United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/9-90/041
Dec. 1990
xvEPA
Bioremediation of
Hazardous Wastes
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EPA/600/9-90/041
Dec. 1990
BIOREMEDIATION OF HAZARDOUS WASTES
BIOSYSTEMS TECHNOLOGY DEVELOPMENT PROGRAM
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
U.S. Environment* r -Jon A^.ncy
Region 5, Lifcr?.r•
77~Wesi; Jc^c' ' "-
Chicago, iL b-. • .
U.S. ENVIRONMENTAL PROTECTION AGENCY
Ada, OK; Athens, GA; Cincinnati, OH; Gulf Breeze, FL;
and Research Triangle Park, NC
Printed on Recycled Paper
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Bioremediafion of Hazardous Wastes
DISCLAIMER
The information in this document has been funded wholly or in part by the U.S. Environmental Protection Agency. It has
been subjected to the Agency's peer and administrative review by the respective laboratories responsible for the research
presented in authors' abstracts, and approved for publication as an EPA document. Mention of trade names or commercial
products does not constitute endorsement or recommendation for use.
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Bioremediation of Hazardous Wastes
EXECUTIVE SUMMARY
Technologies for cleaning up hazardous wastes are often expensive, inappropriate for the site, or ineffective in handling
complex mixtures of pollutants. Some of the most promising of the new technologies for solving hazardous waste problems
involve the use of biological treatment systems. Because biological treatments appear to offer solutions to problems as-
sociated with conventional technologies, EPA's Office of Research and Development (ORD) initiated the Biosystems Tech-
nology Development Program, which is designed to anticipate rapidly increasing research needs that can be applied to our
nation's enormous waste management problems. In February 1990, ORD hosted a Conference on Bioremediation of Hazard-
ous Wastes in Arlington, Virginia, to discuss recent achievements of the Biosystems Technology Development Program and
research necessary to bring biosystem technology into more widespread use. This document outlines the program, its objec-
tives, and accomplishments. Extended abstracts are included to provide information on existing research projects.
Biological treatment systems use microorganisms, such as bacteria or fungi, to transform harmful chemicals into less
toxic or nontoxic compounds. Biotransformation is an attractive option because it depends on natural processes, and process
residues, such as carbon dioxide and water, can be cycled within the biosphere. In many cases, these technologies are also
less expensive and less disruptive than options commonly used to remediate hazardous wastes, such as excavation and in-
cineration. Bioremediation also holds a clear advantage over many technologies relying on physical or chemical processes
because it involves the destruction of contaminants, not merely transference among media.
Early applications of biological treatments on sites containing relatively easily degradable compounds show that
bioremediation technologies may have more widespread application for the cleanup of hazardous waste sites across the
country as well as the 10,000-15,000 oil spills that occur each year. Achievements thus far have only begun to tap the poten-
tial of biological treatments, but they point to the rich opportunities for bioremediation to offer alternative—and often
preferable—technologies for cleaning up hazardous wastes.
For the future, the Biosystems Technology Development Program must focus on a number of key issues. First, most
waste sites contain complex mixtures; yet the ability of biological systems to degrade hazardous chemicals is limited to com-
pounds of low structural complexity or specific chemicals. Research is needed to develop and engineer technologies that are
appropriate for conditions found at different sites. To meet these needs, the biosystems program has identified research goals
in six key areas: process characterization, process development, process engineering, environmental risk, mitigation of ad-
verse consequences, and technology transfer.
Under these six research areas, the Biosystems Technology Development Program has chosen seven media-based or
process-oriented topics for initial efforts:
• Liquid Reactors. In liquid reactors, toxic and hazardous pollutants in liquid form are brought into contact with
microorganisms to accelerate the degradation process. Landfill leachates are a good example of a liquid waste that is
amenable to liquid reactor treatment. In addition, toxic leachates from Superfund sites can be treated in liquid reactors.
EPA researchers are exploring use of liquid reactors at publicly owned treatment works (POTWs) as a model for such
reactors.
• Ground-Water Treatment. Effective bioremediation of ground water is constrained by the geology, hydrology, and
geochemistry of the subsurface environment. Design of any effective remedial action for contaminated ground water
requires a great deal of site-specific information. Ground-water research currently is focused on enhancement of
degradation through injection of substrate, nutrients, and microorganisms into pcrfusion well systems. Furthermore,
aerobic treatment of contaminants is generally preferable to anaerobic treatment because of the more rapid degradation
rate. However, because oxygen has limited solubility in water, contaminated ground water is frequently anaerobic.
This problem can be offset by artificially supplying oxygen to the pollutant plume so that aerobic degradation is not
limited by a lack of oxygen.
• Soil/Sediment Treatment. Decontamination of soils and sediments is one of the most difficult problems found at haz-
ardous waste sites: no current technology is adequate to handle soil cleanup, usually because of the cost involved or
III
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Bioremediation of Hazardous Wastes
the unsuitability of the technology to site environments that are contaminated with a complex mixture of pollutants.
Three industrial chemicals are currently the focus of EPA biosystems research: pentachlorophenol (PCP), polycyclic
aromatic hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs).
• Combined Treatment. Most hazardous waste sites contain complex mixtures of biologically persistent organic and
inorganic contaminants that can be remediated only by a combination of treatment techniques. EPA researchers are
developing methods to combine various physical, chemical, and biological treatment technologies, and comparing the
effectiveness of the various combinations.
• Sequential Treatment. Sequential treatment is generally applied to two waste types: compounds that degrade into
stable intermediates that can be further degraded under different conditions than those used for the parent compound;
and complex mixtures of wastes, which are generally degraded in order of their thermodynamic behavior. EPA re-
searchers are investigating the most effective coupling and sequence of treatments for such wastes.
• Metabolic Processes Research. EPA's metabolic processes research generates a better understanding of the proces-
ses by which microorganisms degrade chemicals, expanding the range of organisms that can be used in biosystems
technologies. Based on the insights gained from this research, scientists can then choose indigenous organisms or en-
hanced organisms to meet needs in pollution cleanup and control.
• Risk Assessment A number of the high-priority compounds that require disposal are known carcinogens or procar-
cinogens (i.e., they must be metabolized to the ultimate carcinogen form before they cause genotoxic effects). In the
area of human health, ORD is focusing on short-term genotoxicity testing to provide the necessary data for risk assess-
ment and on the development of an appropriate animal model to determine fate and effects of carcinogens.
Research on the first six of these topics involves the areas of process characterization, development, and engineering.
The last topic, risk assessment, is a prerequisite for field testing or larger-scale application of any biological treatment.
Through these research projects, the Biosystems Development Technology Program will develop and demonstrate biological
treatments as nondisruptive, cost-effective, efficient control technologies for hazardous wastes.
The Biosystems Technology Development Program draws on ORD scientists who possess unique skills and expertise in
biodegradation, toxicology, engineering, modeling, biological and analytical chemistry, and molecular biology. Participating
laboratories and organizations follow:
Environmental Research Laboratory - Ada, OK
Environmental Research Laboratory - Athens, GA
Environmental Research Laboratory - Gulf Breeze, FL
Health Effects Research Laboratory - RTP, NC
Risk Reduction Engineering Laboratory - Cincinnati, OH
Center for Environmental Research Information - Cincinnati, OH
Research conducted by the EPA laboratories promotes bioremediation technology through the demonstration that
laboratory data can be applied to field situations. In this process, strong interactions are fostered with the private sector,
providing research information that is valuable in the use of bioremediation for site treatment. Future interaction with EPA
programs, such as SITES and private sector efforts, will be an integral feature of the Biosystems Technology Development
Program.
In summary, although the Biosystems Technology Development Program is relatively new, several significant research
milestones have already been met and are presented in the abstracts that follow. Several new approaches developed in the
laboratory are being considered for field demonstration. Conceptually, the program has fostered understanding of new tech-
nology through a series of workshops for EPA regional personnel. The program also develops protocols, methods, and
guidelines for assessing bioremediative technologies.
IV
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Bioremediation of Hazardous Wastes
COlSfTENTS
INTRODUCTION l
SECTION I: TREATMENT OF AQUEOUS WASTES IN A REACTOR
Use of a White-Rot Fungus in a Rotating Biological Contactor 5
Performance of a Recirculating Bioreactor for the Degradation of TCE 6
Treatment of CERCLA Leachates in POTWs: Biodegradation of Volatile Organics
in a Biofilter 8
Treatment of CERCLA Leachates in POTWs: Anaerobic/Aerobic Sequential
Treatment 10
SECTION D: GROUND-WATER TREATMENT
/n&'mBiorestorationofaFuelSpill 13
Enhanced In Situ Biotransformation of Carbon Tetrachloride Under
Anoxic Conditions 15
SECTION ffl: SOIL/SEDIMENT TREATMENT
Use of White-Rot Fungi to Remediate Soils Contaminated with Wood-Preserving
Waste 20
Aerobic Biodegradation of PAHs 23
Anaerobic Degradation of PCP and Other Chlorinated Compounds 25
Aerobic Degradation of Polychlorinated Biphenyls: Biochemistry 26
Aerobic Degradation of Polychlorinated Biphenyls: Genetics 26
Anaerobic Degradation of Phenolic Priority Pollutants: Effects of Different
Reducing Conditions 28
Surfactant/Sorption Effects upon Biodegradation 29
Use of Fluorinated Analogues to Show Anaerobic Transformation of Phenol to
Benzoateviapara-Carboxylation 30
Treatment of CERCLA Leachates in POTWs: Innovative Anaerobic Pretreatment 33
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Bioremediation of Hazardous Wastes
SECTION IV: COMBINED TREATMENT
Use of KPEG and Anaerobic Biodegradation for PCB-Contaminated Soils:
Enrichments for Anaerobes Capable of Degrading KPEG-Treated PCB 35
Use of KPEG and Composting for Treatment of Contaminated Soils 37
Combined KPEG and Biological Treatment of Soils Contaminated with Chlorinated
Dioxins and Phenoxyacetates 37
SECTION V: SEQUENTIAL TREATMENT
Anaerobic Biodegradation of Creosote Contaminants in Natural and Simulated
Ground-Water Ecosystems 39
Development of a Sequential Treatment System for Creosote-Contaminated Soil and
Water: Bench Studies 42
Anaerobic Treatment of Dioxins and Dibenzofurans 46
Degradation of Heterocyclics 46
SECTION VI: METABOLIC PROCESS CHARACTERIZATION
The Involvement of a Toluene Degradative Pathway in the Biodegradation of
Trichloroethylene by Pseudomonas cepacia Strain G4 49
Degradation of Halogenated Aliphatic Compounds by the Ammonia-Oxidizing
Bacterium Nitrosomonas europaea 52
Degradation of Chlorinated Aromatic Compounds Under Sulfate-Reducing
Conditions 54
The Role of Fungal Lignin-Degrading Enzymes in Aromatic Pollutant
Biodegradation 55
SECTION VII: RISK ASSESSMENT
Developing Genotoxicity Risk Assessment Data 57
VI
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Bioremediation of Hazardous Wastes
INTRODUCTION
Background
The U.S. Environmental Protection Agency (EPA) is responsible for protecting public health and the environment from
the adverse effects of pollutants. EPA's authority to develop regulations and to conduct environmental health research is
derived from major federal laws passed over the last 20 years that mandate broad programs to protect public health and the
environment. Each law—including the Clean Air Act, the Safe Drinking Water Act, the Clean Water Act, the Toxic Substan-
ces Control Act, the Federal Insecticide, Fungicide, and Rodenticide Act, the Resource Conservation and Recovery Act, and
the Comprehensive Environmental Response, Compensation and Liability Act (CERCLA, known as Superfund)—requires
that EPA develop regulatory programs to protect public health and the environment.
For the control and cleanup of hazardous wastes, the Superfund law gives EPA broad authority to respond directly to
releases of hazardous materials that endanger public health or the environment. Also, the Superfund Amendments and
Reauthorization Act of 1980 (SARA) expands EPA's authority in research and development, training, health assessments,
community right-to-know, and public participation. EPA's Office of Research and Development (ORD) conducts basic and
applied research in health and ecological effects, hazardous wastes, and remediation development and demonstration of con-
trol technologies. Technologies are designed to provide efficient, cost-effective alternatives for cleaning up the complex mix-
tures of pollutants found at Superfund sites or at other locations, such as oil spills. As the technologies advance, ORD
transfers information on their use and enhancement to groups that apply technologies at specific sites.
Some of the most promising of the new technologies for handling hazardous wastes are biological treatments. In
February 1990, ORD hosted a bioremediation conference to explore recent achievements and to highlight research necessary
to bring these technologies into widespread use. This document highlights the research program and includes extended
abstracts presented by cooperating scientists.
Biosystems: Less Expensive, More "Natural" Control Technologies
Current technologies for cleaning up hazardous wastes are often expensive, inappropriate for the site, or ineffective in
handling complex mixtures of pollutants. Because biological treatments appear to offer solutions to these problems, ORD has
initiated the Biosystems Technology Development Program, which is designed to anticipate research applicable to our
nation's waste management. A promising application of the biosystems program was the cleanup of portions of the shoreline
of Prince William Sound, Alaska, after the March 1989 Exxon Valdez tanker accident. The project demonstrated EPA's
ability to work with private industry1 and academia to develop and transfer expertise in handling hazardous wastes.
Biological treatment uses microorganisms, such as bacteria or fungi, to transform harmful chemicals into less toxic or
nontoxic compounds. To the microorganisms, pollutants can serve as energy sources that they can break down to obtain ener-
gy to live and reproduce. These organisms have a wide range of abilities to metabolize different chemicals; scientists can
tailor the technology to the pollutants at specific sites and in specific media (e.g., contaminated aquifers, waste lagoons, con-
taminated soils) by using an organism in the treatment system that breaks down a particular pollutant. Where possible, tech-
nologies are developed to utilize native microorganisms that have been demonstrated to metabolize the pollutants on the site.
In these cases, the number and/or the rate of degradative activity of the microorganisms—and thus the speed at which a pol-
lutant is broken down—may be increased by adding nutrients or other amendments to the site. In other cases, organisms
known to metabolize the pollutants can be introduced and supplemented if necessary to accelerate biodegradation.
The Federal Technology Transfer Act of 1986 provided a mechanism whereby industry and EPA could share research costs and
results.
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Bioremediation of Hazardous Wastes
Biodegradation is an attractive option because it is "natural," and the residues from the biological processes (such as carb-
on dioxide and water) are usually geochemically cycled in the environment as harmless products. These processes are also
carefully monitored to reduce the possibility of a product of a process being more toxic than the original pollutant. Another in
situ advantage of biological treatments—particularly in situ treatment of soils, sludges, and ground water—is that they can be
less expensive and less disruptive than options frequently used to remediate hazardous wastes, such as excavation followed by
incineration or landfilling. Other methods of applying biological treatments, such as spreading contaminated soils on control-
led land plots or mixing sludges with water for treatment in contained vessels, are currently used with varying degrees of suc-
cess, depending on the chemical contaminants and environmental conditions. Additional research in such technologies will
broaden their applicability and effectiveness. Finally, bioremediation holds another clear advantage over many technologies
relying on physical or chemical processes: Instead of merely transferring contaminants from one medium to another, biologi-
cal treatment can degrade the target chemicals.
Across the United States, state, federal, and private hazardous waste sites are currently slated for cleanup under the Su-
perfund law. In addition, approximately 15% of the nation's four to five million underground tanks that store petroleum, heat-
ing oil, and other materials are leaking. Many more underground tanks will begin to leak in the next 5 to 10 years. Other
sources of contamination are the 10,000-15,000 oil spills that occur each year, requiring cleanup.
Applications of biological treatments on sites that contain relatively easily degradable compounds have demonstrated ef-
fectiveness of bioremediation technologies. In the United States and other countries, organic contamination of soils at hazard-
ous waste sites has already been biologically treated on-site and in situ. Biological systems also have been used to treat soils
and aquifers contaminated by hydrocarbons, phenols, cyanides, and chlorinated solvents, such as trichloroethylene. For ex-
ample, in Conroe, Texas, EPA and Rice University personnel used microorganisms in a disposal pit contaminated with
creosote compounds to degrade five of these compounds. In Arkansas, bioremediation with native organisms on site was used
to clean up three million gallons of acrylic acid and butyl acrylate originating from a railroad spill. So far, biological treat-
ments make up 4% of source control remedies chosen on Superfund sites and 8% of ground-water clean-up remedies chosen.
These achievements have only begun to tap the potential of biological treatments, but they point to the rich opportunities
for the emerging technologies. For this promise to be realized, the Biosystems Technology Development Program must
remain focused on two key issues. First, most Superfund and hazardous waste sites contain complex waste mixtures, and
have the ability of microorganisms that can degrade hazardous chemicals. Second, research is needed to develop and engineer
technologies that are appropriate for the variety of conditions found at different sites. Many forms of wastes are physically
difficult to handle; each site has its own geology and geomorphology; and, while some wastes degrade aerobically, others are
best degraded anaerobically. Therefore, engineering technologies must be developed and applied to accommodate these
biological site variables.
In focusing on these primary issues, the biosystems program has identified research goals in six key areas:
• Process characterization: Isolate and identify microorganisms that carry out biodegradation processes. Search out
and characterize biodegradation processes in surface waters, sediments, soils, and subsurface materials in order to
identify those that may be used in biological treatment systems and develop process-based mathematical models to
evaluate potential treatment scenarios.
• Process development: Develop new biosystems for treatment of environmental pollutants. Biosystems would include
naturally selected microorganisms, consortia, bioproducts, and genetically engineered microorganisms.
• Process engineering: Determine, evaluate, optimize, and demonstrate the engineering factors necessary for applying
biological agents to detoxify or destroy pollutants in situ or at a centralized treatment facility.
• Environmental risk: Determine environmental fate and effects of, as well as the risks involved in the use or release
of, degrading microorganisms or their products to detoxify or destroy pollutants.
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Bioremediation of Hazardous Wastes
• Mitigation of adverse consequences: Develop means to mitigate adverse consequences resulting from the accidental
or deliberate release of microorganisms for pollution control.
• Technology transfer: Transfer information on the technology and its use. Provide technical assistance to appropriate
requestors.
Initially, the program will concentrate on process characterization and development and engineering. Additional
microorganisms that degrade hazardous chemicals must be identified and the metabolic processes by which they work must
be understood. Methods and procedures are also needed to enhance the rate, extent, and scale of such processes. In process
engineering, reactor systems and treatment strategies for optimizing and applying the biodegradation processes must be en-
hanced and expanded.
The program must also acquire experience in taking biological treatment technologies to the field. The first step in field
testing is to characterize the site, its contaminants, and various influential environmental factors, such as hydrology and
geomorphology. A research team would establish the treatment criteria and environmental risk of the treatment technology
proposed for the site. To assess the environmental risk associated with various biological treatments, the biosystems program
must develop toxicity screening procedures and perform further research on the biochemistry of degradation. Finally, the pro-
gram will provide assistance to EPA program offices, regional offices, and commercial firms (through Federal Technology
Transfer Act activities) in efforts to enhance and further evolve practical bioremediation technologies.
Under the six research areas outlined above, the Biosystems Technology Development Program has chosen seven media-
based or process-oriented topics for initial efforts: liquid reactors, ground-water treatment, soil/sediment treatment, combined
treatment, sequential treatment, metabolic processes research, and risk assessment Research on the first six of these topics
cuts across the areas of process characterization, development, and engineering. The last topic, risk assessment, is a prereq-
uisite for field testing or larger-scale application of any biological treatment.
This research program will foster the advances in bioremediation that are expected in the next few years. For example,
enhancing the activities of microorganisms to hasten the degradation of chemicals is expected to become commonplace. One
promising bioremediation procedure under investigation is the isolation of enzymes from microorganisms that catalyze the
conversion/degradation of specific pollutants and their application to break down specific pollutants, such as organophos-
phates. Under the Federal Technology Transfer Act (FTTA), EPA and industry can cooperate in developing and marketing
biological treatment technologies. In this way, industries gain new products and access to new markets, and ORD, through
the Biosystems Development Technology Program, can develop and demonstrate biological treatments as nondisruptive, cost-
effective, efficient control technologies for cleaning up our environment.
This document is divided into seven sections covering the media-based or process-oriented topics that the Biosystems
Technology Development Program has selected for immediate effort. Each section contains abstracts of EPA projects under
topics discussed during the February 1990 ORD Biosystems Conference. A brief description of the research topic precedes
each section of this document.
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TREATMENT OF AQUEOUS WASTES IN A REACTOR
SECTION I
TREATMENT OF AQUEOUS WASTES IN A REACTOR
In liquid reactors, toxic and hazardous pollutants in liquid form are brought into contact with microorganisms
to accelerate the degradation process. Landfill leachates are a good example of a type of liquid waste that is
amenable to liquid reactor treatment. Nearly 160 million tons of solid waste are disposed of each year in landfills
across the nation, and various organic chemicals and ions often leach from the waste into drainage water. Toxic
leachates from landfills and Superfund sites can be treated in liquid reactors. EPA researchers are exploring how
to use liquid reactors at publicly owned treatment works (POTWs) as a model to demonstrate use of these
reactors. POTWs offer several advantages for treating toxic leachates: 1) a diverse biomass that can be easily
acclimated to any waste, 2) dilution, and 3) availability of nutrients (carbon, nitrogen, phosphorus, and other trace
organics required for metabolism).
Many operating POTWs have not taken full advantage of optimizing liquid waste treatment through
bioremediation. Pollutants still pass through the system and are released into salt or fresh water; aeration results
in air stripping of volatile toxicants into the air; and many of the toxicants associated with the residual sludges
are not completely dechlorinated and destroyed. In one current project in this area, EPA researchers are
examining how to pass the leachate through a biofilter, where the pollutants are filtered out and broken down by
microorganisms, before mixing the material with municipal wastewater. Another project demonstrated that a
combination of anaerobic and aerobic methods in a bench-scale treatment system can result in pollutant removal
ranging from 30-70%. Some compounds can be broken down aerobically, while others are best broken down
anaerobically; thus, an aqueous reactor treatment train, combining anaerobic and aerobic degradation, may be
optimal in certain situations.
USE OF A WHITE-ROT FUNGUS IN A
ROTATING BIOLOGICAL CONTACTOR
John A. Closer, Henry M. Tabak, and Edward J.
Opatken, US. EPA, Cincinnati, OH; Thomas W. Joyce,
Hou-min Chang, North Carolina State University,
Department of Wood and Paper Science, Raleigh, NC;
Susan Stroehofer and Carol Hummel, University of
Cincinnati, do US. EPA, Risk Reduction Engineering
Laboratory, Cincinnati, OH.
Since time immemorial, fungi have degraded waste
materials as part of a natural scheme of carbon recycling.
Humans have not made similar use of fungi to degrade
waste for a number of reasons: Many fungi encountered
in waste treatment systems are pathogenic and even the
benign forms disturb the normal processing of waste
materials; and, in activated sludge systems, fungi are
regarded as processing nuisances since they can cause the
precipitation of the sludge blanket
Wood-degrading fungi, however, offer significant
enough advantages that their role as degraders of waste
materials should be reconsidered. The white-rot wood-
degrading fungi, for example, can degrade one of the most
recalcitrant of biogenic molecules—lignin. That ability
can be turned to other uses: Basic similarities in chemical
structure are thought to exist between lignin and the
aromatic organic compounds that make up hazardous
waste; thus, these organisms should theoretically be able
to degrade hazardous waste materials. In fact, the white-
rot fungus Phanerochaete chrysosporium has been the
focus of much research exploring its utility as a degrader
of hazardous waste.
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TREATMENT OF AQUEOUS WASTES IN A REACTOR
This fungus can operate in two distinct metabolic
cycles. The primary, or growth, cycle utilizes carbon
substrates such as sugars or polymeric saccharides such as
cellulose. Depending on growth conditions and the
availability of certain nutrients such as nitrogen, the
fungus may adopt a secondary metabolic cycle in which
the organism secretes a complex mixture of peroxidases
commonly referred to as ligninases. These enzymes have
been shown to be genetically distinct; in other words, they
are not degradation products of one another, but rather
encoded in specific DNA sections. This production of
ligninases is the reason the white-rot fungus can degrade
lignin. The ligninases rely on a supplemental enzyme
system to supply the necessary hydrogen peroxide to start
the oxidation of lignin. Lignin has a random composition
and a high polymeric structure; therefore, the enzymes that
degrade lignin must have wide substrate ranges and low
specificity.
Due to the general aromatic nature of lignin and the
similar aromatic structure of many persistent pollutants,
early research efforts were designed to assess whether any
degrading activity could be elicited from the fungus. Now
a host of pollutant compounds have been tested for
degradation by the fungus, and the results continue to
support research and its extension to treatment technology.
This paper describes a research effort to develop a
biological reactor that incorporates the necessary condi-
tions of cellular morphology and physiology to support the
use of P. chrysosporium. The reactor is configured as a
rotating biological contactor (RBC), consisting of several
rotating disks on which the fungus can attach and grow to
support treatment cycles. One reason for this reactor
design was an observed shock sensitivity of this fungus
that resulted in a reduction of degrading activity. Further
studies reduced the shock sensitivity of the organism so
that other reactor configurations are now possible.
An extensive program has been undertaken to deter-
mine the viability of the RBC reactor for treating hazard-
ous wastes. Target compounds for treatment in this effort
were the organic contaminants associated with the wood-
preserving industry. In the U.S., 700 wood-preserving
sites have been identified that require cleanup; this cleanup
must handle a variety of compounds generated by a
succession of wood-preserving technologies, including
creosote treatment followed by pentachlorophenol and,
more recently, copper chromated arsenite. The tolerance
of the fungus toward the latter is as yet unknown.
The research program introduced here uses bench-
scale studies to determine the operational requirements of
larger-scale reactors. An early step in the program was to
examine the requirements for decolorizing Kraft liquor.
Early reports of efficiency and effectiveness come from
research conducted at North Carolina State University and
the U.S. Department of Agriculture, Forest Product
Laboratory. The program has highlighted interesting
possibilities for operating the reactors under ambient
oxygen concentrations, when reaction rates are slower than
under a pure oxygen atmosphere. Growth substrate has
been a point of major concern for the program since the
current quantities in field operations represent significant
biological oxygen demands and an economic burden for
treatment. Since the carbon growth substrate is glucose,
the problem may be more complex than first envisioned:
Glucose participates as a growth substrate and as an
intermediate supporting the generation of hydrogen
peroxide.
Bolh bench- and pilot-scale results will be presented
at a later time for the degradation of chlorinated phenols
and aromatic hydrocarbons present in creosote.
PERFORMANCE OF A RECIRCULATING
BIOREACTOR FOR THE DEGRADATION
OFTCE
Brian R. Folsom, Technical Resources, Inc., Gulf Breeze,
FL; Peter J. Chapman and Parmely It. Pritchard, U.S.
EPA, Environmental Research Laboratory, Gulf Breeze,
FL.
Of the volatile organic chemicals found as common
ground-water contaminants, trichloroethylene (TCE) has
received significant attention; and, as a result, a number of
bacterial systems with the ability to degrade TCE by
cometabolism are now recognized. In these systems, TCE
is degraded by bacterial enzymes that are typically
expressed following induction with other chemicals.
Previously, one of these TCE-degrading organisms was
isolated by investigators at the Gulf Breeze Environmental
Research Laboratory. This organism has subsequently
been identified as Pseudomonas cepacia strain G4 and
requires either toluene, o-cresol, /n-cresol, or phenol for
induction of TCE dcgradative enzymes(1). A novel
toluene degradative pathway involving sequential
hydroxylation of toluene at ortho and meta positions to
form 3-methylcatcchol has been characterized®. TCE is
completely degraded to CO2 , Cl~, and unidentified,
nonvolatile products by this organism(1).
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TREATMENT OF AQUEOUS WASTES IN A REACTOR
Although readily degraded under laboratory condi-
tions(1'3l4-5l6), TCE's persistence indicates that such activity
is limited in the environment. Limitations to the degrada-
tion need to be identified, characterized, and overcome in
order to promote TCE biodegradation either by supple-
menting contaminated sites(3) or by constructing bioreac-
tors. This research has focused on characterizing the TCE
degradative process, pinpointing the growth characteristics
of P. cepacia, and then testing a bioreactor designed to
degrade TCE. The kinetics of TCE degradation by intact
cells were investigated to establish (he effective range of
degradative activity and to identify potential limitations.
P. cepacia was grown in chemostats on phenol as the sole
carbon source. Cells were harvested and suspended in a
basal salts medium and the concentration dependent rates
of TCE degradation determined. The apparent Kg and
Vmax values for TCE degradation were 3 (iM and 8
nmole/(min-mg cell protein), respectively. Following a
transient lag period, P. cepacia was observed to degrade
TCE with no apparent retardation in rate at initial concen-
trations of at least 300 jiM. Consistent with similarities
in Ks values for TCE and phenol (between 5 and 10 jiM),
a significant inhibition of phenol degradation by TCE was
observed. At equal concentrations of TCE and phenol, the
rate of phenol degradation was
inhibited by about 50%. Effects of
phenol on TCE degradation rates
were much more difficult to quanti-
fy, since phenol was degraded about
50 times faster than TCE. Even so,
phenol was observed to transiently
inhibit TCE degradation until the
concentration of phenol dropped
below that of TCE in solution.
These results demonstrate that P.
cepacia exhibits favorable TCE
degradation kinetics, supporting the
proposal that TCE is amenable to
bioremediation.
The next stage of this investiga-
tion was to characterize the growth
characteristics of P. cepacia in con-
tinuous culture. Chemostats were
operated at a variety of dilution rates
on 5 mM phenol as the sole carbon
source. Specific degradation rates
for phenol and TCE remained con-
stant for residence times ranging
from 5 to 30 hr, with washout oc-
curring at residence times of less
than 5 hr. Several different chem-
ostats were set up and operated with
variable feed concentrations of phenol and/or lactate.
Increased phenol concentrations led to increased biomass
production with specific degradation rates for both phenol
and TCE remaining constant. Although biomass produc-
tion increased for chemostats operated on increasing
concentrations of lactate and fixed concentrations of
phenol, the specific degradation rates for phenol and TCE
decreased. Only at high dilution rates or high phenol feed
concentrations was the steady-state concentration of phenol
in the chemostat detectable (about 1 (iM). These growth
characteristics demonstrated that TCE degradation poten-
tial depended on the total biomass and the level of enzyme
induction. The maximum potential for TCE degradation
was determined to be 300 mg/day for a 1-L reactor operat-
ing with cells grown on 20 mM phenol.
Growth parameters outlined above in conjunction with
the previously determined degradation kinetics were used
to design a bench-scale reactor for testing the performance
of TCE bioremediation. Issues considered in the design
included induction of requisite catabolic enzymes, growth
substrates leading to maximal degradative activity, inhibi-
tion of TCE degradation by phenol, and the volatility of
TCE Reactor
phenol input
TCE input
TCE saturated basal media (8-10 mM)
I TCE layer
Figure 1. Recirculating TCE reactor.
-------
TREATMENT OF AQUEOUS WASTES IN A REACTOR
TCE. A recirculating reactor was constructed and tested.
TCE degradation rates determined for the reactor and for
washed cells gave essentially the same values. In the
reactor, the total amount of TCE that was degraded
increased with increased reaction time and increased
biomass, as expected. TCE degradation was observed up
to 300 \iM TCE with no significant decreases in rates, as
observed with washed cells. At low concentrations, below
5 jiM, there was a decrease in the measured reactor rate
constant that was consistent with concentration dependent
rate changes as Kg for the system is approached. Approx-
imately 40 mg TCE/day was degraded for a 1-L reactor
with the typical biomass of 60 mg cell protein/L.
This recirculating reactor was designed to minimize
limitations established previously and to allow for direct
comparison of results obtained using the washed cell assay
system. These results demonstrated that TCE degradation
kinetics determined for washed cells could be directly used
to design and evaluate the performance of a bioreactor
system. Furthermore, this study demonstrated the feasibil-
ity of TCE bioremediation.
References
1. Nelson, M.J.K., Montgomery, S.O., O'Neill, E.J., and
P.H. Pritchard. 1986. Appl. Environ. Microbiol.
52:383-384.
2. Shields, M.S., Montgomery, S.O., Chapman, PJ.,
Cuskey, S.M., and Pritchard, PJ. 1989. Appl.
Environ. Microbiol. 55:1624-1629.
3. Vogel, T.M., Griddle, C.S., and McCarty, P.L. 1987.
Environ. Sci. Technol. 21:722-736.
4. Wackett, LJP. and Gibson, D.T. 1988. Appl. Envi-
ron. Microbiol. 54:1703-1708.
5. Wilson, J.T., and Wilson, B.H. 1985. Appl. Environ.
Microbiol. 49:242-243.
6. Winter, R.B., Yen, K.M., and Ensley, B.D. 1989.
Biotechnol. 7:282-285.
TREATMENT OF CERCLA LEACHATES
IN POTWs: BIODEGRADATION OF
VOLATILE ORGANICS IN A BIOFILTER
Rakesh Govind and Vivek Utgikar, Department of
Chemical Engineering, University of Cincinnati,
Cincinnati, OH; Richard C. Brenner. US. EPA,
Cincinnati, OH.
The nearly 200 million tons of solid waste disposed
of in landfills is posing problems well beyond those
associated with finding sufficient landfill capacity.
Various organic chemicals and ions leach from the buried
solid waste into drainage water, where they form a
polluted leachate stream. The semi- and nonvolatile
compounds in this leachate stream can be treated in a
conventional activated sludge plant The volatile organic
compounds, however, are normally stripped from an
activated sludge system during the aeration stage and
passed into the atmosphere.
A possible solution to this problem is to treat these
compounds in a biofilter (see Figure 1) before they reach
a publicly owned treatment works (POTWs) and are
mixed with municipal wastewater. In this scenario, the
leachate stream is passed through a stripper, where it is
contacted countercurrently with air or nitrogen. The
volatiles are transferred from the aqueous to the gaseous
phase. The gas stream is then fed to the biofilter, where
the compounds are biodegraded. The biofilter is basically
a packed bed reactor containing biomass supported on a
medium. The support can be inert, such as peat, or it can
have adsorption properties, such as activated carbon. The
nutrients necessary for the growth of biomass are supplied
through an aqueous solution. The biofilter will operate
under steady-state conditions, degrading the volatile
organic compounds and thereby reducing their concentra-
tions in the gas stream to acceptable levels.
The compounds in the gas phase then diffuse through
the bulk gas to the gas-liquid interface and are dissolved
in the liquid phase, which is water. These compounds
diffuse through the water phase to the biomass, which is
a film over the supporting medium. The compounds
undergo diffusion with simultaneous degradation in the
biofilm. A concentration profile is formed for each
compound from the bulk gas phase to the support. A
theoretical model has been developed to describe the
biodegradation of selected volatile organic compounds by
biomass supported on a medium. The model accounts for
various resistances involved in the system, such as gas-
side mass transfer resistance, liquid-side mass transfer
8
-------
TREATMENT OF AQUEOUS WASTES IN A REACTOR
resistance, diffusion of substrate in the biomass, and the
kinetics of biodegradation. The model has been solved to
obtain the concentration profile of each substrate in the
system and the biodegradation rate as a function of system
properties. The various parameters required to describe
the system are:
1. A Thiele type modulus, <|>, to describe the kinet-
ics of biodegradation. <)> is an indicator of
relative rates of degradation and diffusion in the
biofilm.
2. A Thiele type modulus, <|>L, to describe the
external mass transfer resistance. <}>L is an indi-
cator of relative rates of diffusion in the external
film (gas-liquid film) and the biofilm.
3. A ratio of biomass growth to biomass decay, r,
which essentially indicates the magnitudes of
bacterial yield and decay.
This kinetic model is combined with the reactor
model for the biofilter, which is simply the equation for a
packed bed. The resulting equation is used to calculate
the biofilter height required for the desired reduction in
concentration of each volatile organic compound of
interest. Use of the model can be illustrated with the
following example. Degradation of trichloroethylene has
been studied by Folsom et al. (1989). From their data, the
pseudo first-order rate constant for degradation is found to
be 13.85 s"1 (mg protein)"1. Using values of <(> = 10 and
L = 250, the dimensionless degradation rate of 0.01 is
obtained from theoretical calculations. The actual degra-
dation rate of 1.42 x 10"3 p mol/m3s, where p is the partial
pressure of trichloroethylene in the gas in kPa units, is
obtained in the biofilter. At this rate, a biofilter 3 m in
diameter and 5 m in height will reduce the trichloroethyl-
ene concentration in the gas phase from 50 ppm to 18
ppm at a gas flow rate of 10.5 m3/min. Figure 2 shows
the concentration profile of trichloroethylene in the
biofilter.
The research team used this proposed model in
designing a biofilter, as described below. The model has
been solved for the cases of single substrates and multiple
substrates in the gaseous stream. The model predictions
will be verified by experiments with various systems. The
experimental approach is as follows:
1. Experiments to obtain the values of the parame-
ters and r. These values are obtained from the
values of the rate constant, bacterial yield, and
bacterial decay constant The experiments to
determine these values will be carried out in a
sapromat system.
2. Verification of steady-state rate predictions:
these experiments will be carried out to confirm
the model predictions using the values of the
kinetic parameters obtained above. These experi-
ments will be carried out in spouted-bed bioreac-
tors that are essentially continuous stirred tank
reactors.
3. Verification of biofilter design: the values of the
parameters obtained above will be used in the
design of a pilot-scale biofilter (50 mm dia. x
1,000 mm height). This biofilter will be operated
continuously to verify the utility of the design
equations.
Some of the volatile organic compounds involved in
this study are less susceptible to aerobic than to anaerobic
degradation. Consequently, both anaerobic and anaerobic
models of operation will be investigated. Nitrogen gas
and air will be used for stripping of volatiles from leach-
ates during the anaerobic and aerobic studies, respectively.
Leachate
Air
Treated Air
Biofilter
Stripper
To Activated
Sludge Plant
Pump
Figure 1. Schematic of a biofilter system.
-------
TREATMENT OF AQUEOUS WASTES IN A REACTOR
0123456
10
Height, m
Figure 2. Concentration profile of trichloroethylene in biofilter.
toxics into the ambient air, and many of the
toxics may not be completely dechlorinated and
destroyed. By combining anaerobic and aerobic
treatment (see Figure 1), plant operators may be
able to mitigate these problems while still using
existing tankage (with the addition of anaerobic
digestors, if necessary). In addition, this ap-
proach may reduce the costs of treating conven-
tional wastes at the plant.
To achieve the combined treatment, POTW
operators could retrofit the first stage on-line
expanded bed in the existing aeration basin
tankage at the head of the basin just after the
primary clarifier. The effluent from this stage
would then overflow to the remaining portion of
the aeration basin, where it would be polished in
a short-detention-time activated sludge process
before final clarification.
The first stage bed would be operated con-
tinuously as a contact/sorption stage, trapping
organics, colloidal material, and solids in granu-
lar activated carbon (GAC) media. The research
team selected GAC over more inert media (e.g.,
sand) as the best option for adsorbing toxics and
trapping particles. Contacl/sorption in this stage
could reduce the pass-through of volatiles and
toxics to the aeration basin and also reduce the
oxygen demand requiring treatment in the basin.
In addition, by removing particles that were not
removed in the primary clarifier but that could
contain sorbed toxics, this step would separate
these particles from the sludges generated in the
aeration basin so that they could be treated in the
second stage expanded bed.
TREATMENT OF CERCLA LEACHATES
IN POTWs: ANAEROBIC/AEROBIC
SEQUENTIAL TREATMENT
Margaret J. Kupferle and Dr. Paul L. Bishop. Univer-
sity of Cincinnati, Cincinnati, OH; Dolloff F. Bishop,
U.S. EPA, Cincinnati, OH.
Toxics and other pollutants in CERCLA leachates can
often be treated in publicly owned treatment works
(POTWs) with little pass-through into receiving waters.
If the POTW is operated in the conventional aerobic
mode, however, a variety of problems can result from this
form of treatment: Some toxics may still pass through the
system, aeration may result in air stripping of volatile
In the second stage off-line expanded bed
(see Figure 1), the system would anaerobically stabilize
the entrapped-sorbed pollutants associated with the media
from the first stage, thus effectively "regenerating" the
GAC for reuse in the first stage. To reduce heat losses
incurred during the exchange of GAC between the heated
second stage and the unheated first stage, the respective
GAC streams would be separated from their supernatants
and the supernatants would be exchanged before the GAC
exchange is completed. This step would minimize loss of
heated supernatant in the second stage and allow gas
generated in the second stage to be recovered as a source
of energy. The long overall retention time of biomass and
GAC in this system would encourage dechlorination and
destruction of recalcitrant toxics trapped and removed in
the first stage.
10
-------
TREATMENT OF AQUEOUS WASTES IN A REACTOR
The project team is currently conducting proof-of-
concept studies with two 100 L/day bench-scale systems.
The control system is being operated on primary effluent
only, while the test system is being acclimated to primary
effluent spiked with 5% landfill
leachate and a mixture of ten
volatile and five semivolatile
hazardous organic compounds.
Several first-stage experi-
ments have been run at different
hydraulic residence times
(HRT) using GAC from the
stabilization beds. In these
experiments, the team continu-
ously fed the first stage units
with primary effluent for sever-
al days and analyzed samples
for chemical oxygen demand
(COD) and solids removals with
time. Removals for both
ranged from 30% to 70% and
breakthrough was observed up
until termination of the experi-
ment (at approximately 48 hr).
The data suggested a carbon
retention time of 2 d as a prac-
tical upper limit for operation of
the contact/sorption stage to
prevent significant biodegrada-
tion of solubilizing organics
trapped in the first stage, and an
HRT of 30 min for the first
stage as a means of maximizing
COD and solids removals.
The team began operating
the two stages of the control
unit as an integrated system in
September 1989. However,
operational difficulties in the
carbon transfer step between the
first and second stages have
interfered with acclimation.
The affected portion of the
apparatus has been redesigned
and, once fabrication is com-
pleted, acclimation under condi-
tions of integrated operation
will resume. Acclimation of
the test system to the leachate
and spike compounds began in
early 1990. Bench-scale proof-
of-concept testing at the chosen parameters continued
through midyear.
To Aeration
Primary Basin
Effluent/
Leechate
+ TOXICS ' . ; .
3) gSJg; '—]
•o " :»» S
?^fe®>:
rate
rom
rnatant
pollutant-
GAC
35 °C
rnatant
rate
rom
rnatant
enerated
i ambient
r«turc
•natant
SEMI-CONTINUOUS
CARBON EXCHANGE
Figure 1. Conceptual schematic of proposed system.
11
-------
GROUND-WATER TREATMENT
SECTION II
GROUND-WATER TREATMENT
Effective bioremediation of ground water is limited by the geology, hydrology, and geochemistry of the
subsurface environment. A natural system of pores must be used to transport the remedy to the site of
contamination, and engineers must utilize local hydraulic conductivity to direct the biological treatment.
Furthermore, engineers must understand the geochemistry of the site to avoid chelation and precipitation and the
subsequent plugging of aquifers, and must control the hydrology of the system to ensure that the activity of the
microorganism(s) is directed to the site of contamination. Thus, the design of any effective remedial action for
contaminated ground water requires a great deal of site-specific information.
Aerobic treatment of contaminants is generally preferable to anaerobic because of a more rapid degradation
rate. However, because oxygen has limited solubility in water, contaminated ground water is frequently anoxic.
This problem can be offset by artificially supplying oxygen to the pollutant plume so that aerobic degradation is
not limited by a poor oxygen supply. Research in ground-water treatment is focused on these areas: 1)
overcoming mass-transfer limitations to supply oxygen or hydrogen peroxide to contaminated water; and 2) where
appropriate, promoting alternative anaerobic biological processes for destroying contaminants.
In one ground-water study, researchers are assessing whether they can stimulate indigenous microorganisms
in subsurface compartments to remove carbon tetrachloride from contaminated aquifers. Pilot-scale field
evaluations suggest this form of bioremediation will work. Another project deals with how to handle creosote that
has infiltrated into the soil around wood-preserving plants or has moved into the water table. Researchers have
shown that some of the organic compounds in creosote can be anaerobically degraded as they move through the
aquifer.
IN SITU BIORESTORATION OF A FUEL
SPILL
John T. Wilson, Stephen R. Hutchins. Wayne C.
Downs, RS. Kerr Lab, Ada, OK; Rob. H. Douglas,
Bill A. Newman, The Traverse Group, Inc., Traverse
City, MI; DJ. Hendrix, Solar Universal Technologies,
Traverse City, MI.
Nitrate-mediated biorestoration of a fuel-contaminated
aquifer is currently being field demonstrated at a U.S.
Coast Guard facility in Traverse City, Michigan. The
ground water at the site has been contaminated by leaks
from underground storage tanks containing JP-4 jet fuel.
This project is focused on a 30-ft by 30-ft infiltration area
located within the larger area contaminated by the JP-4
spill. An infiltration gallery was installed above the study
area that is part of a closed-loop system designed to
perfuse the study area with ground water supplemented
with nitrate and nutrients. The 30-ft by 30-ft section of
the site was instrumented with cluster monitoring wells
and piezometers. The research team installed a series of
pumping wells down gradient to intercept contaminants,
nutrients, and nitrate and provide hydraulic recirculation
back through the infiltration gallery. In addition, four
interdiction wells are in place to provide a net discharge
from the site and prevent escape of nitrate or contaminants
to regional flow in the aquifer.
To design the system, the team used hydraulic
modeling to evaluate the infiltration rate necessary to raise
the piezometric surface above the contaminated zone, the
13
-------
GROUND"WATER TREATMENT
Table 1. Changes in concentrations of target hydrocarbons in selected cores. Data are in
mg/kg dry weight, mean of two replicate subsamples of single cores. Data in
parentheses arc obtained from GC analyses and have not been quantitated by
GC/MS, and therefore may be subject to interference. The detection limits vary
based on dilutions and methods.
Core Elevation
(ft AMSL) and
Description Compound
607-608, Unsat'd,
Clean
603-604. Unsat'd,
Con lam
600-601, Water Table
599-600, Sat'd,
Conlam.
597-598, Safd, Clean
Benzene
2-Methylhexane
Methylcyclohex.
Toluene
Ethylbenzene
m,p-Xylene
o-Xylene
Pseudocumene
Benzene
2-Methylhexane
Methylcyclohex.
Toluene
Ethylbenzene
m.p-Xylene
o-Xylene
Pseudocumene
Benzene
2-Methylhexane
Methylcyclohex.
Toluene
Ethylbenzene
m^j-Xylene
o-Xylene
Pseudocumene
Benzene
2-Methylhexane
Meihylcyclohex.
Toluene
Ethylbenzene
m,p-Xylene
o-Xylene
Pseudocumene
Benzene
2-Methylhexane
Methylcyclohex.
Toluene
Ethylbenzene
m.p-Xylene
o-Xylene
Pseudocumene
TREATMENT
Prior to
Hydraulic
Loading
<0.003
(0-0)
<0.05
<0.05
<0.05
<0.05
<0.05
<0.05
0.42
(M.O)
373
15.9
14.7
52.1
22.7
74.5
32
(162)
380
184
85
247
116
222
0.059
(0.1)
0.30
0.53
<0.05
0.49
0.24
0.25
0.10
(0.0)
0.12
0.98
<0.05
0.24
0.19
<0.05
After Flood-
ing, Before
Nitrate
<0.002
(0.0)
0.08
<0.05
<0.05
<0.05
<0.05
<0.05
0.024
(12)
7.9
-------
GROUND-WATER TREATMENT
withdrawal rates necessary to retain the contaminants and
nutrients on-site, and the nutrient contact time important
to biological treatment. A tracer study was conducted to
confirm estimated breakthrough times and give a prelimi-
nary evaluation of the performance of the in-situ bioreac-
tor.
The effects of recirculation, wasting, and biodegrada-
tion on the decrease in solution concentration of BTX
compounds within the treatment zone were examined.
The aquifer was cored and analyzed for total petroleum
hydrocarbons and for the quantity of selected fuel hydro-
carbons (see Table 1). Water was recirculated through the
spill for two months before any nitrate or mineral nutrients
were added. This was done to bring the oil and recirculat-
ed waste to chemical equilibrium and allow an estimate of
the reduction in alkylbenzene concentration due to dilution
in the recirculated water, presumably due to biodegrada-
tion supported on ambient concentrations of oxygen and
nitrate in the recirculated water. After the addition of
nitrate, toluene was rapidly removed in the fuel spill, but
not in the recirculated water. Ethylbenzene and m-xylene
were also removed during denitrification; however, there
was little evidence for biodegradation of o- and p-xylene
until the end of the demonstration. As expected, minor
amounts of the alkane fraction were removed.
The reduction in BTX in the aquifer, as a result of the
demonstration, is depicted in Table 2. With the exception
of one of the five monitoring wells, benzene was removed
to concentrations much lower than the federal drinking
water standard. Toluene was removed in all the monitor-
ing wells. Removal of ethylbenzene and the xylenes
follow an interesting pattern. Ethylbenzene and the
different xylene isomers were persistent at low concentra-
tions in different wells. This suggests that there is no one
fixed pattern of alkylbenzene utilization, even at the same
site.
ENHANCED IN SITU
BIOTRANSFORMATION OF CARBON
TETRACHLORIDE UNDER ANOXIC
CONDITIONS
Lewis Semprini, Paul Roberts. Gary Hopkins, and
Perry McCarty, Department of Civil Engineering,
Stanford University, Stanford, CA.
The research team conducted a laboratory study and
pilot-scale field evaluation to assess the efficacy of an in
situ biostimulation approach for biologically removing
carbon tetrachloride (CT) from contaminated aquifers.
The process relies on the ability to biostimulate indigenous
microorganisms in the subsurface that are capable of
transforming CT under anoxic conditions.
Transformation of CT by pure and mixed cultures in
the laboratory has been reported under denitrifying
conditions*1-21 \ fermentative conditions*4'6*, and sulfate-
reducing conditions*2-5*. Generally, a primary substrate for
growth—often acetate—is added, and CT is probably
transformed cometabolically. Endproducts of CT transfor-
mation include carbon dioxide, chloroform (CF), methy-
lene chloride, and unidentified nonvolatile products. This
work was designed to carry out similar transformations in
situ, under conditions representative of contaminated
aquifers.
Prior to the field evaluation, Janssen*7* performed a
microcosm study to simulate field conditions, using
methods described by Siegrist and McCarty*8*. Columns
were packed with test zone aquifer soils and were batch
fed with test zone ground water amended with CT and
growth substrates. Ground water from the field test zone
contained high concentrations of nitrate, but no detectable
oxygen. Columns were operated with different growth
substrates, including acetate, glucose, ethanol, and metha-
nol. Janssen used a column without substrate as a non-
sterile control.
Janssen observed complete nitrate removal after
adding a stoichiometric excess of primary substrates.
Transformation was indicated in columns fed primary
substrates, with lower effluent CT concentrations observed
compared to the nonsterile control. Effluent CT concen-
trations decreased gradually over a 60-day period. The
columns fed ethanol and acetate achieved the highest
transformation, with an 85% reduction in influent concen-
tration compared to 15% in the control. Trace amounts of
CF were observed on the column's effluents, indicating
15
-------
chloroform was the transformation product. These studies
suggested removal of CT would be possible in the field if
acetate were added.
The field study was initiated at a test facility at the
Moffett Naval Air Station, Mountain View, California,
described in detail by Roberts et al.(9). The experiments,
which took the form of a series of stimulus-response tests,
were performed in a shallow sand-gravel aquifer. As a
stimulus, the team injected the subject chemicals—both
substrate and target contaminants—into the test zone as
soluble components of the injected fluid. The response
measured was the concentration as a function of time at
various monitoring locations. Induced gradient conditions
were created by injecting the chemically amended ground
water at 1.5 L/min into a fully penetrating injection well
and extracting ground water at 10 L/min at a fully pene-
trating extraction well located 6 m away. Monitoring
wells SI, S2, and S3 were located in between, at distances
of 1, 2.2, and 3.8 m from the injection well. Chemicals
in or added to the ground water—including acetate as the
growth substrate, nitrate as the electron acceptor, and the
chlorinated aliphatics—were monitored on site using an
automated data acquisition system. The system could
perform a complete analysis every 45 min(9).
The initial experiment assessed CT transport in the
absence of biostimulation (i.e., no growth substrate was
added). CT was continuously injected into the test zone
along with bromide to serve as a conservative tracer. The
transport of CT was retarded by a factor of 2 to 3,
compared to that of the bromide tracer. Approximately
95% of the injected CT broke through at monitoring
locations with continuous injection of 45 ug/L CT,
indicating minimal losses prior to biostimulation. Howev-
er, the formation of approximately 2 ug/L of CF was
observed, indicating some transformation of CT might
have occurred prior to the biostimulation of the test zone.
Biostimulation of the test zone was achieved through
the continuous injection of acetate at an average concen-
tration of 42 mg/L. In order to distribute microbial
growth throughout the test zone, the team added acetate at
321 mg/L for 1 hr of a 12-hr pulse cycle. Nitrate, a
natural contaminant in the ground water, was recycled
continuously at an average concentration of 22 mg/L.
Very rapid stimulation of denitrifying microorganisms was
observed, with complete nitrate consumption occurring
within 2 m of transport, after 70 hr of acetate addition.
Over 90% of the acetate was consumed within the first
meter of transport.
Evidence of CT transformation was observed 300 hr
after acetate injection began. CT concentration gradually
decreased and CF concentration as a transformation
product simultaneously increased over the 1,260-hr period
of acetate injection. In addition to CT, decreases in
concentration of the background contaminants Freon-11
and Freon-113 were observed, following trends similar to
CT.
The percentage transformation of the chlorinated
aliphatics, along with the percentage CF formed from CT
transformation, is presented in Table 1. The extents of
transformation increased with distance traveled, with most
of the transformation occurring after 1 m of transport.
The increase in chloroform concentration mirrored the
trend in CT decrease. Chloroform represented approxi-
mately 55% to 65% of the transformed CT. The rates of
transformation were compound specific, with CT > Freon-
11 > Freon-113 > 1,1,1-TCA. Transformation occurred at
the highest rates in zones that lacked the high concentra-
tions of denitrifying biomass. Some acetate consumption
occurred in the more distant region even in the absence of
nitrate, indicating biological activity in these zones also
occurred.
The team performed a transient experiment to deter-
mine if the zone within 1 m of the injection well could be
stimulated for more rapid CT transformation. This was
accomplished by removing nitrate biologically from the
recycled ground water with a small biological reactor that
was fed acetate. Acetate was then added to the injected
ground water at 20 mg/L. After 20 hr of operation, the
following extents of CT transformation achieved were:
SI, 74%; S2, 95%; S3, 96%; and extraction well, 99%.
Chloroform represented 27% to 36% of the CT trans-
formed—a significant reduction compared to that observed
before the transient. Similar responses were observed for
the other background contaminants.
This field evaluation has demonstrated that in situ
treatment of CT is possible under anoxic conditions. The
denitrifying population itself may not be the main popula-
tion responsible for the CT transformation. This may be
the result of secondary organisms feeding on products of
primary organism growth, as suggested by the pure culture
studies of Galli and McCarty(6). In this study, CF was
formed as an undesirable transformation product. Future
research should therefore focus on reducing CF formation.
Isolation of the microbial population(s) responsible for the
transformation and basic microbial studies with the isolates
would help in optimizing this in situ process.
16
-------
GROUND-WATER TREATMENT
The findings of this study are available as the EPA
report In Situ Biotransformation of Carbon Tetrachloride
Under Anoxic Conditions, Semprini, Hopkins, Janssen,
Lang, Roberts, and McCarty (contact Wayne Downs,
Project Officer, R.S. Kerr Laboratory, Ada, OK).
References
1.
2.
3.
4.
Bouwer, E.J., and P.L. McCarty.
Environ. Microbiol. 1295:1299.
1983. Appl.
Bouwer, E.J., and J.P. Wright. 1988. J. Contaminant
Hydro. 155-169.
Rittmann, B.E., AJ. Valocchi, J.E. Odencrantz, and
W. Bae. 1988. In situ bioreclamation of contaminat-
ed ground water. Illinois Hazardous Waste and
Information Center. H88-026. 121 pp.
Egli, C., T. Tsuchan, R. Scholtz, A.M. Cook, and T.
Liesinger. 1988. Appl. Environ. Microbiol. 2819-
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5. Egli, C., R. Scholtz, A.M. Cook, and T. Liesinger.
1987. FEMS Microbiol. Lett. 257-261.
6. Galli, R., and Pi. McCarty. 1989. Appl. Environ.
Microbiol. 837-844.
7. Dick B. Janssen (personal communication).
8. Siegrist, H., and P.L. McCarty. 1987. J. Contami-
nant Hydro. 31-50.
9. Roberts, P.V., L. Semprini, G.D. Hopkins, D. Grbic-
Galic, P.L. McCarty, and M. Reinhard. 1989. In situ
aquifer restoration of chlorinated aliphatics by meth-
anotrophic bacteria. EPA/600/2-89/033. 214pp.
Table 1.
Percent transformation of the chlorinated aliphatic compounds.
Compound
CT
CF (formed)
Freon-11
Freon-113
1,1,1-TCA
SI
%*
31
20
15
2
0
S2
%
81
43
55
16
0
S3
%
85
52
61
15
4
Extraction
%
97
ND
ND
ND
ND
*Determined from mean concentration estimates over the time period 1160 to 1260 hr.
17
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SOIL/SEDIMENT TREATMENT
SECTION III
SOIL/SEDIMENT TREATMENT
Cleaning up soils and sediments is one of the most difficult problems found at hazardous waste sites. No
current technology is adequate to handle soil cleanup, usually because of the cost involved or the unsuitability
of the technology to the site environment. Soils at industrial sites are contaminated with a complex mixture of
pollutants. Biologically cleaning up these soils in situ is much more effective and inexpensive than excavating
the soils—in itself a major task—before beginning treatment. Three classes of industrial chemicals are currently
the focus of EPA biosystems research on soils and sediments: pentachlorophenol (PCP), pofycyclic aromatic
hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs).
The first of these, PCP, is a common soil contaminant at wood-preserving facilities. EPA researchers have
shown that the white-rot fungus Phanerochaete chrysosporium can reduce the level of PCP in the soil at these
sites. Afield study was set up at a former "tank farm" where for 12 years aboveground storage tanks were filled
with wood preservatives and dieseljuel, which then leaked into the surrounding soil. Two different strains of the
fungus both noticeably depleted the levels of PCP at the field site. Another EPA project showed that, once
organisms known to degrade PCP and other chlorinated compounds were added to freshwater sediments from
areas as diverse as Georgia ponds and the East River in New York City, these compounds were degraded more
rapidly than in reference sediments.
High-molecular-weight PAHs, some of which are known carcinogens, are associated with wood-preserving
plants, coal gasification sites, and petroleum refineries. The estimated 700 wood-preserving plants in the country,
for example, use more than 495,000 tons of creosote per year, some of which leaks from holding tanks and
treatment areas into the soil. Creosote is a complex mixture of over 200 individual compounds. Recent research
on Pseudomonas paucimobilis shows that this bacteria can metabolize a wide array of these chemicals.
Another class of toxic compounds, PCBs, was used heavily for about 50 years in a number of industrial
applications. In the 1960s and 1970s, researchers discovered that PCBs, which resist biological degradation,
were accumulating in the fatty tissues of animals and fish. One EPA project is assessing the rate at which two
different bacterial strains metabolize PCBs and identifying enzymes that initiate degradation.
ORD's research program is exploring three types of solid phase reactors: containerized soil reactors, in situ
treatment, and on-site reactors. Effective reactor design for this type of treatment must allow sufficient contact
between the biomass and pollutants to permit biodegradation in a useful timeframe. Free-standing bioreactors
will be some of the first developed, followed by in situ reactors.
19
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SOIL/SEDIMENT TREATMENT
USE OF WHITE-ROT FUNGI TO
REMEDIATE SOILS CONTAMINATED
WITH WOOD-PRESERVING WASTE
R.T. Lamar and T.K. Kirk, Institute for Microbial and
Biochemical Technology, Forest Products Laboratory,
Madison, WI; JA. Closer, US. EPA, Cincinnati, OH.
The white-rot fungus Phanerochaete chrysosporium
can mineralize a variety of structurally diverse xenobiotics
in liquid culture. These include low-molecular-weight
chlorinated phenolics from kraft pulp mill beach ef-
fluents(6); Arochlor 1254(5); DDT, 3,4,3',4'-tetrachlorobi-
phenyl, 2,4,5,2',4',5'-hexachlorobiphenyl, 2,3,7,8-tetra-
chloro-di-benzo[p]dioxin and lindane(3>4); chlorinated
anilines*1*; pentachlorophenol (PCP)(9); and phenan-
threne(2). The ability of P. chrysosporium to mineralize
such a wide variety of xenobiotics makes it an attractive
organism to test for use in bioremediation of contaminated
media. Information on the ability of white-rot fungi to
degrade xenobiotics to innocuous products in soil does not
exist and is necessary before the organism can be consid-
ered for use in remediating contaminated soils.
Pentachlorophenol is on the EPA's List of Priority
Pollutants and is a common soil contaminant at wood-
preservation facilities where commercial formulations of
technical grade PCP (penta) were used. P. chrysosporium
degrades PCP in liquid culture(9) and in soil<8). Also,
lignin peroxidases and Mn-peroxidases isolated from the
fungus catalyze thep-dechlorination of PCP to tetrachloro-
p-benzoquinone(7). Thus, white-rot fungi may be useful in
remediation of penta-contaminated soils. This paper
describes 1) the results of laboratory-scale studies to
assess the effect of inoculating PCP-contaminated soils
with P. chrysosporium and 2) the status of a field trial in
which the white-rot fungi is being applied in situ to
deplete PCP from contaminated soil.
Laboratory Studies
Inoculation of three soils with P. chrysosporium
resulted in a rapid depletion of extractable PCP (see
Figure 1). The greatest loss of PCP occurred during the
first week. After 2 months, the total amounts of PCP
removed from the three soils (ca. 98%) were similar.
However, initial rates of PCP depletion were dependent on
soil type. The controlled cultures of all three soils also
showed some loss of PCP.
The percentage of PCP mineralized or evolved as
volatile products was small in all three soils (Table 1).
However, mineralization and volatilization from inoculated
cultures was greater than from control cultures, a result
indicating some loss of PCP via these routes due to the
activity of the fungus.
Inoculation of a PCP-contaminated soil with either
Phanerochaete sordida or P. chrysosporium resulted in a
rapid decrease in the level of extractable PCP (see Figure
2). The rate and extent of PCP depletion was slightly
greater in soils inoculated with P. chrysosporium. After
64 d, the level of extractable PCP was reduced by 96% in
soil inoculated with P. chrysosporium and 82.21% in soil
inoculated with P. sordida.
Depletion of PCP by these fungi appeared to occur in
two stages. First, the concentration of extractable PCP in
the soil decreased rapidly. In soil inoculated with P.
chrysosporium, this stage lasted 9 days, during which time
the level of extractable PCP was decreased by 86%. In
soil inoculated with P. sordida, this stage was approxi-
mately twice as long, lasting 21 days, with a 79% de-
crease in extractable PCP. The rapid depletion of PCP in
both cultures coincided with an accumulation of penta-
chloroanisole (PCA) (Figure 2). At the end of the first
stage, approximately 64% and 71% of the PCP was
converted to PCA by P. chrysosporium and P. sordida,
respectively.
The second stage in both fungal cultures was charac-
terized by a slow but steady decline in the levels of
extractable PCP and PCA. During this stage, levels of
PCP and PCA were reduced by 9.6% and 18% in P.
chrysosporium-inocalaled soil and 3% and 23% in soil
inoculated with P. sordida. The level of extractable PCP
in control cultures declined slightly but then returned to a
level that was not significantly different from the original.
No PCA was recovered in extracts from control culture
soils. The data presented here show that the level of PCP
contamination in soil can be greatly reduced by inocula-
tion with white-rot fungi. The initial rate of PCP disap-
pearance is affected by soil type and fungal isolate.
Field Study: In Situ Treatment of a PCP-Contaminat-
ed Soil Using White-Rot Fungi
The team located a site suitable for a field study at
the 8.2-acre Bradley Street Property, 2950 Bradley Street,
Oshkosh, Wisconsin. The site is surrounded. by an
industrial park on the north and west sides, a cemetery on
20
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SOIL/SEDIMENT TREATMENT
the south side, and the Chicago and Northwestern Railroad
on the east.
The area of interest was a former "tank farm" where
aboveground storage tanks were in service from approxi-
mately 1972 until 1984. The tank area is surrounded by
a concrete berm approximately 1 ft high and 1 ft thick.
The tanks were situated on a gravel bed that overlaid the
contaminated soil. The soil is slightly alkaline (pH 7.9),
gravelly sand. There were three 15,000-gallon vertical
steel tanks on the west end of the storage area. These
tanks were reportedly used to store a wood preservative
known as "Woodlife"—a product composed primarily of
mineral spirits (a mixture of high-boiling-point pentanes
and hexanes), but also containing approximately 5% penta.
The number and sizes of aboveground tanks previously on
the eastern two-thirds of the area are unknown. These
tanks were reported to be used for diesel fuel storage. All
tanks have been removed from the site.
The research team chose an area running the width of
the berm and 21 ft down the length of the berm for an
intensive screening of PCP distribution, which was
accomplished by systematically sampling the study area to
a depth of 30 cm for PCP concentrations. PCP concentra-
tions in the soil varied greatly, from 1 to 4,435 ppm.
After pinpointing the distribution of PCP, the team began
efforts to assess the ability of two white-rot fungi to
deplete PCP from the soil.
References
1. Arjmand, M., and Sanderman, H., Jr. 1985. J. Agric.
FoodChem. 33:1055-1060.
2. Bumpus, J.A. 1989. Appl. Environ. Microbiol.
55:154-158.
3. Bumpus, J.A., M. Tien, D. Wright, and S.D. Aust.
1985. Science 228:1434-1436.
4. Bumpus, J.A., and S.D. Aust, 1987. Appl. Environ.
Microbiol. 53:2001-2007.
5. Eaton, D.C. 1985. Enzyme Microb. Technol. 7:194-
196.
6. Huynh, V.-B., H.-M. Chang, T.W. Joyce, and T.K.
Kirk. 1985. Tappi J. 68:98-102.
7. Hammel, K.E., and PJ. Tardone. 1988. Biochem.
27:6563-6568.
8. Lamar, R.T., J.A. Glaser, and T.K. Kirk. Soil Biol.
Biochem. In press.
9. Mileski, GJ., J.A. Bumpus, M.A. Jurek, and S.D.
Aust. 1988. Appl. Environ. Microbiol. 54:2885-
2889.
Table 1. Percent of 14C-PCP evolved as 14CO2 (mineralization) or as 14C-organic volatiles (volatilization)
from soil microcosms of Marsham, Batavia, and Zurich soils that were inoculated with
P'hanerochaete chrysosporium (inoculated) or left non-inoculated (control).
Soil
Marsham
Batavia
Zurich
Mineralization
Inoculated
2.26 (0.66)*
2.02 (0.16)
2.28 (0.53)
Control
1.08 (0.08)
0.67 (0.15)
0.66 (0.13)
Volatilization
Inoculated
0.77 (0.34)
1.80 (0.20)
0.90 (0.18)
Control
0.20 (0.06)
0.43 (0.19)
0.32 (0.22)
*Figures in parenthesis are standard deviations of 3 determinadons for inoculated cultures and 2 determinations
for control cultures.
21
-------
SOIL/SEDIMENT TREATMENT
Table 1. Percent of 14C-PCP evolved as 14CO2 (mineralization) or as 14C-organic volatiles (volatilization)
from soil microcosms of Marsham, Batavia, and Zurich soils that were inoculated with
Phanerochaete chrysosporium (inoculated) or left non-inoculated (control).
Soil
Marsham
Batavia
Zurich
Mineralization
Inoculated
2.26 (0.66)*
2.02 (0.16)
2.28 (0.53)
Control
1.08(0.08)
0.67 (0.15)
0.66 (0.13)
Volatilization
Inoculated
0.77 (0.34)
1.80 (0.20)
0.90 (0.18)
Control
0.20 (0.06)
0.43 (0.19)
0.32 (0.22)
*Figures in parenthesis are standard deviations of 3 determinations for inoculated cultures and 2 determinations
for control cultures.
60
so
40
30
20
10
Marshan
Figure 1.
14 28 42 56
Day
Batavia
, 50
40
CL. 30
U
io
•
Zurich
14 28 42 56
Day
Effect of inoculation with Phanerochaete chrysosporium on the concentration of PCP in three soils.
Vertical bars represent the standard deviations of 4 determinations for inoculated cultures and 2
determinations for control cultures.
14 28 42 56 70
Time (days)
14 28 42
Time (days)
56
Figure 2. Effect of inoculation with Phanerochaete chrysoporiwn or Phanerochaete sordida on the
concentration for pentachlorophenol (PCP) and pentachloroanisole (PCA) in a PCP-amended soil.
22
-------
SOIL/SEDIMENT TREATMENT
AEROBIC BIODEGRADATION OF PAHs
James G. Mueller, Peter J. Chapman, and Parmely H.
Prltchard, US. EPA, Environmental Research Labora-
tory, Gulf Breeze, FL.
Summary
Biological degradation represents the major route
through which polycyclic aromatic hydrocarbons (PAHs)
and other organic chemicals are removed from contaminat-
ed environments*2-3'4*. In most cases, lower molecular
weight PAHs containing two or three rings are readily
degraded biologically^. Conversely, higher-molecular-
weight PAHs are considered resistant to biological action
and tend to persist in contaminated environments^.
Unfortunately, higher molecular weight PAHs represent
the greatest risk to public and environmental health.
For bioremediation to be considered as an acceptable
remedial action alternative for PAH-contaminated sites
(i.e., creosote waste sites, coal gasification sites, petroleum
refineries), biotreatment processes must be capable of
destroying these chemicals. The isolation of microbial
systems capable of utilizing high-molecular-weight PAHs
(e.g., fluoranthene) as sole sources of carbon and energy
for growth partially addresses this challenge.
Results and Discussion
Earlier reports from this laboratory have described the
isolation and characterization of fluoranthene-utilizing
bacteria(6>7). Initially, a stable, seven-member bacterial
community was isolated from a creosote-contaminated
Superfund site at Pensacola, Florida, on the basis of its
ability to degrade PAH components of creosote(7).
Repeated transfer in a fluoranthene-containing medium
successfully reduced the complexity of this community
from seven members to three distinct colony types.
Ultimately, a single member of this community was
isolated in pure culture with an ability to utilize fluoran-
thene as sole source of carbon and energy for growth.
This organism was subsequently identified as a strain of
Pseudomanas paucimobilis, designated EPA505(6).
The ability of strain EPA505 to utilize fluoranthene as
a sole source of carbon and energy for growth was
demonstrated by increased bacterial biomass, changes in
medium coloration, decrease in aqueous fluoranthene
concentration, and the production of transient degradation
products in liquid cultures containing fluoranthene as sole
source of carbon. During growth with fluoranthene, strain
EPA505 produced several metabolites that were extracted
only from acidified culture medium. The properties of
these metabolites were used to project possible pathways
to explain the catabolism of fluoranthene, permitting
growth of the degrading bacterium.
Based on current understanding of bacterial degrada-
tion of PAHs, three possible pathways appear feasible for
the initial oxidation, ring cleavage, and assimilation of
low-molecular-weight metabolites in fluoranthene degrada-
tion (Figure 1). Given the symmetrical nature of the
fluoranthene molecule, initial oxygenation of aromatic
rings may occur at one of three distinct sites (Route I, II,
III). Attack at the 7,8 positions of fluoranthene (Route I)
represents typical biphenyl-23-dioxygenation on a single
aromatic ring*8*. Similar action on the fused ring system
of fluoranthene would result in initial attack at the 12
positions (Route II). Production of 23-dihydroxy-fluoran-
thene would suggest enzymatic activity similar to that of
naphthalene-l,2-dioxygenase(3), which could also be
considered as biphenyl-3,4-dioxygenation. Although 8,9-
dioxygenation is theoretically possible, it is not favored
because of a lack of precedent for this type of attack.
Current knowledge of PAH degradation suggests that
fluoranthene degradation involves dioxygenase activity,
although monooxygenation could also result in the forma-
tion of rra/u-diols of fluoranthene. While previous reports
on PAH degradation describe /rata-cleavage path-
ways(2>3>4), ortfo-cleavage of a dihydroxyfluoranthene is
also possible. Based on precedent, routes for fluoranthene
degradation can be postulated to involve dioxygenation
followed by mefa-cleavage reactions. Continued efforts to
characterize fluoranthene metabolites produced by strain
EPA505 may be useful in the biochemistry facilitating
primary utilization of high-molecular-weight PAHs and
similar chemicals. With a unique ability to utilize high-
molecular-weight PAHs containing four or more fused
rings as primary growth substrates, strain EPA505 and
related organisms provide a possible means of accelerating
the rate of removal of persistent chemicals from contami-
nated environments.
Acknowledgments
We are indebted to Tom Krick (University of Minne-
sota) for the analysis of TMS (trimethylsilyl) derivatives
of fluoranthene metabolites by GC-MS (facilities provided
and maintained by the Minnesota Agricultural Experiment
Station). This work was performed under a cooperative
research and development agreement between the Gulf
Breeze Environmental Research Laboratory and Southern
23
-------
SOIL/SEDIMENT TREATMENT
BioProducts (SBP), Inc. (Atlanta, GA) as defined under
the Federal Technology Transfer Act, 1986 (Contract no.
FTTA-003).
References
1. Bossert, I., and Bartha, R. 1986. Bull. Environ.
Contain. Toxicol. 37:490-495.
2. Cemiglia, CJE. 1984. Adv. Appl. Microbiol. 30:31-
71.
3. Cemiglia, C.E., and Heitkamp, M.A. 1989. In U.
Varanasi (ed.), Metabolism of PAH in the Aquatic
Environment. CRC Press, Boca Raton, FL. pp. 41-68.
4. Chapman, PJ. 1972. In Degradation of Synthetic
Organic Molecules in the Biosphere. Nat. Acad. Sci.,
Washington, D.C. pp. 17-55.
5. McGinnis, G.D., et al. 1988. EPA/600/2-88/055.
6. Mueller, J.G., Chapman, PJ., Blattmann, B.O., and
Pritchard, P.H. 1990. Appl. Environ. Microbiol.
56(4): 1079-1086.
7. Mueller, J.G., Chapman, PJ., and Pritchard, P.H.
1989. Appl. Environ. Microbiol. 55:3085-3090.
8. Smith, MR. and C. Ratledge. 1989. Appl. Microbiol.
Biotechnol. 30:395-401.
CH3
c=o
OH '
COOH
Figure 1. Possible pathways for initial attack on fluoranthene
by Pseudomonas paucimobilis strain EPA505.
24
-------
SOIL/SEDIMENT TREATMENT
ANAEROBIC DEGRADATION OF PCP
AND OTHER CHLORINATED
COMPOUNDS
John Rogers, US. EPA, Athens, GA; Dorothy Hale
and Frank O'Bryant, TAI, Athens, GA; Mahmoud A.
Mousa, UGA, Athens, GA; P. Hap Pritchard, Environ-
mental Research Laboratory, US. EPA, Gulf Breeze,
FL; Barbara Genthner, Technical Resources, Inc., Gulf
Breeze, FL.
The research effort described here is evaluating the
ability of cultures that were adapted to anaerobically
dechlorinate di- and monochlorophenols to enhance the
biodegradation of pentachlorophenol (PCP); 2,4-dichloro-
phenoxyacetic acid (2,4-D); 2,4,5-trichlorophenoxyacetic
acid (2,4,5-T); 2,3,4,5- and 2,3,5,6-tetrachloroanisoles; and
penta- and hexachlorobenzene. In addition, the natural
rates of degradation of these compounds in freshwater
sediments are being evaluated, using sediments from
locations as diverse as ponds in northeast Georgia, the
Wacissa River in north Florida, the East River in New
York City, a pond in the southern Soviet Union, and an
alluvial creek in central Canada. Biodegradation of each
compound occurred under anaerobic conditions and
dechlorinated intermediate products were observed.
Preadapting sediment to dechlorinate dichlorophenols
greatly enhanced ihe degradation of PCP; 2,4-D; and
2,4,5-T. The chloroanisoles and chlorobenzene are
currently being tested in dichlorophenol-adapted sedi-
ments.
Georgia pond sediments adapted to dechlorinate 2,4-
dichlorophenol (DCP) or 3,4-DCP were found to degrade
PCP at faster rates and with shorter lag periods than
nonadapted sediments. A mixture (1:1) of these adapted
sediments dechlorinated PCP, generating 2,3,5,6-tetrachlor-
ophenol; 2,3,5-trichlorophenol; 3,5-dichlorophenol; 3-
chlorophenol; and phenol. The pattern of chlorine remov-
al was para>ortho>meta, a pattern consistent with the
hydroxyl group of phenols being a strong ortho/para
director for electrophilic substitution reactions.
Interestingly, the two dichlorophenol-adapted cultures
showed variable results with 2,4-D and 2,4,5-T. Georgia
pond sediment adapted to 2,4-DCP was able to dechlori-
nate 2,4-D. However, sediment adapted to 3,4-DCP did
not dechlorinate 2,4-D over several months of exposure.
A mixture of these adapted sediments paralleled the
removal of 2,4-D observed with 2,4-DCP adapted sedi-
ment. The only apparent product from adapted or non-
adapled Georgia pond sediment was 4-chlorophenoxyacetic
acid. This product accumulated stoichiometrically as 2,4-
D was removed and the 4-chlorphenoxyacetic acid was not
degraded further. By contrast, 2,4-D dechlorination under
anaerobic conditions in sediment from the Florida river
generated a mixture of products, including 2,4-DCP or 4-
chlorophenoxyacetic acid; 2-chlorophenol (CP); and
phenol. The 4-chlorophenoxyacetic acid did not accumu-
late and apparently was dechlorinated to phenol. Dechlo-
rination of 2,4,5-T in adapted and nonadapted Georgia
pond sediment or Canadian creek sediment produced 2,5-
DCP; 3,4-DCP; or 2,5-D. From any of these intermedi-
ates, 3-chlorophenol and phenol were the only observed
products. Again by contrast, Florida river sediment
biodegraded 2,4,5-T during the initial 2 wk of exposure,
producing 3,4-DCP, 3-CP, and phenol. After 3 wk of
exposure to 2,4,5-T, only 3,4-D; 3-chlorophenoxyacetic
acid; 3-CP; and phenol were produced.
The adapted Georgia pond sediments biodegraded
2,3,4,5-tetrachloroanisol and 2,3,5,6-tetrachloroanisole at
rates commensurate to that of PCP. The 23,5,6-tetra-
chloroanisole was biodegraded faster than 2,3,4,5-tetra-
chloroanisole. Both tetrachloroanisoles produced a 3,5-
dichloroanisole consistent with the O-methyl group being
a ortho/para director similar to the hydroxyl group of PCP.
The fate of chlorinated benzenes has been examined
under anaerobic conditions using sediments collected from
the Georgia ponds. 1,2,4-uichlorobenzene was dechlori-
nated by sediments from two locations, Logan's Pond and
Davidson Mineral Properties, Inc. Pond (DMP), producing
1,2-; 1,3-; and 1,4-dichlorobenzene and monochloroben-
zene. Adapted Logan's sediment to which 1,3-dichloro-
benzene (DCB) had been added showed faster dechlori-
nation of 1,3-DCB than nonadapted sediment. Penta-
chlorobenzene dechlorination by Logan's pond sediment
started after 2 wk, producing 1,2,3,4- and 1,2,4,5-tetra-;
1,2,3-; and 1,2,4-tri- and 1,2-dichlorobenzene. After 4 wk,
all of the pentachlorobenzene was consumed and all
products disappeared except 1,2,3-tri-and 1,2-dichloroben-
zene. The addition of 1,2,4,5-tetrachlorobenzene to
Logan's sediment was followed by a 1-wk lag prior to the
onset of biodegradation. The team identified 1,2,4-tri-;
1,3-; and 1,4-dichlorobenzene as products. 1,2,4,5-tetra-
chlorobenzene was completely removed after 3 wk, as
were the identified intermediate products.
Dechlorination of hexachlorobenzene was observed in
freshwater sediment from Logan's Pond and DMP. After
90 days, the intermediates identified from the DMP pond
were penta; 1,2,3,5-tetra-; 1,3,5-and 1,2,4-tri-; 1,3-di-and
monochlorobenzene. The Logan's Pond sediment pro-
duced penta-; 1,2,3,4- and 1,2,3,5-tetra- and 1,2,3-tri-
25
-------
SOIL/SEDIMENT TREATMENT
chlorobenzene after 110 days of exposure. 1,2-Dichloro-
and monochlorobenzene were identified after 130 days of
exposure in this sediment.
Autoclaved control sediments were inactive in all
sediments.
AEROBIC DEGRADATION OF
POLYCHLORIMATED BIPHENYLS:
BIOCHEMISTRY
Frank J. Mondello and James R. Votes, General
Electric Company. Research and Development Center,
Schenectady, New York; David T. Gibson, Department
of Microbiology, University of Iowa, Iowa City, I A;
Pat R. Sferra, US. EPA, Cincinnati, OH.
An efficient process for the biodegradation of poly-
chlorinated biphenyls (PCBs) requires organisms with high
degradative activity against a wide variety of PCB conge-
ners. In addition, these organisms must survive and
remain active under a wide range of conditions. Recombi-
nant DNA technology is currently being used to develop
bacterial strains with the desired characteristics and to
study the genes involved in PCB degradation. The
organism selected for this work was Pseudomonas sp
LB400, as it is capable of degrading a broad range of PCB
congeners containing up to six chlorines.
PCB degradation by LB400 is an oxidative process
involving two different types of chlorobiphenyl dioxy-
genase attack. At least one of these (biphenyl 2,3-dioxy-
genase) is encoded by the bph genes of the biphenyl
metabolic pathway, and leads to the formation of chloro-
benzoic acids. The second PCB dioxygenase activity
present in LB400 inserts an oxygen molecule at carbon
positions 3,4. This relatively uncommon chlorobiphenyl
3,4-dioxygenase activity may explain the unusual congener
specificity of LB400.
The genes encoding the PCB degradative enzymes
(the bph genes) from LB400 were isolated using the
broad-host-range cosmid vector pMMB34, and found to be
expressed in Escherichia coli. The PCB-degrading ability
of two recombinant strains, FM4110 and FM4560, was
compared to that of LB400. FM4110 is £. coli strain TB1
containing the recombinant plasmid pGEM410 (encoding
bphA, B, C, and D). FM4560 is E. coli strain TB1
containing a derivative of pGEM410. This derivative was
produced by inserting a DN. A fragment encoding bphA, B,
and C, from pGEM410 into cloning vector pUC18. In
resting-cell assays using two PCB mixtures, the ability of
FM4560 to degrade PCBs was similar to that of LB400
and significantly greater than that for FM4110.
The PCB-degrading ability of LB400 is significantly
different from that of most other bacteria. This is reflect-
ed both by its wide range of degradable congeners and its
possession of a 3,4-dioxygenase. DNA-DNA hybridiza-
tion experiments were conducted to determine if the
LB400 bph genes are similar to those in other PCB-
degrading bacteria. Genomic DNAs from nine different
bacterial strains were examined for sequences similar to
the bph genes of LB400. The selected strains varied
widely in PCB degrading ability and included representa-
tives from four different genera and six species. A DNA
probe containing the LB400 bphA, B, C, and D genes
hybridized strongly to genes from Alcaligenes eutrophus
H850. This organism has PCB-degrading abilities that are
very similar to those of LB400. No significant hybridiza-
tion was detected between probe DNA and the genomes
of the other strains tested. These data indicate that there
are at least two distinct varieties of genes for PCB degra-
dation. Southern hybridization analysis was used to
further compare the bph genes of LB400 and H850.
These organisms were found to possess identical restric-
tion maps within their 6p/i-encoding regions (16/16 sites
identical). Since it is extremely improbable that such
similar genes evolved independently in these two organ-
isms, it must be possible for them to be acquired by some
form of DNA transfer.
AEROBIC DEGRADATION OF
POLYCHLORINATED BIPHENYLS:
GENETICS
David T. Gibson, Department of Microbiology, Univer-
sity of Iowa, Iowa City, IA; Pasquale R. Sferra, US.
EPA, Cincinnati, OH.
Introduction
Elegant studies by Bedard, Unterman, and their
colleagues at General Electric Company have led to the
development of assays that can be used to assess the
competence of individual bacterial strains to degrade
polychlorinated biphenyls (PCBs). This work led to the
identification of two organisms, Alcaligenes eutrophus
strain H850 and Pseudomonas sp strain LB400, which are
26
-------
SOIL/SEDIMENT TREATMENT
exceptional in their ability to degrade a wide range of
chlorinated biphenyls.
Several years ago we showed that a Beijerinckia sp
oxidizes biphenyl to c«-2,3-dihydroxy-2,3-dihydrobi-
phenyl, which undergoes an NAD-dependent oxidation to
yield 2,3-dihydroxybiphenyl. Enzymatic ring cleavage of
2,3-dihydroxybiphenyl results in the formation of 2-
hydroxy-6-oxo-6-phenyl-2,4-hexadieneoate. The latter
compound undergoes hydrolytic cleavage to form benzoate
and 2-hydroxy-2,4-pentadienoate. Analogous reactions
appear to be involved in the degradation of several
chlorinated biphenyls.
Congener depletion assays have shown that biphenyl-
grown cells of A. eutrophus H850 and Pseudomonas sp
LB400 degrade 2,5,2',5'-tetrachlorobiphenyl. This
substrate does not have free 2,3-positions for oxygenation,
suggesting that these organisms may have an enzyme that
catalyzes oxidation of the biphenyl nucleus at the 3,4-
position.
Oxidation of 3,4-Dihydroxybiphenyl
Biphenyl-grown cells of A. eutrophus H850 rapidly
oxidized biphenyl, cis-2,3-dihydroxy-2,3-dihydrobiphenyl,
2,3-dihydroxybiphenyl, and benzoic acid. 3,4-Dihydroxy-
biphenyl was oxidized at 20% of the rate observed with
2,3-dihydroxybiphenyl. Similar results were observed with
Pseudomonas sp LB400. Attempts to demonstrate
enzymatic ring fission of 3,4-dihydroxybiphenyl by cell
extracts were not successful.
HPLC analysis of the products formed from 3,4-
dihydroxybiphenyl by A. eutrophus H850 showed the
presence of a major compound that gave an identical mass
spectrum to that given by the methyl ester of protocate-
chuic acid. Minor compounds identified by GC/MS gave
molecular ions at m/e 220 and m/e 202. In this experi-
ment, the 3,4-dihydroxybiphenyl was dissolved in metha-
nol prior to addition to the reaction mixture. When the
experiment was repeated with 3,4-dihydroxybiphenyl
dissolved in deuterated methanol, the major product
detected gave a molecular ion at m/e 171 with a base peak
at m/e 137. These results show that the methyl group in
3,4-dihydroxymethylbenzoate is derived from exogenous
methanol. The addition of 3,4-dihydroxybiphenyl appears
to involve hydroxylation of the unsubstituted ring, ring
fission, and hydrolytic cleavage to yield protocatechuic
acid as the major product Similar results were obtained
when biphenyl-induced cells of Pseudomonas sp LB400
were incubated with 3,4-dihydroxybiphenyl.
Oxidation of 2,5,2',5'-Tetrachlorobiphenyl
Washed cell suspensions of biphenyl-grown A.
eutrophus H850 and Pseudomonas sp LB400 both oxi-
dized 2,5,2' ,5*-tetrachlorobiphenyl to a polar compound
that was isolated by HPLC. The physical properties of the
metabolite (mass spectrum, nuclear magnetic resonance
spectrum, and absorption spectrum) suggested that its
structure was 3,4-dihydro-4,4-dihydroxy-2,5,2',5'-tetra-
chlorobiphcnyl. This was confirmed by acid-catalyzed
dehydration to yield phenolic products. A ctr-relative
stereochemistry of the hydroxyl groups was indicated by
reaction of the metabolite with 2,2-dimethoxypropane to
form an isopropylidene derivative. The dihydrodiol was
further oxidized by both organisms to a more polar
metabolite, which was isolated and identified as cLs-
3,4,3*,4'-tetrahydroxy-3,4,3',4'-tetrahydro-2,5 ,2' ,5'-
tetrachlorobiphenyl. This compound was extremely stable
and acid-catalyzed dehydration to yield phenolic products
required heating to 90°C in 6N
Cell extracts prepared from biphenyl-grown cells of
H850 or LB400 catalyzed the NAD+-dependent oxidation
of 2,3-dihydroxy-2,3-dihydrobiphenyl to 2,3-dihydroxybi-
phenyl. The latter compound was further oxidized to 2-
hydroxy-6-oxo-6-phcnyl-2,4-hexadienoate by the 2,3-
dihydroxybiphenyl oxygenase present in both cell extracts.
No activity was observed when either of the two dihydro-
diols formed from 2,5,2',5'-tetrachlorobiphenyl were used
as substrates.
Oxidation of 2,2'-Dichlorobiphenyl
Biphenyl-grown cells of Pseudomonas sp LB400
rapidly oxidized 2,2'-dichlorobiphenyl. The substrate was
dissolved in methanol. Products identified by GC/MS
were 5 ,6-di hydroxy-5 ,6-dihydro-2,2' -dichlorobiphenyl
(tentative) and the methyl ester of 2-chlorobenzoic acid.
When 2,2'-dichlorobiphenyl was dissolved in dimethyl-
formamide, 2-chlorobenzoic benzoic acid was the major
product formed.
Oxidation of 2,5,2'-Trichlorobiphenyl
Biphenyl-grown cells of Pseudomonas sp LB400
rapidly oxidized 2,5,2'-trichlorobiphenyl. When the
substrate was dissolved in methanol, the methyl ester of
2,5-dichlorobenzoic acid was identified as the major
product. A minor product detected by GC/MS was
tentatively identified as 5,6-dihydroxy-5,6-dihydro-2,5,2'-
trichlorobiphenyl.
27
-------
SOIL/SEDIMENT TREATMENT
Oxidation of Other Chlorinated Biphenyls by Pseudo-
monas sp LB400
Biphenyl-grown cells of Pseudomonas sp LB400
oxidized 4,4'-dichlorobiphenyl, 2,4,4'-trichlorobiphenyl
and 2,43*,4*-letrachlorobiphenyI very slowly. Over a 48-
hr incubation period, 36% of 4,4'-dichlorobiphenyl, 64%
of 2,4,4'-trichlorobiphenyl, and 48% of 2,4,3',4'-tetra-
chlorobiphenyl were removed from the culture medium.
Control experiments with heat-killed cells showed recover-
ies of 60%, 100%, and 100% for the 4,4'-dichlorobi-
phenyl, 2,4,4'-trichlorobiphenyl, and 2,43',4'-tetrachloro-
biphenyl, respectively. At the present time, the identity of
the degradation products from these substrates has not
been determined.
In contrast to the results obtained with 4,4'-chlorinat-
ed substrates, the presence of a 2,5-chlorinated ring
seemed to enhance degradation. For example, biphenyl-
grown cells of Pseudomonas sp LB400 showed significant
degradation of 2,5-, 2,5,2',6-, 2,5,3',4'-, 2,5,2',5'- and
2,5,2',4',5'-chlorinated biphenyls over a 5-hr time period.
The identity of the products formed from these substrates,
with the exception of 2,5,2',5'-tetrachlorobiphenyl, is
currently being determined.
Oxidation of Biphenyl and Chlorinated Biphenyls by
Cell Extracts
In order to determine whether Pseudomonas sp LB400
contains one or more enzymes capable of initiating the
degradation of biphenyl and chlorinated biphenyls, the
research team commenced studies on the purification of
biphenyldioxygenase. This enzyme catalyzes the oxidation
of biphenyl to ciy-2,3-dihydroxy-l,2-dihydrobiphenyl. The
oxygenase, like other aromatic hydrocarbon dioxygenases,
is extremely unstable. At the present time, the optimal
conditions for determining enzymatic activity include
suspension of biphenyl-grown cells in 50mM (2-N-morph-
olino) ethane sulfonic acid (MES) buffer, pH 6.0, followed
by disruption of the cells in a French pressure cell. The
clear supernatant solution obtained after centrifugation at
100,000 x g for 1 hr is taken as the source of crude cell
extract. Enzymatic activity is determined polarographical-
ly. These experiments have yielded cell extracts that only
oxidize biphenyl in the presence of NADH or NADPH.
Cell extracts prepared in this manner also oxidized 2,2'-,
2,5,2'-, 2,5,2',5'-, and 2,4,5,2',5'-chlorobiphenyls. The
major product formed from 2,5,2',5'-tetrachlorobiphenyl
was identified as 3,4-dihydroxy-3,4-dihydro-2,5,2',5'-
tetrachlorobiphenyl.
Future studies will focus on the purification of
biphenyl dioxygenase and determinations of the regio- and
stereospecificity of the purified enzyme with different
chlorinated biphenyl substrates.
Acknowledgments
These studies were supported by Cooperative Agree-
ment No. 812727, EPA Program, 66-507 Toxic Substances
Research Grants for the period 2/10/86 - 6/9/89, and
Cooperative Agreement No. CR-816352, EPA Program
66.802. Research, EPA's Biosystems Technology Devel-
opment Program for the period 10/1/89 - 9/30/91.
ANAEROBIC DEGRADATION OF
PHENOLIC PRIORITY POLLUTANTS:
EFFECTS OF DIFFERENT REDUCING
CONDITIONS
M. HaggUom, MD. Rivera, and L.Y. Young. NYU
Medical Center, Department of Microbiology and
Environmental Medicine. New York. NY; John Rogers,
US. EPA, Athens. GA.
The anaerobic degradation of p-cresol was studied
under three reducing conditions, denitrifying, sulfidogenic,
and methanogenic. Loss of p-cresol (1 mM) in all the
anaerobic systems took initially 3 to 4 wk, but after
acclimation p-cresol was degraded in less than a week.
The sediment has the capacity to utilize the p-cresol, but
an electron acceptor must be available. p-Cresol was
completely metabolized under denitrifying, sulfidogenic,
and methanogenic conditions, with formation of nitrogen
gas, loss of sulfate, and formation of methane and carbon
dioxide, respectively. p-Cresol metabolism proceeded
through p-hydroxybenzaldehyde and p-hydroxybenzoate
under denitrifying, sulfidogenic, and methanogenic
conditions, and these compounds were rapidly degraded
when fed to cultures acclimated to p-cresol. These results
indicate that the pathway of p-cresol degradation is the
same under denitrifying, sulfidogenic, and methanogenic
conditions and proceeds via oxidation of the methyl
substituent top-hydroxbenzoate. Benzoate was additional-
ly detected as a metabolite following p-hydroxybenzoate
in the methanogenic cultures, but not in the denitrifying or
sulfidogenic cultures. This suggests that the degradation
pathway may diverge after p-hydroxbenzoate formation,
depending on which electron acceptor is available.
28
-------
SOIL/SEDIMENT TREATMENT
With the same sediment source the research team
studied degradation of mono- and dichlorophenols under
methanogenic conditions. Degradation of 2,4-dichloro-
phenol and 4-chlorophenol was significantly enhanced
with p-cresol or propionate fed as auxiliary substrate. 2,4-
Dichlorophenol was rapidly dechlorinated to 4-chlorophe-
nol, which was more slowly degraded. Without an
auxiliary substrate, 4-chlorophenol persisted for over a
year. Acclimation for 2,4-dichlorophenol and 4-chlorophe-
nol degradation, with refeeding of the substrate after
depletion of the previous dose, resulted in more than a 10-
fold increase in the rate of degradation. Since the ortho-
position was most readily dechlorinated, enrichments with
2,6-dichlorophenol were set up. 2,6-Dichlorophenol was
rapidly dechlorinated (0.18 mmol/day) with sequential
formation of 2-chlorophenol and phenol. The 2,6-di-
chlorophenol cultures could be subcultured with either p-
cresol or propionate as an auxiliary substrate.
The team studied degradation of mono- and dichloro-
phenols under sulfidogenic conditions using as inoculum
East River sediment or biomass from a bioreactor that
dechlorinates chlorolignin. The cultures with East River
sediment were set up under saline (2.4%) and fresh water
conditions. Complete loss of 2-, 3- and 4-chlorophenol,
and 2,4-dichlorophenol (0.1 mM) in 130 to 200 days after
a lag period of 60 to 100 days was observed. When the
cultures were refed, the chlorophenols were degraded
without a lag in 10 to 20 days. The relative rates of
degradation in the sulfidogenic cultures was 3- and 4-
chlorophenol > 1-chlorophenol. 2,4-Dichlorophenol was
degraded with transient accumulation of 2-chlorophenol.
When sulfidogenesis was inhibited with molybdate, no
degradation of chlorophenols was observed. Cultures set
up with the bioreactor inoculum dechlorinated 2,4-dichlor-
ophenol to 4-chlorophenol, and 2,6-dichlorophenol sequen-
tially to 2-chlorophenol and phenol.
SURFACTANT/SORPTION EFFECTS
UPON BIODEGRADATION
ChadJafvert and John Rogers, U.S. EPA, Athens, GA;
Patricia Van Hoof, University of Georgia, Athens, GA.
Recently, surfactants have received considerable
attention because of their potential to enhance desorption
of pollutants from contaminated soils and sediments as a
stage in decontamination treatment(1"3). Successful analo-
gous uses of surfactant mixtures include the treatment of
open ocean oil spills and the enhanced recovery of oil
using surfactant flooding.
To evaluate the effectiveness of surfactants in remov-
ing contaminants from soils or sediments for biological
treatment, the interactions among the components—surfac-
tant monomer and micelle, pollutants, and soil or sedi-
ment—must be known. Recent investigations generally
have examined interactions between pollutants and
surfactants in clean, well-defined media (e.g., distilled
water). These investigations include the solubilizing
effects of various surfactants or surfactant mixtures on
relatively water-insoluble compounds such as DDT, 1,2,3-
trichlorobenzene(4\ various chloromethanes<5>6\ and
various polycyclic aromatic hydrocarbons*6*. These
studies have been designed to quantitatively characterize
the partitioning of pollutants between their truly dissolved
aqueous phase and the surfactant micellar phase. In one
investigation(4>, partition coefficients for various pollutants
also were given for partitioning to nonmicellar surfactant.
To more fully identify the factors that influence the
effectiveness of surfactants as solubilizing agents, the
research team quantitatively investigated the interactions
within various soil/sediment slurries containing an anionic
surfactant (dodecylsulfate) and various PAH compounds
(naphthalene, phenanthrene, and pyrene). Dodecylsulfate
was used because its properties are well known, a pure
sample (99%) can easily be purchased, it does not inter-
fere with the UV detection of other compounds used in
this study, it is relatively nontoxic, and it is an anionic
surfactant and, therefore, should not adsorb appreciably to
sedimentary materials compared to cationic surfactants.
The three important types of interactions are as
follows:
• Interactions of PAH compounds with sedi-
ments and soils. Considerable work has been
published regarding the partitioning of neutral
hydrophobia organic compounds to soils and
sediments. Generally, the magnitude of partition-
ing to the natural organic carbon of a sediment or
soil (K^) can be related to a chemical's octanol-
water partition coefficient (Kow) or its water
solubility. Karickhoff^ derived a semi-empirical
relationship for the partitioning at equilibrium of
PAH compounds to soils and sediments:
log K^ = 0.989 log Kow - 0.346
(1)
For even moderately hydrophobic compounds
(Kow > 4.0), this relationship implies that these
29
-------
SOIL/SEDIMENT TREATMENT
compounds will partition significantly to sedi-
ments or soils that contain significant organic
carbon, thus reducing their availability to micro-
organisms.
Interactions of dodecylsulfate with sedimen-
tary material. Dodecylsulfate monomers can
adsorb to sediments and soils and, also, can
precipitate from solution as the calcium salt. At
surfactant doses approaching the critical micelle
concentration (CMC), the predominant mecha-
nism of surfactant loss from most natural soils
and sediments will be precipitation. Stellner and
Scamehom(8i9) and Kallay et al.(10) have modeled
the "hardness tolerance" of dodecylsulfate in
well-defined aqueous solutions. The research
team has applied their equations and model
parameters for the precipitation of calcium dodec-
ylsulfate, for the binding of counterions on the
micelle surface, and for the calculation of the
CMC. This allows the determination of the
fraction of surfactant in micellar and monomer
form, and an estimation of surfactant does neces-
sary to obtain a given micellar concentration.
Interactions of dodecylsulfate with PAH com-
pounds. Similar to the partitioning of PAH
compounds to sedimentary organic matter, these
compounds can partition from truly dissolved
aqueous solution to surfactant micelles. This
partitioning, also, is dependent upon the hydro-
phobicity of the specific compounds. Almgren et
al.(I1) have determined the micelle-water partition
coefficients of nine PAH compounds. These
coefficients were correlated to the compounds'
K^ values to predict the partitioning of this class
of compounds among sediment/soil organic
matter, the truly dissolved aqueous phase, and the
dodecylsulfate micellar phase. Results will be
presented, and future research directions will be
discussed.
References
1. Roy, W.R., and Griffin, R.A. 1988. Surfactant- and
Chelate-Induced Decontamination of Soil, Report 21;
Environmental Institute for Waste Management
Studies, The University of Alabama.
2. Nash, J.H. 1987. In Field Studies of In Situ Soil
Washing, U.S. Environmental Protection Agency,
Cincinnati, Ohio. EPA/600/2-87/100.
3. Vigon, B.W., and Rubin, A.J. 1989. Journal W.P.-
C.F. 61:1233-1240.
4. Kile, D.E., and Chiou, C.T. 1989. Environ. Sci.
Tech. 23:832-838.
5. Valsaraj, K.T., Gupta, A., Thibodeaux, L.J., and
Harrison, DP. 1988. Water Res. 22:1173-1183.
6. Valsaraj, K.T., and Thibodeaux, LJ. 1989. Water
Res. 23:183-189.
7. Karickhoff, S.W. 1981. Chemosphere 10:833-846.
8. Stellner, K.L., and Scamehom, J.F. 1989. Langmuir
5:70-77.
9. Stellner, K.L., and Scamehorn, J.F. 1989. Langmuir
5:77-84.
10. Kallay, N., Fan, X., and Matijevic, E. 1986. Acta
Chem. Scand. A40:257-260.
11. Almgren, M., Grieser, F., and Thomas, J.K. 1979. J.
Am. Chem. Soc. 101:279-291.
USE OF FLUORINATED ANALOGUES
TO SHOW ANAEROBIC
TRANSFORMATION OF PHENOL TO
BENZOATE VIA para-CARBOXYLATION
Barbara R. Sharak Genthner, Technical Resourc-
es, Inc., Gulf Breeze, FL; Peter J. Chapman, U.S.
EPA, Environmental Research Laboratory, Gulf
Breeze, FL.
Summary
Isomeric fluorine-substituted phenols were used as
analogues of phenol to investigate the transformation of
phenol to benzoate by an anaerobic, phenol-degrading
consortium derived from freshwater sediment. Transfor-
mation of 2-fluorophenol and 3-fluorophenol led to the
accumulation of 3-fluorobenzoate and 2-fluorobenzoate,
respectively, while 4-fluorophenol was not transformed
(Figure 1). These data indicate that the carboxyl group was
introduced para to the phenolic hydroxyl group in the
fluorophenol isomers (Figure 2). Para-carboxylation was
confirmed when the fluorinated analogue of pora-hydroxy-
benzoate was dehydroxylated to 3-fluorobenzoate (Figure 3).
30
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SOIL/SEDIMENT TREATMENT
By analogy, we presume that phenol is transformed to
benzoate by a similar mechanism.
Results and Discussion
Transformation of phenol to benzoate in sewage sludge
was recently reported*1J. While studying anaerobic degra-
dation of monochlorophenols by freshwater and estuarine
sediments, we observed the conversion of phenol to
benzoate after reductive dechlorination and before mineral-
ization to CO2 and CH4(4>5). To investigate this phenome-
non, a phenol-degrading anaerobic consortium was derived
and used as inoculum.
2-Fluorophenol was transformed to a compound
tentatively identified, by retention time, as 3-fluoroben-
zoate. The UV spectrum of the transformation product was
identical to authentic 3-fluorobenzoate. GC/MS analysis of
culture supernatant ether extracts confirmed its identity as
a fluorobenzoate. 3-Fluorophenol was stoichiometrically
transformed to 2-fluorobenzoate. 2-Fluorobenzoate was
not further metabolized by this consortium. The presence
of phenol (500 fiM) enhanced the rate and degree of
transformation. Transformation of phenol to benzoate
occurred in the presence of 2-fluorophenol, but > 250 nM
2-fluorophenol inhibited the rate of phenol transformation.
A limited amount of 3-fluorophenol (2-3%) was
transformed to a single fluorobenzoate product that was
identified by retention time, UV spectrum, and mass
spectral analysis as 2-fluorobenzoate. Transformation was
not observed in the absence of phenol and the 2-fluoroben-
zoate product was not further metabolized. No other
fluorobenzoate products were detected in culture superna-
tant or ether extracts. As a result of limited transformation,
we were unable to determine the stoichiometry of this
transformation.
In contrast to 2- and 3-fluorophenol, 4-fluorophenol
was not transformed at any concentration tested (50 nM -
2 mM), either in the presence or absence of phenol.
The transformation pattern observed with the three
isomers of fluorophenol indicated carboxylation at the
position para to the phenolic hydroxyl group (Figure 1).
The stoichiometric transformation of 2-fluorophenol to 3-
fluorobenzoate indicated that carboxylation occurred at
either the ortho or para position. Detecting 2-fluoroben-
zoate as the only product of 3-fluorophenol transformation
and observing no transformation of 4-fluorophenol elimi-
nated orr/io-carboxylation, leaving para-carboxylation as
the only feasible alternative. This is analogous to the
industrial chemical carboxylation of phenol (Kolbe-Schmitt
reaction) in which salicylate and pora-hydroxybenzoate
(PHB) are formed via ortho- and pora-carboxylation,
respectively. By analogy, our data suggest that phenol is
carboxylated to PHB and subsequently dehydroxylated to
benzoate (Figure 2). Knoll and Winter*2* have concluded
that PHB was not an intermediate in a similar phenol
transformation carried out by a sewage sludge consortium.
To confirm our findings, we examined the transforma-
tion of 3-fluorosalicylate (3FS) and 3-fluoro-4-hydroxyben-
zoate (3F4HB), which are fluorinated analogues of the
intermediates resulting from ortho- andpora-carboxylation,
respectively. 3FS was not transformed. 3F4HB was
converted to 3-fluorobenzoate and 2-fluorophenol indicat-
ing that both dehydroxylation and decarboxylation occurred
(Figure 3). While 3F4HB was present, both 2-fluorophenol
and 3-fluorobenzoate accumulated. After 3F4HB had been
depleted, 2-fluorophenol declined at a slow rate and 3-
fluorobenzoate continued to accumulate but at a similar
slow rate. Decarboxylation of 3F4HB was not entirely
unexpected, as decarboxylations of phenolic acids having
a free para-hydroxy group are commonly observed reac-
tions of anaerobes(3).
Confirmation of para-carboxylation was obtained,
using a consortium which transformed phenol to benzoate
without complete mineralization of benzoate. Both phenol
and benzoate were detected as intermediates of PHB
degradation. In addition, PHB was detected as an interme-
diate of phenol transformation in cultures containing 6-
hydroxynicotinic acid, a structural analogue of PHB, or
high initial concentrations of phenol (>5 mm) or benzoate
(>10mm).
References
1. Knoll, G., and Winter, J. 1987. Appl. Microbiol.
Biotechnol. 25:384-391.
2. Knoll, G., and Winter, J. 1989. Appl. Microbiol.
Biotechnol. 30:318-324.
3. Scheline, R. 1973. Pharm. Rev. 25:451-532.
4. Sharak Genthner, B.R., Price, W.A., and Pritchard,
P.H. 1989. Appl. Environ. Microbiol. 55:1466-1471.
5. Sharak Genthner, BJR., Price, W.A., and Pritchard,
P.H. 1989. Appl. Environ. Microbiol. 55:1472-1476.
31
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SOIL/SEDIMENT TREATMENT
A.
COQH
B.
OH
2-Ftuoroph«nol
OH
3-Fluoroplwnol
COON
x^\»'
(Oj
2-FlMrotMnzo«te
NO PRODUCT
OH
4-ftoxofhtnct
Figure 1. Transformation products from conversion of monofluorophenols
to monfluorobenzoates via pora-carboxylation.
F-) Phenol
+C02
COOH
— «_
F-) 4-Hydroxybenzoat*
COOH
-) B«nzoate
Figure 2. Possible pathway for the transformation of phenol
(fluorophenols) to benzoate (fluorobenzoates).
o
o
M
c
HI
.a
x
x
O
I
a
6 8
TIME (DAYS)
10
Figure 3. Transformation of 3-fluoro-4-hydroxybenzoate to
2-fluorophenol and 3-fluorobenzoate. Insert expanded
time scale.
32
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SOIL/SEDIMENT TREATMENT
TREATMENT OF CERCLA LEACHATES
IN POTWs: INNOVATIVE ANAEROBIC
PRETREATMENT
M.T. SuidanandA.T. Schroeder, Department of Civil
and Environmental Engineering, University of Cincin-
nati. Cincinnati, OH; ML. Taylor andA.G. Safferman,
PEI Associates, Inc., Cincinnati, OH; R.C. Brenner,
US. EPA, Cincinnati, OH.
The objective of this research is to assess the effective-
ness of fixed-film anaerobic biological processes in treating
and decontaminating CERCLA leachates containing
synthetic organic chemicals (SOC) representative of both
the volatile and organic groups. Use of such processes is
expected to reduce problems associated with air stripping
of volatiles, pass-through of semi-volatiles, and incomplete
degradation of chlorinated compounds, which can occur
when CERCLA leachates are discharged with no pretreat-
ment to publicly owned treatment works (POTW).
Leachate Characteristics
Three municipal landfill leachates are being used in
this study. Two of the leachates are shipped in stainless
steel drums on a monthly basis from landfills operated by
the Delaware Sob'd Waste Authority. One of these leach-
ates is representative of a weak and stabilized leachate
(COD levels of approximately 1,000 mg/L) typical of those
that emanate from old landfills, while the second is moder-
ate in strength (COD levels of approximately 2,500 mg/L)
and contains many of the volatile acids present in active
landfill leachates. The third leachate is used in very large
volumes and is obtained locally by tank-truck from the
Rumpke municipal landfill in Georgetown, Ohio. This
leachate is very weak and variable in composition (COD
levels of 200 to 1,600 mg/L) and is supplemented with a
mixture of acetic, propionic, and butyric acids as needed to
increase its total COD to approximately 1,600 mg/L. All
leachates are rendered hazardous by combining them with
a mixture of nine volatile and five semivolatile organic
compounds, shown with their corresponding target concen-
trations in Table 1. Selection of these compounds and their
respective spike concentrations was based on a review of
published data on CERCLA leachates. Chloroform was not
included in the spike mixture due to potential toxicity
problems. The tolerance of anaerobic cultures to chloro-
form is being evaluated in a separate anaerobic filter.
Treatment Systems
The anaerobic pretreatment of CERCLA leachates is
being investigated in fluidized-bed reactors charged with
16x20 U.S. Mesh granular activated carbon (GAC) and in
upflow anaerobic filters packed with 1-in. Pall Rings. All
anaerobic pretreatment systems are operated at 35°C. The
empty-bed contact time in the fluidized-bed reactors is
maintained at 6 to 8 h, while a longer detention time of 48
to 96 hr is used in the anaerobic filters. The anaerobically
pretreated Rumpke leachate is then diluted at a ratio of 15%
pretreated leachate to 85% raw municipal wastewater, and
the mixture is treated in bench- or pilot-scale POTW units
consisting of primary clarification (2-hr detention) followed
by an activated sludge treatment with 6 or 8 hr aeration
time.
The treatment of the Rumpke leachate is evaluated in
three parallel trains. One train consists of leachate pre-
treated in a bench-scale (4-in. diameter x 42-in. high)
fluidized-bed GAC anaerobic reactor followed by mixing
with raw municipal wastewater and treatment in a bench-
scale activated sludge POTW unit. Another train is
identical to the first one, with the exception that the fluid-
ized-bed reactor is replaced by a bench-scale (6-in diameter
x 48-in. high) upflow anaerobic filter packed with 1-in. Pall
Rings. The third treatment train consists of a pilot-scale (2-
ft diameter x 6.4-ft high) anaerobic filter followed by a
41.2-ft3 primary clarifier, a 35.2-ft3 aeration tank, and a
7.0-ft2 surface area final clarification tank. The objective
of this process train is to evaluate the scale-up of anaerobic
filters and to observe potential problem areas such as bed-
plugging and wall effects. Comparison of the performance
characteristics of the two process trains containing the
anaerobic filter will provide guidelines as to the utility of
results obtained from bench-scale studies.
The two Delaware leachates are treated in two parallel
bench-scale (4-in. diameter x 42-in. high) fluidized-bed
GAC anaerobic reactors. Volatile fatty acids in the moder-
ate-strength leachate constitute approximately 80% of the
total COD, and biological activity within the reactor
treating this leachate is primarily methanogenic. The
weaker leachate contains a negligible amount of volatile
acids, and biological activity within the second reactor is
primarily due to sulfate reduction. After acclimation to
their respective unspiked Delaware leachates for several
months, the two anaerobic reactors have been in operation
for five months on leachate supplemented with the SOC
spike. Data collected after spiking of the SOC's to the feed
had reached their target levels (Table 2) suggest that the
fluidized-bed anaerobic GAC reactors are capable of
affecting high degrees of removal of these compounds, with
33
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SOIL/SEDIMENT TREATMENT
the exception of bis (2-ethylhexyl) phthalate. Furthermore,
the volatile fatty acid content of the moderate strength
leachate appears to be almost completely converted to
methane and carbon dioxide.
Due to previous toxicity problems with chloroform, the
impact of this compound on anaerobic systems is being
evaluated in a separate 6-in. diameter anaerobic filter. The
concentration of chloroform in the Rumpke leachate
(supplemented with a mixture of volatile fatty acids to a
COD of approximately 1,600 mg/L) has been gradually
increased to 5,000 ng/L with an ultimate target level of
10,000 p.g/L. To date, effluent chloroform levels have
ranged from 15 to 62 |ig/L. Once the tolerance of the
anaerobic system to chloroform has been established, this
compound will be added to the SOC spike for the remain-
ing anaerobic reactors.
Table 1. SOC supplement to simulate CERCLA leachates.
Volatile Organic Compound
Methylenc chloride
Acetone
Trichloroethylenc
Toluene
Elhylbenzene
1,1-Dichloroelhanc
Chlorobenzene
Methyl ethyl kelone
Methyl isobutyl ketone
Semi- Volatile Organic Compounds
Dibutyl phthalate
Bis (2-ethylhcxyI) phihalatc
Nitrobenzene
1 .2,4-trichlorobenzene
Phenol
Target Concentration, ifg/L
1,200
10,000
400
8.000
600
100
1.100
5,000
1,000
200
100
500
200
2,600
Table 2. SOC effluent concentrations - anaerobic GAC reactors.
Compound
Melhylene chloride
Acetone
Trichloroethylene
Toluene
Ethylbenzene
1,1-DichIoroethane
Chlorobenzene
Methyl ethyl ketone
Methyl isobutyl ketone
Dibutyl phthalate
Bis (2-ethylhexyl) phthalate
Nitrobenzene
1 ,2,4-Trichlorobenzene
Phenol
Effluent Concentration, «g/L
Weak Leachate
72 (41)*
196(222)
9(6)
222(34)
16(1)
35 (13)
26(3)
41 (18)
49(7)
13(6)
72(45)
3(3)
3(3)
42(29)
Moderate Leachate
46 (14)
327 (465)
8(3)
403(164)
30(6)
26(10)
53 (10)
141 (113)
37(8)
10(3)
51 (43)
8(14)
1(2)
49 (23)
'Numbers in parentheses represent standard deviations.
34
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Combined Treatment
SECTION IV
COMBINED TREATMENT
Most hazardous waste sites contain complex mixtures of persistent organic and inorganic contaminants that can
be cleaned up only by a combination of treatment techniques. Researchers are developing methods to combine
various physical, chemical, and biological treatment technologies, and comparing the effectiveness of the various
combinations. For example, a chemical treatment—adding potassium polyethylene glycol (KPEG) to the soil at a
site—may be used to dechlorinate PCBs. Then, biological treatment, which is more effective after dechlorination,
can be used to complete the soil restoration. To handle a full range of compounds, this biological treatment can
be both anaerobic and aerobic, or a combination of the two. In one project, researchers are combining KPEG
treatment with composting to enhance biodegradation in contaminated soils. These initial studies will indicate how
effectively composting enhances the metabolic activity of the organisms at the site.
The KPEG process is chiefly effective against halogenated compounds. Other chemical and physical processes
are also being developed to handle other types of compounds. Combined treatments offer great promise for
remediating many of the contaminated soils polluting our environment.
USE OF KPEG AND ANAEROBIC
BIODEGRADATION FOR
PCB-CONTAMINATED SOILS:
ENRICHMENTS FOR ANAEROBES
CAPABLE OF DEGRADING
KPEG-TREATED PCB
JA. Krzycki. US. EPA, Cincinnati, OH.
Polychlorinated biphenyls were heavily used for about
50 years in a number of industrial applications. In the
1960s and 70s, scientists recognized that PCBs were
recalcitrant to biological degradation and accumulated in
the adipose tissues of organisms high in the food chain.
Though their use has been largely discontinued, an estimat-
ed half million tons of PCBs are buried in landfills and
closed systems awaiting detoxification(1). Cleanup is
accomplished either by removing the contaminated materi-
als to landfills or incinerators or by chemically treating the
PCBs with a formulation of KOH and polyethylene glycol
400 (KPEG). Since this method does not completely
dechlorinate higher substituted PCBs, the feasibility of
combining the KPEG treatment with biological remediation
needs to be explored. Experiments are currently being
initiated to establish enrichment cultures capable of degrad-
ing KPEG-modified PCBs under anaerobic conditions.
KPEG is usually added to contaminated soils at
approximately equal weights and heated from 60°C to
140°C over a period of several hours. A recent field test
demonstrated that soil contaminated with 1900 ppm
Arochlor 1260 could be detoxified by two treatments,
resulting in less than 1 ppm per resolvable congener*2'. The
mechanism of dechlorination by KPEG is nucleophilic
attack at a chlorine-bearing carbon. The products are KCI
and a polyethylene glycol-PCB adduct. With continued
heating, the adduct decomposes to allene glycol and the
corresponding biphenylol(3'4). Higher chlorinated bi-
phenyls are more susceptible to attack, and mono- and
dichlorinated hydroxylated biphenyls remain following
treatment of a highly chlorinated PCB mixture. Treatment
of a single hexachlorinated congener with KPEG is expect-
ed to yield a mixture of hydroxylated, chlorinated bipheny 1
isomers. When the reaction takes place at lower tempera-
tures, PEG-substituted biphenyls will also be present.
KPEG formulations may function as bifunctional reagents,
leading to PCB-PEG complexes containing two or more
biphenyls.
35
-------
Combined Treatment
In order to determine if this mixture of products can be
degraded by enrichment cultures under anaerobic condi-
tions, the research team will react KPEG with a 14C
labelled PCB congener. Several are commercially avail-
able, and both hexachlorinated (such as 2,2',4,4',5,5') and
tetrachlorinated (such as 2,2',4,4') biphenyls will be tested.
Mineralization of KPEG-PCB products in enrichment
cultures will be followed by evolution of 14C carbon
dioxide and methane, measured by a gas chromatograph
coupled to a gas proportional counter. Aliquots of cultures
will be removed and extracted with hexane. Changes in the
hexane/aqueous partitioning of radioactivity will be taken
as presumptive evidence of metabolism. The hexane and
water-soluble fractions of the KPEG/14C-PCB products
will be analyzed using HPLC. Changes in radioisotope
elution profiles during incubation will serve to identify
metabolizable and recalcitrant products of the KPEG
reaction.
Enrichment cultures suitable for use as a second stage
treatment following KPEG treatment of PCB-contaminated
soils would use PEG as the primary carbon and energy
source. PEG will be formed by hydrolysis of the KPEG
reagent and be at a much higher concentration than the
modified PCB target compounds. Thus, degradation of the
modified PCBs is expected to be the result of co-metabo-
lism by members of a PEG-degrading consortia. Recently,
several enrichments of anaerobes have been described that
form products such as acetate, ethanol, hydrogen, carbon
dioxide and propionate from PEG(5>6). Present in the
enrichments were methanogenic, sulfidogenic, and aceto-
genic bacteria. Representatives of each trophic group have
been observed to conduct dehalogenation of chlorinated
aromatics and or aliphatics(1). Hydroxylation of aromatics
under anaerobic conditions is an endergonic process;
pretreatment of PCB with KPEG will yield hydroxylated
aromatics that may be more readily degraded under anaero-
bic conditions.
Inocula for enrichment cultures will be obtained from
anaerobic environments that have a history of PCB expo-
sure, such as contaminated river or lake sediment Anaero-
bic sludge from municipal waste digesters will also be
used. Both PEG depolymerization and aromatic dechlori-
nation have been observed by organisms obtained from or
in sludge(7). Each source will be introduced into enrich-
ment media supplemented with different electron acceptors
such as sulfate, nitrate, and carbon dioxide under either a
hydrogen or nitrogen gas phase. PEG will be present at
0.2% in all enrichments. Autoclaved inocula and unmodi-
fied congeners will serve as controls for each variation in
enrichment media.
Successful enrichments will be stabilized, and individ-
ual organisms responsible for modified PCB degradation
isolated and characterized.
References
1. Reineke, W., and Knackmuss, H.J. 1988. Ann. Rev.
Microbiol. 42:263.
2. Komel, A., Rogers, CJ., and Sparks, H. 1988. Field
experience with the KPEG reagent, EPA/600/9-88/021
pp. 403-413.
3. Brunelle, D.J., and Singleton, D.A. 1985.
Chemosphere 14:173.
4. Komel, A., and Rogers, C. 1985. J. Haz. Mat. 12:161.
5. Wagener, S., and Schink, B. 1988. Appl. Environ.
Microbiol. 54:561.
6. Dwyer, D.F., and Tiedje, J.M. 1983. Appl. Environ.
Microbiol. 46:185.
7. Fathepure, B.Z., Tiedje, J.M. and Boyd, S.A. 1988.
Appl. Environ. Microbiol. 54:327.
36
-------
Combined Treatment
USE OF KPEG AND COMPOSTING FOR
TREATMENT OF CONTAMINATED
SOILS
O. Karl Scheible, Talbert N. Eisenberg, HydroQual,
Inc., Mahwah, NJ; Albert D. Venosa, US. EPA,
Cincinnati, OH.
While the KPEG process has proved effective in
reducing the concentration of halo-organic soil contami-
nants, the questions of restoration and return of an environ-
mentally sound treated soil remain to be answered. PCB-
contaminated soils, following treatment with KPEG, still
contain unreacted KPEG reagent, reacted PEG, residual
PCBs, and hydroxychlorobiphenyls. The residual PCBs are
predominantly the lower chlorinated isomers (mon, di, and
tri), which are more amenable to biodegradation. The
objective of this study is to determine the biodegradability
of the residual PCBs, the KPEG reactants, and the hydro-
xychlorobiphenyls.
Specific objectives of the phase 1 project are to: 1)
determine if the products resulting from KPEG treatment of
PCB-contaminated soils are biodegradable in a composting
environment; 2) determine approximate design parameters
with respect to organic loading, soil loading, temperature,
residual PCB concentrations, etc.; 3) determine require-
ments for and develop the design of a pilot-scale compost-
ing system to handle KPEG-treated soils; and 4) determine
capital and O&M costs for a compost operation under a site
remediation scenario. The project is currently in startup,
focusing on the development of analytical procedures and
the assembly of laboratory compost reactors.
Initial laboratory-scale batch reactor studies are being
performed to determine the biodegradation of the KPEG-
pretreated PCBs in a compost environment. The initial
studies employ the reaction product from a KPEG-treated
mixture of octa-, hepta-, and pentachlorobiphenyl in "neat"
form (i.e., a liquid or nonsoil matrix). Results from the
initial studies are expected to indicate the relative effective-
ness of compost treatment and direct the final laboratory
studies, which will determine the feasibility of composting
a KPEG-pretreated soil matrix. A secondary element of the
phase 1 effort is to develop and demonstrate analytical
procedures to monitor the fate of the reaction products.
COMBINED KPEG AND BIOLOGICAL
TREATMENT OF SOILS
CONTAMINATED WITH CHLORINATED
DIOXINS AND PHENOXYACETATES
Richard Haugland, University of Cincinnati, OH;
Charles Rogers. Al Kernel. Pasquale Sferra, US. EPA,
Cincinnati. OH; ThomasTier nan, Wright State Univer-
sity, Dayton, OH.
Studies have been initiated to evaluate the combination
of biological treatment with the KPEG chemical process for
treating soils contaminated with mixtures of chlorinated
phenoxyacetates and chlorinated dioxins. The rationale for
this combined approach stems partially from projections
that the KPEG process may incur unacceptably high
reagent costs when used to treat chlorinated dioxins in soils
containing high concentrations of phenoxyacetate herbi-
cides. This is due to the tendency of the KPEG reagent to
react with the herbicides as well as dioxins and thus be
consumed in high quantities under such conditions. In
addition, chemical modification of chlorinated phenoxyace-
tates by the KPEG reaction is considered to be a less
desirable alternative than biologically mediated mineraliza-
tion for eliminating these compounds from soils. Largely
as a result of the recent development in our laboratory of a
Pseudomonas cepacia strain, RHJ1, with the capability to
mineralize high-concentration mixtures of 2,4,5-trichloro-
phenoxyacetate (2,4,5-T) and 2,4-dichlorophenoxyacetate
(2,4-D), the potential now exists for biologically eliminat-
ing excessive amounts of these herbicides from soils in a
relatively rapid and inexpensive manner. Combined KPEG
and biological treatment thus has the potential to lower
overall remediation costs of soils contaminated with
chlorinated dioxins and high concentrations of phenoxya-
cetate herbicides and eliminate unwanted byproducts.
Experiments were performed to evaluate 2,4,5-T and
2,4-D degradation by RHJ1 both before and after treatment
of these compounds with KPEG. The latter approach was
based on previous studies with contaminated oil samples
which indicated that at relatively mild reaction tempera-
tures of 100°C or less, KPEG reagent selectively attacks
chlorinated dioxins and shows minimal activity toward
chlorinated phenoxyacetates. Mixtures of 2,4,5-T and 2,4-
D treated at 70°C for 24 hr showed no detectable modifica-
tion by the KPEG reagent and were readily degraded by
RHJ1 following neutralization and dilution in a mineral
salts medium to approximately 2,000 ug/mL (100 ug/mL
ea. of 2,4,5-T and 2,4-D). The same mixture treated at
100°C for 24 hr showed partial modification (18-30%) by
the KPEG reagent Similar incubations of these KPEG-
37
-------
Combined Treatment
treated mixtures with RHJ1 revealed that the KPEG-
modified compounds not only registered degradation
themselves, but also apparently inhibited the degradation of
unmodified 2,4,5-T and 2,4-D.
Soil samples were obtained from a former herbicide
manufacturing site and analyzed for chlorinated dioxin and
phenoacetate herbicide content Each of these samples
showed measurable levels (5.8-9.6 ng/g) of 2,3,7,8-tetra-
chlorodibenzodioxin (TCDD) but contained unexpectedly
low concentrations of 2,4-D (0-260 ug/g) and 2,4,5-T (0-15
ug/g). Treatments of these samples with KPEG at 100°C
resulted in only partial eliminations of TCDD and substan-
tial formations of modified 2,4-D and 2,4,5-T. These
results indicated mat treatment of soil with RHJ1 after
KPEG treatment is not a practical approach since KPEG
modification of the phenoxyacetates appears to be unavoid-
able at temperatures required to obtain complete elimina-
tion of TCDD.
In evaluations of biological degradation of chlorinated
phenoxyacetates prior to KPEG treatment, soil samples
from the herbicide manufacturing site were either directly
treated or amended with 1000 ug/g each of 2,4,5-T and 2,4-
D (to raise the concentration of these compounds) prior to
treatment. Treatments consisted of suspending the soil
samples at a ratio of 1:1 (w/v) with either uninoculated
mineral salts medium or the same medium containing 108
cells per mL of RHJ1, followed by incubation of the
suspensions with shaking at 29°C. Analyses of 2,4,5-T and
2,4-D concentrations in the soil suspensions inoculated
with RHJ1 over a five-day period showed that these
compounds were eliminated irrespective of their starting
concentrations. Similar analyses of an uninoculated soil
suspension initially containing low concentrations of 2,4,5-
T and 2,4-D also showed elimination of both compounds;
however, an uninoculated suspension containing a relative-
ly high initial concentration of these compounds showed
losses of 2,4,-D but not 2,4,5-T.
These results suggest that indigenous microorganisms
with degradative activity toward 2,4-D were present in the
soil samples. The apparent stimulation of these organisms
by the nutrient, water, and/or oxygen amendments associat-
ed with the incubations resulted in significant reductions of
2,4-D, even when this compound was present at the higher
concentration. The results further suggested, however, that
indigenous microorganisms from the soil did not have the
capability to rapidly degrade high concentrations of 2,4,5-
T. Inoculation of soils containing high concentrations of
this compound therefore still appears to be warranted.
While testing with additional field samples is needed,
our studies indicate that biological treatment of soils to
eliminate chlorinated phenoxyacetates is a feasible option
prior to KPEG treatment. This approach offers the advan-
tage that temperatures in the KPEG reaction can be raised
to whatever level is required to allow complete elimination
of the chlorinated dioxins present. Analyses of the effects
of KPEG treatment on TCDD concentrations in biological-
ly pretreated soil samples are currently in progress and the
results will be presented.
38
-------
SEQUENTIAL TREATMENT
SECTION V
SEQUENTIAL TREATMENT
Sequential treatment is generally applied to two types of waste: compounds that degrade into stable
intermediates that can be further degraded under different conditions than those used for the parent compound; and
complex mixtures of wastes, which are generally degraded in order of their thermodynamic behavior. The classic
example of the first type of waste is DDT, which is effectively degraded under alternating anaerobic and aerobic
conditions. For mixed wastes, an understanding is needed of the degradation pathways for the waste components
and for intermediate compounds, the sequence by which complex mixtures are degraded under field conditions, the
physical and chemical factors that can influence the degradation pathways, and the availability of organisms or
adapted bacterial communities that are able to degrade or transform components of the mixture or intermediate
degradation products.
Sequential technologies can vary from the simple coupling of sequential aerobic and anaerobic processes for
the degradation of a single compound to the use of sequential reactors containing bacterial cultures adapted for
degrading specific compounds or compound classes that are components of a hazardous organic mixture.
ANAEROBIC BIODEGRADATION OF
CREOSOTE CONTAMINANTS IN
NATURAL AND SIMULATED
GROUND-WATER ECOSYSTEMS
Edward M. Godsy, Donald F. Goerlitz. US. Geological
Survey, Menlo Park. CA; Dunja Grbic-Galic, Stanford
University, Stanford, CA; John Rogers, US. EPA,
Athens, GA.
Creosote is the most extensively used insecticide and
industrial wood preservative today. It is estimated that
there are 600 wood-preserving plants in the United States,
and their collective use exceeds 4.5 million kilograms per
annum* l\ Creosote is a complex mixture of over 200 major
individual compounds with differing molecular weights,
polarities, and functionalities, along with dispersed solids
and products of polymerization(2). Chemical analysis of
creosote shows a composition of approximately 85% by
weight polynuclear aromatic compounds, 12% phenolic
compounds, and 3% heterocyclic nitrogen, sulfur, and
oxygen- containing compounds.
The research field site is located within the city of
Pensacola, Florida, at an abandoned wood-preserving plant.
The 7.3-hectare site is situated approximately 550 m north
of Pensacola Bay and near the entrance to Bayou Chico.
The preservation of pine poles at the site consisted of
removing as much of the cellular moisture as possible from
the poles and replacing the moisture with either creosote
and/or pentachlorophenol. Effluent from the treatment
process consisted of water, cellular debris, diesel fuel,
creosote, and pentachlorophenol. The chemically complex,
organic-rich mixture was discharged to the shallow, unlined
waste disposal ponds, and large-but-unknown quantities of
the waste infiltrated the soil and moved downward toward
the water table. Upon reaching the water table, the creosote
mixed with the ground water, resulting in two distinctive
phases: a denser-than-water hydrocarbon phase that moved
vertically downward somewhat perpendicular to ground-
water flow, and an organic rich aqueous phase. The latter
phase is enriched in organic acids (35%), phenolic com-
pounds (36%), single- and double-ring aromatic com-
pounds (4%), and single- and double-ring nitrogen, sulfur,
and oxygen-containing aromatic compounds (25%). These
dissolved contaminants are subject to physical, chemical,
and biological processes that will tend to retard their
movement relative to the ground water.
39
-------
SEQUENTIAL TREATMENT
The overall research objectives of this study were: 1)
to establish laboratory microcosms simulating the subsur-
face environment in order to determine which compounds
in the aqueous phase were anaerobically biodegradable, 2)
to determine the nature of the microbial population in the
aquifer sediments, and 3) to compare the results of labora-
tory transformations to field observations of temporal and
spatial changes in subsurface distribution of major com-
pounds during downgradient movement from Site 3 to Site
37 (Figure 1).
The existence of microbes in the subsurface has only
recently come to light. Several reports have clearly
demonstrated the existence of active microbes in these
environments^. Determination of the physiological groups
in the resident population often helps in determining the
ultimate fate of the contaminants during downgradient
movement in the aquifer and was instrumental in determin-
ing the fate of coal tar derivatives at a contaminated water
site at St. Louis Park, Minnesota*4*. Microbiological
analysis of the aquifer material and the corresponding pore
water at the six sites has revealed that a small, but apparent-
ly active, population of bacteria reside in the subsurface
and that at least 90% of this population is attached to the
aquifer material. These analyses revealed that the only
apparent difference between the contaminated sites and the
uncontaminated site was a 10- to 100-fold increase in the
number of methanogenic bacteria throughout the contami-
nated area.
Various pathways have been proposed for the anaero-
bic degradation of aromatic compounds depending on the
nature of the substrate, but they are all similar in the major
steps involved: 1) the removal of the substituent side
groups by reduction, demethylation(5i6), demethoxylation,
dehydroxylation, and decarboxylation'7^ 2) incorporation
of oxygen from water into the ring structure of compounds
that do not already have oxygen incorporated* -9); 3)
hydrogenation of the aromatic structure with the formation
of an alicyclic intermediate; 4) fission of the alicyclic
compounds to straight chain acids; 5) conversion of these
acids to acetate, hydrogen, and carbon dioxide; and 6)
under methanogenic conditions, conversion of these
endproducts to methane.
This study indicates that, along with the phenolic
compounds and organic acids previously reported to be
anaerobically biodegradable(10), the nitrogen-containing
compounds quinoline, isoquinoline, and 2(lH)-quinolinone,
and l(2H)-isoquinolinone are also anaerobically biodegrad-
able to methane and carbon dioxide. A three-step sequen-
tial degradation of these compounds occurred in which the
compounds in group 1 were degraded before phenol and
those in group 3 were degraded after phenol:
1. Quinoline, isoquinoline, benzoic acid, and 3
through 6 carbon volatile fatty acids
2. Phenol
3. 2-, 3-, 4-methylphenol, 2(lH)-quinolinone, and
1 (2H)-isoquinolinone
Acetic acid is the major intermediate compound in the
conversion of the biodegradable compounds in the aqueous
fraction. This is very similar to conventional anaerobic
domestic sewage sludge digesters, where acetic acid is the
major intermediate compound in the conversion of complex
substrates to methane and carbon dioxide. The volatile
fatty acids are rapidly converted to acetate in the digester
along with those degradable compounds in step 1. An
increase in acetate is also observed during each of the other
steps of the sequential degradation.
Inorganic analyses of water samples from the contami-
nated area clearly demonstrate that anoxic conditions
prevail. The contaminated ground water is devoid of
dissolved oxygen, is approximately 60-70% saturated with
respect to methane, and contains a relatively high concen-
tration of hydrogen sulfide. Sufficient nitrogen and
phosphorus are present to allow for microbial degradation
of susceptible organic compounds. The concentration of
phenol, 2-, 3-, 4-methylphenol, quinoline, isoquinoline,
2(lH)-quinolinone, and l(2H)-isoquinolinone decrease
disproportionately downgradient when compared to the
conservative tracer 3,5-dimethylphenol.
The ground-water velocity in the area of the plant has
been determined to range from 0.3 to 1.2 m/day. Flow
velocities at the 6.1 m sampling depth have been deter-
mined to be on the order of 1 m/day. This value allows for
easy comparison of laboratory and field data: Residence
time in the aquifer approximately equals the downgradient
distance traveled. The approximate downgradient distance
of contaminated sites from the source ponds are as follows:
Site 3,6 m; Site 39,53 m; WP 34,99 m; Site4,122 m; and
Site 37, 150 m. Comparison of the composition of the
major compounds in the aqueous fraction of the ground-
water samples to the digester composition after a compara-
ble time of residence facilitates recognition of the disap-
pearance pattern observed in the ground-water samples.
During the first 50 days of residence in the digestors, or
travel from Site 3 to Site 39, the 3 through 6 carbon volatile
fatty acids are rapidly converted to acetic acid and
ultimately to methane and carbon dioxide. Benzoic acid,
quinoline, and isoquinoline are also biodegraded. Phenol
occurs between days 50 and 99 in the digestors, and phenol
40
-------
SEQUENTIAL TREATMENT
EXPLANATION
e • SITE ANIO NUMBER
(5 AL.TITUOC Of WATER TABLE.
COIMTOUR INTERVAL. O.S
METER. DATUM IS SEA UEVEl.
A' OCOI.OO4C SECTtOM UIIMC A-A*
Figure 1. Site location and configuration of the water table in the vicinity of the creosote works site.
41
-------
SEQUENTIAL TREATMENT
also disappears from the ground water during transit from
site 39 to WP 34. After 100 days in the digestor, 2-, 3-, 4-
methylphenol, 2(lH)-quinolinone, and l(2H)-isoquino-
linone are biodegraded and are removed from the system
after about 200 days. A similar pattern of disappearance is
observed during travel from WP 34 to Site 4 for 2-, 3-, and
4-methylphenol; however, the degradation of 1(2H)-
isoquinolinone is delayed somewhat compared to the
digesters.
Integration of both laboratory results and field observa-
tions helps in determining the ultimate fate of the subsur-
face contaminants. Results demonstrate that the dispropor-
tionate decrease of selected organic compounds observed
during downgradient movement in the aquifer can be
attributed to microbial degradation of the compounds. The
anoxic conditions in the contaminated area, high concentra-
tions of dissolved methane, and the increased numbers of
methanogenic bacteria suggest that methanogenic fermenta-
tion was the predominant microbiological process. This
observation was confirmed in the laboratory digesters.
References
1. Von Rumker, R., Lawless, E.W., and Meiners, A.F.
1975. Production, distribution, use and environmental
impact potential of selected pesticides. EPA 540/1 -74-
001,439pp.
2. Novotny, M., Strand, J.W., Smith, S.L., Weisler, D.,
and Schwende, FJ. 1981. Fuel 60:213-220.
3. Harvey, R.W., Smith, R.L., and George, L. 1984.
Appl. Environ. Microbiol. 48:1197-1202.
4. Ehrlich, G.G., Godsy, E.M., Goerlitz, D.F., and Hult,
M.F. 1983. Develop. Indus. Microbiol. 24:235-245.
5. Young, L.Y., and Rivera, M.D. 1985. Wat. Res.
19:1325-1332.
6. Szewzyk, U., Szewzyk, R., and Schink, B. 1985.
FEMS Microbiol. Ecol. 31:79-87.
7. Grbic-Galic, D., and Young, L.Y. 1985. Appl.
Environ. Microbiol. 50:292-297.
8. Vogel, T.M., and Grbic-GaliC, D. 1986. Appl.
Environ. Microbiol. 52:200-202.
9. Berry, D.F., Madsen, E.L., and Bollag, J.M. 1987.
Appl. Environ. Microbiol. 52:180-182.
10. Godsy, E.M., and Goerlitz, D.F. 1986. In Mattraw,
H.C., Jr., and Franks, B.J., eds., Movement and Fate of
Creosote Waste in Ground Water, Pensacola, Florida:
US Geological Survey Toxic Waste-Ground Water
Contamination Program. U.S. Geological Survey
Water Supply Paper 2285, Alexandria, Virginia, 63pp.
DEVELOPMENT OF A SEQUENTIAL
TREATMENT SYSTEM FOR
CREOSOTE-CONTAMINATED SOIL AND
WATER: BENCH STUDIES
James G. Mueller, Peter J. Chapman, and Parmely H.
Pritchard, US. EPA, Environmental Research Labora-
tory, Gulf Breeze, FL; Ron Thomas, Ellis L. Kline,
Suzanne E. Lantz. SBP, Inc., Pendelton, SC.
Summary
A triphasic, sequential treatment system for the
potential remediation of environments contaminated by
mixtures of hazardous wastes has been co-developed by
Southern BioProducls (SBP), Incorporated, Atlanta, GA
and the Gulf Breeze Environmental Research Laboratory,
Gulf Breeze, FL. This approach integrates established soil-
washing technology (phase I) with dewatering and chemi-
cal fractionation through a patented filtration process
(phase II). Together, these steps reduce the volume of
material to be subsequently treated from 100 to 100,000
fold. Volume reduction of this magnitude facilitates phase
III of this system: biodegradation of susceptible pollutants
employing EPA/SBP-patented bacteria capable of utilizing
high-molecular-weight creosote constituents as growth
substrates. The efficacy of this treatment system on
creosote-contaminated soil and water is currently being
evaluated at the bench-scale level.
Results and Discussion
A newly described, triphasic sequential treatment
system has been proposed for the efficient remediation of
creosote- and similarly-contaminated soil and water (Figure
1). The steps in this process entail: soil washing, mem-
brane extraction and biodegradation of extracted pollutants.
Essentially, all phases of this system are viewed as volume
reduction steps which significantly reduce the amount of
material requiring subsequent treatment (e.g., incineration).
Bench-scale studies are currently underway to assess the
efficacy of this system on creosote-contaminated soil and
water from the American Creosote Works Superfund site at
42
-------
SEQUENTIAL TREATMENT
Pensacola, Florida, to predict performance under field
conditions, and to evaluate the system from an economic
perspective.
Component portions of the treatment system have been
developed and shown to function according to design
criteria in individual tests. Previous studies have shown
each phase of this process to perform successfully when
operated independently from other components of the
system. Soil-washing technologies have demonstrated the
ability to remove creosote constituents from the surfaces of
larger-grained soil particles (>2.0 mm diam) and transfer
them to the wash water (aqueous phase)(1). Cleaned soil
particles are then separated from highly contaminated soil
fines (<2.0 mm diam). Depending on the texture of the
starting material (% sand, silt, clay) and pollutant
concentration, soil washing can reduce the amount of
contaminated soil that requires further treatment to as little
as 10% of the initial volume.
Whereas soil washing significantly reduces the volume
of material requiring subsequent treatment and removes
contaminants from the environment, the process generates
large amounts of contaminated wash water, and highly
contaminated fine soil particles accumulate. To address
these problems, a unique process of dewatering and
concentrating pollutants has been developed. Reverse
osmosis hyperfiltration through porous stainless steel
membranes (exclusive to SBP) has been shown to remove
>99% of creosote constituents present in creosote-contami-
nated groundwater (Table 1). Whereas the effectiveness of
this system on soil wash water is currently being evaluated,
it is expected to be equally efficient.
Creosote constituents removed from soil, wash water,
and groundwater and concentrated by reverse osmosis
hyperfiltration are subsequently fed to specially enriched
microbes housed in continuous flow bioreactors. The
ability of these organisms to degrade artificial creosote
mixtures has been demonstrated(2). However, their activity
on actual material produced through soil washing and
membrane extraction of field materials is currently being
evaluated.
Integration of these three phases in a sequential
treatment process circumvents some of the main problems
which currently limit the performance and applicability of
bioremediation technologies: 1) since pollutants are
removed from the environment and sequestered for subse-
quent destruction or recycling, application of large numbers
of organisms to field environments is not required, 2) the
system is designed to treat contaminated soil, groundwater,
and surface water simultaneously, 3) mixtures of chemicals
(as often found at hazardous waste sites) can be fractionat-
ed into related groups thereby enhancing their biodegrada-
tion by specially selected microorganisms, 4) toxic or
inhibitory compounds (i.e., heavy metals, radioactive
isotopes) can be removed (recycled) prior to biodegrada-
tion, and 5) by housing degradative microbes in monitored
bioreactors their catabolic activities can be optimized hence
accelerating the rate and extent of biodegradation. By
overcoming these barriers, bioremediation becomes
applicable to an increasing number of hazardous waste
problems.
Acknowledgments
This work is performed under a cooperative research
and development agreement between the Gulf Breeze
Environmental Research Laboratory and SBP, Inc. (Atlan-
ta, GA) as defined under the Federal Technology Transfer
Act (contract no. FTTA-003). We also acknowledge the
cooperation and support from Beverly Houston, U.S. EPA
Region IV.
References
1. Esposito, M.P., Locke, B.B., Greber, J., and Traver,
R.P. 1988. EPA/600/9-88/021 pp. 177-192.
2. Mueller, J.G., Chapman, P.J. and Pritchard, P.H. 1989.
Appl. Environ. Microbiol. 55:3085-3090.
43
-------
SEQUENTIAL TREATMENT
Table 1. Removal (membrane extraction) of creosote constituents from creosote-contaminated ground water collected
from the American Creosote Works site, Pensacola, Florida.
Compound
Naphthalene
l-Methylnaphthalene
2-Methylnaphthalene
Biphenyl
2,6-Dimethylnaphthalene
2,3-Dimethylnaphthalene
Acenaphthalene
Fluorene
Phenanthrenc
Anthracene
2-Methylanthracene
Anthraquinone
Fluoranthene
Pyrene
Benzo(b)fluorene
Chrysene
Benzo(a)pyrene
TOTALS
Number of peaks
Total peak area
Concentration of Pollutants (Mg/ml)
Feed
1.8
0.4
0.3
0.2
0.1
0.1
1.1
1.3
1.8
3.9
0.1
0.7
2.7
1.4
0.3
0.5
0.2
112
721,721
Permeate 1
ND1
ND
ND
ND
ND
ND
ND
ND
0.3
0.02
ND
0.05
ND
ND
ND
ND
ND
8
15,670
Permeate 2
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
ND
0
0
*ND = Not detected (<10 ng/mL).
44
-------
SEQUENTIAL TREATMENT
CONVENTIONAL
SOIL
WASHING
SBP
MEMBRANE
EXTRACTION
SBP / EPA
BIOREACTOR
I.
Figure 1. Schematic diagram of 5BP/EPA tri-phasic treatment process: soil washing, membrane
extraction/concentration, biodegradation.
45
-------
SEQUENTIAL TREATMENT
ANAEROBIC TREATMENT OF DIOXINS
AND DIBENZOFURANS
Dunja Grbic-Galic, Stanford University, Stanford, CA;
John A. Closer, U.S. EPA, Cincinnati, OH.
Introduction
A major hazardous waste problem confronting authori-
ties in the United States is the waste associated with the
wood treatment industry. The remediation efforts required
for an estimated 700 wood-treating sites serve to emphasize
the need for reliable remediation technology.
Pentachlorophenol, a potential fungicide, is a major
contaminant at wood-treating sites that exhibits significant
toxicity toward microflora. Chemical production of
pentachlorophenol has been shown to include contamina-
tion composed of higher chlorinated dioxins and dibenzofu-
rans. In a general sense, these components of the contami-
nation represent a small portion of the total waste material
but are regarded as highly toxic. As impurities in the
synthesis methodology forming chlorinated phenols,
especially pentachlorophenol, they are present in trace
quantities. The higher chlorinated dioxins and dibenzofu-
rans have physical properties, low water solubilities, and
low vapor pressures that ensure their environmental
persistence. The low chemical reactivity of these pollutants
results in concern regarding the eventual treatment of this
material.
Polychlorinated dioxins and dibenzofurans, due to their
high toxicity, carcinogenicity, bioaccumulation, and
persistence in various ecosystems, belong to a very special
class of hazardous environmental pollutants. Due to the
high degree of chlorination of these compounds, they tend
to be poor substrates for biological oxidation. Individually,
they are rather insoluble and are largely lipophilic.
The recent success of several research groups investi-
gating the anaerobic transformation of polychlorinated
biphenyls (PCBs) has prompted investigation of similar
transformations of dioxins and dibenzofurans. In the case
of the PCBs, transformation of highly chlorinated biphenyl
mixtures such as Arochlor 1260 to lower-chloririe-content
biphenyls has been observed. The well-advanced aerobic
transformation of PCBs has been judged limited to biphen-
yls with seven and fewer attached chlorines. The strengths
of the aerobic and anaerobic processes are seen as comple-
mentary. One direction of potential treatment development
has been to treat lower-chlorinated mixtures by aerobic
means and to use a sequential treatment of aerobic and
anaerobic processing for the higher-chlorinated congeners.
This same sequence of treatment may be useful for the
treatment of dioxins and dibenzofurans.
This project began by studying the anaerobic transfor-
mation of single substrates from the class of dioxin and
dibenzofuran congeners of five and greater attached
chlorines as cometabolic substrates for mixed anaerobic
microbial cultures growing on easily degraded organic
substrates. Screening of anaerobic consortia from a variety
of environmental sources served to identify and select the
desired activity. The research team then extended the basic
study to congeners mixtures and then to additional compo-
nents of the wood-treating waste in an effort to determine
applicability to realistic treatment conditions. As part of
this effort, pathways, rates, intermediates, and products of
the transformation process will be determined. Where
incomplete dehalogenation is encountered, conditions will
be explored to switch the treatment to aerobic conditions to
possibly permit the conversion of substrate to carbon
dioxide.
Conclusion
Ultimate development of anaerobic treatment systems
capableof degrading dibenzo[p]dioxins and dibenzofurans
with five or more attached chlorines is the objective for this
project. The importance of this class of pollutants is best
understood in the context of wood-treating waste. Penta-
chlorophenol is a major component of this generic waste.
Depending on the chemical synthetic means leading to the
formation of pentachlorophenol, these higher-chlorinated
dioxins and dibenzofurans accompany the waste. Toxico-
logical considerations often identify this small yet impor-
tant fraction of the total waste burden as a major objective
for the remediation of such waste sites.
DEGRADATION OF HETEROCYCLICS
Richard W. Eaton, Peter J. Chapman, and Parmely H.
Pritchard, US. EPA, Environmental Research Labora-
tory, Gulf Breeze, FL.
Many contaminants of ground water are heterocyclic
compounds that are potentially hazardous to human health
and to the environment. These classes of compounds along
with many others are formed as combustion products in the
generation of synthetic fuels from fossil fuels. Ground-
water contamination by these chemicals has been shown in
the vicinity of coal gasification plants, oil shale-retorting
46
-------
SEQUENTIAL TREATMENT
facilities, and at wood treatment facilities where coal-tar
derived creosote has been used. To develop biological
approaches to treat sites contaminated with these and other
chemicals, the microbiology and biochemistry of their
biodegradation must be understood. This research is
designed to provide such insights.
Compared to the wealth of information available from
studies of the bacterial degradation of aromatic compounds,
comparatively little is known of the transformation and
biodegradation of heterocyclic compounds. Recent con-
cerns about the persistence of hazardous pollutant chemi-
cals have led to a renewal of interest in the biodegradation
of this class of chemicals. Organisms have been isolated
from a variety of sources for their ability to utilize different
N-, O-, and S-heterocycles. For example, bacteria with the
ability to utilize such compounds as thiophene, dibenzo-
thiophene, dibenzofuran, pyridine, quinoline, isoquinoline,
and indole have been described in recent years. From these
and earlier studies of the pathways and enzymes employed
by axenic cultures of bacteria utilizing specific hetero-
cycles, certain features characteristic of the biochemistry of
their degradation can be discerned. Where aerobic bacteria
introduce hydroxyl groups into the ring systems of their
aromatic growth substrates by means of oxygenase en-
zymes (enzymes that catalyze incorporation of one or both
atoms of molecular oxygen into their substrates), bacteria
that utilize various heterocyclic compounds frequently
initiate attack on these ring systems by using water as a
source of oxygen for hydroxylation. First recognized in the
conversion of nicotinic acid to 6-hydroxynicotinic acid by
a Pseudomonasw, this type of hydroxylation has also been
noted for compounds with oxygen- or sulfur-containing
ring systems (Figure la). Thus, bacteria that utilize furan-
2-carboxylic and thiophene-2-carboxylic acids for
growth*2'3* use water as a source of oxygen to hydroxylate
the ring systems of these compounds as their coenzyme A
derivatives (Figure Ib). Quinoline-degrading bacteria*4*
evidently use the same mechanism in the formation of the
first metabolite, 2-hydroxyquinoIine (Figure Ic). Not
surprisingly, the anaerobic degradation of a number of
these compounds has been shown to involve the same
initial reactions as those taken by aerobic organisms.
Transformation of heterocyclic compounds by cometa-
bolic attack represents an alternative mode for their degra-
dation. For example, bacteria with the ability to oxygenate
phthalic acid have been shown to transform the correspond-
ing pyridine dicarboxylic acids such as quinolinic acid*5*
presumably to hydroxylated products. Compounds pos-
sessing both homocyclic and heterocyclic ring systems,
such as dibenzothiophene, are also subject to cometabolic
oxidation by organisms able to utilize biphenyl as growth
substrate*6*. Only one aromatic ring of dibenzothiophene
is attacked in these situations by reactions that are directly
analogous to the reactions undergone by naphthalene
(Figure 2a). A number of dibenzothiophene-utilizing
bacteria have been isolated and shown to employ the same
reactions, bringing about only partial degradation of this
compound*7*. By contrast, recent studies with bacteria
isolated for their ability to utilize the corresponding O-
heterocycle, dibenzofuran*8*, have shown that a novel
dioxygenation mechanism leads to cleavage of the furan
ring system before subsequent degradation of the aromatic
rings (Figure 2b).
Three approaches are currently being developed to
obtain microorganisms that can degrade heterocyclic
compounds:
1. Use of enrichment culture techniques with hetero-
cyclic compounds as sole sources of carbon and
energy or nitrogen. Enrichment experiments have
only been recently initiated, but successful isolation of
pure cultures of organisms using dibenzofuran, diben-
zothiophene, carbazole, and fluorene as sole sources of
carbon and energy has been achieved. In all of these
enrichments, dimethylsulfoxide was used (200 «g/mL)
to increase the solubility of the heterocyclic com-
pound. Soils contaminated with creosote have proved
to be valuable sources of such organisms.
2. Identification of aromatic hydrocarbon-degrading
microorganisms that can transform (cometabolize)
heterocyclic compounds. Certain bacterial strains
originally isolated for their ability to degrade either
cumcne (isopropylbenzene) or toluene, when induced
by growth with those compounds, are able to transform
dibenzofuran or dibenzothiophene. The colored
products of these transformations appear to have
resulted from oxygenase-catalyzed hydroxylation and
cleavage of one of the aromatic rings. It is also inter-
esting to note that only certain strains selected from
many aromatic hydrocarbon-degrading bacteria have
been shown to transform these heterocyclic com-
pounds.
3. Construction of hybrid catabolic pathways. Using
the information obtained from degradation studies, it
is planned to construct organisms and mixed cultures
of organisms that degrade heterocyclic compounds by
novel pathways. One interesting strategy for the
construction of dibenzothiophene-degrading organisms
was suggested by the work of Mormile and Atlas*9*.
This involves a mixed culture consisting of an aromat-
ic hydrocarbon-degrading bacterium such as the
47
-------
SEQUENTIAL TREATMENT
Beijerinclda sp. of LaBorde and Gibson(6), which
transforms dibenzothiophene to 3-hydroxy-2-formyl-
thianaphthene (Figure 2a) and an organism isolated
from enrichment cultures with the latter compound and
capable of using it as sole source of carbon and energy.
Subsequently, the cloned genes of these strains will be
introduced into the other to create a stable, pure culture
capable of extensive degradation of dibenzothiophene.
References
1. Hunt, A J... Hughes, D.E., and Lowenstein, J.M. 1958.
Biochem.J. 69:170-173.
2. Trudgill, P.W. 1969. Biochem. J. 113:577-587.
3. Cripps, R.E. 1973. Biochem. J. 134:353-366.
4. Pereira, W.E., Rostad, C.E., Leiker, TJ., Updegraff,
D.M., and Bennett, J.L. 1988. Appl. Environ. Micro-
biol. 54:827-829.
5. Taylor, B.F., and King, C.A. 1987. FEMS Microbiol.
Lett 44:401-405.
6. LaBorde, A.L., and Gibson, D.T. 1977. Appl. En-
viron. Microbiol. 34:783-790.
7. Monticello, DJ., Bakker, D., and Finnerty, W.R.
1985. Appl. Environ. Microbiol. 49:756- 760.
8. Engesser, K.H., Strubel, V., Christoglou, K., Fischer,
P., and Rast, H.G. 1989. FEMS Microbiol. 65:205-
210.
9. Mormile,M.R., and Atlas, R.M. 1988. Appl. Environ.
Microbiol. 54:3183-3184.
CT
^ki^
,COOH
• H2O
HO
COOH
COOK
-2H ,
HO
(S)
COOH
•HjO
(S)
-2H
COSCoA HO
(S)
OSCoA
OH
-2H
OH
Figure 1. Water as a source of oxygen for hydroxylation.
OH
CHO
H°H OH
OH
OH |
OH
Figure 2. Comparison of degradation routes for dibenzothiophene and dibenzofuran.
48
-------
METABOLIC PROCESS CHARACTERIZATION
SECTION VI
METABOLIC PROCESS CHARACTERIZATION
EPA's metabolic processes research generates a better understanding of the processes by which microorganisms
degrade chemicals, expanding the range of organisms that can be used in biosystems technologies. Based on the
insights gained from this research, scientists can then choose indigenous organisms or enhanced organisms to meet
needs in pollution cleanup and control. In one research project, EPA is currently developing systems tobiodegrade
pollutants formed as combustion products of synthetic fuel generated from fossil fuels. These heterocyclic
compounds are known to be persistent in the environment, but researchers are examining the metabolic processes
by which bacteria can break them down. Another study is focusing on how to handle a single but ubiquitous
pollutant: trichloroethylene (TCE). Examination of bacteria already shown to degrade TCE led to a better
understanding of the metabolic process by which TCE is oxidized. As a result, researchers have identified additional
strains that can more rapidly mineralize this pollutant.
THE INVOLVEMENT OF A TOLUENE
DEGRADATIVE PATHWAY IN THE
BIODEGRADATION OF
TRICHLOROETHYLENE BY
PSEUDOMONAS CEPACIA STRAIN G4
Malcolm S. Shields and Stacy O. Montgomery, Techni-
cal Resources Inc., Gulf Breeze, FL; Peter J. Chapman,
Stephen M. Cuskey, and Parmely H. Pritchard, U.S.
EPA, Environmental Research Laboratory, Gulf
Breeze, FL.
Summary
The oxidation of trichloroethylene (TCE) by bacteria
possessing enzymes that act on toluene is now established
for three species: P. cepacia^, P. putida(2\ and P. mendo-
c/W4). In all cases an oxygenase acting directly on the
aromatic ring of toluene is responsible for an attack on the
alternate substrate TCE. P. cepacia, strain G4, was shown
to catabolize toluene by a novel pathway. HPLC and
GC/MC analysis indicated toluene hydroxylations at first
the ortho and then meta positions to sequentially form o-
cresol and 3-methylcatechol. Analysis of pathway interme-
diates (by GC/MS) formed in a defined atmosphere of 18O2
and 16O2 confirmed the sequential nature of this monooxy-
genation activity and the source of the oxygen as O2(3).
The generation of several mutants unable to metabolize a
variety of related aromatic compounds—toluene, phenol,
cresol, catechol and hydroxymuconic semialdehyde
(HMS)—revealed that the only mutants that also suffered
loss of TCE-metabolizing ability were those that coinciden-
tally lost the ability to hydroxylate toluene, phenol, o- or m-
cresol. The principle of toluene ring oxygenation has led
to the identification of strains ofNocardia, Alcaligenes, and
Acinelobacter as capable of TCE degradation.
Results and Discussion
Mutants of P. cepacia, strain G4, obtained following
both N-methyl-N'-nitro-N-nitrosoguanidine (NTG) treat-
ment and transposon insertion (Tn5) were analyzed for
their ability to grow on toluene, phenol, and o- and m-
cresol as sole carbon sources; the accumulation of products
from toluene and phenol and enzyme activities are normal-
ly associated with these transformations (Table 1). One
NTG-induced mutant, G4 102, was shown to accumulate 3-
methylcatechol from toluene. Subsequent HPLC analysis
indicated that o-cresol was also transiently present.
Previously published workw has established a pathway of
toluene catabolism via 3-methylcatechol by P. putida that
involves a toluene-2-3,-dioxygenase. Enzymatic activities
from cell free extract of P. cepacia G4 indicated induced
levels of catechol-2, 3-dioxygenase and HMS hydrolase,
49
-------
METABOLIC PROCESS CHARACTERIZATION
CH,
OH
CH,
s- ^
a
1/2 O5
tomA
CH2COOH
CHP COOH
Hms
NAD
NADH
;)!
40c
Figure 1. Summary of the observed catabolic transformations of toluene, o-cresol, m-cresol, and phenol. The
proposed genes responsible are tomA. tomB, semialdchyde; 40C, 4-oxalocrotonate; Kpc, 2-ketopent-4-
cnoate.
50
-------
METABOLIC PROCESS CHARACTERIZATION
Table 1.
Comparison of mutant and wild phenotypes of P. cepacia G4.
P. cepacia G4
Derivatives8
G4
G4 100 (tomA)
G4 102 (tomB)
G4 103 (tomC)
Growth on
Toluene
+
_
_
-
Phenol
+
_
+
+
Activities of Enzymes from
Phenol-Induced Cellsb
(nmol/min/mg protein)
C23O
6,550
5,290
0.3
1,680
HMSH
112
195
148
0.2
HMSD
52
122
74
34
"Derivative cultures of G4 were obtained by mutagenesis with NTG.
bCatechol-2,3-dioxygenase (C23O), hydroxymuconic semialdehyde hydrolase (HMSH), hydroxymuconic
semialdehyde dehydrogenase (HMSD).
Figure 2. Toluene degradation pathways.
CH,
CH,
P. putlda F1
CH,
CH,
P. cepacia (G4)
CH,
OH
OH
CH
P. mendoclna
COOH
(Q)
TOL
51
-------
METABOLIC PROCESS CHARACTERIZATION
but no toluene-2,3-dioxygenase. HPLC and GC/MS
analyses of 3-methylcatechol produced by G4 102 from
toluene in an atmosphere containing a defined mixture of
stable oxygen isotopes clearly indicate that the catabolism
of toluene proceeds through the independent stepwise
addition of a single atom of oxygen from diatomic oxygen
molecules, first at the ortho then at the meta position. This
activity is consistent with the action of a toluene monooxy-
genase. These analyses have led to a more complete
understanding of aromatic catabolism by P. cepacia G4
(Figure 1).
Enzymatic and biodegradative analyses indicate that
the ability to degrade TCE is directly correlated with the
ability of G4 to hydroxylate toluene or phenol. Under-
standing the relationship of the toluene ring oxygenases to
TCE catabolism (Figure 2) has allowed the identification of
several other TCE-degrading bacteria: Acinetobacter (two
strains), Alcaligenes eutrophus (two strains), Nocardia
corallina (one strain), and two more strains of P. cepacia.
Current efforts are directed toward genetic alteration of
the control mechanism that requires the use of an aromatic
inducer to achieve TCE catabolism. This will greatly
simplify bioreactor design constraints and scenarios for the
in-situ treatment of TCE-contaminated environments.
References
1. Gibson, D.T., Hensley, M., Yoshioka, H., and Mabry,
T.J. 1970. Biochem.9:1626-1630.
2. Nelson, M.J.K., Montgomery, S.O., Mahaffey, W.R.,
and Pritchard, P.H. 1987. Appl. Environ. Microbiol.
55:949-954.
3. Sheilds, M.S., Montgomery, S.O., Chapman, P.J.,
Cuskey, S.M., and Pritchard, P.H. 1989. Environ.
Microbiol. 55:1624-1629.
4. Winter, R.B., Yen, K.M., and Ensley, B.D. 1989.
Bio/Technology. 7:282-285.
DEGRADATION OF HALOGENATED
ALIPHATIC COMPOUNDS BY THE
AMMONIA-OXIDIZING BACTERIUM
NITROSOMONAS EUROPAEA
Todd Vannelli, Myke Logan, David M. Arclero, and
Alan B. Hooper, University of Minnesota, St. Paul, MN;
Peter J. Chapman, US. EPA, Environmental Research
Laboratory, Gulf Breeze, PL.
The ubiquitous soil- and water-dwelling nitrifying
bacteria are obligate autotrophic aerobes that depend for
growth on the oxidation of ammonia by ammonia monoxy-
genase (AMO); 2H+ +2e + O2 + NH3-»NH2OH + H^O.
Two electrons for the reaction originate in the subsequent
reaction of hydroxylamine oxidoreductase; H2O+NH2OH— »
4e' + 5H+ + NO'.
Halogenated substrates were measured by electron
capture detector after gas chromatography*1*. Substrates,
except vinyl chloride (200 ^M), were at a concentration of
Ippm (-10 pM). All 16 compounds tested, except for
tetrachloromethane, tetrachloroethylene, and trans-dibro-
moethylene, were degraded at rates comparable to values
for trichloroethylene catalyzed by ammonia-, toluene- and
methane-oxidizing bacteria. Acetylene or 2-chloro-6-
trichloromethyl pyridine (1 mM), which specifically inhibit
the ammonia-oxidizing system in Nitrosomonas, inhibited
degradation of halogenated aliphatics by at least 70%.
Thus, the reaction is at least dependent on and probably
catalyzed by the ammonia oxygenase. In all cases in which
the test compound was degraded, its presence also de-
creased the rate of nitrite production from ammonia (Table
1), consistent with competition for active site. Degradation
was not accompanied by inactivation of the enzyme, and
within 24 hr most or all of the test compound had disap-
peared. Degradation of substrate was observed in the
absence of added ammonia although the rates and extent
were always greater with ammonia (Table 1).
Nitrosomonas will oxidize with cis- and trans-2-
butene, which are structural analogs of the dibromoethyle-
nes; 2-butene-l-ol and lesser amounts of an epoxide are
produced from /ran5-2-butene, whereas c/s-2-butene is
converted to an equal mixture of both. We rationalize our
observations with a model for an active site containing an
oxygen-activating region and a hydrophobic pocket. The
cw-butene may be positioned by the hydrophobic pocket
for easy reaction of double bond with the oxygen-activating
site, whereas frans-butene would be better positioned for
hydroxylation at the methyl group. The unreactive Br of
f rans-dibromoethylene would be positioned at the oxygen-
52
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METABOLIC PROCESS CHARACTERIZATION
Table 1.
Oxidation of halogenated aliphatic compounds by Nitrosomonas.
Substrate1
Ammonia
Dichloromethanea
Dibromomethaneb
Trichloromethanec
Tetrachloromethaned
Bromomethane0
1 ,2-Dibromomethanef
1 , 1 ,2-Trichloroethane
1,1,1 -Trichloroethane
Chloroethylene8
gem-Dichloroethyleneh
cw-Dichloroethylene
/ra/is-Dichloroethylene
cw-Dibromoethylene
frans-Dibromoethylene
Trichloroethylene
Tetrachloroethylene1
1 ,2,3-Trichloropropane
Cells
A Subst./
A Time2
.
6.5
5.3
3.3
0.0
5.5
1.0
1.9
1.6
15
1.1
0.7
2.2
0.4
0.0
2.8
0.0
0.9
Cells + NH3
A Subst./
A Time2
-
9.8
7.2
4.5
0.0
7.3
11
4.2
2.2
57
3.9
8.8
3.9
12
0.0
6.7
0.0
2.0
A Nitrite/
A Time2
970
650
590
600
620
880
820
660
600
230
780
400
850
690
690
790
140
660
Substrate
Remaining3
-
0.0
4.2
41
110
25
16
35
80
23
53
9.1
25
3.0
100
6.4
120
77
'Common names:
"Methylene chloride
bMethylene bromide
cChloroform
""Carbon tetrachloride
^thylbromide
2Initial rates (^imoles/hr/g of wet weight cells).
3Value at 24 hr (% of initial value).
fEthylene dibromide
8Vinyl chloride
hVinylidene chloride
'Perchloroethylene
53
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METABOLIC PROCESS CHARACTERIZATION
activating site and block reaction with the double bond,
whereas the double bond of cw-dibromoethylene would be
positioned for easy reaction.
In nature or in pollution treatment, actively nitrifying
Nitrosomonas would appear to have a potential role in the
degradation of halogenated aliphatic compounds. With
compounds such as dichloromethane ("methylene chlo-
ride") where the compound is volatile or gaseous, cycling
pollutant-laden air over beds of Nitrosomonas has the
potential of effectively removing the compound. We note
that decontamination of soils after sterilization with 1,2-
dibromoethane ("ethylene dibromide") may be promoted by
Nitrosomonas.
Reference
1. Arciero, D., Vannelli, T., Logan, M, and Hooper, A.B.
1989. Biochem. Biophys. Res. Commun. 159:640-
643.
DEGRADATION OF CHLORINATED
AROMATIC COMPOUNDS UNDER
SULFATE-REDUCING CONDITIONS
Patricia JS. Colberg, Department of Molecular Biolo-
gy, University of Wyoming, Laramie, WY; John Rogers,
US. EPA, Athens, GA.
Microbially mediated sulfate reduction has long been
known to be responsible for virtually all mineralization of
organic matter in marine sediments. Laanbroek and Pfen-
nig(1) were the first to postulate that sulfate-reducing
bacteria may be able to replace methanogens at the terminal
end of the microbial food chain in other environments. We
now know that the role of dissimilatory sulfate-reducing
bacteria in the global cycling of carbon is, in fact, analo-
gous to that of the methanogenic bacteria and that sulfate
respiration and methanogenesis may be viewed as alterna-
tive terminal processes(2).
The number of dissimilatory sulfate- and sulfur-
reducing genera described has rapidly grown from only two
to ten over the last decade. The recent expansion of our
knowledge of dissimilatory sulfate respiration has been
coincidental with progress made in understanding other
anaerobic transformations.
Even though the transformation of both haiogenated
and nonhalogenated aromatic compounds is thermodynami-
cally favorable under sulfate-reducing conditions, our
understanding of the extent to which sulf idogenic environ-
ments may contribute to the turnover of such substrates is
severely limited. This may be due, in a large part, to the
long-held view that anaerobic microorganisms are able to
degrade only a limited number of simple substrates and that
the sulfate-reducers are restricted to using lactate, ethanol,
and pyruvate. In fact, the first study to use an aromatic
compound (benzoate) for initial enrichment of sulfate
reducers from an environmental sample was not published
until 1983(3).
The number of aromatic substrates known to be
amenable to microbial sulfate reduction is still compara-
tively small, but includes a growing list of nonhalogenated
compounds such as phenol, catechol, p-cresol, 4-hydro-
xybenzoate, and benzoate. It is perhaps fortuitous that the
only known obligately anaerobic dehalogenating bacterium
happens to be a sulfidogen. It was originally isolated from
an anaerobic consortium that mineralized 3-chlorobenzoic
acid and was recently named Desulfomonile tiedje^.
Except for some laboratory evidence that suggests
sulfate-mediated degradation of 2-chlorophenol, 4-chloro-
phenol, and 3-chlorobenzoate(5) and the well-documented
removal of halogen substituents by D. tiedje, there are few
other published reports of transformations of chlorinated
compounds by dissimilatory sulfate-reducing bacteria. The
concepts of "microbial food chains" and syntrophic associa-
tions have already gained acceptance as applied to mixed
methanogenic systems and have enhanced our expectations
of defining equally significant role for sulfate-reducing
consortia in anoxic environments that have been contami-
nated with hazardous substances.
The objectives of our three-year project are to:
1. Develop cultures or consortia that are able to dehalo-
genate five classes of chlorinated aromatic compounds
(i.e., chlorinated benzenes, phenols, benzoates, ani-
lines, biphenyls)
2. Determine the activity of the cultures obtained over a
range of environmental conditions (e.g., pH, tempera-
ture, salinity)
3. Evaluate the substrate specificity of the cultures
toward all of the compound classes
4. Evaluate in microcosm and field experiments the
ability of these cultures to enhance degradation of
hazardous chlorinated aromatic compounds in contam-
inated soils and sediments.
54
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METABOLIC PROCESS CHARACTERIZATION
Ackno wledgmen t
This project is being supported by the EPA Environ-
mental Research Laboratory, Athens, GA, through Cooper-
ative Agreement CR-816398-01-0.
References
1. Laanbroek, H.J., and Pfennig, N. 1981. Arch.
Microbiol. 128:330-335.
2. Widdel, F. 1988. In A.J.B. Zehnder (ed.), Biology of
Anaerobic Microorganisms, John Wiley and Sons,
Inc., New York.
3. Widdel, R, Kohring, G.W., and Mayer, F. 1983.
Arch. Microbiol. 134:286-294.
4. DeWeerd, K.A., Mandekco, L., Tanner, R.S., Woese,
C.R., and Suflita, J.M. Arch. Microbiol. 154:23-30.
5. Genthner, B.R., Sharak, W.A. Price II, and Pritchard,
P.H. 1989. Appl. Environ. Microbiol. 55:1466-1477.
THE ROLE OF FUNGAL
LIGNIN-DEGRADING ENZYMES IN
AROMATIC POLLUTANT
BIODEGRADATION
K.E. Hammel, Dept. of Chemistry, State University of
New York, College of Environmental Science and For-
estry, Syracuse,NY; JA. Closer, U.S. EPA, Cincinnati,
OH.
The ligninolytic fungi that cause white rot of wood
have recently become the object of increasing attention
from workers in the hazardous waste field. These fungi
normally grow on decaying wood and forest litter, and
appear to be unique among microorganisms in that they can
rapidly depolymerize and mineralize lignin, a complex,
irregular, nonhydrolyzable, and environmentally persistent
wood polymer of phenylpropane subunits. Lignin contains
numerous substructures that are also found in common
organic pollutants, and its exceptional recalcitrance to
attack by most microbes has led many researchers to
surmise that any organism capable of mineralizing it must
have highly nonspecific oxidizing systems that could be
applicable to aromatic pollutants as well. Recent work has
shown that ligninolytic fungi can degrade a wide variety of
aromatic pollutants, and that some of these compounds,
certain polycyclic aromatic hydrocarbons in particular, are
also substrates for the unique extracellular lignin peroxi-
dases (LiP's) that are instrumental in lignin degradation by
these organisms. It remains unclear, however, whether
LiP's are actually necessary in vivo for the biodegradation
of aromatic pollutants by Phanerochaete, and the goal of
this work is to elucidate their role in fungal xenobiotic
metabolism.
Anthracene is the simplest PAH that is a substrate for
LiP. The LiP-catalyzed spectral changes obtained with this
PAH showed disappearance of the starting material and
accumulation of a product with a peak at 253 nm and
shoulder at 275 nm, which was consistent with the produc-
tion of anthraquinone. Thin-layer chromatography and gas
chromatography showed only one product, which co-
chromatographed with authentic anthraquinone, and
GC/MS analysis confirmed this identification. Mass
spectrum: m/z (relative intensity) 208 (M+, 100), 180 (-CO,
85) 152 (-2CO, 35), 151 (15), 76 (15). No products were
detected that would indicate a direct LiP-catalyzed ring
cleavage of anthracene.
A key question, then, was whether anthraquinone
would occur as a metabolite of anthracene in fungal
cultures, and an HPLC analysis was therefore done of the
ethyl acetate-extractable neutral fraction from nitrogen-
limited P. chrysosporium cultures that had been given
anthracene (10 joM) 4 days after inoculation, at which time
LiP activity was already present. The results showed that
all of the added anthracene disappeared from the culture
fluid within 24 hr, and that anthraquinone was a prominent
neutral metabolite. Ethyl acetate extracts of the fungal
mycelium gave the same result, and, in all, about 40% of
the added anthracene was recovered as extractable anthra-
quinone after 1 d of incubation.
The occurrence of anthraquinone as an anthracene
metabolite does not necessarily show that the quinone is an
intermediate in anthracene mineralization, unless it can be
shown that the two compounds are mineralized at compara-
ble rates in culture. To address this problem, we next
compared the ability of P. chrysosporium to mineralize
[14Cj.8]anthracene and [14C,.8]anthraquinone. The
anthraquinone was purchased and chromatographically
repurified to give a radiochemical purity of 99%, whereas
the anthracene was synthesized by reduction of the quinone
with hydriodic acid and recrystallized, also to a
radiochemical purity of 99%. Both compounds (3.2 mCi
mmol"1* 0.4 jxM final concentration) were added to rotary-
shaken fungal cultures (25 mL) at the time of inoculation,
and 14CO2 evolution was then monitored daily by trapping
55
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METABOLIC PROCESS CHARACTERIZATION
and scintillation counting. Both the PAH and its quinone
were mineralized by Phanerochaete at approximately the
same rate, giving approximately 11% of anthracene, and
10% of anthraquinone, oxidized to CO2 after 14 d. For
both anthracene and anthraquinone, the day of onset of
mineralization occurred between days 2 and 3, which
agrees well with the time at which LiP activity was first
detectable in these cultures (day 3) and with the previously
reported day of onset of lignin degradation in agitated
cultures (also day 3). These results, taken together with the
finding that anthraquinone is a major anthracene metabolite
in fungal culture, support the hypothesis that the quinone,
produced in an LiP-catalyzed reaction, is an intermediate in
anthracene mineralization.
Certain PAH that are not LiP substrates have been
reported by other workers to undergo mineralization in
P. chrysosporium cultures that were not grown in the usual
defined ligninolytic culture media. For example, [14C9]-
phenanthrene was mineralized (ca. 1% in 30 days) in
nitrogen-limited medium to which 50 mg I"1 of anthracene
oil, a crude PAH mixture similar to creosote, had been
added. However, no control experiment was described that
would have determined whether phenanthrene can be
mineralized by P. chrysosporium without the addition of a
crude PAH mixture. Similarly, Phanerochaete was
reported to mineralize naphthalene in the presence of coal
tar (another crude PAH mixture), soil, and 3,4-
dimethoxybenzyl alcohol. It was stated that mineralization
was negligible in the absence of these additives, but no
experiments were reported that would have shown which of
them was essential.
To address this problem, we next examined the ability
of P. chrysosporium to mineralize the LiP nonsubstrate
naphthalene. We found that in low nitrogen culture
medium, which induces the production of LiP's, virtually
no [14Cj]naphthalene was mineralized. By contrast, a
concurrently run low nitrogen positive control with
[ Ci_8]anthracene showed that the course of mineralization
for this LiP substrate was the same as in the
anthracene/anthraquinone experiment that had been done
previously. Finally, the use of high nitrogen culture
medium, which inhibits the expression of LiP activity,
suppressed anthracene mineralization almost entirely. In
interpreting this experiment, we have considered two
additional points: 1) the inability of the cultures to
mineralize naphthalene cannot be attributed to problems
with the solubility of this PAH, since anthracene, which is
mineralized, is even less soluble; and 2) there is no reason
to suspect that the lack of 14CO2 evolution from
[14C,]naphthalene is due merely to a general inability of
Phanerochaete to mineralize angular carbons in PAH:
Although our mineralization results with [14C1_8]anthracene
do not prove that CO2 is necessarily released from Cl of
this compound, previous results with
[14C7 10]benzol[a]pyrene have specifically shown that
angular carbons can be mineralized.
Radiolabeled positions in the anthracene and
naphthalene used in this study, and in the
benzo{a]pyrene employed for previous work.
We conclude, therefore, that there are only two
explanations for the grossly qualitative difference obtained
with anthracene vs. naphthalene in mineralization
experiments: 1) either naphthalene is not oxidized at all in
standard low-nitrogen fungal cultures, because it is not a
LiP substrate and no other route for its oxidation is
available, or 2) naphthalene is oxidized, e.g., in cytochrome
P-450-catalyzed reactions, but the products of these
reactions are not substrates for subsequent ring cleavage.
56
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RISK ASSESSMENT
SECTION VII
RISK ASSESSMENT
A number of the high-priority compounds that require disposal are known carcinogens orprocarcinogens. Since
biodegradation in the field (and well as sometimes in reactors) does not necessarily result in total degradation to
carbon dioxide and water, researchers need to assess whether ultimate orprocarcinogens are created by a given
biosystem technology. Because low risk is associated with using indigenous organisms, the potential ecological
effects of nonindigenous organisms have received the most attention. In the area of human health, ORD is focusing
on short-term genotoxicity testing to provide the necessary data for risk assessment and on the development of an
appropriate animal model to determine fate and effects of these microorganisms in human digestive and
reproductive tracts.
Before environmental engineers can adopt a biologically based remediation technology, they need to know
which alternative is most feasible, most efficient, and provides the least possible hazard/risk to the environment and
human health. Work in this research area, therefore, is designed to develop comparative risk assessment methods
to evaluate and compare mutagenicl carcinogenic products generated by different microbial treatment processes in
different environmental settings.
DEVELOPING GENOTOXICITY RISK
ASSESSMENT DATA
L.D. Claxton, U.S. EPA, Research Triangle Park, NC.
Microorganisms are already used for large-scale
environmental treatments of hazardous substances, and this
type of treatment is expected to increase in frequency. The
greatest risks from these activities are anticipated to be
ecological, rather than health, because known human
pathogens will not be approved for use in bioremediation
technologies. Nevertheless, the Environmental Protection
Agency is responsible for examining the potential human
health hazard because a finite risk does exist. The Office
of Health Research (OHR) in the Office of Research and
Development (ORD) is improving EPA's capabilities to
protect the population against health risks by identifying
possible risks, prioritizing them by the magnitude of the
risks, and developing recommendations for research and
risk assessment based on this analysis.
A survey of past and current efforts to systematically
define health risks associated with the environmental
release of biotechnology products is reported in a document
entitled "A Survey of Research Efforts Associated with the
Health Effects of Microorganisms Used in the Field of
Biotechnology." The document, which examines both the
scientific and regulatory aspects associated with biotech-
nology health issues, makes clear that studies involving
biotechnology-related organisms have focused on ecologi-
cal, food, drug, pollution control, and other areas not
directly related to detecting and evaluating human health
effects. In part, this focus stems from the recognition that
the introduction of nonindigenous organisms into novel
ecosystems can result in serious adverse ecological conse-
quences. EPA and other groups associated with the
environmental release of biotechnology products, therefore,
have concentrated their efforts on ecological rather than
health effects. As the document makes clear, however,
health effects should not be ignored. In the unlikely event
that an environmental release did cause human disease,
significant adverse impacts would occur. Health effects
due to bioremediation efforts also are not limited to the
effects of the microorganisms. As toxic compounds and
other co-contaminants are metabolized, additional toxic
materials may be produced. This can occur within the
57
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RISK ASSESSMENT
environmental situation itself or may occur within an
exposed organism. The risk assessment process for
bioremediation, therefore, must assess the involved micro-
organisms, known toxicants, potentially produced toxi-
cants, and potential interactions associated with human
exposure.
Although biotechnology health research needs can be
grouped in several ways, the research can be categorized
into four major categories:
1. Infectivity/pathogenicity questions
2. Allergic reaction issues
3. Toxicity
4. Modulation effects
The Health Effects Research Laboratory is establishing a
prioritized research program that examines these issues
while supporting the immediate needs of the Biosystems
Technology Development Program. This presentation will
include a discussion of on-going research that supports the
Biosystems Program. The type of human health risk
assessment research needed in support of bioremediation
efforts, development of a semi-automated genotoxicity
assay for use with environmental mixtures, the effect of
intestinal metabolism upon multiple chemical exposures,
and the development of an animal model for determining
the survival of environmentally released engineered
bacteria within Ihe human intestinal tract also will be
discussed. The role of these research activities in the risk
evaluation of bioremediation approaches and strategies will
be illustrated.
This is an abstract of a proposed presentation and does
not necessarily reflect the views of the U.S. EPA.
58
U.S. GOVERNMENT PRINTING OFFICE: 1991-548-187/20527
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