bioavailability (Allard
and Neilson 1997). Abiotic processes, such as evaporation, transformation and adsorption, played an important role only in the decontamination of an artificial pollution.
Biostimulation of the indigenous soil microorganisms with inorganic
N-P-K nutrients resulted in a significant increase in oil biodegradation
in artificially contaminated soils, but the effect of bio stimulation
declined with time. However, fertilization had no significant effect on
TPH loss in seven soil samples from a former tank farm, containing
0.3-10.5 g/kg soil (mean contamination of 4.7 g/kg), despite the nutrient
deficiency of the soil. Natural attenuation (abiotic loss plus biodegradation by indigenous soil microorganisms) played a major role. After 5
months at 10 oe (this temperature corresponded to the mean annual
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
environmental temperature of the area and to the groundwater temperature), a TPH loss of 20-49 and 23-66% was found in unfertilized
and fertilized soil sampies, respectively. There was also a negative effect
of fertilization on the number of cold-adapted hydrocarbon degraders
and on soil respiration, which was explained by the oligotrophic nature
of the soil microorganisms (Margesin and Schinner 1999b).
Biostimulation enhanced bio degradation to a significantly greater
degree than bio augmentation with cold-adapted oil-degrading inocula.
The use of surfactants, to increase mobility and surface area available
for microbial cell contact with hydrocarbons, inhibited hydrocarbon
bio degradation, which could not be attributed to residual surfactant
pollution or inhibition of microbial activity (Margesin 2000).
Field Bioremediation
Oil pollution in ski res orts is caused by the use of motor vehicles for
the preparation of ski runs and also by leaks and storage tank ruptures. Bioremediation may be the logistically and economically most
favourable solution, as the use of conventional techniques requires
costly excavation. The feasibility of bioremediation as a treatment
option for a diesel oil-polluted soil in a Tyrolean glacier area was determined during a 3-year field study (Margesin and Schinner 2001a). The
mean annual air temperatures ranged from 0.6 to -1.8 oe and annual
levels of precipitation were 1250-1800mm. The short annual thaw
season is between the end of June and September. Summer temperatures vary greatly from near freezing to above 20 oe at the soH surface.
The soil collected from the contaminated zone (2.6 g TPH/kg soH) in the
motor pool area was an alkaline (pH 8) mixture of carbonaceous gravel
and sand, and contained low levels of nutrients. To examine the
efficiency of natural attenuation and bio stimulation, soil in mesocosms
remained untreated or was amended with an inorganic slow-release Np-K fertilizer. Most of the hydrocarbon loss occurred during the first
summer season (42% loss) with fertilization, and during the second
summer season (41 % loss) without fertilization. At the end of the third
summer season, the initial contamination was reduced by 70 and 50%
in the fertilized and unfertilized soil, respectively, resulting in a still high
residual contamination of 1.3 and 0.77 g TPH/kg soil. Despite repeated
nutrient amendment the bio stimulation effect decreased with consecutive treatments. Nonetheless, the initial fertilization treatment was
appropriate in terms of accelerated bio degradation.
In the fertilized soil, all biological parameters measured at regular
intervals during the study (numbers of culturable microorganisms, soil
R. Margesin
respiration, soil enzyme activities) had increased significantly and correlated weIl with each other and with the residual hydrocarbon concentration, indicating the importance of biodegradation. The positive
correlation of the available nutrient contents (N, P), with the hydrocarbon content, the microbial counts, and the activities in the fertilized soil,
indicated the relevance of nutrients. Microbial activities in the unfertilized soil fluctuated around background levels and did not correlate with
the residual contamination, which led to the conclusion that a considerable part of the decontamination had to be attributed to nonbiological processes.
Antarctic Soils
Antarctica is the coldest, driest and most barren continent on Earth. The
soils of the ice-free areas of Antarctica are formed under the most
extreme conditions. Due to the extremely slow rate at which soil
processes operate,Antarctic soils are particularly susceptible to humaninduced damage. The interest in exploring Antarctica is increasing and
with it the volume of petroleum transported. Terrestriallife is concentrated in ice-free regions along the coastline. Most of these regions are
located in the Ross Sea region and on the Antarctic Peninsula, where
many scientific stations are based and field operations such as storage
and refueling of aircraft and vehicles, can result in oil spills. Oil spills
have been reported at most major bases, resulting in TPH concentrations up to 48 g/kg soil (Aislabie et al. 1999,2001; Balks et al. 2001; Snape
et al. 2001).
Chemical characterization of hydrocarbon contaminants in Antarctic and subantarctic soils indicates slow weathering of aliphatic and
aromatic compounds (Aislabie et al. 1998, 1999; Delille and Pelletier
2002). Hydrocarbon contamination in the Ross Sea region was limited
to surface soils and was not detected below 27 cm at one site, whereas,
at another site, high contamination levels were detected in the subsurface but not in the surface (Aislabie et al. 2001). Polycyclic aromatic
hydrocarbons (PAHs) were found at lower depths in contaminated soils
(Aislabie et al. 1999).
Limiting Factors tor Biodegradation
Antarctic terrestrial ecosystems differ from those in the Arctic as they
are colder, drier (the moisture content of the active layer of Antarctic
soils is typicaIly between 1 and 10%), have lower levels of available
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
nutrients, and are often alkaHne with pH values up to 9.9 (Kerry 1993;
Wardelll995; Aislabie et al. 1998,2001; Balks et al. 2001). Antarctic soils
experience large temperature fluctuations with summer freeze-thaw
cycles and desiccation. Summer soil surface temperatures ranging from
15-18 to -5 oe and even to -35 oe have been reported at sites in the Ross
Sea region. In summer, the soil surface temperature of soils, contaminated for more than 30 years, was often high er (by up to 7 Oe) than that
of the adjacent control site. This was evident at least to 20 cm depth.
This temperature increase could be the result of decreased surface
albedo caused by a dark surface coating due to the presence of oil, which
increases the soil capacity to absorb heat (Balks et al. 2001). The
increased temperature in the soil may lead to an increase in the depth
to permafrost and lower, or even to the destruction of the glaciallayer
at a historically contaminated site (Aislabie et al. 2001). Soils sampies,
from five sites along the Antarctic east co ast, were characterized by a
poor microflora (bacteria, fungi, and algae). Indicators of vital and
enzymatic activities showed the existence of potentials for mineraHzation and biosysnthesis, but the environmental conditions do not permit
the humification of organic material (Negoita et al. 2001).
Methods for remediation of oil-polluted soil are limited in this
environment. Technological and logistical problems, as well as high
expenses associated with excavation and removal, indicate that in situ
treatment of petroleum contaminants is probably the only viable management strategy (Snape et al. 2001). In situ bioremediation is based on
the presence of oil degraders that are able to utilize the contaminants
as a carbon source. Therefore, it is important to know the biodegradative potential of the indigenous microorganisms, also because the
Antarctic Treaty prohibits the introduction of foreign organisms.
Furthermore, it is necessary to understand the factors that control
microbial activity at a specific site.
A number of studies have focused on the characterization of the
microbial community present in contaminated Antarctic and subantarctic soils (Kerry 1993; Wardell 1995; Tumeo and Guinn 1997;
Aislabie et al. 1998, 2001; Delille 2000; DeHlle et al. 2002). All these
studies demonstrated the ability of the indigenous microbial population
to adapt and utilize the contaminants. Numbers and diversity of culturable microorganisms were found to vary depending on the level of
contamination. The number of hydrocarbon degraders were generally
elevated, often by several (up to 7) orders of magnitude, in contaminated
as compared with control soils, while numbers of heterotrophie bacteria were enhanced only in some contaminated soils (Aislabie et al. 1998,
2001; DeHlle and Pelletier 2002). Significant numbers of oil degraders
were also detected below lOcm in oil-impacted soils, although total
R. Margesin
microbial counts decreased generally down the soil profile. The dominant speeies were bacteria. The inereasing number of yeasts and
filamentous fungi in eontaminated soils may indicate their role in the
degradation ofhydrocarbons or their metabolites (Aislabie et al. 2001).
The decrease in soil nitrate levels in contaminated soils was taken
as indireet evidence of the in situ aetivity of indigenous oil degraders
(Aislabie et al. 2001). Recently, free-living heterotrophie nitrogen-fixing
bacteria could be isolated from fuel-contaminated Antaretic soils
(Eckford et al. 2002). Heterotrophie nitrogen fixation may be important
for the natural attenuation of Antaretie fuel spills by providing nitrogen
to hydroearbon-degrading microorganisms, thereby indireedy enhaneing bioremediation.
Field Bioremediation
Litde is known about the impact of hydroearbons on land, whereas a
number of studies have investigated the effect of oil spills on Antarctic
marine eeosystems. Unfortunately, little information is available for the
treatment of historically contaminated Antaretie soils.
In Situ Bioremediation
Field studies, on the potential application of in situ bioremediation
in Antarctic soils, have been carried out in pristine regions that were
intentionally eontaminated to simulate oil spills (Table 1).
Kerry (l993) applied Antarctie Blend Distillate, the fuel used on
Australian stations, to field plots to produee an artificial moderate oil
spill (511m 2 , >20 g/kg soil). The pristine soil was coarse, saline mineral
sand with an extremely low water-holding capacity. Bioremediation
treatments inc1uded N-P-K fertilization and proeedures aimed at water
retention use of xanthan gum and plastic covering. After 1 year, a
signifieant part of the decontamination had to be attributed to natural
attenuation in both surface and subsurface soils. Surfaee control soils
(0-3 cm) had a residual hydrocarbon concentration of 4.3 g/kg soil,
while fertilized plots contained only 1.1 g/kg. Microbial activity was
generally low, but highest in fertilized plots. Higher temperatures were
reeorded under the plastie eover but had no stimulating effect on
biodegradation. Subsurfaee (3-6em depth) control soils contained 3.4g
hydroearbons/kg soil after 1 year, and none of the tested treatments
resulted in a significantly higher decontamination. Temperatures lower
than those in the surface soils and the retention of toxic volatile
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
compounds were considered as the major limiting factors, besides the
low oxygen concentration.
Ornithogenic (nutrient-rich organic) soil from the Antarctic penguin
rookeries was artificially contaminated with Arabian crude oil or diesel
fuel (Delille 2000). Mean monthly air temperature in this area rises
above freezing only during December-January and reaches -20°C in
July-August. After 9 months, no significant biodegradation had
occurred. Crude oil induced an initial increase in both heterotrophie
and oil-degrading bacteria, while diesel caused a decrease in heterotrophs and an increase in oil degraders. A general decrease in
microbial numbers was observed after 9 months.
Field bioremediation in Antarctic intertidal sediments was studied by
Delille et al. (2002). Light Arabian crude oil was intentionally released
(2.851/m 2 ) along a subantarctic sandy beach. Sediment temperatures
fluctuated considerably and ranged from 18 in summer to 2°C in winter.
The treatments of the fine and very fine sand included the addition of
a slow release oleophilic fertilizer and three different fish composts.
After 6 months, nearly complete degradation of aliphatic hydrocarbons
was observed in all plots including untreated plots. Biodegradation was
faster in treated than in untreated plots. The addition of fish composts
resulted in a higher acceleration of bio degradation than the oleophilic
fertilizer. There was a low correlation between hydrocarbon -degrading
bacterial abundance and bio degradation, but there was a clear correlation between the microbial production (mean daily growth rate of
hydrocarbon degraders) and hydrocarbon consumption (apparent
degradation rate). Nonetheless, high toxicity levels were noted in all
sand sampies, regardless of the treatment, even 300 days after the contamination event, whereas no toxicity was detected in the interstitial
On-Site Bioremediation
Tumeo and Guinn (1997) recommended ex situ bioremediation because
this treatment allows greater control over limiting environmental variables. Land farming retains the benefits of simplicity and low cost. The
results obtained from land farming during the Antarctic summer
demonstrated hydrocarbon biodegradation, but the reduction in contamination was well below that required to meet regulatory standards
(Turneo and Gawde 1997). Treatments involved the addition of inorganic N-P-K nutrients, wastewater as an alternative nutrient source and
commercial surfactants. After 5 weeks, the contamination (3.6g/kg soil
from jet fuel spills) was reduced by 43% in the nutrient-amended soil,
R. Margesin
and by 17-53% in soil treated with surfactants. Apparent increases in
hydrocarbon concentration were correlated with a drop in temperature
to below freezing. It was most likely that the contamination was
excluded from a freezing front that was moving from the bottom to the
top as the ambient temperature dropped.
Bioremediation of Desert Soils After the Gulf War
As a consequence of one of the most massive oil spills (estimated
0.95-1.27 Mm3 ) during the Gulf war in 1991, large parts of the western
coast of the Arabian Gulf and the Kuwaiti desert were contaminated
with crude oil from over 700 damaged oil weHs. Most of the crude oil
has been pumped out of the oil lakes, but the bottom sand remains
heavily polluted to depths of 30cm or more (Al-Daher et al. 1998).
Attention has focused on the consequences of hydrocarbon pollution
in this environment. The depth of oil penetration depended on the soil
characteristics. The downward oil movement and salt accumulation
were restricted to the upper (25-95 cm) layer in soil sampies from one
site, whereas the spreading over greater depths (I50-270cm) has
occurred in another soil profile (Massoud et aL 2000). Saeed et al. (I998)
demonstrated the effect of weathering on the chemical composition of
oil in Kuwait oil lakes. After 4 years, the asphaltene content of the oil
had increased significantly, especially in lakes that had been dried out.
Saturates had increased, while the aromatic fraction had decreased. The
concentration of resins and of high er PAHs continued to increase.
Ten years after the oil spills, the sulfate concentrations in the surface
layers of two soils were relatively small compared with the large
amounts of sulfur incorporated in the soils. erude oil from Kuwait contains high levels of organic sulfur (2.5-4.2% w/w). A layer of soil aggregates coated with oil and mixed with tarry residue covers the surface of
soils in the Greater Burgan oil field in Kuwait. As a consequence, the
oxygen content of soil air is reduced and an aerobic conditions are
created, which may result in the formation of sulfide and hydrogen
sulfide from sulfate. Oxidation of these reduced sulfur compounds
results in the production of sulfuric acid and sulfur. This continuous
release of sulfur in acidic form, in the long term, may convert presently
alkaline soils (pH 7.8) into acidic soils (Massoud et al. 2000).
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
Limiting Faetors for Biodegradation
The Arabian Gulf is situated in a semi-arid region, eharaeterized by
high evaporation rates (200em/year) and high salinities (4-7%). The
atmospherie temperature frequently exeeeds 50 oe in summer, and
the surfaee-soil temperature is even higher (Sorkhoh et al. 1993). Soil
temperatures above 50 oe prevail in Kuwait over aperiod of 5 months
(Obuekwe et al. 2001). The desert soils are alkaline (pH 7.8-8.6) and
eontain aeeumulated salt in the surfaee layer. The partly extremely high
salinity is attributed to the large volume of seawater used to extinguish
the oil well fires, and the evaporation of this water (Balba et al. 1998).
The biologieal deeontamination of hydroearbon pollution in such an
environment requires the presenee and aetivity of indigenous mieroorganisms able to degrade hydroearbons under the prevailing eonditions.
Elevated temperatures have a signifieant influenee on the bioavailability and solubility of organie eompounds. Diffusivity, desorption, mobilization and re action rates are inereased and viseosity is deereased.
A number of studies have shown the presenee of oil-degrading thermophilie mieroorganisms in eontaminated desert soils. Thermophiles
are designated as mieroorganisms that grow optimally above 40 oe.
Desert soil from an oilfield eontained more mesophilie than thermotolerant hydroearbon-utilizing baeteria (Al-Daher et a1.1998). No thermophilie oil-utilizing fungi were found (Sorkhoh et al. 1993). All of
more than 300 bacterial erude-oil utilizing isolates belonged to the
genus Bacillus, the species B. stearothermophilus being predominant.
Two strains degraded about 80-89% of erude oil (5 g/l) within 5 days at
their optimum growth temperature of 60 oe, whereas bio degradation
was three- to four-fold lower at 40 oe. Both strains utilized only middleehain n-alkanes. AI-Maghrabi et al. (1999) isolated two thermophilie
Bacillus strains that degrade erude oil (20% vIv) at 40-45 oe, one of
whieh eould grow with erude oil even at 80 oe. The isolates survived
under saline eonditions typieal of seawater and reservoir eonnate water.
Optimal growth in the presenee of 10% vIv erude oil and 2% salinity
was observed at 60-80 oe.
Obuekwe et al. (2001) monitored the presenee of thermophilie
mieroorganisms (55 Oe), and their hydroearbon degradation potential
in desert soil sampies from the Burgan South oil field of Kuwait over
1 year. Numbers of heterotrophie baeteria were highest from February
through April, the cool and wet period of the year, and deereased by two
orders of magnitude in February and November. Thermophilie fungi
were not deteeted, while aetinomyeetes were present oeeasionally. erude
R. Margesin
oil degradation potential did not vary significantly in sampies collected
in winter (November-March) and summer (April-October). Bacteria
were better degraders of crude oil than fungi, especially in the dry, hot
months of the year. Middle-(C I6 ) and long-chain (C32 ) n-alkanes were
utilized effectively, but the capacity to utilize phenanthrene was low.
Field Bioremediation
Various bioremediation treatments in the field, including bioventing,
land farming and composting, have been used to decontaminate oilpolluted desert soils (Margesin and Schinner 2001b). These studies
demonstrated the essential role of moisture and fertilization and the
significance of indigenous oil-degrading bacteria compared to inoculation. An important aspect is irrigation. Saltwater accelerates the soil
water evaporation; as a consequence salt remains and accumulates at
the soil surface, which inhibits biodegradation. The use oflarge volumes
of freshwater, in bioremediation of desert soils, leads to the leaching of
salts and the reduction of soil salinity (Balba et al. 1998).
Balba et al. (1998) compared the efficiencies of land farming,
windrow composting and bioventing in an oil lake area. The soil was
amended with N, P, compost and wood chips. Land farming involved
the treatment of 400 m 3 soil, using regular tilling and soil irrigation with
fresh water to obtain 8-10% soil water content. Windrow composting
piles were used to treat 1l00m3 soil, had a continuous supply of water
and nutrients, and were turned over monthly. Static bioventing piles,
used for the treatment of 500 m 3 soil, had a continuous supply of air,
water and nutrients. At the beginning, the soil scraped from the surface
layer contained 133-694g TPH/kg soil and was extremely salty (about
30-40% w/w NaCl). Contamination and salinity were gradually reduced
with soil depth, and reached levels of 0.2-0.7 g TPH/kg soil and
0.03-0.3% w/w NaCl. After 1 year, a significant contaminant reduction
was detected in all treatments. The TPH content was reduced by 64%
(static bioventing piles), 74% (windrow piles) and 83% (land farming).
The loss of total alkanes was 91 % (land farming) and 82% (other two
treatments), compared to a loss of 16-26% in untreated controls.
Despite the lower bioremediation efficiency of bioventing piles, this
technique requires considerably lower operation and maintenance costs
than other techniques; it also requires a much smaller site area, and less
water for irrigation.
Al-Daher et al. (1998) tested composting as a treatment option for
oil-polluted (33-55 g/kg soil), N-P-K-fertilized desert soil. The applied
windrow soil pile system, subjected to regular irrigation and turnover,
Bioremediation of Petroleum Hydrocarbon-Polluted Soils
resulted in a significant reduction of up to 60% of the contamination
within 8 months. The PAH concentration was reduced from 11-16 to
5-6 mg/kg soil, but PAHs with five or more rings resisted
Phytoremediation could be a feasible approach for bioremediation of
oil-polluted desert soil. The rhizosphere of wild desert plants and crop
plants contained more hydrocarbon-utilizing bacteria than the general
soil. This effect was much more pronounced for plants growing in
oil-polluted than in uncontaminated soil (Radwan et al. 1998).
Bioremediation, the process whereby natural bio degradation rates are
accelerated through stimulation of the indigenous microorganisms,
is an effective, ecologically and economically acceptable reclamation
alternative. Limiting environmental factors need to be overcome if
microbial breakdown of contaminants is to be used effectively. Sitespecific (soil properties, geology, hydrogeology, geochemistry) and
contaminant -specific (composition, concentration, age, and bioavailability of hydrocarbons) factors must be considered.
Limiting factors for in situ bio degradation in polar soils include
reduced enzymatic reaction rates, increased viscosity of liquid hydrocarbons, reduced volatility of toxie compounds, and limited bioavailability of nutrients and contaminants. Antarctic soils differ from those
in the Arctic as they are colder, drier, lower in available nutrients,
and often alkaline. High numbers of culturable hydrocarbon-degrading
microorganisms, the prevalence of genotypes involved in the degradation of representative fractions of petroleum hydrocarbons, and high
hydrocarbon mineralization potentials in contaminated soils provide
evidence for the bio degradation potential of indigenous soH microorganisms. In situ bioremediation is often preferred because many contaminated sites are remote. On-site treatments include land farming and
biopiles and thermal insulation systems have recently been developed.
Depending on the local conditions, natural attenuation may playa major
role. Biodegradation of many contaminants may be accelerated by
nutrient and oxygen supply, however, there is a tendency for the nutrient biostimulating effect to decline with time.
Various field bioremediation approaches, including land farming,
bioventing and composting, have been tested to treat desert soils contaminated from oil spills. These studies demonstrate the significant role
played by indigenous thermophilie and thermotolerant hydrocarbon
degraders and the essential roles of irrigation and fertilization.
R. Margesin
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Innovative Methods in Biofiltration
of Air Contaminants
Zarook Shareefdeen I
Biofiltration is an effective, economical and environmentally benign
technology for treating contaminants from industrial airstreams. It is
based on the ability of microorganisms to oxidize (degrade) toxic and
odorous air contaminants. The contaminants treated in a bio filter are
biodegradable volatile organic compounds (VOCs) or inorganic odiferous compounds. These compounds are released from various industrial
sources such as wastewater and petrochemical industries, rendering
plants, composting facilities, coating industries, and food processing
industries. These chemicals are emitted in large amounts and create
hazards to the ecosystem and to human health. The VOCs, many of
which are carcinogens, are also precursors to smog formation. Inorganic contaminants (hydrogen sulfide, ammonia) and reduced sulfur
compounds (methyl mercaptan, dimethyl sulfide and dimethyl
disulfide), prevalent in municipal wastewater collection and treatment
systems, are highly odorous, and thus create a nuisance in the environment. The regulatory agencies have set strict controls on emission of
VOCs and odorous gases. The United States Environmental Protection
Agency (USEPA) enforces strict controls on more than 180 hazardous
air pollutants (Guilbault 2002). The market for the removal of odor and
VOCs is driven by regulatory issues, generally enforced as a result of
public complaints about poor local air quality and through emissions
monitoring by the enforcement agencies. International protocols, such
as those for the reduction of greenhouse gases, are resulting in increased
regulations and enforcement of the regulations (Shareefdeen et al.
1 BIOREM Technologies Inc., 7496 Wellington Road 34, R.R. #3, Guelph, Ontario
N1H 6H9, Canada, e-mail: [email protected], Tel: +1-519-7679100, Fax:
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
Z. Shareefdeen
Many soil organic contaminants are volatile. Indeed, in soil bioremediation processes large quantities of VOCs may get transferred to the
atmosphere, thereby creating hazards to human health and contributing to tropospheric ozone production. It has previously been noted that
volatilization of organics occurs during natural attenuation and landfarming processes (see Chap. 9). In addition, the efftuent air streams
from bioventing, biosparging and engineered soil pile systems are often
heavily contaminated with VOCs. In order to prevent atmospheric pollution from these sources surface containment strategies are required
together with processes for decontamination of these VOC-containing
air streams. Air bio filtration provides a cost-effective biological remedy
to this problem and, as such, represents an important process component in many soil bioremediation systems.
In recent years, bio filtration has received increased acceptance for
odor removal in the wastewater treatment industry. Increasingly, other
industrial sectors including rendering, food processing, petrochemical
industries, fiavor manufacturers, and composting facilities, are also
selecting bio filter systems for odor and VOC rem oval in their facilities.
Devinny et al. (1999) reported a projection of a 100 million dollar
biofilter technology market for the year 2000. In the last 2 years, the
bio filter market has grown rapidly. It is estimated that a 1 billion dollar
market for odor and VOC treatment will develop in the next 5 to 7 years
made up of: rendering plants, $75 million; pet food plants, $75 million;
food processing, $300 million; sewage sludge treatment, $400 million;
composting, $50 million and industrial wastewater, $100 million. Many
of these markets are in their early growth stages and will continue to
increase in size. In this chapter classical and emerging technologies in
odor and VOC control methods will be introduced and recent advances
in bio filtration technology will be discussed.
Non-biological Methods
Treatment of off-gases has been practiced for years and is primarily
based on non-biological methods such as condensation, activated
carbon adsorption, absorption/scrubbing and incineration. In condensation processes, cooling and compression condense contaminant
vapors from air. This process is economical for higher boiling point
compounds and more concentrated vapors (Devinny et al. 1999). In
adsorption process, pollutants are adsorbed onto adsorbents such as
activated carbon or zeolite. As with bio filtration, this process is effective
when the concentration of contaminants is low. Once the adsorption
Innovative Methods in Biofiltration of Air Contaminants
bed is saturated, regeneration of the adsorbents can be done using
steam or hot air. Often, the recovery of compounds is costly and thus
spent adsorbent becomes solid waste, which needs to be landfilled or
incinerated. In absorption/scrubbing processes, pollutants from the air
are absorbed into a scrubbing solution such as water or chemicals. By
changing the pH of the scrubbing solution, the mass transfer rate of contaminants from the air to water phase can be adjusted. Often chemical
costs are high and the generated liquid waste needs treatment. Frequently, nozzle maintenance, complex feed systems and high operating
costs are problems associated with the absorption method.
Incineration technology has been widely used due to its high efficiency. This process involves combustion of contaminants at a high
temperature. To increase the efficiency of incineration and to reduce
fuel requirement for combustion, several forms of this technology, such
as recuperative, regenerative and catalytic oxidation, are practiced.
Although incineration leads to the complete destruction of contaminants, it is an expensive method due to its high energy and fuel requirements. Incineration is not economical if contaminant concentration
levels are low and large airflow volumes need treatment. This process
also pro duces the highest amounts of CO 2 and NOn which contribute to
greenhouse gases.
Biological Methods
Biological treatment provides significant economical advantages over
conventional methods. It is effective and economical for low concentrations of biodegradable contaminants and for large flow volumes of
air. In general, highly soluble and low molecular weight compounds
such as methanol, ethanol, aldehydes, acetates, ketones and some
aromatic hydrocarbons are easily biodegraded. Inorganic compounds,
such as hydrogen sulfide and ammonia, are also easily biodegraded
in bio filters. Low molecular weight aliphatic hydrocarbons such as
methane, pentane and some chlorinated compounds are more difficult
to be degraded in bio filters.
Biofiltration technology is an adaptation of the process by which
the atmosphere is cleaned naturally. Plants and soil adsorb VOCs
and odorous contaminants from the atmosphere, and degrade them.
Inefficient contact of soils and plants with VOCs, which are in the
atmosphere, leads to relatively low reaction rates. Biofiltration provides
maximal contact and allows sufficient time for VOCs to react (Bohn
1992). There are basically three types of biofilters: classical biofilters,
Z. Shareefdeen
biological trickling filters and bioscrubbers. Recently, novel bio filters
including rotating drum biofilters, horizontal fiow biofilters, foamed
emulsion bioreactors, short contact time biotrickling filters (Gabriel et
al. 2002), higher plant-based bio filters (Guilbault 2002) and microwave
concentrator/biofilter integrated systems (Webster et al. 2002) have
been introduced. These bioreactors are in the early stages of development and several pilot -scale studies are in progress.
Classical Biofilters
In classical bio filters, as the contaminated air is passed through a bed
of media, the contaminants and oxygen are first transferred to the
biofilms formed on the surface of the media particles and then metabolized by bacteria. In order to sustain bacterial growth on the media
particles moisture is provided by saturating the process air before it
enters the biofilter unit, and by occasional irrigation. The medium
within a bio filter is normally composed of material such as peat, wood
bark, soil, compost, coated ceramic particles, synthetically manufactured media, or a combination of these products. If the empty bed
residence time (EBRT) is large enough, complete removal of the
contaminants in the bio filter can be achieved.
Media selection for a bio filter is based on the ability to support bacterial growth and adsorption of contaminants. Compost-based medium
is cheap and has diversified microbial communities capable of degrading many pollutants (Torkian et al. 2002). Naturally occurring media,
such as peat and compost, contain organisms capable of biodegrading
some VOCs. In some applications, activated sludge suspensions from
sewage treatment plants are used as inocula (Ottengraf 1990). Limiting
constraints ofbiofilter applications in the past have been the large space
requirements (or high EBRT) and frequent media replacements due to
media deterioration or failure.
For effective performance, bio filter media must provide an excellent
environment for bacterial growth while maintaining large reactive
surface, good absorption capacity, adsorption property, pH buffering
capacity, low pressure drop, good pore structure, very low compaction
over time and low residence time (Leson and Winer 1991; William 1992).
Although a selection of media should be based on all of these parameters, frequently media with good biodegradation properties (i.e. peat,
compost, soil, chicken manure) are selected without giving consideration to structural, mass transfer characteristics or adsorption
Innovative Methods in Biofiltration of Air Contaminants
Poor performances were generally attributed to biofilter technology
rather than the improper media; hence, the reputation of this technology has suffered for years. Some of these issues have been resolved
with the introduction of inorganie media such as BIOSORBENS
(Shareefdeen et al. 2002a, b). This medium consists of hydrophilie
mineral cores, which are coated with sorption materials, with a very
high specific surface area for efficient adsorption that compliments the
biologieal oxidation process in removing odors and VOCs. In addition,
the coatings include nutrient-rich organic material for microbe hosting
and suitable binders to provide product stability. Unlike wood-based
media, BIOSORBENS particles are synthetically manufactured to give
uniform size particles, whieh offer consistent and more predictable performance. These media particles are packed in compact and durable
modular biofilter treatment systems.
Biotrickling Biofilters
Biotrickling filters are bioreactors in whieh contaminated air streams
are passed through the media. The major difference between biotrickling filters and classieal bio filters is the presence of a continuous water
flow in the reactor. The water phase carries nutrients for the microorganisms, and is usually neutralized before recirculation, for pH-control
purposes. Microbial oxidation takes place in the water phase as well as
in the immobilized bio films attached on the media particles. Microorganisms in the bio films degrade absorbed contaminants into harmless
products. Excessive biomass growth and clogging are major problems
encountered in biotrickling filters. Biotriekling bio filters are more
complex to construct and operate than classieal bio filters. However, for
chlorinated VOCs and compounds, which produce intermediate acidic
by-products, biotrickling filters are very effective.
While classieal and biotrickling-bed biofilters employ immobilized
organisms, bioscrubbers utilize dispersed (suspended) cultures. Bioscrubbers consist of two units: a usual scrubber in which VOCs and
odorous compounds are transported from the air to a water phase, and
a classieal bioreactor where the water exiting the scrubber is subjected
to biological treatment in the liquid phase. The bioreactor, which contains suspended cultures, requires sufficient oxygen through aeration to
Z. Shareefdeen
maintain a high level of biodegradation. Due to volatilization from the
bioreactor and scrubber parts of operations, emissions may not be completely eliminated. Thus, a secondary polishing unit may be required in
many cases. As in biotrickling filters, the water phase allows for the addition of nutrients to the system. Bioscrubbers can handle varying inlet
concentrations and flow rates, or shock-loading conditions, more easily.
Biofilter Terminology and Parameters
This section briefly discusses important terminology and bio filter
parameters used in engineering calculations and field scale biofilter
Empty Bed Residence Time
Devinny et al. (1999) defines empty bed residence time (EBRT) as the
ratio of volume of biofilter to the volumetrie air flow rate. This is an
important parameter, which needs to be determined for the actual
sizing of biofilter units. Due to the different porosities of the media and
void fraction, actual residence time of contaminated air in the bed
varies. In industrial designs, EBRT is often calculated as the ratio of
ac tu al medium volume to the volumetrie air flow rate. EBRTs of 30-60
s are common and economical in industrial designs. However, depending on the compounds being treated and concentration levels in the
airstreams, EBRT varies. Very low EBRT leads to higher velocity through
the bed, higher pressure drop across the bed of media and higher power
consumption. On the other hand, very high EBRT leads to a large footprint, a large volume of media and high capital cost.
Elimination Capacity or Removal Rate
Elimination capacity (EC) or removal rate is defined as the mass of contaminants removed per volume of media per unit time. Usually, EC has
a unit of gram removed per cubic meter of media per hour. In order to
evaluate different media and the biodegradability of different contaminants, EC values are used in designing biofilters. The EC values ranged
from 8 to 229 g/m3 of media per hour depending on the compounds
(Devinny et al. 1999). These values can also be used to determine the
biodegradability of compounds in bio filters. However, these values are
Innovative Methods in Biofiltration of Air Contaminants
media specific and not applicable if different media are used in a
bio filter system. Often, EC and EBRT need to be determined through
laboratory and pilot-scale studies, especially for applications involving
new compounds.
Removal Efficieney
Removal efficiency is the percent reduction in concentration levels.
Often, removal efficiency is an important specification for the design.
Medium, which has a high adsorption capadty, will give higher removal
efficiency of the contaminants at the start-up of the biofilter due to
physical adsorption and also due to absorption in water. Similarly,
during the transient process, contaminants may desorb from the media
leading to a negative removal.Removal effidency should be determined
under steady-state conditions. Accurate determination should exclude
removal due to physical processes such as adsorption/desorption and
absorption. Through complete mass balance of contaminants in all
phases (air, biofilm, water and solid media), removal effidendes can
be accurately determined. The removal effidency can be easily
predicted through mathematical modeling of the bio filtration process
(Shareefdeen and Baltzis 1994a).
Mass Loading
Mass loading is the mass of contaminants applied per volume of
medium per unit time. Usually, load has units of grams applied per
cubic meter of medium per hour. If removal is complete (or removal
effidency is 100%), mass loading is equal to EC. Mass loading is therefore related to the concentration of contaminants and flow rate. Variations in mass loading conditions refer to shock loadings. Biofilters are
effective in handling concentration and shock loading conditions when
manufactured media are used as bio filter media. In full-scale designs,
information on shock loading conditions plays an important role.
Media Volume, Media Depth and Footprint
Once EBRT is. determined through laboratory or pilot-scale
experiments, media volume can be easily estimated. Media such as
compost, peat, soil and wood te nd to compact over time, and channeling problems and odor breakthrough can take place. Thus, these types
z. Shareefdeen
of media cannot be placed to larger depths in a bio filter. In general,
media depth in a bio filter ranges from lOS-180cm. Structural properties are important in determining depth of the bed and hence the necessity to define footprint requirements. For many industrial applications,
space saving is critical and biofilters are placed on smaller footprints
and modular bio filter units may be stacked in stages.
Media Properties
Media materials used as packing materials or as bulking agents include
activated carbon, bark, ceramics monoliths coated with activated carbon, coconut carbon, compost, crushed oyster shells, extruded diatomaceous earth, glass wool, peat, manure composts, perlite, polypropylene
rings, polystyrene coated with powdered activated carbon, polyurethane foam, sewage sludge, silica, soil, vermiculite, wood chips and
yard waste. For efficient performance of bio filter operations, media
must have a good nutrient content (nitrogen, phosphorous etc.), organic
content, moisture retention capacity, ability to biodegrade contaminants
at neutral pH, good adsorption properties must also provide a good
environment for bacterial growth, have the ability to sustain bacterial
activity during shut-down and have low pressure drop and good structural properties. The survival of microorganisms, and thus bio filtration,
depends not only on oxygen and carbon sources (a role played by the
VOCs), but also on other nutrients such as nitrogen and phosphorus
sources. These additional nutrients are not needed when materials,
such as peat, compost and bark, are used as solid support for the organisms. The latter materials contain nutrients, which can be supplied
to the microorganisms. In synthetically manufactured media such as
in BIOSORBENSTM, nutrients are added during media preparation
(Shareefdeen et al. 2002b).
In classical bio filters, the absence of a continuous water phase may
create problems during operation. Moisture addition and temperature
effects are critical in winter operation of bio filters (Shareefdeen et al.
2002b). Water is required for biological activity and is retained in the
biofilm and in the pore structures of the media. Due to the high air
volumetric rates used during bio filtration, the bed can dry out. Also,
bio degradation is an exothermic oxidative process. Temperature rises
in the bio filter bed, which induces evaporation of water from the pores
of the solid packing. Leson et al. (1993) discussed the problem of temperature and moisture content in relation to a bio filtration demonstration project. Increasing temperatures and material dry-out can also lead
to channeling effects, which could lead to considerable reduction in
Innovative Methods in Biofiltration of Air Contaminants
removal rates (Ottengraf et al. 1983) due to a decrease in the gas-solid
interfacial area. Compost media are transformed from a hydrophilie to
a hydrophobie state when dried. Thus, surface irrigation does not bring
the media to required moisture content levels. Water-holding capacity
of compost media can be as high as 80% (Devinny et al. 1999) and about
20% for synthetically manufactured media. Due to lower moisture
retention and hydrophilie properties, synthetically manufactured media
can be easily irrigated without creating pressure drop problems, and can
be washed out to remove intermediate by-products and excess biomass.
Performance of media is also dependent upon the specific surface area
available for adsorption and biological degradation of contaminants.
The long-term use of media may lead to loss of active adsorption sites
and available surface area for bio film growth. The Langmuir method,
through nitrogen gas absorption onto media, is used to determine
surface area (Shareefdeen et al. 2002b).
Feasibility of VOC and Odor Removal in Biofilters
Biofiltration feasibility studies are reported for a number of compounds, induding butanol, ethylacetate, butyl acetate, toluene
(Ottengraf and Van den Oever 1983), propane (Ebinger et al. 1987),
tetrahydrofurane (Paul and Roos 1989), methylformiate, methanol!
isobutanol mixtures, styrene, vinylcydohexene, butadiene (Van Lith
1989), phenol (Zilli et al. 1993), toluene, dichloroethane, trichloroethene
(Ergas et al. 1993), methyl ethyl ketone (MEK) and methyl isobutyl
ketone (Deshusses and Hamer 1993), ethanol (Leson et al. 1993),
methanol (Shareefdeen et al. 1993), benzene and toluene mixtures
(Baltzis and Shareefdeen 1994),MEK and toluene (Llewellyn et al. 2002),
ammonia (Nicolai et al. 2002) and hydrogen sulfide (Aroca et al. 2002;
Sologar et al. 2002).
Although different biofilter configurations and media have been
used, most studies were done using a dassical bio filter configuration
(Ottengraf and van den Oever 1983; Ebinger et al. 1987; van Lith 1989;
Paul and Roos 1989; Zilli et al. 1993; Ergas et al. 1993; Deshusses and
Dunn 1993; Togna and Frisch 1993; Leson et al. 1993). Ergas et al. (1993)
resolved a pH -drop problem, by adding crushed oyster shells, a source
of calcium carbonate, in the packing material. Ebinger et al. (1987) used
a soil bed without external microbial inoculation. Deshusses and Hamer
(1993) used packing material consisting primarily of day spheres.
Fewer studies exist regarding VOC removal in biotrickling filters,
mostly utilizing inorganic support materials. Utgikar et al. (1991)
reported results from the biotrickling filter for the removal of landfill
Z. Shareefdeen
leachate off gases. Sorial et al. (1993) treated toluene vapor in a trickling bed having a monolithic channelized support. Biotrickling filters,
with plastic packing beds have been used to remove chlorinated compounds (Ottengraf et al. 1986; Diks 1992; Diks and Ottengraf 1989,
1991), gasoline fractions (Phipps and Ridgeway 1993) and methanol
(Overcamp et al. 1991). These studies demonstrate that biofiltration is
feasible for a wide range of compounds in different bio filter
Modeling of Biofiltration Processes
Modeling of biofiltration processes involves determinations of the mass
balance of contaminants and oxygen in air, biofilms and solid phases.
Accumulation of contaminants in all phases, dispersion effects in the
air, mass transfer between air and bio film phases, diffusional mass
transfer in the biofilm, consumption due to biological oxidation,
adsorption of contaminants onto solid matrix, biomass growth and
clogging in pore spaces, are all considered in the modeling of biofiltration processes. The resulting equations are frequently a complex set of
partial differential equations with interactive expressions for the kinetics of biological oxidation and adsorption isotherms (Dharmavaram
The complexity of the model is substantially reduced by the use of
assumptions such as equilibrium of contaminants and oxygen at the
biofilm/air interface, plug flow for air, constant biofilm thickness, no
biomass growth, steady-state, negligible oxygen limitation and simple
zero-order kinetics. These assumptions make model equations very
simple and allow users to calculate design parameters easily. However,
in some cases, simplified assumptions make models unrealistic and
often lead to errors in estimations. The most widely used model of
Ottengraf and van den Oever (1983) deals with steady-state
biofiltration, using zero- and first -order kinetic expressions, for biological oxidation under diffusion- and reaction-limited regimes. With
single VOCs, Andrews inhibitory kinetics (methanol, toluene) and
Monod kinetics (benzene) are also used (Shareefdeen 1994). In the case
of contaminant mixtures, biodegradation rate expressions may be
more complex, as kinetic interactions such as competitive inhibition
(benzene/ toluene mixtures ) arise.
Although detailed kinetic studies may be impractical in some cases,
e.g., multicomponent gasoline emissions, some understanding of the
kinetics is essential. The availability of oxygen in the biofilm is impor-
Innovative Methods in Biofiltration of Air Contaminants
tant for complete biological oxidation of contaminants. The assumption
that, since there is plenty of oxygen in the air (relative to the VOCs presence), the same should also hold in the bi9film, is not correct. In fact,
with water-soluble VOCs (e.g., methanol), oxygen is depleted in the
bio film before VOC consumption. This oxygen limitation should be
expected to happen at relatively high VOC concentrations in airstreams.
The model for steady-state bio filtration of VOC mixtures (Baltzis and
Shareefdeen 1994) takes into account potential oxygen effects as weIl as
potential interactions among VOCs du ring their biodegradation.
Ozis et al. (2002) have recently presented a discrete percolation model
to describe biomass-clogging effects on bio filter performance. There are
onlya few models that describe the biofiltration of highly odorous compound such as hydrogen sulfide. Similarly, there are few published
studies on the transient performance and response of bio filters (Baltzis
et al. 1993; Shareefdeen and Baltzis 1994b). Questions as to how weIl a
biofilter can respond to variations in volumetrie flow rate, concentration, and composition, obviously are of importance for commercial
application of this technology.
Selected Industrial Applications
Industrial operations, such as meat rendering, wastewater treatment
and printed circuit board manufacturing facilities, produce nuisance
and, in some cases, toxic odor constituents. In meat rendering and
wastewater treatment plants, major odorous chemicals are hydrogen
sulfide (H 2S), methyl mercaptan, dimethyl sulfide (DMS), ammonia and
ethylamine (Shareefdeen et al. 2002a). In addition to nuisance odor
characteristics, several health effects, including headaches, nausea, eye
irritation, paralysis and even death are associated with these compounds when exposed to high concentrations. Similarly, printed circuit
board manufacturing facilities emit odorous volatile organic compounds (VOCs) including propylene glycol monomethyl ether acetate
(PGMEA). Overexposure to glycol ethers can cause anemia, intoxication
similar to the effects of alcohol, and irritation of the eyes, nose or skin
(Shareefdeen et al. 2002a). In this section, three case studies dealing with
different applications, bio filter configurations, media type and airflow
directions, are discussed.
A lS,OOOcfm, open-bed biofilter system, fabricated from reinforced
concrete with a strong, slotted support floor to handle the weight of a
skid-Ioader, was instaIled at an animal rendering plant located in
Hickson, Ontario, Canada (Fig. 1). A humidification manifold located in
Z. Shareefdeen
Fig. 1. A 15,000 efm eapacity open-bed eommercial biofilter system whieh treats air
emissions from a meat rendering plant (eourtesy of BIO REM Teehnologies Ine., ON,
the plenum, conditioned and humidified the air stream. Misting spray
nozzles are configured and sized appropriately to ensure that relative
humidity is greater than 98% at all times. The biofilter was filled with
wood-based media (BIOMIX), produced as nuggets and screened to
provide a consistently sized coarse product with a surface coating of
nutrient enriched fines.
The air stream of the open-bed bio filter system was evaluated using
a Fourier transform infrared (FTIR) mobile monitoring system and olefactometer analysis for odor. The common constituents were ammonia,
methyl mercaptan, hydrogen sulfide, ethylamine and di-methyl sulfide.
Loadings and elimination capacities (ECs) in g/m 3 media/h were: for
ammonia (0.62,0.62); methyl mercaptan (0.08,0.06), hydrogen sulfide
(0.13,0.12); ethylamine (0.14,0.14) and di-methyl sulfide (93.27,93.27).
The load and EC values for di-methyl sulfide were higher than the
expected range. For all the compounds (except methyl mercaptan), the
average removal efficiency exceeded 96%. The biofilter treated a contaminated air stream of 24,544 odor units with an average of 96.6%
removal efficiency in 30-s EBRT. Odor rem oval efficiency over the cross
section of the media varied between 90 and 99%.
Innovative Methods in Biofiltration of Air Contaminants
A full-scale compact and modular bio filter, treating air emissions
from a wastewater pumping station where hydrogen sulfide is the main
odor-causing compound, was installed in the Greater Toronto Area
(GTA), Ontario, Canada (Fig. 2). This modular biofilter system was
packed with the manufactured synthetic media (BIOSORBENS) and
operated under negative pressure. The H2S concentration data were collected by continuous measurements of inlet and outlet concentrations
at I-min intervals using a H2S data logger and compared with laboratory data for removal efficiencies (Fig. 3). The loadings to the bio filter
varied from 0.7-9.2g/m 3 media/h and the bio filtration system at 30-s
EBRT effectively eliminated varying H2S loads (peak concentration of
80 ppm H2 S) with >99% removal efficiency.
A printed circuit board manufacturing facility was subjected to odor
complaints from neighbors due to the odorous emissions of PGMEA
from a formulated chemical mixture for surface coating. A 7500cfm
capacity dual modular biofilter system (Fig. 4), combined with impinge-
Fig.2. A 1,500 cfm eapaeity modular biofilter system whieh treats air emissions from
a wastewater pumping station (eourtesy of BIOREM Teehnologies Ine . , ON, Canada)
Zarook Shareefdeen
110 ,-------------------------------------------------T 110
»j ,
Hours of operation
Fig. 3. H2S rem oval using a modular biofilter system with synthetic media installed a
municipal wastewater pumping station
Fig.4. A 7,500 cfm capacity biofilter system at a printed circuit board manufacturing
facility. One unit is filled with the wood media and the other with manufactured inorganic media (courtesy of BIOREM Technologies Inc., ON, Canada)
Innovative Methods in Biofiltration of Air Contaminants
45 -
wood-based media
... synthetic media
Load (g/m 3-media/hr)
Fig.5. Comparison of synthetic and wood-based media for PGMEA elimination capacity as a function of inlet loads to the laboratory pilot bio filters
ment scrubber for aerosol removal, was installed in the year 2000. This
system operated under negative pressure in the down-flow mode of
operation. One biofilter unit was packed with wood-based (BIOMIX)
and the other with inorganic synthetic (BIOSORBENS) media. Due to
the unique characteristics of the solder masking formulation used
in printed circuit board manufacturing, a laboratory pilot-scale
bio filtration study was needed to obtain the required engineering data
for full-scale design (Shareefdeen et al. 2002b). PGMEA elimination
capacities obtained from the laboratory pilot bio filter units were plotted
against loads to the laboratory-scale pilot biofilters. While elimination
capacity dropped at a loading rate greater than 20gm-3 h- 1 in woodbased media, the elimination rate was equal to the loading rate even
for an inlet load of 40 gm-3 h- 1 (>99.9% removal) for synthetic media
(Fig. 5).
Other interesting applications include biofiltration of candle fragran ces, glycerin manufacturing process emissions, and waste oil
recycling. A list of biofilter installations by some leading biofilter manufacturing companies are shown in Table 1. For different applications,
different media are generally used, e.g. sewage waste, pine bark, softwood chips, soil, compost, perlite, inert materials, and synthetically
manufactured media.
Z. Shareefdeen
Table 1. Applications of biofilters in various industries
Mount Pleasant,
South Carolina
Guelph, Ontario
Louisville, Kentucky
Guelph, Ontario
Jefferson County,
Guelph, Ontario
City of Broomfield,
Guelph, Ontario
City of Poughkeepsie,
Pewee Valley,
Softwood chips
Animal feed
St. Joseph, Montana
Guelph, Ontario
City of Toronto,
Gue1ph, Ontario
Candle wax
Brampton, Ontario
Gue1ph, Ontario
Toronto, Ontario
Guelph, Ontario
Fish feed
West Co ast Rendering,
British Columbia
Guelph, Ontario
Food and
Allen, Ilinois
PPC, Longview,
Mixture of
inert! organic
Cast Alloy Inc.,
Bio Filtration,
Aachen, Germany
Mixture of
Particle board
PPC, Longview,
Mixture of
inert! organic
Pet food
Gue1ph, Ontario
Printed circuit
Scarborough, Ontario
Guelph, Ontario
Hanceville, Alabama
Guelph, Ontario
Canada Composting,
Guelph, Ontario
Synthetically manufactured inorganic media.
Wood-based proprietary media.
Innovative Methods in Biofiltration of Air Contaminants
This chapter presented various aspects of physico-chemical and
biological odor and VOC control technologies. Three case studies of
large-scale applications utilizing biofilters to deal with emissions from
a meat rendering plant, municipal wastewater treatment plant and a
printed circuit board manufacturing facility demonstrated the robustness of bio filtration technology in treating a wide variety of compounds
using different media and biofilter configurations. The evaluation of feasibility studies and the data from these biofilter applications confirm
that this technology can be successfully applied to different industries
in eliminating nuisance odors and pollutants from waste gas streams.
In future, advances in biofilter technology will allow for the efficient
elimination of toxic and odorous chemicals at municipal and industrial
sites and will provide cleaner and more breathable air.
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Risk-Based Remediation
of Contaminated SoH
Tahir Husain 1
Risk assessment is an emerging multi-disciplinary scientific discipline
used to evaluate health and ecological risks posed by the chemicals of
concern (COC), also known as "risk agents". This evaluation helps in
formulating remedial actions and risk-based management plans to
achieve target risk reduction. However, to comply with the regulations
and to develop a cost-effective remedial action plan, there is a need
to introduce a systematic and scientifically sound methodology. The
methodology should not only help in assessing the extent of contamination and associated risks at a site, but it should also be able to identify appropriate remediation technologies. This paper intro duces the
concept of risk-based site remediation. The decision on the extent of
cleanup at a site is based on environmental and human health risks to
potential receptors through possible exposure pathways. In those locations where exposure of contamination to human health and ecosystems is low, a complete cleanup may not be required. The proposed
methodology also allows "natural attenuation" as one of the remediation alternatives. If the contamination at a site can be contained and
if the natural attenuation rates are promising, "no action" may be considered as one of the remedial options for site remediation. Many
hydrocarbon-contaminated sites can be identified under the "natural
attenuation" option using this approach.
Methodology proposed in this chapter has several potential advantages. It can speed up the process of remediation by focusing only on
those exposure pathways that are of concern to human health and
ecosystems. Unlike conventional remedial plans, which are designed to
attain a more stringent cleanup for the whole site, the proposed method
1 Department of Environmental Engineering, Faculty of Engineering and Applied
Science, Memorial University ofNewfoundland, St. John's, Newfoundland, Canada, AlB
3XS, e-mail: [email protected], Tel: +1-709-7378781, Fax: +1-709-7374042
Soil Biology, Volume 1
Applied Bioremediation and Phytoremediation
(ed. by. A. Singh and O. P. Ward)
© Springer-Verlag Berlin Heidelberg 2004
T. Husain
concentrates on a cleanup process concerned with the protection of
human health and the environment.
Regulatory Standards
Since the exposure mechanisms and intake rates of chemieals in air and
water follow relatively similar patterns from place to pi ace and are easily
quantifiable, it is simple to develop unified national standards for water
and air quality. However, since soil is relatively immobile and since the
intake rate of soil contaminants through ingestion and dermal contacts
varies considerably, unified cleanup levels are difficult to develop. As a
result, the agencies responsible in the Uni ted States are in a process of
developing soil and groundwater cleanup standards for each State separately. A review of the standards being proposed shows significant variations in the action levels due to limited guidance by federal agencies,
variations in the compositions of petroleum products, and uncertainties in the fate and transport variables affecting distribution of hydrocarbon constituents in soil and groundwater (Bell et al. 1991). Table 1
lists cleanup levels adopted in ühio, West Virginia, New York, lowa, and
Missouri. As shown, the cleanup level for benzene in potable groundwater varies from 0.7 jlg/l in New York to 5 jlgll in other States. For
non-potable groundwater, Missouri has established the cleanup level at
50 jlg/l. Similarly for toluene, the variation is from 5 jlg/l in New York to
12,000 jlg/l in ühio. For total petroleum hydrocarbons (TPH) and
benzene, toluene, ethylbenzene, and xylene (BTEX), there are no unified
cleanup levels for groundwater contamination. High variability in soil
cleanup levels is also observed.
In order to define site-specific cleanup levels, exposures and conditions unique to the sites are considered for each contaminant. For
carcinogens, the cleanup levels are usually established to represent an
excess cancer risk between 1 in 10,000 and 1 in 1,000,000. For noncarcinogens, these levels are developed using a reference dose. In both
cases, factors such as land use, site locations with reference to potential
receptors, hydrological and geological settings, and site sensitivity with
respect to receptors, are also considered (Gaudet et al. 1992). For these
factors, depending upon their relative potential of risk to receptors,
scores are assigned on a subjective basis and total scores are calculated.
The resultant scores are then used to determine cleanup levels. Table 2
lists the scoring criteria used in Missouri for soil cleanup. Site-specific
factors in this case include groundwater depth and its uses, geological
settings at the contaminated site, types of receptors and their locations
Groundwater (agil)
Cleanup level
50 (Fluorine)
50 (Pyrene)
50 (MTBE)
380-1,156 100
Soil (mg/kg)
1 (Fluorine)
1 (Pyrene)
1 (Anthracene)
1 (MTBE)
"Please refer to Table 2 for variations due to land use, geology, soil permeability, receptor locations, etc.
Table 1. Comparative evaluation of groundwater and soil cleanup levels for petroleum-contaminated sites
>1000 ft.
Depth to groundwater?
1s groundwater potable?
Drinking water supply proximity
Distance to surface water?
Geological feature present?
Man-made vertical conduit
Man-made horizontal conduit?
Soil permeability?
Soil thickness?
Environmentally sensitive receptors?
Surrounding land use distance?
Future land use?
Off-site impact?
Total score
Soil cleanup level (mg/kg)
Score 15 if true
Site features
1000-2500 ft.
500-1000 ft.
1000-2501 ft.
501-1000 ft.
2501-5000 ft.
1001-2000 ft.
501-1000 ft.
Score 5 if true
Score 10 if true
Table 2. Missouri cleanup level for remediation of hydrocarbon contaminated soil
Score 0 if true
Risk-Based Remediation of Contaminated Soil
with reference to the site, land use plan and soH properties. If the site is
located in an environmentally sensitive area and if soH has high porosity, the score counts will be low and hence more stringent regulations
would be adopted for soH cleanup. If the site is located in a remote area
with deep non-potable groundwater in low soil permeability, the score
counts will be high and in that situation high cleanup levels can be used.
In the case of Missouri, the variations in the TPH cleanup level are from
50 to 1000 mg/kg depending upon the sensitivity and remoteness of the
site. Chemical and toxicological information on BTEX is provided in
Table 3.
Risk-based corrective action (RBCA) integrates site assessment and
remediation with the United States Environmental Protection Agency
(EPA) approved risk assessment and exposure practices. It is a tiered
approach and involves very little site-specific information at Tier 1 to
an enormous level of data collection and analysis at Tier 3 (ASTM 1995).
The following sections summarize the concept and data requirements
in each tier.
Table 3. Chemical and toxicological information on BTEX
Major hydrocarbon constituents
Chemical properties
Molecular weight (glmol)
Density (g/cm 3)
Diffusion coefficient (cm 2/s)
Log Koc
1.1 x 10-5
9.4 X 10-6
8.5 X 10-6
8.5 X 10-6
Henry's law constant
Atm-m 3/mol
Vapor press ure (mmHg)
Solubilityat 20-25°C (mg/I)
6.29 x 10-3
6.25 X 10-3 7.69 X 10-3
6.97 X 10-3
T. Husain
Table 3. Continued
Major hydrocarbon constituents
Risk assessment
Reference dose (mg/kg/day)
Slope factor (mg/kg/dayr 1
Unit risk factor (URF)
(mg/m 3r 1
EPA weight of evidence
8.3 x 10-6
Regulatory data
Max concentration (mg/I)
- ingestion
Permissible exposure
limit (mg/m 3 )
Surficial soil (mg/kg)
- residential
- industrial
Subsurface soil (mg/kg)
Ingestion/leachate to
Soil vapor (JIg/m 3)
Indoor air (inhalation)
- residential
Indoor air (inhalation)
- industrial
Groundwater (mg/I)
Indoor air (inhalation)
- residential
Indoor air (inhalation)
- residential
Half-life (days)
VP denotes that even at a concentration equal to the vapor pressure of a chemical, the
hazard quotient is less than 1; Sol denotes that even at a concentration equal to the
solubility of a chemical, a hazard quotient is less than one.
Risk-Based Remediation of Contaminated Soil
Tier 1 Approach
A Tier 1 evaluation is a risk-based analysis to develop non-site specific
risk based screening levels (RBSL) of contaminants for direct and indirect exposure pathways. Direct exposure pathways are inhalation of
volatile gases and/or particulate matter from surface soil, subsurface soil
and groundwater; ingestion of soil, dust and groundwater; and dermal
contacts with surface soil. Indirect exposures, such as risk-causing
vapor emission from soil to air, leaching of risk agents from soil to
groundwater, and from soil to enclosed spaces are also considered in
developing look-up tables for RBSLs. Other than pathways, all possible
types of receptors are also considered. These receptors are residents in
different age groups in the vicinity of the site, commercial and industrial settings and site cleanup workers exposed to risk agents. Using the
quantitative risk assessment methodology (US EPA 1991a), RBSLs for
all possible combinations of chemicals, receptors, pathways, and exposure factors are determined. The typical equations used for some
selected pathways are presented in the following sections. The variables
used in the equations are defined in Tables 4 and 5. The default values
of each variable and their units are also listed in the tables.
Direct Inhalation - Carcinogenic Risk Agents
TR*BW*AT *365 days *10 3 J1g
RBSL . [ J1g ] =
aIr m 3 _ air
SI; * IRair * EF * ED
Direct Inhalation - Non-Carcinogenic Risk Agents
m3 _ air
THQ* RfD· * BW* AT * 365 days *10 3 J1g
mg (lb)
IR air * EF * ED
Direct Ingestion (Groundwater) - Carcinogenic Risk Agents
RBSL w [
TR *BW * ATe * 365 days
mg ] =
L - water
SFo * IR w * EF * ED
Receptor scenarios
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer Risk = 10-6
Chronic HQ = 1
Cancer Risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Cancer risk = 10-6
Chronic HQ = 1
Target level
2.08 x 10 5
1.45 x 106
2.08 x
1.45 x 106
RES, Selected risk level is not exceeded to pure compound present at any concentration; S, selected risk level is not exceeded for all
possible dissolved level.
Surficial soil ingestion/dermall
Surficial soil ingestion/dermall
Soil vapor intrusion to
enclosed space
Soil (mg/kg)
volatilization to outdoor air
Vapor ambient air
Vapor (enclosed space)
Groundwater (mgll)
Exposure pathway
Table 4. Tier 1 risk-based screening level (RBSL) look-up table. (ASTM 1995)
Risk-Based Remediation of Contaminated Soil
Table 5. RBSL Estimates for some selected pathways and medium of exposure
Variables (unit)
Definition (default ASTM value)
ATe (year)
ATn (year)
Averaging time - carcinogen (70)
Averaging time - non-carcinogen. On-site commercial worker
(25); on/offsite resident - adult (30); on/off-site resident child (6); construction worker (1)
Body weight adult (70); child (15)
Exposure duration. On-site commercial worker (25); on/off site
resident - adult (30); on/off-site resident - child (6);
construction worker (1)
Exposure frequency. On-site commercial worker (250); on/off
site resident (350); construction worker (183)
Soil ingestion rate. On-site commercial worker (50); on/offsite
resident - adult (100); on/off-site resident - child (200);
construction worker (100)
Daily outdoor inhalation rate (20)
Daily water ingestion rate. On/offsite resident - adult (2);
on/off-site resident - child (1); construction worker (1)
Soil-skin adherence factor (0.5)
Oral relative absorption factor (1)
Dermal relative absorption factor - volatile (0.5)
Skin surface area for dermal contact with water. Construction
worker (3160)
Risk-based screening level for media i (mg/kg for soil; mg/l for
water or j1g/m3 for air
Inhalation chronic reference dose (chemical-specific)
Oral chronic reference dose (chemical-specific)
Oral cancer slop factor (chemical-specific)
Target hazard quotient (~1)
Target excess individuallifetime cancer risk 10-4 to 10-6
Volatilization factor (groundwater to ambient air (mg/m 3 air per
mg/l water)
Volatilization factor (surface soil to ambient air (mg/kg air per
mg/l soil)
Volatilization factor (subsurface soil to ambient air (mg/kg air
per mg/l soil)
Particulate emission factor - surface soil to ambient air
(g/cm 2/s)
Leaching factor: soil to groundwater (mg/l water per mg/kg soil)
BW (kg)
ED (year)
EF (days/year)
IRsoi1 (mg/day)
IRair (m 3/day)
IRw (I/day)
M (mg/cm2 )
RAF o (unitless)
RAF d (unitless)
SA (cm 2)
RBSl i
RfD i (mg/kg/day)
RfD o (mg/kg/day)
SF o (mg/kg/dayt 1
VF ss
H, Henry's law constant; Uair> wind speed above ground surface in ambient mixing zone
(cm/s); 4ir> ambient air mixing zone height (ern); LGw, depth to groundwater (ern) and
is equal to thickness of capillary fringe (h eap ) = thickness of vadose zone (hv ); W, width
of source area parallel to wind or groundwater fiow direction; D:ff, effective diffusivity through vadose zone (cm 2 /s); Dw: ff, effective diffusivity from groundwater (cm 2/s);
Ow" volumetrie water content in vadose zone soil (cm3 of water per cm3 of soil); k" soilwater sorption coefficient (g of water per g of soil); Ba" volumetrie air content in vadose
zone soil (cm 3 of air per cm 3 of soil); Ugw, groundwater Darcy's velocity (cm/s); Sgw,
groundwater mixing zone (ern); I, infiltration rate of water through soil (cm/year).
T. Husain
Direct Ingestion (Groundwater) - Non-Carcinogenic Risk Agents
RBSL w [
THQ* RfDo * BW* ATn * 365 days
mg ] =
IR w * EF* ED
Direct Ingestion and Dermal Contacts (Soil) - Carcinogenic Risk
s kg-soil
TR * BW * AT.: * 365[ days ]
EF* ED[(SFo *10-6 ~~ *{A+B})+C ]
Direct Ingestion and Dermal Contacts (Soil) - Non-Carcinogenic Risk
THQ * BW * ATn * 36S[ day~ ]
where A = IRson*RAFo; B = SA*M*RAF d; and C = IRai/(VFss*PEF).
Indirect Inhalation
mg . ] =
kg - SOll
m 3 - air * 10-3 mg
Indirect Ingestion
RBSL [ m g
s kg - soil
mg ]
]= RBSL [ L-water
Risk-Based Remediation of Contaminated Soil
Using the above equations, RBSL levels are calculated in air, water, and
soil. Besides exposure and receptor-specific data as listed in Table 4,
chemical-specific information such as the reference dose for noncarcinogens and slope factors for carcinogens are also required. Such
information is available in chemical and toxicological databases (ASTM
1995). Some of the information extracted from databases for BTEX is
listed in Table 5.
For example, if we are interested in developing a look-up table on
RBSLs for benzene and toluene for the case of direct ingestion of
groundwater, and the receptor in this case is an adult resident in the
area. Since benzene is a human carcinogen, Eq. (2a) will be used to
calculate the RBSL while toluene, being non-carcinogen, Eq. (2b) will
be applied. Assurne that the excess individuallifetime cancer risk is 1
in 1 million and that default values, as listed in Table 4 are used for other
variables. The numerical values required to calculate RBSLs are
extracted from Tables 4 and 5 and are listed below:
Variables Units
Table Risk agent( s)
IR w
Dimensionless 10-6
[mg/kg/dayt 0.029
Dimensionless 1
Benzene, toluene
Benzene, toluene
Benzene, toluene
benzene, toluene
Substituting the above values in Eq. (2a), RBSL for benzene in potable
groundwater considering residential adult as a receptor is calculated to
be 0.00294mg/I. For toluene (non-carcinogen), Eq. (2b) is used which
gives RBSL as 7.3 mg/I. Using the above equations and similar formulations (all not listed above), RBSLs for specific chemical, pathways and
receptor are calculated for soil, groundwater, and air. Table 5 lists these
values for BTEX. However, if the calculated RBSL values do not exceed
the corresponding levels listed in the look-up table, a further compliance monitoring program should be required to confirm that the corrective action goals are satisfied and no further action on remediation
would be required. Where the calculated values exceed the RBSLs listed
in the look-up table, remediation options need to be identified and a
Tier 2 approach would be implemented.
T. Husain
Tier 2 Approach
In the case of Tier 1 evaluations, the data requirements are limited to
on-site land use with the maximum concentration levels of risk agents
within the source zone media. However, a Tier 2 evaluation is more sitespecific and is extended to evaluate affected zones. A partial list of the
data requirements for a Tier 2 evaluation is as follows:
1. Lateral and vertical extent of affected soil and groundwater zones
2. Groundwater flow gradient, seepage velo city, hydraulic conductivity
and flow direction
3. Average annual climatic conditions
4. Leaching potential through overlying soil zones
5. Soil conditions such as porosity, thickness, and surface and subsurface soil types
6. Attenuation factors such as electron acceptors, retardation factors
and decay rate coefficients
7. Anticipated receptor types and their locations
In a Tier 2 evaluation, the contaminant levels in affected soil and
groundwater are compared to site-specific target levels (SSTLs) for
applicable exposure pathways and risk agents. SSTL values represent the
upper boundary of risk agents that, if achieved throughout the source
zone, will ensure the applicable risk limits at the potential point of exposure are not exceeded (Connor et al. 1995). These values are calculated
using simple analytical models and can be derived by either of the
following methods:
1. Method 1 (SSTL at the source zone): in this method, modified Tier 1
RBSL values using site specific input values at the source zone are
2. Method 2: SSTL at point of population exposure (POPE) for individual chemical). Natural attenuation factors (NAFs) for individual risk
agents are calculated for indirect exposure using empirical equations, analytical transport models (Domenico 1987; Farmer 1980), or
point of exposure site data obtained at a specific distance from the
source. Empirical equations are developed by fitting regression lines
between the measured concentration and source data measured at
known distance from the source. SSTL is then calculated using the
following relationships:
Risk-Based Remediation of Contaminated Soil
SSTL = NAF * ~ (carcinogens)
(non - carcmogens)
SSTL = NAF * (POE exposure limit)(applicable standards)
where E = exposure rate for specified pathway
=------BW* AT* 365 days
The point of exposure (POE) limit is the applicable POE concentration
limit, e.g. maximum concentration limit (MCL). All other variables are
defined in Table 5. For direct exposure, NAF will be one .
• Method 3 (SSTL at POE for cumulative chemieals): under this option,
values are calculated to prevent exceedance of both individual and
cumulative constituent risk limits at the point of exposure.
* Cumulative TR (
SSTL = Cs +
Baseline RT
(non - carcinogens)
Baseline HI
= representative COC concentration in source medium = Cpoe *
NAF, where
Cpoe = representative chemical concentration in the exposure
medium at POE
TR = target risk limit for carcinogenic effects of multiple chemicals
(10-4 to 10-6 )
cumulative carcinogenic values for pathways = IR!
R! = individual risk for with chemical which is equal to CDli * SFi
CD Ii is average chronic daily intake of carcinogens
SF i is slope factor of carcinogenic chemicals in (mg/kg/dayt 1
THI = target hazard index for toxic effects of multiple chemie als
HI = hazard index value for pathway (summation of hazard quotient values for individual chemieals = lliQ!
T. Husain
= hazard quotient for non-carcinogenic chemical = IIRfD
= average non-carcinogenic chemical intake in [mg/kg/day]
RfD = reference dose in [mg/kg/day]
For sites subject to individual and cumulative chemical risk limits, SSTL
values for each pathway must be calculated using Eq. (5a, b) and c and
(6a, b), and the lesser values are identified for each risk agent which are
then used to identify the minimum soil and groundwater cleanup
In calculating SSTL, it is important to estimate site-specific NAFs.
These factors are based on cross-media transfer of contaminants and
lateral transport and dilution. Cross-media transfer of contaminants
include migration of contaminants from soil to air, soil to groundwater
and groundwater to surface water. Both monitoring and modeling tools
can be used to estimate these factors. Monitoring techniques include
measurements of soil-vapor concentration, surface soil-vapor flux, soil
leaching tests, etc. Conservative steady-state analytical models can also
be employed to estimate the transfer rates of contaminants from one
medium to another medium. Some of the analytical multi-media model
to estimate cross-media transfer factors used in a Tier 2 RBCA are as
follows (ASTM 1995):
Vapor Emission Factor from Groundwater to Ambient Air
[ mg/rn 3 - air] =
wamb mgll- water
* Dair * L Gw ]
* 10 3
W* Deff
Vapor Emission Factor from Soil to Ambient Air
• ]
H* ps
*10 3
[ m g/rn-aIr =
samb mglkg - soil
H8] [1 Uair * Da ir * Ls ]
ws + sPs + as + +
Leaching Factor from Soil to Groundwater
LF[mgll- water] = 8ws + ksPs + H8as
mglkg - soil
[1 + U gw * Dgw ]
Risk-Based Remediation of Contaminated Soil
The variables used in the above equations are defined in Table 5. The
measured concentration in soil and groundwater at monitoring locations are then compared with the corresponding 88TLs. However, if the
measured value for a specific risk agent and pathway exceeds the SSTL
value, subsequent action will be required to remediate the site to "sitespecific Tier 2 cleanup goals" or further evaluation using the Tier 3
approach. Another option may be to go for interim response measures
targeted at principal risk sources by partial source reduction and site
classification adjustment.
Tier 3 Approach
In Tier 3 RBCAs less conservative assumptions are used, i.e. the exposure and risk computations are evaluated for conditions as they exist
on the site rather than using simplified general exposure scenarios. It
involves rigorous and intensive site data collection, additional information on model parameter, model parameter uncertainty, and fate and
transport models applicable for highly complex, heterogeneous, and
cost-significant sites. A detailed investigation on the selection and use
of fate and transport models would be required in a Tier 3 approach.
An integration of remediation technology with the fate and transport
models would make the process in a Tier 3 approach more costeffective.
Transport Models and Remediation Technology Database
Analytical models use mathematical solutions to govern mass flow and
constituent transport processes that are continuous in the space and
time domains. These models are generally based on the assumptions of
uniform properties and regular geometry. For multiple sources with
complex and irregular boundaries, simple analytical models may not be
applicable. For such cases, more sophisticated analytical and numerical
models are used.
The selection of an appropriate model for a site-specific case is not a
simple task. Their selection should be based on a careful review of the
models and underlying assumptions in solving transport equations.
Other factors to be considered in the model selection are model features, site-specific input data requirements, model prediction accuracy,
flexibility to incorporate modifications, availability of technical support,
and model validation his tory. There are numerous fate and transport
T. Husain
models available in the literature to estimate hydrocarbon constituents
(Foster Wheeler 1998). However, for site cleanup purposes, the model(s)
should be able to estimate (1) dispersion and migration of hydrocarbon
vapors and particulate from surficial and subsurface soils to ambient air
and enclosed spaces; (2) transport and migration of hydrocarbon constituents in groundwater with multiphase, multicomponent processes;
and (3) biodegradation and distribution ofhydrocarbon constituents in
the vadose and unsaturated zones. The models should also have the
capability to evaluate the effectiveness of site remediation technologies.
Considering the above minimum requirements, both analytical and
numerical models were evaluated and most relevant models for the
remediation ofhydrocarbon sites were identified, these are summarized
in Table 5. These models will be evaluated further before implementing
them in the Tier 3 RBCA process.
Remediation technologies, as described in Fig. 1, are broadly
classified as in situ or ex situ. In situ remediation is conducted at the
contaminated site, but in the case of ex-situ remediation, the contaminated medium is removed to the site of remediation for treatment. In
both cases, the principle of remediation is based on physical/chemical
treatment, biodegradation using microorganisms and heat to remove
hydrocarbon vapors. Some of these processes, relevant for the remediation of hydrocarbon contaminated soil and groundwater, are defined
below and are described further in other chapters of this book.
• Vapor extraction: using vacuum extraction weHs, contaminated
vapors from soil can be removed using this technology.
• Air sparging: air at high pressure is injected into the ground below
the contaminated area, which helps in the formation of bubbles that
rise and carry trapped and dissolved contaminants to the surface. In
combination with a soil vapor extraction system, the trapped and dissolved contaminants are captured.
• Bioremediation: this process uses microorganisms to biodegrade contaminants into less harmful substances. Using physical/mechanical
pro ces ses such as aeration and bioventing, biodegradation rates can
be enhanced.
• Thermal desorption: by heating the soil at relatively low temperatures
contaminants with low boiling points can be vaporized; they are then
captured and removed for further treatment or destruction.
• Soil washing: uses water or a washing solution and mechanical
processes to scrub excavated soils and remove hazardous contaminants.
After comprehensive evaluation, the most appropriate fate and transport model(s) from Table 6 can be selected and linked with the REMTEC
Soil Vapor
Fig. 1. Remediation technologies for petroleum-contaminated sites
Siurry Phase
Hot Water/
Air Sparging
Ground Water
Vapor diffusion
Gases and particulate
Farmer; 1-D Analysis (Farmer 1980)
ISCST; 1-D Gaussian (US EPA 1992)
Leaching factor
SAM; 1-D Analysis (Connor et al. 1996)
Multiphase transport
MOFAT 2-D finite element (USEPA 1991b)
NAPL simulator; finite element
Flux to air and
SESOIL 1-D hybrid/analysis (Bonazountas
and Wagner 1984)
Soil to groundwater
Vapor (ambient air)
SCREEN 3; 1-D analysis (US EPA 1995)
Soil to air
Models; numerical techniques
Transport up to five non-inert chemieals of DNAPL/LNAPL in
water, oil, and gas phase in saturated porous media; complex
but gaining popularity
Multiphase fate and transport in vadose, capillary and watertable zones
LNAPL through vadose zone, capillary fringe, and water-table
Fate and transport in vadose zone with dissolution, diffusion,
absorption, dispersion, bio degradation, and volatilization.
Widely used and readily available
Dilution, evap-transpiration, sorption, and bio degradation with
time for leaching
Area, point and volume sourees, multiple receptors; extensive
Instantaneous soil surface fiux and vapor fiux with time;
diffusion and advection
Multiple sources and receptors; complex geometry; decay term;
extensive testing
Model features/limitations/validation
Table 6. An overview of potential fate and transport models for remediation of hydrocarbon contaminated sites
Groundwater with remediation technologies
REMTEC; relational database
Remediation technology database
RBCA Tier 2 analyzer; 2-D analytical
AIRFLOW/SVE, finite difference
MARS; 2-D/3-D finite element
BIOPLUME; 2-D finite difference
(Rifai et al. 1997)
BIOSCREEN; Analytical
Over 500 innovative site remediation technologies in userfriendly environment:Lists technology description,
regulations, environmental concerns, and projects
Natural attenuation
Remediation through natural attenuation; bioremediation using
of organics in water
aerobic and anaerobic electron receptors
HC remediation
Advection, dispersion, adsorption, aerobic, and anaerobic decay;
saturated zone
Screening mode
Soil vapor extraction, biventing and bio degradation of up to 250
HC components Contaminant partitioning among water,
vapor, free HC, and solid phase
LNAPL - multiphase
Recovery/migration of LNAPL in groundwater; multiple
areal remediation
pumping and injection weHs; Cleanup of heterogeneous and
anisotropic fractured porous media
Vapor flow and
Vapor flow and transport in unsaturated heterogeneous and
transport in soil
anisotropic soil Useful for soil vapor extraction and
bioventing remediation system
Natural attenuation of contaminated groundwater and pumpBiodegradatio
and-treat remediation system; prediction of downstream
impacts for risk assessment
T. Husain
database through a user-friendly interface. This will help in making site
remediation more scientific and cost-effective for complex sites.
The following conclusions can be drawn from this study:
1. Risk-based remediation is a cost-effective method for contaminated
sites. However, its application should be based on well-defined exposure scenarios.
2. Site-specific information should be used on exposure related parameters such as frequency and duration of exposure, chemical intake
rates, expected life expectancy and body weight to develop RBSLs
and SSTLs.
3. There is a need to develop soil and groundwater cleanup levels for
Saudi Arabian conditions. These levels should reflect site features
such as geology and hydrology of the area, soil properties, aquifer
characteristics, sensitivity of the site and its future use.
4. For effective assessment and remediation of sites, remediation database and some selected models should be integrated with RBCA. The
model users should be familiar with the basic assumptions used in
the model development and extent of input data requirements.
Acknowledgment. The financial supports provided by the NSERC,
Canada Research Grant 3-36215 and Saudi Aramco, Saudi Arabia
Research Contract 3-38747 are highly appreciated.
ASTM (1995) Standard guide for risk-based corrective action applied at petroleum
release. American Society for testing and materials, ASTM E-1739
Bell CE, Kostecki PT, Calabrese EJ (1991) Review of state clean up levels for hydrocarbon contaminated soils. In: Kostecki PT, Calabrese EJ (eds) Hydrocarbon contaminated soils and groundwater, volL Lewis, Boca Raton, pp 77-89
Bonazountas M, Wagner JM (1984) SESOIL: a seasonal soil compartment model; prepared for the US EPA. Arthur D. Litde, Cambridge, MA
Connor JA, Nevin JP, Malander M, Stanley C, DeVaull G (1995) Tier 2 guidance manual
for risk-based corrective action. Groundwater Services, Houston, TX
Connor JA, Bowers RL, Paquette SM, Newell CI (1996) Soil attenuation model for derivation of risk-based soil remediation standards. Groundwater Services, Houston,
Domenico PA (1987) An analytical model for multidimensional transport of a decaying contaminant species. J Hydrol 91:49-58
Farmer WJ (1980) Hexachlorobenzene: its vapor pressure and vapor phase diffusion
in soi!. Soil Sci Soc Am J 44:445-450
Risk-Based Remediation of Contaminated Soil
Foster Wheeler (1998) RB CA Fate and transport models: compendium and selection
guidance. Prepared by Foster Wheeler Environmental Corporation for ASTM
Gaudet C, Brady A, Bonnell M, Wong M (1992) Canadian approach to establishing
cleanup levels for contaminated sites. In: Kostecki PT, Calabrese EJ (eds) Hydrocarbon contaminated soils and groundwater, voll. Lewis, Boca Raton, pp 49-66
Newell CJ, McLeod KR, Gonzales JR (1996) BIOSCREEN Natural attenuation decision
support system. User's Manual Version 1.3
Rifai HS, Bendient PB, Wilson JT, Miller KM, Armstrong JM (1988) Biodegradation
odeling at aviation fuel spill site. J Environ Eng 114:1007-1029
US EPA (1991a) Risk assessment guidance for superfund, voll. Human Health Evaluation Manual. Part B: Development of risk-based preliminary remediation goals,
EPA 540/R-92/003
US EPA (1991 b) MOFAT: a two-dimensional finite element pro gram for multiphase flow
and multicomponent transport, US EPA Robert S. Kerr Environmental Research
Laboratory, Ada Oklahoma, EPA/600/2-91/020
US EPA (1992) User's guide for the industrial source complex (ISC2) dispersion models.
Office of Air Quality Planning and Standards, Technical Support Division, Research
Triangle Park, North Carolina, EP4.8: IN2/V,2
US EPA (1995) SCREEN3 model user's guide. Office of Air Quality Planning and Standards, Emissions, Monitoring and Analysis Division, Research Triangle Park North
Carolina, EP4.8: SCR2/2
US EPA (1996) Soil screening guidance: technical background document, Office of
Emergency and Remedial Response, Washington DC, EPA/540/R-95/128
Subject Index
degradation 162
factors 103
processes 160
Absorption 236
Accelerated bioremediation 189
Adsorption 236
Advection 162
Aerobic bio degradation ofRDX 65
bio transformation of TNT 58, 59
Aerosol 249
Aged soils 26, 27
Aging 26,27
Agrobacterium radiobacter 47
Alcaligenes denitrificans 48
Alcaligenes eutrophus 39
Alpine soil 221
Anaerobic bio degradation 60
respiration 83, 86, 88
Antarctic soil 224
Aquifer 168
Arctic soil 216
Atrazine 46
Atrazine chlorohydrolase 47
Azoxy metabolites of TNT 72
Bedrock 178
Bioaccumulation 35
Bioattenuation screening
protocol 182
Bioaugmentation 90,100,194,217,223
Bioavailability 152, 188
limitations 26, 27
Biodegradation potentialIS, 25, 26
Biodegradability 57
Biodiversity 10, 135
Bioenergetics 86
Biofilter 237
Biofiltration 235
Bioinorganic chemistry 84
Biological diversity 152
ligands 87
treatment 2
wastes 5
Biologically relevant elements 83
Biomass yield 139, 244
Biomix 246
Biopiles 16, 196,220
Bioreactor 204
Bioremediation technologies 187
Bioremediation process development
configuration 7,8
cost estimates 109
field testing 106-107
monitoring 107-108, 111
requirements 89-90
site restoration 90
tools 100
Bioremediation of metals
abiotic influences 89, 103
hydrological influences 90-94
geochemical influences 95-98
physicochemical 98-100
Bioscrubber 238
Biosparging 198
Biostimulation 90,105,108,176,
Biotrickling bed 239
filter 238
Biotreatment technologies 15
Bioventing 8, 15, 198,230
Breeding 145
BTEX 168, 256
isopleth map 170
Bulking agents 194
Subject Index
Cadmium 147
Callus 143
Carbamate hydrolase 45
Carbamates 45
Carbon dioxide 83,97,99,110
Carcinogenic risk agents 261
Cell suspension 143
Chelatobacter heintzii 47
Chemical remedia ti on 3
structure 56
weapon 142
Chiral herbicides 48
Chlorinated ethanes 179
solvent 166
Chlorophenoxy acids 38
C:N:P ratio 20
Cold-adapted 215
Cometabolic substrate 191
Cometabolism 6, 24, 25
Community dynamics 101
ecology 106, 110
evolution 101-103, 111
Commercial bacteria 24
Composting 16,199,230
Condensation 236
Constructed wetlands 201
desorption 26, 27
loading 22
plumes 91
sequestration 26, 27
Controlled bio stimulation 108
Critical micellar concentration,
CMC 192
Cyclic nitramine explosives 60
Cytochrome p450 inhibition 66
DDD 41
DDT 40
Decontamination 35,116
Denitration 64
Denitrification 165
Dense non-aqueous phase liquid
Dermal contacts 264
Desert soil 228
Destructive mechanisms 160
Detoxification 35, 45
Diffusion 162
Dimethyl sulphide 246
Direct ingestion 264
inhalation 261
Dispersion 162
Dissimilatory metal-reducing
bacteria 88
reduction 86
Dissolved oxygen 164
Electron acceptor 164, 190
Empty bed residence time (EBRT) 240
Endosulfan 43
Energy yielding reactions 164
Enrichment cultures 24
Environmental contaminants 4
Environmental Protection Agency
(EPA) 195,259
Environmental toxicity 151
endpoints 27,28
Enzymes in RDX biodegradation 66
Ethylene 149
Evolutionary microbiology 102
Explosive-contaminated soil 73
Explosives 55, 56
Expressed assimilative capacity 175
Exudate 141,150
Fate and Transport model 272
Field capacity 18
Foot print 241
Freeze-thaw-cycles 217
Frozen soi! 216
FTIR 246
Fungi as TNT degrader 71,72
Gamma-hexachlorocyclohexane 42
Genetic engineering 138, 142, 145, 152
Genetically engineered Rhodococcus 46
Geochemical indicators of
bio degradation 171
Geochemistry 175
Geology 168
Gibbs free energy 163
Green rust 88
Groundwater flow 92
Hairy root 144
Harvested biomass 139
Hazard index 267
quotient 260
Subject Index
Histidine 148
HMX 60
Hydrocarbon degraders 23,24
Hydrocarbon spills 13,14
Hydrocarbon-contaminated soil 2
Hydrogen peroxide 191
Hydrogeologie setting 93
Hydrogeological test 169
Hydrogeology 168
Hydrolases 36
Hydrophobic-lipophilic balance,
HLB 192
Hyper-accumulator 136,144
Immobilized cells 45
In situ bioremediation 218,226
subsurface bioremediation 197
In vitro cultures 143
Incineration 236
Inhibition ofbiodegradation 22
by salts 23
by heavy metals 23
Inoculation 24
Inorganic media 242
pollutant 135
Intrinsic bioremediation 160
Iron ores 89
reduction 165
respiration 88
Iron-respiring bacteria 88
Klebsiella pneumoniae
63, 64
Laboratory treatability study 176
Landfarming 196,202,220,227
Land treatment 8
Land Treatment Units (LTU) 16
Leaching 17
factor 268
Lewis acids 85
Lignin peroxidase 43
Limiting factors 231
Lindane 42
Look-up table 265
Magnesium peroxide 191
Manganese reduction 165
Mass loading 241
Mecoprop 48
Media 238
Meisenheimer complex of TNT 69
Mercury 146
Metal accumulation 137
bioremediation 106
binding 139
biotransformation 100
detoxification 87
ore 136
respiration 86
Metallic pollutants 81
Metallothionein 147
Methanogenic granular sludge 43
Methylenedinitramine 62
Microbial transformation 5
communities 128
Microcosm studies 176
Microorganisms in sediment 67
Microtox 28
Mineralization ofRDX 61
ofTNT 72
Modeling 244
Modular biofilter 239
Moisture content 188
Monitored natural attenuation 159
Monitoring well 177
Munitions-contaminated soil 200
Mycorrhizal fungi 141,149
Natural attenuation 159, 197,223
factors 266
Natural organic material 95, 104
4-Nitro-2,4-diazabutanal 65
Nitro reduction 69
Nonylphenol ethoxylate
Nitrogen 20
Nutrients 217,227
Odor 235
Oil spill 2
Oleophilic fertilizers 201
On site bioremediation 220, 227
Organic pollutants 125
Organochlorines 38
Organophosphate hydrolase 44
Organophosphates 44
Oxidative dehalogenation 36
Oxidation-reduction potential 86,
Oxyanions 84
Subject Index
Oxygenases 36
Oxygen concentration
Public attitude 9
Push-puH test 106
Periodic table 81, 82, 94
Peroxidase 71
Persistent chemieals 189
Pesticide-contaminated soils 35
Petroleum composition 14, 189
contaminated soils 187
hydrocarbons 215
Phanerochaete chrysosporium
Phosphorus 20
Photooxidation 17
40, 71
Physical remediation 3
Physicochemical methods 187
Physiological adaptation 102
Phytorextraction 118
Phytofiltration 119
Phytodegradation 120
Phytocover 120
Phytoremediation 1,3, 115, 120, 129,
processes 206
technologies 7
Phytovolatilization 121
Pesticides 123
Phytoche1atin 137
Phytomining 152
Plume 168
Point of exposure 267
Polychlorinated biphenyls PCBs
Radioactive elements 81,90
Radionuclide 4,81, 138
Ralstonia eutropha 6, 39
RDX biodegradation 63
degraders 61
metabolism 64
metabolit es 62
Recalcitrance to biodegradation 25, 26
Recalcitrant 189
Recharge rate 92
Redox reactions 83
Reduction of aromatic ring ofTNT 69
Reductive dechlorination 41
dehalogenation 166
Regulations 9
Regulatory standards 256
Remediation technologies 271
Remediation technology database 269
Removal rate 240
Rendering 245
Rhizodegradation 119
Rhizoremediation 140
Rhizosphere 140
Rhodococcus 65
Risk agents 255
assessment 28, 260
Risk-based corrective action 259
remediation 255
screening levels 261
126, 187
Polychlorinated dibenzodioxins,
PCDD 187
Polychlorinated dibenzofurans,
PCDF 187
Polycyclic aromatic hydrocarbons,
Population exposure 266
Prediction of bio degradation
potential 25, 26
Pr in ted circuit board 245
Production weH 167, 169
Promutagen 140
Pseudomonas paucimobilis 42
Pseudomonas pseudoalcaligenes
Pseudomonas putida 39,44
Psychrophilic 215
Psychrotrophic 215
Safety procedure 73
Sampling and monitoring
protocols 195
Secondary metabolites 150
Selenium 147
Sequestration 94
Siderophore 88, 137, 141, 145, 149
Site characteristics 188
restoration 90
Site-specific factors 256
target levels 266
Slurry bioreactor 16,202
Soil cleanup levels 257
aeration 19
cation exchange 95
composition 95-96
moisture 18
Subject Index
Thermal insulation system 220
Thermophilie biodegradation 21,229
TNT biotransformation 68
metabolites 58, 59, 68, 70
Transgenic plants 146
Transport models 269
Treatment train 6
Trichloroethane 177
particle size 95,98
pH 95,99
pile 199
quality 28
scaling 98-99
toxicity 28
type 188
water activity 96
Sorption 17,162
Solubilizing agents 193
Sorptiveness 94
Sphingomonas herbicidivorans 39
s-Triazines 46
Structural property of explosives 57
Subsurface environment 99
microbiota 111
monitoring 107
water 81,91-93
Sulfate reduction 165
Surfactants 192
Sustainable development 135, 152
Symbiotic bacteria 148
Temperature 21
Terminal electron acceptor
Upflow anaerobic sludge blanket
Uranium 137
Vapor emission factor 268
Vinyl chloride 180
Volatile organic carbons (VOCs)
Volatilization 162, 190
Warfare agent 142
Wastewater treatment
Wetlands 142
Wood media 249
8, 136
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