United States Environmental Protection Agency
Office of Research and Development
SYMPOSIUM ON BIO REMEDIATION
OF HAZARDOUS WASTES:
EPA's Biosystems Technology
Development Program
ABSTRACTS
Fairview Park Marriott
Falls Church, Virginia
April 16-18, 1991
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United States Environmental Protection Agency
Office of Research and Development
SYMPOSIUM ON BIOREMEDIATION
OF HAZARDOUS WASTES:
EPA's Biosystems Technology
Development Program
ABSTRACTS
Fairview Park Marriott
Falls Church, Virginia
April 16-18, 1991
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Disclaimer
The information in this document has been funded wholly or in part by the U.S. EPA. It has not
been subjected to Agency review and is intended for internal use only. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
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SYMPOSIUM ON BIOREMEDIATION OF HAZARDOUS WASTES:
U.S. EPA'S BIOSYSTEMS TECHNOLOGY DEVELOPMENT PROGRAM
April 16-18, 1991
Falls Church, Virginia
AGENDA
April IS. 1991
7:00 - 9:00 p.m. Early Onsite Check-in
April 16. 1991
7:30 a.m.
8:30 a.m.
9:00 a.m.
Session I:
9:15 a.m.
9:25 a.m.
9:50 a.m.
10:05 a.m.
10:30 a.m.
Session II
Onsite Check-in/Registration
Introduction
Walter Kovalick, Director, Technology Innovation Office, U.S. EPA,
Office of Solid Waste and Emergency Response, Washington, D.C.
E. Timothy Oppelt, U.S. EPA, Risk Reduction Engineering Lab,
Cincinnati, OH
Overview of the Biosystems Program
John A. Glaser, U.S. EPA, Cincinnati, OH
Ground-water Treatment
Session Chairperson: John Wilson, U.S. EPA, Ada, OK
Introduction to Bioventing
John Wilson, U.S. EPA, Ada, OK
Bioventing of an Aviation Gasoline Spill: Design and Operation of a Field
Demonstration
John Armstrong* and Christopher Griffin,* Traverse Group, Inc.,
Ann Arbor, MI
BREAK
Bioventing of an Aviation Gasoline Spill: Performance Evaluation of a
Field Demonstration
Don Kampbell, U.S. EPA, Ada, OK
Laboratory and Field Studies of the Kinetics of Bioventing
David Ostendorf* and Ellen Moyer, University of Massachusetts,
Amherst, MA
Don Kampbell, U.S. EPA, Ada, OK
Treatment of Aqueous Wastes in a Reactor
Session Chairperson: To Be Announced
Indicates presenter when there are several authors.
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Page 2
10:55 a.m. Treatment of Wastewater with the White Rot Fungus Phanerochaete
Chrysospormm
Thomas Joyce* and Hou-min Chang, North Carolina State University,
Raleigh, NC
11:35 a.m. Anaerobic/Aerobic Sequential Treatment of CERCLA Leachates in POTWs:
An Innovative Treatment Approach
Margaret Kupferle and Paul Bishop, University of Cincinnati, OH
Dolloff F. Bishop* and Steven Safferman, U.S. EPA, Cincinnati, OH
12:15 p.m. LUNCH
1:30 p.m. Manipulation of TCE-degradative Genes of Pseudomonas cepacia
Malcolm Shields, TRI, Gulf Breeze, FL
1:50 p.m. Treatment of TCE and Degradation Products Using Pseudomonas cepacia
Malcolm Shields, TRI, Gulf Breeze, FL
2:10 p.m. Treatment of CERCLA Leachates by Carbon-Assisted Anaerobic
Fluidized Beds
Makram Suidan,* R. Nath and A.T. Schroeder, University of
Cincinnati, Cincinnati, OH
E.R. Krishnan, PEI Associates, Cincinnati, OH
R.C. Brenner, U.S. EPA, Cincinnati, OH
2:50 p.m. BREAK
3:05 p.m. Improved Prediction of GAC Capacity in a Biologically Active Fluidized
Bed
Makram Suidan* and R.D. Vidic, University of Cincinnati,
Cincinnati, OH
R.C. Brenner, U.S. EPA, Cincinnati, OH
Session 111 Soil/Sediment Treatment
Session Chairperson: John A. Glaser, U.S. EPA, Cincinnati, OH
3:45 p.m. Use of Lignin-Degrading Fungi in the Remediation of Pentachlororophenol-
Contaminated Soils
Richard Lamar,* D.M. Dietrich and T.K. Kirk, USDA,
Madison, WI
4:25 a.m. Anaerobic Degradation of Chlorinated Aromatic Compounds
John Rogers,* U.S. EPA, Athens, GA
Dorothy Hale, Wren Howard, and Frank Bryant, TAJ, Athens, GA
Mahmoud Mousa and Shiu-mei Hsu, UGA, Athens GA
4:45 p.m.
6 00 p.m.
ADJOURN
Cash Bar Reception and Banquet (Dinner at 7:00 p.m.)
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Page 3
April 17, 1991
8:30 a.m. para-Hvdroxvbenzoate as an Intermediate in the Anaerobic
Transformation of Phenol to Benzoate
Barbara R. Sharak Genthner, TRI, Gulf Breeze, FL
8:50 a.m. Aerobic Degradation of Chlorinated and Non-chlorinated PAHs
Peter Chapman, U.S. EPA, Gulf Breeze, FL
9:30 a.m. Influence of Nonionic Surfactants on the Anaerobic Dechlorination of
Hexachlorobenzene
Patricia Van Hoof* and Chad Jafvert, U.S. EPA, Athens, GA
10:10 a.m. BREAK
10:25 a.m. Anaerobic Degradation of Chloroaromatic Compounds Under Different
Reducing Conditions
M.M. Haggblom, M.D. Rivera, and Lily Young,* New York
University, New York, NY
John E. Rogers, U.S. EPA, Athens, GA
Session IV Combined Treatment
Session Chairperson: John Rogers, U.S. EPA, Athens, GA
11:05 a.m. Onsite Biological Pretreatment Followed by POTW Treatment of
CERCLA Leachates
Rodha Krishnan,* R.C. Haught, and M.L. Taylor, PEI Associates,
Inc., Cincinnati, OH
M.T. Suidan and M. Islam, University of Cincinnati, Cincinnati, OH
R. C. Brenner, U.S. EPA, Cincinnati, OH
11:45 a.m. LUNCH
1:15 p.m. Bacterial Degradation of KPEG-modified PCBs In Anaerobic
and Aerobic Enrichment Cultures
Wale Adewunmi and Joseph Krzycki,* Ohio State University,
Columbus, OH
1:55 p.m. Aerobic Biodegradation of Volatile Organic Compounds in a Biofilter
Vivek Utgikar and Rakesh Govind, University of Cincinnati,
Cincinnati, OH
Dolloff F. Bishop,* U.S. EPA, Cincinnati, OH
Session V .Sequential Treatment
Session Chairperson: P. Hap Pritchard, U.S. EPA, Gulf Breeze, FL
2:35 p.m. Methanogenic Degradation of Kinetics of Phenolic Compounds
Michael Godsy* and Donald Goerlitz, U.S. Geological Service,
Menlo Park, CA
Dunja Grbic-Galic, Stanford University, Palo Alto, CA
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Page 4
3:15 p.m. BREAK
3:30 p.m. Aerobic Biodegradation of Creosote
James Mueller,* Suzanne Lantz, Ron Thomas, and Ellis Kline,
Southern Bio Products, Inc., Gulf Breeze, FL
Peter Chapman, Douglas Middaugh, and P. Hap Pritchard, U.S.
EPA, Gulf Breeze, FL
Richard Coivin, Allan Rozich, and Derek Ross, Environmental
Resources Management, Exton, PA
4:10 p.m. Anaerobic Degradation of Highly Chlorinated Dioxins and Dibenzofurans
Peter Adriaens* and Dunja Grbic-Galic, Stanford University,
Palo Alto, CA
4:50 p.m. Degradation of Naphthalene, PAHs, and Heterocyclics
Richard Eaton* and Peter Chapman, U.S. EPA Gulf Breeze, FL
5:30 p.m. ADJOURN
April 18. 1991
Session VI Metabolic Process Characterization
Session Chairperson: Albert D. Venosa, U.S. EPA, Cincinnati, OH
8:30 a.m. Degradation of Halogenated Aliphatic Compounds by the Ammonia-
Oxidizing Bacterium Nixrosomonas Europaea
Alan Hooper* and Todd Vanelli, University of Minnesota,
Minneapolis, MN
Peter Chapman, U.S. EPA, Gulf Breeze, FL
9:10 a.m. Degradation of Chlorinated Aromatic Compounds Under Sulfaie-Reducing
Conditions
Patricia Colberg,* University of Wyoming, Laramie, Wyoming
John E. Rogers, U.S. EPA Athens, GA
9:35 a.m. Ring-Fission of Polycyclic Aromatic Hydrocarbons by White Rot Fungi
Kenneth Hammel, State University of New York, Syracuse, NY
10:15 a.m. BREAK
10:30 a.m. Aerobic Biodegradation of Polychlorinated Biphenyls: Genetic and Soil
Studies
Frank Mondello* and Bruce Erickson, General Electric, Schenectady, NY
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Page 5
11:10 a.m.
11:50 p.m
1:15 p.m.
1:55 p.m.
2:35 p.m.
3:00 p.m.
Genotoxicity Assays of Metabolites from Biological Treatment Processes
Larry Claxxon, U.S. EPA, Research Triangle Park, NC
LUNCH
Results of the Bioremediation Field Inititative
Fran Kremer, U.S. EPA, Cincinnati, OH
John Wilson,* U.S. EPA, Ada, OK
Walter Kovalick and Nancy Dean, U.S. EPA, Washington D.C.
Oil Spill Bioremediation Project: Summary of Laboratory and Field Studies
P. Hap Pritchard, U.S. EPA, Gulf Breeze, FL
Future Directions of the Biosystems Program
John A. Glaser, U.S. EPA, Cincinnati, OH
ADJOURN
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TABLE OF CONTENTS
Page
Bioventing of an Aviation Gasoline Spill:
Design and Operation of a Field Demonstration 1
Bioventing of an Aviation Gasoline Spill:
Performance Evaluation of a Field Demonstration 5
Laboratory and Field Studies of the Kinetics of Bioventing 7
Treatment of Wastewater with the White Rot Fungus
Phanerochaete Chrysosporium 11
Anaerobic/Aerobic Sequential Treatment of CERCLA Leachates
in POTWs: An Innovative Treatment Approach 13
Manipulation of TCE-degradative Genes of Pseudomonas cepacia 17
Treatment of TCE and Degradation Products Using Pseudomonas cepacia 21
Treatment of CERCLA Leachates by Carbon-Assisted Anaerobic
Fluidized Beds 25
Improved Prediction of GAC Capacity in a Biologically Active
Fluidized Bed 29
Use of Lignin-Degrading Fungi in the Remediation of
Pentachlororophenol-Contaminated Soils 35
Anaerobic Degradation of Chlorinated Aromatic Compounds 39
para-Hydroxvbenzoate as an Intermediate in the Anaerobic
Transformation of Phenol to Benzoate 43
Influence of Nonionic Surfactants on the Anaerobic
Dechlorination of Hexachlorobenzene 47
Anaerobic Degradation of Chloroaromatic Compounds Under
Different Reducing Conditions 51
Onsite Biological Pretreatment Followed by POTW Treatment of
CERCLA Leachates 55
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Table of Contents cont.
Page
Bacterial Degradation of KPEG-modified PCBs in Anaerobic and
Aerobic Enrichment Cultures 59
Aerobic Biodegradation of Volatile Organic Compounds in a BioQlter 65
Methanogenic Degradation of Kinetics of Phenolic Compounds 67
Aerobic Biodegradation of Creosote 71
Anaerobic Degradation of Highly Chlorinated Dioxins and Dibenzofurans 77
Degradation of Naphthalene, PAHs, and Heterocyclics 81
Degradation of Halogenated Aliphatic Compounds by the Ammonia-Oxidizing
Bacterium Nitrosomonas Europaea 85
Degradation of Chlorinated Aromatic Compounds Under
Sulfate-Reducing Conditions 89
Ring-Fission of Polycyclic Aromatic Hydrocarbons by White Rot Fungi 93
Aerobic Biodegradation of Polychlorinated Biphenyls: Genetic
and Soil Studies 97
Genotoxicity Assays of Metabolites from Biological Treatment Processes 101
Results of the Bioremediation Field Initiative 105
Oil Spill Bioemediation Project 107
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BIOVENTING OF AN AVIATION GASOLINE SPILL:
DESIGN AND OPERATION OF A FIELD DEMONSTRATION
By: John M. Armstrong, Ph.D.
Christopher J. Griffin, P.E.
The Traverse Group, Inc.
Ann Arbor, Michigan
The Bioventing reclamation pilot study is designed to evaluate the biodegradation ol
hydrocarbon contaminated vapors within the unsaturated zone during induced volatilization.
This study is being conducted at the U.S. Coast Guard Air Station in Traverse City,
Michigan, which is the site of a spill of about 35,000 gallons of aviation gasoline which
occurred in 1969. After 20 years a major portion of the spill still persists in the subsurface
as a plume which is about 1,100 feet long and 250 feet wide. This study is being conducted
as a cooperative effort between the United States Coast Guard and the U.S. Environmental
Protection Agency's Robert S. Kerr Groundwater Research Lab.
The subsurface conditions at the site consist of a uniform beach sand extending to
depths of about 50 feet, underlain by a gray glacial silty clay. The water table is located at
a nominal depth of about 15 feet below the ground surface but over the past six years the
water table elevation has fluctuated six to eight feet.
The 90 by 75-l'eet study area has been divided into two equal areas of 45 by 75 feet
to evaluate the effects of different flows and extraction patterns. The northern area lias an
injection system while the southern area has an injection and extraction/reinjection system.
The pneumatic properties of the unsaturated zone were evaluated by the performance ol
a pneumatic pump test, resulting in a design radius of influence of ten feet. The work plan
calls for ambient air to be injected into both areas at an initial rate which replaces the
calculated volume of air-filled pore space over 24 hours. The flow rate will be increased to
a vapor recharge rate of eight hours or less as the system becomes acclimated.
The blower package, therefore, has to be capable of extracting vapors in the south
study area, at depths of 15 to 18 feel (depth of the water table) and flow rates ranging from
5 to (S3 cubic feet per minute (cfm), then reinjecting the vapors at the same rate, at a depth
of 10 feet. Additionally, the system has to be able to inject ambient air at the same flow
rate within both the extraction/reinjection plot (south area) and the air injection plot (north
area). Accordingly, because the ambient air injected will be placed in twice the area (two
test plots), the blower has to be able to inject air at flow rates ranging from 10 cfm to I2K
cfm.
The construction of the Bioventing project consisted of installingjin the north area
15 aeration injectior points placed on 10-foot centers in a three-by-five grid and screened
just above the water table. In the south area, eight sets of injection points coupled with
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Biovcnting of an Aviation Gas Spill:
Design and Operation of a Field Demonstration
seven extraction points, 10 feet on center, were installed with screens placed just above the
water table. Eight reinjection wells were installed with the screens placed at a depth of H)
feel.
The blower package used is a 45 URAI Roots vacuum pump with a maximum flow
rate of 130 cfm. This system extracts vapors at a vacuum of 4 to 6 inches of mercury and
reinjects the vapors at a pressure of 6 pounds per squre inch (psi). The vacuum pump is
driven by a 10 HP 3-phase electric motor. Similarly, an additional 45 URAI Roots pump,
also with a maximum tlow rate of 130 cfm, is used for the ambient air injection. This blower
is driven by a 7.5 HP 3-phase electric motor at an operaling pressure of 6 psi. All the
equipment is explosion-proof.
The monitoring requirements of the EPA Work Plan called for the installation ol
several different types and depths of monitoring equipment and/or sample points. To
monitor vapor hydrocarbon and oxygen concentrations, six 5-point cluster wells were
installed with three cluster wells per plot. The cluster wells consisted of 1/4-inch diameter
tubing with a wire mesh screen covering the tip. The five points of each cluster well were
installed at 3.28 feet (1 meter) depth increments throughout the unsaturated zone.
Additionally, we installed three 14-point cluster monitoring wells (well screens at 1.5 foot
intervals from ground surface to 21 feet - one per plot and one at an upgradiem locution)
and one set of moisture/temperature probes per plot were installed. The
moisture/temperature probes are Soil Test Series 300 moisture/temperature cells consisting
of thermistor soil cells buried at depths of 5, 10 and 15 feet below grade.
The development of a sufficient microbial population to degrade the hydrocarbon
vapors requires adequate quantities of nitrogen, phosphorous and potassium. The EPA
Bioventing work plan called for an initial application of these nutrients consisting of M
pounds of nitrogen, 13 pounds of phosphorus and 5 pounds of potassium to be applied to
each area prior to start-up. Additionally, during the growing season, 10 pounds of nitrogen,
2 pounds of phosphorous and 1 pound of potassium was to be applied to each area monthly.
These nutrients were applied in an aqueous solution by sprinklers until they were detected
in the ground water, indicating they had moved completely through the treatment zone.
The Biovent Project is sampled and/or monitored daily, biweekly and monthly. Daily
monitoring consists of measuring the blower's operating parameters such as How rate,
pressure, and vapor temperature. Combustible gas concentration within the vapor
reinjection flow line is determined daily utilizing a Bacharach Treshold Limit Value (TLV)
combustible gas meter.
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Biovcnting of an Aviation Gas Spill:
Design and Operation of a Field Demonstration
Biweekly monitoring includes determining the combustible gas and oxygen
concentration within the three 5-point cluster wells located in each plot. The combustible
gas concentration is determined using the TLV gas meter and the oxygen concentration is
determined using a Bacharach Oxygen indicator. Additionally the soil moisture content and
soil temperature is measured biweekly in the moisture/temperature probes.
The surface emissions are sampled at two locations within each of the study areas
biweekly and two upj^radient locations weekly. The samples, taken over a four-hour period,
are pulled using an Ismatec peristaltic pump set for a flow of approximately one liter per
hour. A 19-inch diameter stainless steel bowl which has a volume of 4.3 gallons (16 liters),
is inverted and placed flush on the ground. The sample is pulled from the bowl through
flexible vinyl tubing that is attached to the bowl by a 1/4-inch diameter steel ball valve
tapped into the bottom of the bowl. Any water which collects within the emission chamber
is removed by a water trap, located upstream from the sample trap, consisting of a flask
containing a drying agent (Drierite).
Water quality data is obtained by sampling at two depths in each of the 14-point
monitoring wells. The water samples are analyzed for nutrients and BTEX.
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BIOVENTING OF AN AVIATION GASOLINE SPILL:
PERFORMANCE EVALUATION OF A FIELD DEMONSTRATION
by: Don Kampbell
U.S. EPA
Ada, OK 74820
ABSTRACT
A spill of about 35,000 gallons of aviation gasoline
occurred in 1969 at a U.S. Coast Guard air station. Much of the
spill persists after 22 years as an oily phase residue at the
water table near a depth of five meters. The subsurface matrix
is a fairly uniform beach sand to 15 meters.
Aerobic soil microcosms were used in the laboratory to
simulate the ability of the spill site soil to biodegrade
aviation gasoline vapors. Reaction rates were rapid using
acclimated microcosms with disappearance curves showing typical
first-order kinetics. Degradation rates within a temperature
range of 12 to 23°C were high. A nutrient addition of ammonia,
nitrate, phosphorus, and potassium increased by several fold
bacteria count, degradation rate, and active biomass. A
suppressive effect was shown on degradation of dimethyl and
trimethyl pentanes mixtures when compared to singular components.
A Lineweaver-Burk reciprocal plot was used to calculate a
biochemical reaction kinetic maximum velocity value of 5.7g
fuel/kg soil-hour and half saturation constant of 7.0g fuel/kg
soil for aviation gasoline vapors. Extrapolation of the rates to
field conditions suggest consumption of the gasoline vapors
during bioventing within eight hours in the unsaturated zone.
The pilot demonstration systems have been operational
slightly more than three months. One area has injected aeration
only. The second area has both injection , extraction and
reinjection. Turf was established to cover both treatment areas.
A nutrient solution of nitrogen and phosphorus was dispersed
throughout the unsaturated zone. Subsurface flow characteristics
will be defined with a sulphur hexafluoride tracer test. Core
material, soil gas, and underground water are being analyzed to
determine the extent of remediation. Core material fuel
hydrocarbons have been reduced about 40% in both treatment areas
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during the first three months of operation. Surface emissions
have been less than one percent of total volatile hydrocarbons
detected in the soil gas aeration stream at a one meter depth.
Objectives of the project will be to demonstrate:
- Remediation will be completed in a reasonable
time.
- Surface emissions of gasoline do not occur.
- Remediated core material will be < 10 mg fuel
carbon/kg.
- Final benzene levels in the groundwater will not
exceed 5 /ig/liter.
- Performance and economical advantages will be
applicable to full-scale remediation.
The project described in this presentation is jointly funded
by the U.S. Environmental Protection Agency and the U.S. Coast
Guard. The work described is in the initial phase of data
compilation. Therefore the contents have not been subjected to
the Agency's review policy and no official endorsement should be
inferred.
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LABORATORY AND FIELD STUDIES OF THE KINETICS OF BIOVENTING
by: David W. Ostendorf, Associate Professor
Civil Engineering Department, University of Massachusetts
Amherst, MA 01003
Don H. Kampbell, Research Chemist
RS Kerr Environmental Research Laboratory, USEPA
Ada, OK 74820
Ellen E. Moyer, Research Assistant
Civil Engineering Department, University of Massachusetts
Amherst, MA 01003
ABSTRACT
The interaction of laboratory and field investigations of biodegraded
hydrocarbon vapors are discussed under naturally diffusive and engineered
advective conditions.
The coupled transport of aviation gasoline and oxygen vapors has been
measured and modeled at the US Coast Guard Air Station in Traverse City,
Michigan as part of a series of research projects managed by the RS Kerr
Environmental Research Laboratory of the USEPA. The data consist of a
group of stainless steel tubing clusters set at one meter depth increments
over the 5 m thick unsaturated zone in the uniform sand at the site. The
clusters have been sampled over a 13 month period with hydrocarbon and
oxygen meters, calibrated against known headspace standard gases. The model
is a time averaged steady balance of diffusion and M i chae 1 i s - Me n t on
kinetics, coupled stoichiometrically under the assumption of abundant
oxygen. An implicit solution is put forth by Ostendorf and Kampbell (1991)
z " (diffusion) (1)
with elevation z above the contaminated capillary fringe, soil moisture dif-
fusivity D, half saturation constant K, maximum reaction rate V, and
hydrocarbon concentration H. The integral function I(H/K) is evaluated by
Ostendorf and Kampbell (1991), and the corresponding oxygen concentration
follows from the stoichiometry of the reaction. This natural diffusion
model is calibrated with the field soil gas data at four clusters, yielding
the following kinetics
Unacclimated sandy soil
V - 8.6x10*^ kg/m^-s
K - 0.10 mg/1
Bioventing is the engineered advection of air blown through con-
taminated soil, subject to aerobic biodegradation. Ostendorf and Kampbell
(1990) model this process as a balance of advection and Michael is-Menton
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kinetics, yielding a simple prediction for the effluent hydrocarbon vapor
concentration H_
E
tpv H H -He
— - ln(—)+—~— (advection) (2)
E
with pneumatic residence time tp and influent concentration Hj . Thus the
reaction kinetics play a major role in determining the removal efficiency of
the bioventing reactor. The foregoing theory describes field data [Kampbell
et al. (1987)] from an acclimated clay soil bioreactor treating a
propane/butane waste gas mixture at Racine, Wisconsin.
Ostendorf and Kampbell (1990) apply laboratory microcosm data to the
diffusion and advection transport models cited above. Aseptic soil samples
from the Wisconsin and Michigan sites were dosed with appropriate gases in
headspace vials, and the subsequent decay of concentration was measured in a
gas chromatograph (subject to abiotic control). The temporal decay of the
initial concentration Hq is described by a balance of storage, headspace
sorption, and Michaelis-Menton kinetics
V H0 Ho"H
t - -^-[ln(—)+——] (microcosm) (3)
with retardation factor R^. The microcosm based kinetics for Traverse City
agree fairly well with the field calibrated values listed above. The
Wisconsin soil kinetics, obtained from the microcosms, are
Acclimated clay soil
V - 4.1x10*^ kg/m^-s
K - 0.088 mg/1
These values, when substituted into the bioventing advection model (2) yield
a 92% predicted removal rate that compares favorably with the observed range
of 90-99% [Kampbell et al. (1987)]. We conclude that laboratory microcosms
yield kinetics that are consistent with field values, implying no transport
limitations on the field or microcosm scale in unsaturated soil.
We note the close correspondence of half saturation constants for two
dramatically different soil types. The wide range of maximum reaction rates
may be due to biomass variation in clays and sands, and also suggests that
biostimulation may be possible in natural soils. Ostendorf and Kampbell
(1990) explored the latter possibility by running a series of microcosm
studies for stimulated soil samples at Traverse City with the results
Acclimated sand with nutrients
V - 2.8x10 ^ kg/m^-s
K - 0.25 mg/1
These kinetics are input to a series of simulations of bioventing effective-
ness at the site. The most optimal conditions, consisting of a high
concentration influent and a biostimulated soil, result in very effective
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bioventing, with complete removal of product in about one month of system
operation. Natural kinetics are too slow for effective bioventing at
Traverse City, and an auxilliary exhaust air treatment process would have to
be considered at the site.
REFERENCES
Kampbell, D.H., Wilson, J.T., Read, H.W., and Stocksdale, T.T. (1987),
"Removal of Volatile Aliphatic Hydrocarbons in a Soil Bioreactor, " Journal
Air Pollution Control Association. 37: 1236.
Ostendorf, D.W. and Kampbell, D.H. (1990), "Bioremediated Soil Venting of
Light Hydrocarbons," Hazardous Waste and Hazardous Materials. ]_\ 319.
Ostendorf, D.W. and Kampbell, D.H. (1991), "Biodegradation of Hydrocarbons
in the Unsaturated Zone," Water Resources Research, in press.
This paper has been reviewed in accordance with
the U.S. Environmental Protection Agency's peer
and administrative review policies and approved
for presentation and publication.
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TREATMENT OF WASTEWATER WITH THE WHITE ROT
FUNGUS PHANEROCHAETE CHRYS0SP0R1UM
THOMAS W. JOYCE and HOU-MIN CHANG
Department of Wood and Paper Science
North Carolina State University
Raleigh NC 27695
The purpose of this summary is to provide an overview of the use of the
white rot fungus Phanerochaeie chrysosporium in the treatment of aqueous
wastewaters with emphasis on wastewaters from paper manufacture bleach
plants.
White rot fungi, and in particular P. chrysosporium, are able to degrade
a wide variety of organic pollutants (for example, 1, 2, 3). Those studies were,
however, extensions of research which found that the fungus was able to
decolorize the bleach plant effluent resulting from the manufacture of white
paper (4). As the chromophobe material in the effluent was destroyed, it was
found that the organic chlorine containing compounds were also degraded (5).
The degradation is believed to occur as the result of secondary metabolic
activity, i.e., the fungus cannot utilize the substrate for carbon or energy
purposes (6). The degradation of pollutants is now known to be mediated by a
family of enzymes excreted by the enzymes (7).
The initial treatability studies on bleach plant effluent were done using
125 mL shake flasks. After evaluation of numerous reactor designs, it was
determined that a rotating biological contactor (RBC) was most suited for the
physiological needs of the fungus. Our standard laboratory-scale reactor has a
capacity of about 2.1 L. It contains 8 partially wetted (40%) disks rotating at 1
rpm (8). It has also been found that a higher-than-atmospheric oxygen tension
promotes organic pollutant destruction. Typically, the fungus is grown at 37 °C
and, after substantial mycelial growth, the temperature is lowered to 30 °C for
enzyme production (9). The pH is maintained in the range of 3 to 5. After 2-4
days of fungal growth, the nitrogen is removed from the wastewater and the
fungus soon enters secondary metabolism when the enzymes actually responsible
for organic pollutant destruction are excreted. Other researchers have found
that the fungus can be immobilized on porous plastic (10) or in alginate (11), or
that the enzymes themselves can be immobilized and remain active (12).
In respect to bleach plant effluents, the use of P. chrysosporium may be
particularly advantageous. Not only does the fungus reduce the color of the
effluent, but it also can dechlorinate the organic chlorine compounds as
measured by AOX. At present, typical biological treatment systems, aerated
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lagoons and activated sludge, do not remove substantial amounts of color or
AOX from the effluent. The immobilized fungi can remove some 70% of the
color and about 50% of the AOX. We have found that the high molecular
weight organic compounds in bleach plant effluent can be transformed to lower
molecular weight compounds which exhibit more toxicity as measured by the
MICROTOX bioassay (13).
We believe the most logical use for the immobilized fungus may be in a
two stage system whereby the fungus is used as a pretreatment prior to
conventional secondary treatment (14). The fungus would thus be used to
dechlorinate the AOX compounds and reduce the average molecular weight of
the remaining lignin compounds. The conventional secondary treatment could
then remove a significant portion of the remaining organic compounds.
In some cases it may be economically advantageous to concentrate
effluent to be treated by the fungus by ultrafiltration. We have found that the
rate of color removal is generally proportional to color concentration. In one
example case, the cost of fungal treatment for a bleach plant effluent was
$12.37/ton pulp; when ultrafiltration was added to the treatment flowsheet, the
cost decreased to $4.46/ton pulp despite the added cost of the UF system (15).
To date the immobilized fungus has not been tested at even pilot plant
scale. We believe, however, that it can be an effectively integrated into an
existing treatment scheme to pretreat a wide variety of wastewaters to make
them more amenable to conventional treatment.
1. Bumpus, J. et al. 1985. Science, 228: 1434.
2. Messner, K. et al. 1988. Forum Mikrobiologie, 11: 492.
3. Pellinen, J. et al. 1988. J. Biotech., 8: 67.
4. Eaton, D., et al. 1982. TAPPI J., 65: 89.
5. Huynh, B. etal. 1985. TAPPI J., 68: 98.
6. Kirk, T. and Farrell, R. 1987. Ann. Rev. Microbiol., 41: 465.
7. Kirk, T. et al. 1986. Enzyme Microb. Technol., 8: 27.
8. Chang, H-m. etal. 1985. U.S. Patent 4,554,075.
9. Asther, M. etal. 1988. Appl. Environ. Microbiol., 54: 3194.
10. Messner, K. et al. 1988. S. Eur. Patent Al-0 286 630.
11. Livernoche, D. et al. 1983. Biotech. Bioeng., 25: 2055.
12. Presnell, T. et al. 1990. Proc. Cellucon '90, Bratislava, Czechoslovakia,
28-31 August.
13. Fukui, H. etal. 1991. Proc. Symp. Cellulose and Lignocellulosic Chem.,
Guangzhou, China, 13-15 May.
14. Yin, C-f. etal. 1989. J. Biotech., 10:77.
15. Yin, C-f. et al. 1990. Proc. 24th EUCEPA Conf., Stockholm, Sweden,
8-11 May.
-12-
-------
AN AEROBIC/AEROBIC SEQUENTIAL TREATMENT OF CERCLA LEACHATES
IN POTW's: AN INNOVATIVE TREATMENT APPROACH
Margaret J. Kupferle
Paul L. Bishop
University of Cincinnati
Cincinnati, Ohio
Dolloff F. Bishop
Steven I. Safferman
U.S. Environmental Protection Agency
Risk Reduction Engineering Laboratory
Cincinnati, Ohio
The presented innovative approach to anaerobic/aerobic sequential treatment of
hazardous waste leachates in a POTW is proposed to migitate some of the problems
with conventional aerobic treatment of these wastes, i.e., pass-through of toxics, air
stripping of volatiles and lack of sufficient anaerobic contact for dechlorination of toxics.
Figure 1 is a conceptual schematic of the process. A contact/ sorption stage consisting
of a granular activated carbon (GAC) expanded bed which would be placed between
the primary clarifier and the aeration basin of the POTW is designed to reduce
pass-through of toxics, retaining them for treatment in an off-line reactor. The
exhausted GAC from the contact/sorption stage is exchanged with "bioregenerated"
GAC from an off-line anaerobic stabilization stage, conserving the GAC in the system.
In order to reduce heat losses incurred during the exchange of GAC between the
heated stabilization stage and the ambient temperature contact/sorption stage, the
respective GAC streams are separated from their supernatants and the supernatants
are recycled. This minimizes the loss of heated supernatant in the second stage. In
addition, methane gas, a recoverable source of energy, is generated in the second
stage. The long overall retention time of the biomass and GAC in this system
encourages dechlorination and destruction of recalcitrant toxics trapped and removed
in the first stage.
Two 87 L/day bench-scale systems are in operation for proof-of-concept studies.
Each system has two stages. The first stage is operated as a contact/sorption unit and
the second stage is operated as a stabilization unit. The carbon retention time (CRT) in
-13-
-------
the first stage is 2 days with a hydraulic retention time (HRT) of 30 minutes. The target
CRT in the second stage is 15 days. One system (control) is treating primary effluent
only and one system (test) is treating primary effluent spiked with 5% landfill leachate
and a mixture of nine volatile and five semivolatile hazardous organic compounds.
Concentrations of the spiked organics as well as conventional wastewater treatment
parameters such as chemical oxygen demand (COD) are routinely monitored. Results
to date indicate little or no pass-through to the aeration basin of most of the spiked
organics, and average COD removals in the 40% to 50% range.
-14-
-------
Primary
Effluent/
Leachate
+ Toxics
¦o
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0)
CD
O
3)
O
CD
To Aeration
Basin .
Sss#
mum:
Ttstittk
CONT ACT/SORPTION
STAGE
* On-line expanded bed
* Ambient temperature
<_>
<
o
TD
CD
—>
CD
O)
c
O)
CJ)
0)
Q£
c
3
—>
d)
CkL
Make-up
Carbon
Pollutant-Laden GAC
a)
o
o
y,
>•/
Mix pollutant-
laden GAC
with 35 °C
supernatant
Separate
GAC from
supernatant
Mix regenerated
GAC with ambient
temperature
supernatant
KEY
Gas
Liquids
GAC/Liquid
Slurries
Semi-Continuous
Flow
SEMI-CONTINUOUS
CARBON EXCHANGE
FIGURE 1
CONCEPTUAL SCHEMATIC OF PROPOSED SYSTEM
-15-
-------
-16-
-------
Manipulation of TCE-degradative genes of Pseudomonas
cepacia
Malcolm Shields
TRI/US EPA, Gulf Breeze, FL
SUMMARY
Pseudomonas cepacia G4, like several other toluene-utilizing
bacteria, can degrade trichloroethylene (TCE) only when the
requisite oxygenase is expressed in response to aromatic inducers.
A major limitation to the application of these organisms for TCE
bioremediation is their requirement for exogenous inducers. A
mutant of G4 (G4 Phel) was selected for its ability to degrade TCE,
1,1-dichloroethylene, cis- and trans-l.2-dichloroethvlene without
requiring induction. Experiments to determine the genetic basis for
this altered regulation indicated that enzyme involved in the
conversions of cresols to ring cleavage products were
constitutively expressed. Genetic stability of the strain was
assessed following 100 generations of non-selective growth. The
constitutive phenotype was completely stable under these
conditions. G4 Phel was anticipated to be highly sensitive to
chlorinated aromatics as a result of the constitutive aromatic
meta-fission pathway. This anticipated problem was partially
ameliorated through the introduction of a derivative of the pJP4
2,4D-degradative plasmid, pROlOl (Tnl721:pJP4), which encodes an
ortho-fission chlorocatechol pathway.
RESULTS AND DISCUSSION
Pseudomonas cepacia strain G4 possesses a novel pathway of
toluene catabolism (1). Mutants unable to hydroxylate toluene, o-
cresol, m-cresol and phenol, were also shown to be incapable of the
degradation of TCE, suggesting that an ortho-acting toluene
monooxygenase may be responsible for its degradation (2) . Pursuant
to a more complete genetic description several mutant classes were
analyzed. A variant was produced from one class lacking detectable
toluene monooxygenase activity, that constitutively degraded TCE,
1,l-dichloroethylene, cis- and trans-l.2-dichloroethvlene (Table
1). G4 and its constitutive derivative G4 Phel, were examined for
their ability to convert 3-trifluoromethyl phenol (TFMP) to 2-
hydroxy-7,7,7-trifluoroheptadienoic acid (TFHA) during growth on
lactate. This chromogenic reaction requires two enzymes: a
toluene/cresol monooxygenase and a catechol-2,3-dioxygenase (2)
(Fig. 1). G4 clearly exhibited an inducible response in its rate
of production of TFHA (measured by absorbance at 385 nm). G4
Phel, however, lacked any detectable inductive response,
demonstrating instead a fairly consistent level of activity
throughout growth. The high specific activity of the second enzyme
of TFMP metabolism, Catechol-2, 3-dioxygenase (C230) , was determined
under both induced and non-induced conditions. Table 2 clearly
indicates that C230 was constitutive in G4 Phel. In addition the
-17-
-------
evidence suggests that C230 in G4 is under the control of more than
one promoter/operator.
Stability of the constitutive property of G4 Phel was assessed
after growth under non-selective conditions (i.e. basal salts
medium with sodium lactate at 20 mM as the sole carbon source) and
serial dilution (allowing ca. 10 generations per transfer) for 100
generations. All independent isolates examined (>1000) were found
to have maintained constitutivity as determined by their ability
to immediately metabolize TFMP to TFHA.
Since chloroaromatics are frequent co-contaminants with TCE
at waste sites the constitutive aromatic monooxygenase of G4 Phel
might cause serious problems since it can produce toxic
chlorocatechols from these chloroaromatics. G4 Phel metabolized
chlorobenzene to 2-chlorophenol and then to 3-chlorocatechol (not
further metabolized by the C230 of G4) . Plasmid pROlOl
(pJP4::Tnl721), which encodes a chlorocatechol-1,2-dioxygenase
(that accepts both mono- and di-chlorinated catechols) was
introduced into this strain. G4 Phel (pROlOl) utilized 2,4-
dichlorophenoxyacetic acid as a sole carbon source, was resistant
to phenylmercuric acetate, and did not accumulate 3-
chlorocatechol. Concentrations of 2-chlorophenol and chlorobenzene
that interfered with the degradation of TCE by G4 Phel (pROlOl)
(100/xM) were ten fold higher that those with an inhibitory effect
on the plasmid-free strain G4 Phel (1-10 /xM) .
REFERENCES
1 Shields, M. S., S. 0. Montgomery, P. J. Chapman, S. M. Cuskey,
and P. H. Pritchard. 1989. Appl. Environ. Microbiol. 55:1624-
1629.
2 Shields, M. S., S. 0, Montgomery, S. M. Cuskey, P. J. Chapman,
and P. H. Pritchard. 1991. Appl. Environ. Microbiol.
Submitted.
TFHA Production by
I I I P
200 400 600
Minutes
l
i
©
o
<
<
X
ih
H
c
X eH
o
u
a
E
V
o 2H
E
E
TFHA Production by
G4 5223 Phel
too
1 I
200 300
Minutes
500
Figure 1. Rates of conversion of TFMP to TFHA during growth.
-IS-
-------
Table l. Action of P. cepacia G4 Phe(l) on chlorinated ethylenes.
% Chloroethylene Remaining8
Strain
1.1-DCE
cis-
1.2-DCE
trans-
1.2-DCE
TCE
PCE
Uninoculated
100
100
100
100
100
± 2
± 4
± 9
± 3
± 7
G4 Uninduced
104
69
107
133
103
± 3
± 19
± 5
± 5
± 7
Phe(l)
50
12*
o"
2
104
± 3
± 9
± 2
± 3
8 Substrate remaining expressed as percentage of that determined in
uninoculated controls. Abbreviations: 1,1-DCE, 1,1-
Dichloroethylene; cis-1.2-DCE, cis- 1,2-Dichloroethylene, trans-
1,2-DCE, trans-1,2-Dichloroethylene; TCE( Trichloroethylene; PCE,
perchloroethylene; M, Metabolite detected by GC.
Table 2 Differences in specific activity of catechol-2,3-
dioxygenase in wild type and mutants of G4.
Catechol-2,3-dioxygenase
nmoles min'1 mg protein'1
Substrate
Cat8 3mCat
Strain
Inducer
G4
none
2.1
3.7
G4
phenol
53.6
62.5
G4 5223
none
0. 07
2.2
G4 5223
phenol
13.1
31.4
Phe(l)
none
156. 0
50.4
Phe(l)
phenol
48.0
34.9
8 Cat, catechol; 3raCat, 3-methylcatechol.
-19-
-------
-20-
-------
Treatment of TCE and degradation products using Pseudomonas
cepacia
Malcolm Shields
TRI/US EPA, Gulf Breeze, FL
SUMMARY
The constitutive trichloroethylene (TCE) -degrading Pseudomonas
cepacia strain G4 Phel was found to be very stable, with no
detectable loss of activity (or marker) after 100 generations of
non-selective growth. Toluene monooxygenase activity towards
trifluoromethyl phenol was constant throughout the growth curve as
measured by the rate of production of 2-hydroxy-7,7,7-
trifluoroheptadienoic acid. TCE degradation rates measured for G4
Phel (0.5 - 1 nmole TCE min"1 mg'1 protein) are approximately 15
to 30% of the maximal wild type G4 phenol induced activity. G4
Phel was found to degrade TCE over a fairly wide range of physical
conditions: 4 to 3 0 °C, pH 4 to 9, 3 to 30 mg l"1 oxygen, and 0 to
20 0/00 salinity. A slight lowering of growth rate in G4 Phel was
observed in the presence of TCE at an aqueous concentration of 53 0
/iM. Chlorobenzene and 2-chlorophenol were found to be inhibitory
to TCE degradation at 1-10 /iM. Following introduction of pROlOl
(a transposon derivative of pJP4) 2-chlorophenol and chlorobenzene
levels inhibitory to TCE degradation were increased to IOOjiM.
RESULTS AMD DISCUSSION
Groundwater contamination by organic pollutants is a subject
of overwhelming concern throughout the industrialized world. Chief
among these pollutants are those categorized as volatile organics.
These include the chloroethylenes: trichloroethylene (TCE),
tetrachloroethylene, trans-1.2-dichloroethylene (DCE), 1,1-DCE, and
vinyl chloride (ranked first, second, third, fifth and more than
tenth respectively of all volatiles detected as groundwater
contaminants in the United States) (3) . The constitutive TCE
degrader, Pseudomonas cepacia G4 Phel, that employs a unique
toluene monooxygenase (4) for the degradation of TCE (5) , was
investigated for its bioremediation potential.
Rates of TCE degradation were determined using a glass syringe
with a Teflon plunger and no air headspace as the bioreaction
chamber as previously described (1). This assay was adopted for
rate analysis because it more closely resembles a contaminated
aquifer and multiple samples can be taken without the introduction
of an air headspace. The range of certain physical variables
(temperature, oxygen concentration, pH, and salinity) likely to
affect the rate of TCE degradation by P. cepacia G4 Phel under
environmental conditions was investigated using this technique.
The most profound effect on the rate of TCE degradation was
found at the lowest temperatures used (Fig. la) where despite
cooling to 4°C, G4 Phel maintained approximately 30% of the rate
of TCE degradation measured at 30°C. TCE degradation rates of
approximately 20 and 4 5% (relative to the maximal pH 7 value) were
-21-
-------
evident at pH extremes of 4 and 9 respectively (Fig lb). Little
effect on TCE degradation rates was observed over the ranges of
Oxygen concentration (2.8 - 31.3 mg L*1) (Fig. lc) or salinity (0 -
20 0/00) tested (Fig. Id).
Possible toxicity of a metabolite of TCE towards G4 Phel was
investigated. G4 does not affect TCE in the absence of an aromatic
inducer (2). Therefore, any toxicity in the presence of TCE would
be due to direct toxicity. G4 Phel differs only in one
significant respect with G4 (i.e. the constitutive expression of
the toluene and TCE degradative enzymes). Any toxic effects beyond
those seen with G4 would therefore be attributable to the active
metabolism of TCE. A slight toxic effect was measured as a
depression of growth rate for G4 Phel on glucose and yeast extract
as a result of the metabolism of TCE at levels c.a. 2mM as if all
were in aqueous solution (measured at 530 /iM) .
REFERENCES
1 Folsom, B. R., P. J. Chapman, and P. H. Pritchard. 1990. Appl.
Environ. Microbiol. 56:1279-1285.
2 Nelson, M. J. K., S. O. Montgomery, W. R. Mahaffey, and P. H.
Pritchard. 1987. Appl. Environ. Microbiol. 53:949-954.
3 Rajagopal, R. 1986. Environ. Prof. 8:244-264.
4 Shields, M. S., S. 0. Montgomery, P. J. Chapman, S. M. Cuskey,
and P. H. Pritchard. 1989. Appl. Environ. Microbiol. 55:1624-
1629.
5 Shields, M. S., S. 0. Montgomery, S. M. Cuskey, P. J. Chapman,
and P. H. Pritchard. 1991. Appl. Environ. Microbiol.
Submitted.
-22-
-------
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Fig 1. Effects of varying temperature, pH, dissolved oxygen and
salinity on the rate of TCE degradation by Pseudomonas
cepacia G4 Phel.
-23-
-------
-24
-------
Treatment of CERCLA Leachates by Carbon-Assisted
Anaerobic Fluidized Beds
M. T. Suidan, R. Nath, and A. T. Schroeder
Department of Civil and Environmental Engineering
University of Cincinnati
Cincinnati, OH 45221-0071
E. R. Krishnan
PEI Associates, Inc.
Cincinnati, OH 45204
R. C. Brenner
U.S. Environmental Protection Agency
Cincinnati, OH 45268
Rainfall and surface runoff percolating through a landfill are contaminated
with a number of organic and inorganic compounds, and direct discharge of exiting
leachate to municipal wastewater treatment plants can result in inadequate
removal of many hazardous substances. Many volatile and semivolatile synthetic
organic chemicals (SOCs) used as solvents, degreasers, and components in
industrial products are present in leachates originating from hazardous waste
sites. These SOCs are often inadequately treated in aerobic wastewater treatment
processes as volatiles are subject to air stripping, many semivolatiles simply
pass through untreated, and highly chlorinated compounds are difficult to degrade
aerobically.
In this study, two anaerobic granular activated carbon (GAC) expanded-bed
bioreactors were tested as pretreatment units for the decontamination of
hazardous leachates. Two municipal leachates, rendered hazardous with the
addition of ten volatile and four semivolatile organic compounds commonly found
in leachates, were fed to two identical bench-scale (10.2 cm diameter x 96.5 cm
high) expanded-bed reactors. One leachate, with a COD of approximately 1,100
mg/L, was representative of those that emanate from old, stabilized waste
-25-
-------
landfill sices where only a small fraction of the leachate COD is made up of
volatile fatty acids. Sulfate concentrations in this leachate averaged 89 mg
SOt/L. The second leachate, with a COD of approximately 3,800 mg/L, was typical
of younger, more active landfill sites, where the majority of the COD is
attributable to volatile fatty acids. This leachate had a very low sulfate
content. The different characteristics of the two leachate feed streams resulted
in one reactor operating in a sulfate-reducing mode and the second in a strictly
methanogenic environment. Both reactors were operated with a 6-hr unexpanded
empty-bed contact time and achieved SOC removals acceptable for pretreatmenc
units. In both reactors, the majority of the SOCs were removed by biological
activity, with GAC adsorption providing stability to each system during startup
and by buffering against load fluctuations during long-term operation.
The SOCs and their respective concentrations added to the two leachates are
summarized in Table 1. During the first phase of the study, chloroform vau
deleted from the SOC supplement because of toxicity problems associated with this
volatile organic compound. The effect of chloroform on the performance of the
treatment systems was evaluated during the second phase of the study.
Phase One (No Chloroform Addition)
The first phase of the study extended over a period of 400 days during
which both reactor systems exhibited excellent removal efficiencies for all the
SOCs listediin Table 1. A summary of the performance of the two systems averaged
over the last 260 days of operation is given in Table 2. A comparison of the two
systems indicates that the sulfate reducing environment may produce equal or
better performance than a methanogenic environment in removing a consortium of
hazardous chemicals from waste streams. All three volatile aromatic compounds
(toluene, chlorobenzene, and ethylbenzene) in the SOC consortium were removed at
higher rates in the sulfate reducing environment. Also, the persistence of the
-26-
-------
intermediate biodegradation product of nitrobenzene, aniline, was only
significant in the methanogenic reactor.
Table 1: Composition of SOC Supplement Added to the Leachates
Compound Concentration (ug/L)
VOLATILE ORGANIC COMPOUNDS
Acetone 10,000
Methyl Ethyl Ketone 5,000
Methyl Isobutyl Ketone 1,000
Trichloroethylene 400
1,1-Dichloroethane 100
Methylene Chloride 1,200
Chloroform 0 to 3,500
Chlorobenzene 1,100
Ethylbenzene 600
Toluene 8,000
SEMIVOLATILE ORGANIC COMPOUNDS
Phenol 2,600
Nitrobenzene 500
1,2,4-Trichlorobenzene 200
Dibutyl Phthalate 200
Table 2: Summary of SOC Concentrations in Reactor Effluents
Effluent Sulfate Effluent Methanogenic
Compound Reducing Reactor Reactor
Acetone
189*
(216)*
410'
'(577)
Methyl Ethyl Ketone
70
(68)
220
(150)
Methyl Isobutyl Ketone
35
(16)
58
(25)
Trichloroethylene
3
(10)
5
(3)
Methylene Chloride
65
(50)
46
(44)
1,1-Dichloroethane
20
(17)
14
(8)
Chlorobenzene
67
(42)
165
(86)
Ethylbenzene
34
(19)
85
(41)
Toluene
436
(303)
1,102
(744)
Phenol
22
(23)
93
(63)
Nitrobenzene
6
(16)
8
(17)
1,2,4-Trichlorobenzene
10
(15)
14
(16)
Dibutyl Phthalate
26
(29)
36
(30)
Ail concantrations in ut/L.
* Standard deviation.
-27-
-------
Phase Two (Chloroform Addition)
During the second phase of the study, chloroform was added to the SOC
supplement at an initial feed concentration of 2 mg/L. This feed concentration
was maintained for a period of 105 days. During this period, chloroform was not
observed to exert an adverse effect on the performance of either the sulfate
reducing or the methanogenic reactors. Encouraged by the ability of both
reactors to handle this feed level of chloroform, the influent chloroform
concentration was raised to 3.5 mg/L. The methanogenic reactor responded with
a rapid decline in methane production and COD reduction and correspondingly
increases in effluent SOC concentrations. The sulfate reducing reactor, on the
other hand, was able to withstand this increased feed concentration of chloroform
and continued to maintain excellent removal of all the SOCs for a period of 156
days. It is presently planned to continue increasing the feed concentration of
chloroform to the sulfate reducing reactor. Recovery of the methanogenic reactor
was rapidly achieved after chloroform addition was terminated.
-28-
-------
Improved Prediction of GAC Capacity in a Biologically Active Fluidized Bed
M. T. Suidan and R. D. Vidic
Department of Civil and Environmental Engineering
University of Cincinnati
Cincinnati, OH 45221-0071
R. C. Brenner
U.S. Environmental Protection Agency
Cincinnati, OH 45268
The first applications of activated carbon for water quality control involved the
addition of powdered activated carbon (PAC) to drinking water treatment plants for taste
and odor control. In recent years, granular activated carbon (GAC) has received increasing
attention for removing synthetic organic compounds from water. Moreover, the elimination
of a sludge disposal problem and the possibility of GAC reuse have accentuated the
applicability of granular activated carbon to drinking water treatment.
Interest in removing biologically resistant organic contaminants lead to the
application of activated carbon in wastewater treatment PAC has been used in activated
sludge systems (Weber and Jones, 1986) while GAC found application in aerobic fixed film
processes (Lowry and Burkhead, 1980) and in anaerobic wastewater treatment (Suidan gt
sL 1983). In addition to its adsorptive properties, activated carbon has been reported to
provide an excellent surface for microbial attachment (Characklis, 1973; Suidan et al.. 1987).
Nakhla (1989) found the adsorptive capacity of GAC for o-cresol in expanded-bed
anaerobic GAC bioreactor treating a synthetic mixture of phenol, acetic-acid, and o-cresol
to be much lower than the capacity determined from adsorption isotherm conducted on
virgin carbon using the standard bottle-point technique. Further experimentation revealed
that the adsorptive capacity obtained from the biological anaerobic reactor agreed very well
-29-
-------
with capacities of virgin GAC determined from isotherm experiments conducted in the
absence of molecular oxygen. The adsorptive capacity of virgin GAC exhibited for o-cresol
in the presence of molecular oxygen (oxic conditions) was almost 200% above the adsorptive
capacity of virgin GAC attainable in the absence of oxygen (anoxic conditions). Since both
aerobic and anaerobic biological GAC reactors are in use, this study was designed to further
investigate the effects of molecular oxygen on the adsorptive capacity of activated carbon
for different organic compounds.
To evaluate the effect of the same functional group substituted at different positions
on the parent phenol molecule, both oxic and anoxic adsorption isotherm experiments were
conducted using 16x20 U.S. Mesh virgin F-400 GAC (Calgon Carbon, Pittsburgh, PA) as an
adsorbent and 2-, 3-, and 4-methylphenol as adsorbates. Experimentally determined data
on adsorption capacity for these three compounds on GAC are presented in Figure 1. All
the adsorption isotherms were found to be well described with the Freundlich isotherm
equation qe = K * Ce1/n. The straight lines on Figure 1 were obtained by nonlinear least
square fit of experimental data to the Freundlich isotherm equation. As is apparent from
this figure, the absence of molecular oxygen from the test environment resulted in almost
identical GAC adsorptive capacities for all three compounds. On the other hand, the
presence of molecular oxygen had a diverse impact on the adsorptive capacity of GAC for
different compounds. This is well illustrated in Figure 2 where the ratio of oxic and anoxic
adsorptive capacities for all three compounds are plotted as functions of the respective
equilibrium adsorbate liquid phase concentrations. The most significant increase in oxic
adsorptive capacity was observed in the case of 2-methylphenol followed by 4-methylphenol,
while the adsorptive capacity of GAC for 3-methylphenol was the least affected by the
-30-
-------
. OX1C
anoxic
o 2-Methylphenol
o 3-Methylphenol
o 4-Methylphenol
10'
C, mg/L
Figure 1. Oxic and Anoxic Adsorption Isotherms of Methylphenol on GAC.
3.0
2-Methylphenol
3-Melhylphenol
— 4-Methylphenol
2.5
9* 2.0
C, mg/L
Figure 2. Ratio of Oxic to Anoxic Adsorprive Capacities of GAC for Methylphenol.
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presence of molecular oxygen. Another important conclusion from this figure is that the two
adsorption isotherms (oxic and anoxic) have very different values of the coefficient 1/n in
the Freundlich isotherm equation since the ratio of the two adsorptive capacities depends
on the liquid phase concentration of the adsorbate.
Further investigation of the observed phenomenon was continued by extracting GAC
that was preloaded with each of the adsorbates during the adsorption isotherm experiments.
Extraction was performed in a soxhlet extraction apparatus using methanol for one day
followed by further extraction with methylene chloride for an additional period of three
days. The results of these experiments are presented in Figure 3. Extraction efficiency was
calculated as the ratio of the mass of adsorbate in the extract to the mass of adsorbate
loaded on the carbon during the adsorption isotherm experiment.
The extraction efficiencies obtained for the carbons used in anoxic isotherm
experiments showed very little dependence on the carbon loading. On the average, 90% of
the adsorbed compound was extracted from these carbons. On the other hand, the
extraction efficiencies for the carbons used in the oxic isotherm experiments were
significantly lower and exhibited strong dependency on the equilibrium carbon loading.
Furthermore, the least amount of adsorbate in the oxic isotherm experiments was extracted
from the GAC in the case of 2-methylphenol followed by 4-methylphenol and 3-
methylphenol. This arrangement is in agreement with the degree of influence of molecular
oxygen on the adsorptive capacity of GAC for these three adsorbates.
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>»
o
c
V
a
o
—i
o
«>
u
X
0)
100
90
80
70
60
50
40
30
20
10
~ o
u
„ n t> o
o « o
0O° o 0 o
anoxic
o 2-Methylphenol
o 3-MethyIphenol
o 4-Methylphenol
100
~ a
o o
o o
~ a 0
o
oxic
200 300
q. ing/g
400
500
Figure 3. Extraction Efficiency of Methylphenol from GAC.
-33-
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USE OF LIGNIN-DEGRADING FUNGI IN THE
REMEDIATION OF PENTACHLOROPHENOL-CONTAMINATED SOILS
R. T. Lamar, D. M. Dietrich and T. K. Kirk
Institute for Microbial and Biochemical Technology
Forest Products Laboratory
Madison, WI 53705
ABSTRACT
Recently we reported that a lignin-degrading fungus, Phanerockaeie chrysosporium
was capable of depleting pentachlorophenol (PCP) from three soils in the laboratory. This
depletion was found to result primarily from transformation of PCP to nonvolatile products.
The nature of these products-whether they were soil-bound or extractable—was greatly
influenced by soil type. We believe that the soil-bound products are humic acid-PCP hybrid
polymers that are formed as a result of the activity of extracellular enzymes (lignin
peroxidases, Mn peroxidases, laccases) with phenol-oxidizing activity, that are produced by
the lignin-degrading fungi. Our objective is to determine the feasibility of using
lignin-degrading fungi to remediate soils that are contaminated with wood-preserving
chemicals. The results of some of our studies are summarized below.
In investigations of white-rot fungal degradation of hazardous compounds,
P. chrysosporium strain BKM-F-1767 has been used almost exclusively as the experimental
organism. However, there are an estimated 1,400 species of lignin-degrading fungi and
there is great diversity among these organisms in their ability to degrade lignin. Thus, there
is reason to believe that this same diversity will be seen in xenobiotic degradation.
Several studies were conducted to determine the degree of interspecific and
intraspecific variation among selected Phanerochaete spp. in growth rate and in their ability to
tolerate and degrade the wood preservative PCP. Mycelial extension rates of selected strains
of Phanerochaete chrysorhiza, Phanerochaete laevis, Phanerochaete sanguinea,
Phanerochaete filamentosa, Phanerochaete sordida, Inonotus circinatus, and Phanerochaete
chrysosporium and the ability of these organisms to tolerate and degrade the wood
preservative pentachlorophenol (PCP) in an aqueous medium and in soil, were measured.
Most of the tested species had maximum mycelial extension rates in the range of < 0.5 to 1.5
cmd"', but there were large interspecific differences. A notable exception, P. sordida, grew
very rapidly, with an average mycelial extension rate of 2.68 cm d"^ at 28°C. There were
also significant intraspecific differences in mycelial extension rates. For example, mycelial
extension rates among strains of P. sordida ranged from 1.78 to 4.81 cm d"^.
Phanerochaete spp. were very sensitive to PCP in 2% malt agar. Growth of several
species was prevented by the presence of 5 ppm PCP. However, P. chrysosporium and P.
sordida grew at 25 ppm PCP, albeit at greatly decreased mycelial extension rates. We have
observed lignin-degrading fungi in PCP-contaminated soils with PCP concentrations
exceeding 500 ppm. Therefore, growth of lignin-degrading fungi in the presence of PCP on
malt agar should be used as a screening tool for determining relative sensitivities and not as
an indication of growth performance of these organisms in PCP-contaminated soil.
In an aqueous medium, mineralization of PCP by P. sordida strain 13 was
significantly greater than that by all other tested P. sordida strains and P. chrysosporium
(Table 1). litis strain was tested in a field investigation with P. chrysosporium to determine
the ability of these fungi to deplete PCP in contaminated soils.
Inoculation of a field soil contaminated with a commercial wood preservative product
and containing 250-400 jig g*1 PCP with either Phanerochaete chrysosporium or P. sordida
resulted in an overall decrease of 88% to 91% of PCP in the soil in 6.5 wk. This decrease
was achieved under suboptimal temperatures for the growth and activity of these fungi, and
without the addition of inorganic nutrients. A small percentage (8% to 13%) of the decrease
in the amount of PCP was a result of fungal methylation to pentachloroanisole (PCA).
However, both of these organisms can also transform PCA (Table 2) and thus it would also
be expected to be depleted from soil over time.
For soil studies, inocula have consisted of aspen chips (0.65 cm-1.3 cm) preinfested
with pure cultures of a single fungus. We have found in both laboratory studies and in a field
study that aspen chips absorb PCP. Therefore, we have also investigated the metabolism of
PCP by lignin-degrading fungi in wood chips.
-35-
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Table 1. Percentage of total [14C]PCP mineralized and volatilized in liquid cultures of
P. chrysosporium and several strains of P. sordida.sSter 30 days.3
Strain Mineralization Volatilization Total 14C evolved
(%)
P. chrysosporium
1.97 (0.22)a
13.82 (1.20)a
15.79
P. sordida 1
2.67 (1.09)a
12.91 (2.79)a
15.58
P. sordida 8
1.92 (0.44)a
8.88 (2.2l)a
10.80
P. sordida 9
1.22 (0.54)a
11.92 (2.35)a
13.14
P. sordida 13
11.64 (2.54)b
8.48 (0.52)a
20.12
Control
0.17 (0.03 )c
0.06 (0.03)b
0.23
aFigures in parentheses represent the standard deviation of three observations. Means
followed by the same letter are not significantly different.
Table 2. Percentage of mineralization and volatilization of pentachlorophenol and
pentachloroanisole in liquid cultures of P. chrysosporium or P. sordida 13 and from control
(noninoculated) cultures.
Culture
Compound
Mineralization
Volatilization
(%)
(%)
P. chrysosporium
PCP
8.91
7.95
PCA
7.95
9.54
P. sordida
PCP
16.13
8.25
PCA
13.07
10.02
Control
PCP
1.05
0.08
PCA
0.40
15.92
Inoculation of PCP-contaminated softwood or hardwood chips, that had been sterilized
or not, with either P. chrysosporium or P. sordida resulted in a decrease in the PCP
concentration of the chips. No decrease in the PCP concentration was observed in
noninoculated chips, indicating that the PCP decreases observed were due to the activities of P.
chrysosporium or P. sordida. Depletion in hardwood and softwood chips inoculated with P.
chrysosporium was rapid and extensive (63%-72% decrease after 6 weeks), except in nonsterile
softwood chips. In nonsterile softwood chips, depletion of PCP was very slow and resulted in
only a 30% decrease after six weeks. This lower rate of PCP depletion may have been the
result of a lower rate of colonization of these chips by P. chrysosporium due to competition
from indigenous microbes.
Depletion of PCP by P. sordida was also affected by sterilization. Inoculation of
nonsterile softwood and hardwood chips resulted in only a 50% and 45% decrease in the PCP
concentration, respectively, after 42 days. However, the PCP concentration in both hardwood
and softwood chips that had been sterilized was decreased by ca. 66% by P. sordida after 42
days. Again, this lower rate of decrease was probably due to competition from indigenous
microbes.
Depletion of PCP was always accompanied by an increase in the concentration of PCA
Accumulation of PCA in sterile cultures was much greater than in nonsterile cultures of both
fungi. This was particularly true for cultures inoculated with P. sordida. Only 7% and 19% of
the PCP decrease in nonsterile softwood and hardwood cultures, respectively, was due to
conversion of PCP to PCA. However, this low rate of conversion was associated with
relatively low amounts of PCP depletion.
In nonsterile hardwood and softwood chips inoculated with P. chrysosporium, 65%
and 72%, respectively, of the PCP decrease was due to conversion of PCP to PCA. In sterile
chips inoculated with either fungus, virtually all of the PCP decrease was due to conversion to
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PCA. The difference between sterile and nonsterile chips in the amount of PCP loss due to
methylation to PCA suggests that the indigenous microbes inhabiting the nonsterile chips,
which were not able to metabolize PCP, were able to metabolize PCA and thus prevent its
accumulation.
Inoculation of PCP-contaminated softwood chips with the lignin-degrading fungus
RAB-83-12 also resulted in a rapid and extensive removal of PCP (Figure 1). Approximately
62% of the PCP was removed after 28 days of incubation. This is similar to and greater than
the amount of PCP removed by P. chrysosporium and P. sordida, respectively, after the same
incubation rime. However, the removal was not due to conversion of PCP to PCA. Since no
PCA accumulated in cultures of this fungus we are in the process of determining the fate of PCP
in soils supporting growth of RAB-83-12 to determine the ability of this organism to transform
PCP in soils to innocuous products.
500
Control
- 400
CO
I 300
a. 200
u
ft.
100
Inoculated
0
20
10
30
Time (days)
Figure 1-. Effect of inoculation with lignin-degrading fungus RAB-83-12 on the PCP
concentration of PCP-contaminated softwood chips.
The results of these and other studies have shown that lignin-degrading fungi can effect
rapid and extensive depletion of PCP from soils and other contaminated media (i.e. wood
chips),and that there is a great diversity among fungal species in their ability to effect
decreases of PCP and in their metabolism of PCP. Further studies are needed to confirm the
incorporation of PCP into humic materials in soils and to assess the stability of these hybrid
polymers, and to continue the screening process to identify fungi with superior
PCP-degrading abilities.
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ANAEROBIC DEGRADATION OF CHLORINATED AROMATIC COMPOUNDS
jJohn E. Rogers, U.S. Environmental Protection Agency, Athens, GA;
Dorothy Hale, Wren Howard and Frank Bryant, TAI, Athens, GA; Mahmoud Mousa and
Shiu-Mei Hsu, UGA, Athens, GA.
Over the last few years, our laboratory has been investigating the
anaerobic degradation of a variety of nitrogen heterocyclic and chlorinated
aromatic compounds. These compounds have included pentachlorophenol, 2,4-D,
2,4,5-T, DDT and chlorinated anisoles, anilines, benzenes, benzoic acids,
biphenyls, and phenols. In 1991, we reported on the reductive dechlorination
of pentachlorophenol by sediment microbial communities acclimated to
dechlorinate 2,4- and 3,4-dichlorophenol. We report this year on the
degradation of 2,4-D, 2,4,5-T, DDT, hexachlorobenzene,and chloroanisoles.
The ability of unacclimated sediment microorganisms and sediment
microorganisms acclimated to dechlorinate 2,4- and 3,4-dicnlorophenol (DC?) to
degrade 2,4-D and 2,4,5-T was investigated. Acclimated sediment
microorganisms were prepared from sediment collected in November 1987 and
March 1990. When the experiments were conducted using sediments from November
1987, the 2,4-DCP acclimated microorganisms dechlorinated 2,4-D without a lag
to 4-chlorophenoxyacetic acid, the 3,4-DCP acclimated microorganisms did not
dechlorinate 2,4-D over several months of exposure, a mixture of the two
acclimated microbial populations paralleled the 2,4-DCP acclimated
microorganisms, and the unacclimated sediment microorganisms paralleled the
3,4-DCP acclimated microorganisms. The 4-chlorophenoxyacetic acid was
produced in stoichiometric quantities and was stable for the duration of the
experiment. Similar results were observed with sediments collected in March
1990; however, 4-chlorophenoxyacetic acid was readily degraded in the 2,4-DCP
acclimated sediment. 2,4,5-T was degraded at the same rate with acclimated
sediment microorganisms and unacclimated sediment microorganisms; no
degradation intermediates were detected and dechlorination followed a lengthy
lag period (> 14 days).
Previous studies in our laboratory (Struijs and Rogers, 1989) have shown
that sediment microorganisms acclimated to dechlorinate 2,4-DCP or 3,4-DCP
could dechlorinate the respective chlorinated anilines but not the the
respective chlorinated benzoic acids. A partial explanation for the
dechlorinating specificity was that the phenol and aniline are ortho/para
directing, whereas benzoic acid is meta directing. To further test this
hypothesis we have examined the dechlorination of chlorinated anisoles.
Because of the similarity of the sigma values for the methoxy substituent of
anisole and the hydroxy and amine substituents of phenol and aniline, the
methoxy substituent should also be ortho/para directing for dechlorination.
Therefore, we have examined the degradation of 2,4-dichloroanisole (DCAn) by
the 2,4- and 3,4-DCP acclimated sediment microorganisms (March 1990). When
added Co unacclimated sediment microorganisms, 2,4-DCAn was slowly
demethylated (50% in 24 days) to 2,4-DCP following a 5-day lag period. The
2,4-DCP was subsequently degraded over the next 2 weeks. Following a second
addition of 2,4-DCAn, a 50% loss of 2,4-DCAn was observed in 2 days; however,
the product observed was 4-chloroanisole. When 2,4-DCAn was added to an equal
mixture of 2,4-DCP and 3,4-DCP acclimated sediment microorganisms or to
separate 2,4- and 3,4-DCP acclimated sediment microbial communities, greater
than 50% of the 2,4-DCAn was lost in 2 days in all cases. The only products
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identified were 4-chlorophenol and phenol. Unfortunately, we were unable to
determine whether dechlorination and demethylation occurred concomitantly or
sequentially. The finding that 4-chloroanisole was the primary product from
the second addition of 2,4-DCAn, however, is indicative of rapid reductive
dechlorination of 2,4-DCAn by a microbial community acclimated to dechlorinate
2,4-DCP.
To date we have investigated the reductive dechlorination of several
compounds having some structural resemblance to the acclimating substrates.
Recently we have investigated the ability of acclimated sediment
microorganisms to degrade compounds with little structural resemblence. DDT
added to unacclimated sediment microorganisms, both autoclaved and non-
autoclaved, was rapidly converted to DDE within 4 days. Similar results were
observed for DDT added to 2,4- and 3,4-DCP acclimated sediment microorganisms,
both autoclaved and non-autoclaved. However, 4-chlorophenol was observed as a
significant product with acclimated microorganisms (autoclaved and non-
autoclaved) but not with unacclimated microorganisms. The mechanism of 4-
chlorophenol formation is currently under investigation. Hexachlorobenzene
(HB) was investigated in sediments collected from two different ponds. In
both cases, a loss of HB (70 ppm) was observed after a substantial lag phase
(90 to 100 days). Less than 20% of the HB remained after 170 days of
incubation. A mixture of lesser chlorinated congeners were identified as
intermediate degadation products. A different mixture was observed in the two
sediments and no one product dominated in either case. Hexachlorobenzene was
stable in autoclaved sediment.
The reductive dechlorination of chlorinated aromatic compounds requires
a source of reducing equivalents. Dolfing and Tiedje (1986) have identified
molecular hydrogen as a source of reducing equivalents for the dechlorination
of 3—chlorobenzoate by an anaerobic consortium. Gibson and Suflita (1990) and
Nies and Vogel (1990) have shown that organic substrates can also be the
source of reducing equivalents for the reductive dechlorination of 2,4,5-T and
PCBs. We are currently testing molecular hydrogen and a number of organic
substrates as sources of reducing equivalents for the reductive dechlorination
of 2,4- and 3,4-DCP by microorganisms in sediment slurries. In some cases,
our results are different from previous reports. We found, for example, that
changes in head space gas composition resulted in the following order of
increasing reductive dechlorination: C02/N2 > N2 > Hz/N2. We also observed
that propionate, formate and butyrate, which have been shown to stimulate the
degradation of 2,4,5-T, inhibited or did not enhance the reductive
dechlorination of 2,4-DCP. In a parallel study, we have tested a number of
complex organic substrates as possible sources of reducing equivalents.
Addition of sterile sediment or a sediment extract supported reductive
dechlorination of 2,4-DCP. Lake water collected from just above the sediment
source only marginally supported 2,4-D degradation. Organic mixtures such as
landfill leachate and rumen fluid did not stimulate dechlorination of 2,4-DCP
when added to sediments at low concentrations (1 to 2%) and inhibited activity
at higher (15 to 20 %) concentrations.
References:
Dolfing J. and J. M. Tiedje. 1991. Influence of substituents on reductive
dehalogenation of 3-chlorobenzoate analogs. Appl. Environ. Microbiol. 57: 820-
824.
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Gibson, S. A. and J. M. Suflita. 1990. Anaerobic biodegradation of 2,4,5-
trichlorophenoxyacetic acid in samples from a methanogenic aquifer:
stimulation by short-chain organic acids and alcohols.
Nies, L. and T. M. Vogel. 1990. Effects of organic substrates on
dechlorination of Aroclor 1242 in anaerobic sediments. Appl. Environ.
Microbiol. 56:2612-2617.
Struijs, J. and J. E. Rogers. 1989. Reductive dechlorination of
dichloroanilines by anaerobic microorgqanisms in fresh and dichlorophenol-
acclimated pond sediment. Appl. Environ. Micrbiol. 55: 2527-2531.
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-42-
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para-Hvdroxvbenzoate as an Intermediate
in the Anaerobic Transformation of Phenol to Benzoate.
Barbara R. Sharak Genthner
Technical Resources, Inc.
Environmental Research Laboratory, U.S. EPA
Gulf Breeze, FL. 32561
SUMMARY
Anaerobic transformation of phenol was studied using a
bacterial consortium which transformed phenol to benzoate without
complete mineralization of benzoate. Products of monofluoro-
phenol transformation indicated para-carboxvlation. Phenol and
benzoate were detected during para-hvdroxvbenzoate (jd-OHB)
degradation. p-OHB was detected in phenol-transforming cultures
containing 6-hydroxynicotinic acid (6-OHNA), a structural
analogue of p-OHB, or at elevated initial concentrations of
phenol (> 5 mM), or benzoate (> 10 mM).
RESULTS AND DISCUSSION
The original phenol-degrading consortium (2) first
transformed phenol to benzoate followed by complete
mineralization of benzoate. Subculture B-l transformed all of
the phenol to benzoate, but failed to completely mineralize the
resulting benzoate.
Subculture B-l stoichiometrically transformed 2-fluorophenol
(2FP) to 3-fluorobenzoate (3FB) in the presence or absence of
phenol. The rate of 3FB formation (12.9 moles"l~1'd~1) was similar
to the rate of 2FP decline (13.5 moles"l'1'd*1) . In the presence
of phenol, a small amount (3%) of 3-fluorophenol was transformed
to 2-fluoro-benzoate. 4-Fluorophenol was not transformed in the
presence or absence of phenol. Neither 2-, 3-, nor 4-fluoro-
benzoate was used as an energy source. These results are the
same as those obtained with the original phenol consortium (3).
Thus, transformation was via para-carboxvlation as previously
concluded for the original phenol consortium.
Para-carboxylation of phenol implies the formation of e~0HB
as an intermediate of transformation. Therefore, we examined the
degradation of p-OHB by both the original phenol consortium and
subculture B-l. After three days, e~ohb was completely degraded
in both cultures. Phenol was the only compound detected during
p-OHB degradation by the original phenol consortium (Fig. 1A).
By contrast, both- phenol and benzoate were detected in subculture
B-l (Fig. IB). After two weeks, phenol was no longer detected in
either consortium, but 900 uM benzoate was present in subculture
B-l. Thus, the original phenol consortium had completely
degraded the phenol formed via decarboxylation of jd-OHB, whereas
subculture B-l had subsequently transformed phenol benzoate.
Failure to detect benzoate during e-OHB degradation by the
-43-
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i 30
z
o
2.0
i-
<
cr
i—
0.5
z
UJ
u
z
o
o
\
\ a'l'IB
0.0« '
50 0 1C
TIME (HOURS)
20
30
20
30
40
10
Figure 1. Degradation of para-Hydroxybenzoate by the Original
Phenol Consortium (A) and Subculture B-l (B).
Symbols: 0, Phenol; A, para-Hvdroxybenzoate; Benzoate.
2
E
<
o
M
z
UJ
CD
O
z
LJ
X
0.
25 50 75
TIME (DAYS)
100
x
-<
o
70
O
S
CD
m
Z
ISI
O
5
m
Z
Figure 2. Detection of para-Hvdroxvbenzoate (A) as an
Intermediate in the Transformation of Phenol (Q)
to Benzoate ( ~ ) in the presence of (A) 10 mM Phenol
(B) 6-Hydroxynicotinic Acid (1 mM) or (C) 10 mM Benzoate,
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original phenol consortium may be a result of rapid benzoate
turnover. The detection of benzoate during p-OHB degradation by
subculture B-l may be the result of diluting out a bacterial
species essential for complete benzoate degradation.
E~OHB was not detected in the original phenol consortium,
but small amounts (2 uM) of a compound with the retention time of
p-OHB were detected in subculture B-l during transformation of
phenol. This suggested that benzoate accumulation inhibited ja-
OHB breakdown. Consequently, studies were devised to enhance p-
OHB formation using analogue-, product- and substrate-inhibition.
The highest concentrations of e~OHB (14 ~ 43 uM) were detected in
the original phenol consortium (data not shown) and in subculture
B-l at initial phenol concentrations > 5 mM (Fig. 2A). The
intermediate was also detected (6-8 uM) in the presence of >
100 uM 6-OHNA or > 10 mM benzoate (Figs. 2B & 2C, respectively).
The intermediate in subculture B-l grown with 10.5 mM phenol
cochromatographed with authentic p-OHB under the two sets of
separation parameters. Its UV spectrum matched that of authentic
jd-OHB. Its identity was confirmed by GC/MS analysis. The mass
spectrum matched that of the TMS-derivative of authentic p-OHB
with peaks at m/e 282 and m/e 267 (-CH^) . Neither 2-, nor 3-
hydroxybenzoate was detected during this analysis. Identifying
p-OHB in cultures that were transforming phenol provided a direct
indication of para-carboxvlation. Since p-OHB was not detected
in the original phenol consortium with initial phenol
concentrations < 5 mM, it .is possible that p-OHB was not detected
in other studies (1,4,5) because of the use of lower phenol
concentrations (< 2 mM).
REFERENCES
[1] Knoll, G. and Winter, J. (1989) Appl. Environ. Biotechnol.
30, 318-324.
[2] Sharak Genthner, B.R., Townsend, G.T. and Chapman, P.J.
(1989) Biochem. Biophys. Res. Comm. 162, 945-951.
[3] Sharak Genthner, B.R., Townsend, G.T. and Chapman, P.J.
(1990) Biodegradation 1, 65-74.
[4] Zhang X., Morgan T.V. and Wiegel, J. (1990) FEMS Microbiol.
Lett. 67, 63-66.
[5] Zhang, X. and Wiegel, J. (1990) Appl. Environ. Microbiol.
56, 1119-1127.
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INFLUENCE OF NONIONIC SURFACTANTS ON THE
ANAEROBIC DECHLORINATION OF HEXACHLOROBENZENE
by: Patricia Van Hoof
U. of Georgia/U.S. EPA
College Station Road
Athens, GA 30613
Chad T. Jafvert
Environmental Research Laboratory
U.S. EPA
Athens, GA 30613
ABSTRACT
BACKGROUND
Surfactants can solubilize pollutants into micellar solution in the
presence of soil or sediment solids, effectively desorbing them from
these natural media (1,2). This phenomena has been suggested as a
possible tool to enhance the treatment of contaminated sediments or
soils (3-5). The rationale is that surfactant micelles (or monomers or
emulsions) will solubilize precipitated or sorbed compounds, making them
more readily available for biological remediation, pump-and-treat
operations, or soil washing operations.
The effectiveness of surfactants in removing contaminants from
soils or sediments is largely a function of (i) the sorption reactions
of pollutants to the sedimentary materials, (ii) the solubilization of
pollutants by the surfactant micelles (and/or monomers), and (iii) the
interactions of surfactant monomers and micelles with sediment or soil
components. These processes recently have been examined in several
freshwater sediments spiked with various polycyclic aromatic
hydrocarbons (PAHs) and the anionic surfactant sodium dodecylsulfate,
SDS (1,2). Also, we recently have examined the solubilization of PAHs
and other compounds in solutions of various nonionic surfactants,
including Brij 35, Tween 80, and Tween 20, as well as examination of
some of the interactions of these surfactants with sediment solids.
The structures of Tween 80 and Brij 35 are shown in Figure 1.
Unlike SDS, which can be purchased in pure form (> 99X), these
surfactants are homolog mixtures of differing ethoxy chain lengths.
Also, unlike SDS, whose primary interaction with sediment components
under conditions of interest is the precipitation of its calcium salt,
nonionic surfactants sorb to sediment solids -- possibly through both
hydrophobic and hydrophilic mechanisms. At surfactant concentrations
around their critical micelle concentration (cmc), and at sediment
-47-
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TWEEN 80
o
Rl= -(CHjCHp)^
Rj = -(CHjCHjO)^
Ra= - 952) of the surfactant
is sorbed. However, because of their lower toxicity to microorganisms
(relative to anionic surfactants), we chose to examine the effects of
enhanced solubilization by these nonionic surfactants on microbially
mediated transformation reactions of hydrophobic organic pollutants.
Hexachlorobenzene (HCB) was chosen as a test compound, primarily because
of its low water solubility, 5.0 /ig/L (6), (i.e., high sediment-water
partition coefficient) and because it is known to degrade relatively
slowly in anaerobic sediments (7,5).
ANAEROBIC DECHLORINATION OF HCB
The dechlorination of low levels of hexachlorobenzene (2.8 x 10"3
mM) incubated in anaerobic pond sediments (72 solids) occurred after an
initial lag period of 10 days at a rate of 5.2 x 10~5 mM/day . Fresh
sediments, inoculated with acclimated sediments, dechlorinated HCB (2.2
x 10~3 mM) at a slightly faster rate of 6.5 x 10"5 mM/day with no initial
lag phase. Initially, only one dechlorination pathway was observed with
pentachlorobenzene (PCB), and 1, 2,3,5-tetrachlorobenzene (1,2,3,5-TTCB)
as intermediates and 1,3,5-trichlorobenzene (1,3,5-TCB) as a possible
end product. After 55 days, a second dechlorination pathway was
observed with 1,2,4,5-TTCB and 1,2,4-TCB as intermediates, and all three
dichlorobenzenes (DCBs) as possible end products. Both pathways have
been observed during anaerobic incubation of fresh digester sludge (7).
After 75 days, degradation of HCB is complete, leaving 1,3,5-TCB,
1,2,4,-TCB and 1,3-DCB as the major products. In acclimated sediments,
a shorter delay of 10 days occurs before onset of the second pathway.
With additions of Tween 80, the polyoxyethylene sorbitan monooleate
surfactant, the aqueous phase concentration of HCB were increased by one
to two orders of magnitude over that of controls, which were generally
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at half aqueous saturation (2-3 jig/L) in the sediment slurry. After 1
week, however, the surfactant was degraded, and the concentration of
aqueous HCB decreased to control levels. Initially, the rates of HCB
dechlorination and product formation were similar to that of controls
for low levels of Tween 80 (1500 mg/L). After 40 days, however,
dechlorination ceased. As Tween 80 concentrations were increased to
5000 mg/L, initial rates of dechlorination decreased to almost
imperceptible levels. In addition, there was no evidence of the second
pathway. In acclimated sediments exposed to low levels of Tween 80 (900
and 1200 mg/L), rates of HCB dechlorination are similar to those of
controls. At even lower levels of Tween 80 (300 and 600 mg/L),
dechlorination rates are slightly faster. Both pathways are evident in
these acclimated sediments.
Contrasting the influence of Tween 80 on HCB dechlorination is the
effect of Brij 35, a polyoxyethylene alcohol surfactant. HCB
dechlorination is completely suppressed in acclimated sediment slurries
in the presence of low levels of Brij 35 (1000 mg/L). In fresh sediment
slurries, the surfactant addition results in longer lag periods. Again,
the surfactant appears to be readily degraded, thus preventing aqueous
solubility enhancement effects from being observed. Currently, another
alcohol surfactant, Brij 30 [C12H250(CH2CH20)^H] , is being tested and
shows no signs of being degraded, itself, to date (after 4 weeks).
This paper has been reviewed in accordance with the U.S.
Environmental Protection Agency's peer and administrative review
policies and approved for presentation and publication.
REFERENCES
1. Jafvert, C.T.; Heath J.K. 1991. Environ. Sci. Tech. In press.
2. Jafvert, C.T. 1991. Environ. Sci. Tech. In press.
3. Roy, W.R.; Griffen, R.A. 1988. Surfactant- and Chelate-Induced
Decontamination of Soil, Report 21; Environmental Institute for
Waste Management Studies, The University of Alabama.
4. Nash, J.H. 1987. In, Field Studies of In Situ Soil Washing, U.S.
Environmental Protection Agency, Cincinnati, Ohio. EPA/600/2-
87/100.
5. Vigon, B.U.; Rubin, A.J. 1989. Journal W.P.C.F. 61:1233-1244.
6. Chiou, C.T.; Schmedding D.W. 1982. Environ. Sci. Technol. 16:4-
10.
7. Fathepure, B.Z.; Tiedje, J.M.; Boyd, S.A. 1988. Appl. Environ.
Microbiol. 54:327-330.
8. Mousa, M. A.; Rogers, J.E., personal communication.
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ANAEROBIC DEGRADATION OF CHLOROAROMATIC COMPOUNDS
UNDER DIFFERENT REDUCING CONDITIONS
M.M. Haggblom, M.D. Rivera and L.Y. Young
Depts. of Microbiology and Environmental Medicine,
NYU Medical Center, 550 First Avenue, New York, NY 10016
and
J. Rogers
U.S. Environmental Protection Agency, Athens, GA 30613
CHLOROPHENOL DEGRADATION UNDER METHANOGENIC CONDITIONS
If only dechlorination of a chloroaromatic substrate is taking place and the aromatic
ring remains intact, no carbon is provided for microbial growth. In the environment carbon
sources in addition to a chloroaromatic are likely to be available, which may affect the
metabolism of the chloroaromatic. In order to examine this, sediment enrichment cultures were
set up with 2,4-dichlorophenol and 4-chlorophenol under methanogenic condition with and
without the addition of a supplementary carbon source. Propionate was chosen as a readily
utilizable carbon source, and para-cresol was used as a structurally similarly non-chlorinated
substrate.
2,4-Dichlorophenol was dechlorinated to 4-chlorophenol without a lag in 25 to 60 days
in the presence or absence of an auxiliary carbon source. It took approximately 50 days before
the onset of 4-chlorophenol degradation. Degradation of 4-chlorophenol was stimulated by
either p-cresol or propionate. In cultures without an auxiliary carbon source 4-chlorophenol
persisted. By repeated feedings of the chlorophenols and auxiliary substrates the degradation
rates were significantly enhanced. After 6 feedings of/>-cresol and chlorophenol over a period
of 450 days, the rate of 2,4-dichlorophenol and 4-chlorophenol degradation was enhanced over
ten fold to 40 and 10 /jmol liter"1 day'1, respectively.
Dechlorination at the orr/io-position could be sustained and by repeated dilution into
fresh medium and refeeding, a stable microbial enrichment culture free of sediment, which
degraded 2,6-dichlorophenol was established. 2,6-Dichlorophenol was sequentially dechlorinated
to 2-chlorophenol and phenol and ultimately mineralized to CH4 and CO,. On the other hand,
the ability to degrade 4-chlorophenol was lost by transferring the culture into fresh medium,
but by refeeding the 4-chlorophenol degrading culture could be maintained.
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Cultures adapted to 2,4- and 2,6-dichlorophenol readily dechlorinated other
dichlorophenols containing an ort/io-chlorine. Dechlorination of 2,3- and 2,5-dichlorophenol
yielded 3-chlorophenol. Dichlorophenols with no ort/20-chlorines persisted. 2,3,6-
Trichlorophenol was dechlorinated at the ort/io-position yielding first 2,3- and 2,5-
dichlorophenol, and then 3-chlorophenol. Similarly 2,4,6-trichlorophenol was sequentially
dechlorinated to 2,4-dichIorophenol and 4-chlorophenol. This preferential removal of ortho-
chlorines, with meia-or para -chlorines removed at slower rates, appears to be characteristic for
methanogenic cultures.
CHLOROPHENOL DEGRADATION UNDER SULFIDOGEN1C CONDITIONS
Degradation of chlorophenols under sulfate-reducing conditions was studied with an
estuarine sediment inoculum (East River). After an initial lag period of approximately 50 to
100 days 2-, 3- and 4-chiorophenol and 2,4-dichlorophenol (0.1 mM) were completely removed
in 120 to 220 days. 4-Chlorophenol was detected as a transient metabolite of 2,4-
dichlorophenol, but no metabolites of the monochlorophenols were detected. The rate of
chlorophenol degradation was greatly enhanced after repeated refeeding of the substrate to the
sediment cultures. In acclimated cultures the monochlorophenols (0.16 mM) were degraded in
6 to 20 days, corresponding to rates of 8 to 40 /xmol liter'1 day1, which are similar to the
degradation rates in methanogenic cultures. The relative rates of degradation were
4-chlorophenol > 3-chlorophenol > 2-chlorophenol, 2,4-dichlorophenol. No degradation of
chlorophenols was observed in sterile controls.
During degradation of all three monochlorophenol isomers in the sediment cultures
there was a concomitant loss of sulfate, corresponding to the stoichiometric values expected
for complete oxidation of the chlorophenol to COb according to the following equation:
C6HsC10 + 3.25 SO,2" + 4 H:0 —> 6 HC03 + 3.25 H,S + CI + 0.5 H+
Formation of sulfide was confirmed with 4-chlorophenols using a radiotracer technique.
No methane was produced in the cultures, verifying that sulfate reduction was the main
electron sink. Addition of molybdate, a specific inhibitor of sulfate reduction, inhibited
chlorophenol degradation completely. These results indicate that chlorophenols can be
mineralized under sulfidogenic conditions and that oxidation of the chlorophenol is coupled to
sulfate reduction.
The sulfidogenic cultures were propagated by repeated refeeding of chlorophenols and
dilution into fresh medium, and are able to utilize the chlorophenol as a source of carbon and
energy. This is the opposite of what was observed under methanogenic conditions, where 4-
chlorophenol degrading cultures could not be subcultured without loss of activity. The
sulfidogenic cultures were very specific and only degraded the monochlorophenol isomer to
which they were acclimated. However, all the cultures rapidly degraded phenol, suggesting that
it may be an intermediate in chlorophenol degradation.
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DEGRADATION OF CHLORINATED PHENOLS AND BENZOIC ACIDS UNDER THREE
REDUCING CONDITIONS
Anaerobic enrichment cultures, under methanogenic, sulfidogenic and denitrifying
conditions were established on each of the three monochlorophenol and monochlorobenzoate
isomers with Hudson River sediment from two different sites (HR1, HR2). In addition
denitrifying cultures were also established on the same compounds with East River sediment.
Initial results monitoring substrate loss indicated that all three monochlorophenols and 3-
chlorobenzoate were degraded under methanogenic conditions in HR1 cultures. Sulfidogenic
HRl cultures were active against all three monochlorophenols and 3- and 4-chlorobenzoate.
Transient accumulation of phenol was detected in some of the chlorophenol amended cultures
under both methanogenic and sulfidogenic conditions, indicating that reductive dechlorination
is taking place. HR2 cultures, in general, showed the same activity but at a much slower rate.
The rate of degradation was enhanced with refeeding of the substrates. Under denitrifying
conditions degradation of 3- and 4-chlorobenzoate and 2-chlorophenol was observed in HRl
sediment cultures. These results indicate that degradation of these chlorinated aromatic
compounds can take place under more than one reducing conditions.
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Onsite Biological Pretreatment Followed by
POTW Treatment of CERCLA Leachates
E.R. Krishnan, R.C. Haught, and M.L. Taylor
PEI Associates, Inc.
Cincinnati, OH 45246
M.T. Suidan and M. Islam
Department of Civil and Environmental Engineering
University of Cincinnati
Cincinnati, OH 45221-0071
R.C. Brenner
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
The objective of this research is to assess the effectiveness of fixed-
film anaerobic biological processes in treating and decontaminating leachates
containing synthetic organic chemicals (SOCs) that may be regulated under the
Comprehensive Environmental Response, Compensation, and Liability Act
(CERCLA). This is an attractive proposition because anaerobic processes
result in lesser air stripping of volatile organic compounds compared to
aerobic processes due to their lower gas production rates. Anaerobic
pretreatment processes are also expected to reduce problems associated with
incomplete degradation of chlorinated compounds as well as passthrough of
semivolatile organic compounds, which can occur when CERCLA leachates are
discharged without any pretreatment to publicly owned treatment works (POTWs).
Two types of anaerobic pretreatment processes are being evaluated in this
study: an upflow anaerobic filter reactor and a granular activated carbon
(GAC) anaerobic fluidized-bed reactor. The tests are being conducted at U.S.
EPA's Test and Evaluation (T&E) Facility in Cincinnati, Ohio.
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Leachate Characteristics
The leachate for the experiments is obtained from a large commercial
municipal landfill in Georgetown, Ohio. The leachate is highly variable in
composition (COD levels ranging from 300 to 2500 mg/L) with low to moderate
levels of biodegradability. Sulfate concentrations in the leachate range from
3 to 300 mg/L. Due to its relatively low biodegradable content during the
first 6 months of the project, the leachate was supplemented with a mixture of
acetic, propionic, and butyric acids to increase its total COD to approxi-
mately 1600 mg/L. Later, the total COD of the leachate increased signifi-
cantly to a maximum of 2500 mg/L, at which time volatile acids addition was
discontinued. The leachate was fed without any addition of volatile acids
during the remainder of the project. During most of the latter part of the
project, the total COD of the raw leachate remained relatively low, ranging
between 400 to 1000 mg/L. The leachate is rendered hazardous by supplementing
it with a mixture of ten volatile and four semivolatile organic compounds,
shown with their corresponding target concentrations in Table 1. Chloroform
was not added to the leachate until the later stages of the project because of
its potential toxicity.
Treatment Systems
The treatability of the leachate is being evaluated in three parallel
trains. One train consists of leachate pretreatment in a bench-scale upflow
anaerobic filter reactor (6-in. diameter x 48-in. high) packed with 1-inch
Pall rings, followed by mixing with raw municipal wastewater in a ratio of 95%
wastewater to 5% leachate and treatment in a bench-scale activated sludge POTW
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Table 1. Composition of SOC Supplement to the Leachates
Compound
Concentration (ua/U
VOLATILE ORGANIC COMPOUNDS
Acetone
Methyl Ethyl Ketone
Methyl Isobutyl Ketone
Trichloroethylene
1,1-Dichloroethane
Methylene Chloride
Chloroform
Chlorobenzene
Ethyl benzene
Toluene
0 to 2,000
10,000
1,100
600
8,000
5,000
1,000
400
100
1,200
SEMIV0LATILE ORGANIC COMPOUNDS
Phenol
Nitrobenzene
1,2,4-Trichlorobenzene
Dibutyl Phthalate
2,600
500
200
200
unit. The second train is similar in scale to the first, with the exception
that a fluidized-bed reactor (4-in. diameter x 42-in. high) filled with 16 x
20 U.S. mesh granular activated carbon is used for leachate pretreatment
instead of the upflow filter reactor. The third treatment train consists of a
pilot-scale anaerobic filter reactor (4.25-ft diameter x 7.5-ft high) followed
by a pilot-scale activated sludge POTW unit. 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. The anaerobic
pretreatment systems are operated at 35°C. The empty bed contact time in the
GAC fluidized-bed reactor is maintained between 6 and 8 hrs, while a longer
detention time of 48 to 96 hrs is employed in the anaerobic filter reactors.
The COD removal efficiency in the bench-scale anaerobic filter and GAC
fluidized-bed system averaged 42% and 48%, respectively. The average influent
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Results
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COD was about 1,100 mg/L. The COD removal efficiency for the systems was
found to increase with increasing influent COD. During the period of the
volatile acids addition, the primary COD removal mechanism was methanogenic.
After the volatile acids addition was stopped and the COD of the leachate
decreased, the COD removal mechanism was due to a combination of
methanogenesis and sulfate reduction. The average sulfate reductions in the
6AC fluidized-bed and anaerobic filter reactors were 71% and 65%,
respectively, corresponding to an average influent sulfate concentration of
116 mg/L. In the bench-scale anaerobic filter reactor, most of the organic
compounds were removed in excess of 90%. Chlorobenzene, ethyl benzene, 1,1-
dichloroethane and trichlorobenzene exhibited gradual degradation, indicating
the need for longer acclimation periods. In the bench-scale GAC fluidized-bed
reactor, all organic compounds exhibited removal efficiencies exceeding 95%
with the exception of 1,1-dichloroethane, which required a longer acclimation
period. Despite some differences in the design and operation of the pilot-
scale anaerobic filter reactor compared to the bench-scale anaerobic filter
reactor, the performance of both systems were similar prior to chloroform
addition to the leachate. Within 3 weeks after the addition of chloroform,
however, the pilot-scale system showed a decline in the removal of some of the
SOCs (including chloroform). SOC removals continued to decline over a period
of 4 months, at which time chloroform addition was discontinued. The pilot-
scale system is now being monitored to determine if performance can be
recovered to pre-chloroform addition levels while continuing to add the SOC
consortium (excluding chloroform) to the leachate feed.
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BACTERIAL DEGRADATION OF KPEG-MODIFTED PCBS IN
ANAEROBIC AND AEROBIC ENRICHMENT CULTURES
by: Wale Adewunmi and Joseph A. Krzycki
Department of Microbiology
Ohio State University
Columbus, OH 43210
In the 1960s and 70s it became apparent that poly chlorinated biphenyls
were recalcitrant to biological degradation, and were accumulating in
organisms throughout the food chain. Subsequently they were phased out of
use; but now it is estimated that at least a half million tons of PCB now exists
in landfills and closed systems awaiting detoxification (1). Destruction is
accomplished by transport of contaminated materials to incinerators. An
alternative method is treatment of contaminated materials with a
formulation of KOH and polyethylene glycol 400 (KPEG). KPEG attacks
chlorine bearing carbons, resulting in the formation of chloroarylpolyglycols
(2,3,7). These compounds resemble PEG based non-ionic surfactants, and it
appeared probable that PCB reacted with KPEG would be rendered soluble in
water. Considerations such as this led to the suggestion that KPEG treatment
coupled with biological degradation could be an alternative method of PCB
disposal. We have been testing the feasibility of this idea by studying
metabolism of a KPEG treated PCB congener in both aerobic and anaerobic
primary enrichment cultures.
Reaction of KPEG with a polychlorinated biphenyl may yield
numerous products, since nucleophilic attack of KPEG can occur at several
different chlorine bearing carbons and result in a mixture of
chloroarylpolyglycols with different degrees of PEG substitution, as well as
small amounts of chlorobiphenylols (2, 3). Therefore, we elected to carry out
our studies using uniformly labelled 2,2',4,4',5,5' hexaclorobiphenyl (14C-
HCBP) in order to facilitate identification of the products of the initial
chemical reaction, as well as biodegradation products. This congener is 8.2%
of Arochlor 1260, and 3.3% of Arochlor 1254 (4); and has been identified as a
major congener found in both human adipose tissue (5) and milk (6).
KPEG was formulated as described by Kornell and Rogers (7) except that
1.3 mol of KOH was dissolved with heating into 1.0 mol of PEG 400. The
mixture was added to ^C-HCBP absorbed to glass at a ratio of 100 |ig congener
to 1 ml KPEG. The reaction was incubated at 84°C and quenched by addition of
water. Figure 1 details a timecourse of the reaction which demonstrates KPEG
effects a rapid phase transfer of 14C-HCBP.
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4000
Aqueous
3000
2
a.
Q 2000
1000
Hexane
0
10
20
30
40
50
60
Time (min)
Fig. 1. Phase transfer of HCBP during reaction with KPEG. Vials containing
100 p.g of 14C-HCBP (20 }iCi/mg) and 1 ml of KPEG reagent were incubated at
84°C before the reaction was quenched with water at the indicated times.
After the reaction had proceeded for sixty minutes, 95% of recovered
radioactivity had been converted to a form which remained in the aqueous
phase following repeated extractions with hexane. The products formed from
l^C-HCBP after 48 h reaction with KPEG remained soluble in water. In
control experiments where l^c-HCBP was incubated with PEG, rather than
KPEG, phase transfer was not observed and the congener remained hexane
soluble. This experiments indicated that KPEG reaction did convert HCBP to a
water soluble form, presumably the chloroarylpolyglycols first described by
Brunelle and Singleton (3). Reverse phase HPLC analysis (figure 2) using an
instrument equipped with an on-line radioactivity detector demonstrated
increasing product complexity with time. Two major water soluble products
were formed within minutes, at later times these were converted to forms
which eluted relatively quickly from the reverse phase column. These
products accumulated and with 12 h reaction time six products from the
aqueous phase could be resolved. These may represent chloroarylpolyglycols
with higher degrees of PEG substitution. On the basis of these experiments a
reaction time of one hour was chosen in order to minimize product
homogeneity in the subsequent biodegradation studies.
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0.04
0.03-
*230
0.02-
aoi-
1 HOUR
Time (min)
1500
- 1000
CPM
. soo
0.0*1
12 HOURS
0.03-
1
1
\ * \ f/
—• cpm
¦~A230
0.02-
;i
\ h 1; \n
> M./ I
0.01-
j"
0-
-
¦ 400
CPM
- 200
30
40 60
Time (min)
80 IX
Figure 2.14C-HCBP / KPEG products after 1 and 12 h of reaction time analyzed
on a silica based C-18 HPLC column eluted at 0.35 ml /min with a 30-50%
isopropanol gradient.
Biodegradation experiments were initiated with anaerobic sewage
sludge taken from the Jackson Pike municipal waste treatment plant
(Columbus, OH). arylpolyglycols were incubated with either undiluted
sludge or primary enrichment cultures. Enrichment cultures were made
using a mineral medium, and incubated under a nitrogen headspace. The
KPEG /1 ^C-HCBP reaction mixture was neutralized with HCL and added to
cultures to final concentrations of 0.2 mM total arylpolyglycols and 25 mM
PEG. Enrichments were set up in duplicate under methanogenic (no
additions), sulfidogenic (15 mM Na2S04), iron reducing (0.1% amorphous
iron oxide), or denitrifying (20 mM KNO3) conditions. Controls were
established for each condition investigated by inoculating cultures with
autodaved sludge. The fate of the labelled arylpolyglycols was followed by
periodic ethyl acetate extraction of aliquots from each culture. Both
arylpolyglycols and unreacted ^C-HCBP could be extracted using this solvent.
The ethyl acetate extract was then evaporated and partitioning of radioactivity
in the residue between hexane and water determined. Figure 3 shows the
average results obtained under methanogenic conditions for active cultures
and controls. At the start of incubation 80% of expected counts could be
recovered in the ethyl acetate fraction, 19% was left in the culture
supernatant. Of radioactivity in the ethyl acetate fraction, only 4% was soluble
in hexane, the remainder was in the aqueous fraction. After incubation for 28
days at 37°C, 71% of expected counts could be recovered by ethyl acetate
extraction of the cultures. However, 60% of radioactivity in the ethyl acetate
fraction was now soluble in hexane; the remaining radioactivity was still
soluble in water. In contrast, killed controls showed no change in the phase
distribution of label. Similar results were obtained with enrichments cultures
established under sulfate, nitrate, or iron reducing conditions, as well as with
l^C-chlorobiphenylpolyglycols incubated in undiluted sludge. These results
indicated that in this anaerobic sludge bacteria were capable of converting
chlorobiphenylpolyglycols into a water insoluble form. A possible
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mechanism is hydrolysis of the PEG moiety from the aromatic ring, leaving a
relatively insoluble polychlorobiphenylol. An analogous reaction has been
documented with nonoxynol (Triton N), a PEG based non-ionic surfactant (S).
Biodegradation in sewage sludge resulted in cleavage of the PEG moiety from
nonoxynol and liberation of the recalcitrant, hydrophobic compound,
nonylphenol.
Autodaved
control
Active
culture
DAY DAY
Figure 3. Phase transfer of ^4C-HCBP/KPEG reaction products during
incubation in methanogenic enrichment cultures. The partitioning of
extracted radioactivity from 0.5 ml of culture between hexane and water is
shown. The amount of radioactivity not extracted from the culture by ethyl
acetate is also indicated.
Biodegradation of the labelled arylpolyglycols was also tested in aerobic
primary enrichments cultures. Metabolism of the compounds did occur;
however, unlike anaerobic cultures, the products remained water soluble.
Aerobic primary enrichments were established in triplicate sealed 160 ml
vials which were periodically replenished with oxygen. Each culture received
0.1 mM labelled arylpolygycol and 12 mM PEG. Duplicate enrichment cultures
were also established with autodaved inocula as controls. Results are
illustrated in figure 4. At the start of incubation an average of 66% of total
radioactivity added to active cultures could be recovered by ethyl acetate
extraction, while 16% was not extractable and remained in the culture
supernatant. 90% of radioactivity in the ethyl acetate fraction was water
soluble. After 41 days of incubation, only 19% of radioactivity added to the
culture could be recovered by ethyl acetate extraction, while 57% was not
removed from the cultures after repeated extraction with ethyl acetate. The
amount of radioactivity which was soluble in both ethyl acetate and hexane
did not increase. This indicated that aerobic biodegradation of the labelled
arylpolyglycols had resulted in products which were soluble in water, but no
longer soluble in ethyl acetate.
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Active
culture
^ tqueoiu
n orglnlc
not extracted
Autoclaved
ng Autoclaved
H^fffliU control Iran
LL
0 43 0 43
day day
Figure 4. Phase partitioning between hexane and water of ^C-HCBP/KPEG
reaction products extractable by ethyl acetate after incubation for 43 days in
aerobic enrichment cultures and killed controls. "Not extracted" refers to
radioactivity left in the culture sample after four sequential 1 ml extractions
with ethyl acetate.
Reverse phase HPLC analysis of the aqueous supernatant of cultures (figure 5)
confirmed metabolism of the products occurred. Killed controls had
radioactive HPLC elution profiles similar to that of the starting material,
however the supernatant of active cultures no longer contained any of the
KPEG/l^C-HCBP reaction products. Instead, nearly all the radioactivity
present in the sample eluted immediately from the column. Using a different
HPLC methodology we have identified two major radioactive peaks with
absorbance at 230 nm, indicating the breakdown products still contain an
aromatic ring. Currently we are determining if these compounds are
endproducts or transient intermediates. They may represent products in
which a PEG substituted aromatic ring has been cleaved following hydrolysis
of the PEG moiety from the ring, resulting in a chlorinated aromatic
carboxylic acid.
0.08
^ 200
800
Active
culture
230
0.06 -
CPM
0.04 -
- 400
0
20
40
BO
60
100
20
40
0
60
80
100
Time (min) Time (tnin)
Figure 5. Reverse phase HPLC of aerobic culture supernatants after 41 days of
incubation. The column was eluted with a 30-50% isopropanol gradient at
0.25 ml/min.
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In summary, we have examined both aerobic and anaerobic
degradation of the arylpolyglycols produced by KPEG reaction with 2,2',
4,4^5,5' hexachlorobiphenyl in enrichment cultures established with inocula
from a municipal waste treatment plant. Under anaerobic conditions the PEG
moiety was apparently cleaved from the chloroarylpolyglycols, resulting in
hydrophobic products. In contrast, aerobic enrichments carried out
degradation of the ethyl acetate and water soluble chloroarylpolyglycols into
two water soluble, but ethyl acetate insoluble, aromatic compounds; possibly
chlorinated arylcarboxylic acids. Our data supports the possibility of
biodegradation of PCB/KPEG products as part of a combined
chemical/biological treatment of PCBs. However, the process may be feasible
under only aerobic, and not anaerobic, conditions.
This paper has been reviewed in accordance with the U.S.
Environmental Protections Agency's peer and administrative review policies
and approved for presentation and publication.
References.
1. Reineke, W., and Knackmuss, H.-J. (1988). Ann. Rev. Microbiol. 42:263.
2. Brunelle, D.J., and D.A. Singleton. (1983). Chemosphere 12:183.
3. Brunelle, D.J., and D.A. Singleton. (1985). Chemosphere 14:173.
4. Albro, P.W., J.T. Corbett, and J.L. Schroeder. (1981). J. Chromatog. 205:103.
5. Jensen, S., and G. Sundstrom. (1974). Ambio 3:70.
6. Safe, S., Safe, L., and M. Mullin. (1985). J. Agric. Food Chem. 33:24.
7. Kornel, A., and C. Rogers (1985). J. Hazard. Mater. 12:161.
8. Stephanou, E., and W. Giger. (1982). Environ. Sci. Technol. 16:800.
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AEROBIC BIODEGRADATION OF VOLATILE ORGANIC
COMPOUNDS IN A BIOFILTER
Vivek Utgikar and Rakesh Govind
Department of Chemical Engineering
University of Cincinnati
Cincinnati, Ohio 45221
Fred Bishop
Chief, Biosystems Branch
Risk Reduction Engineering Laboratory
U.S. EPA
Cincinnati, Ohio 45268
ABSTRACT
In recent years the emission of volatile organic compounds
(VOCs) has received increased attention from EPA, OSHA and other
government agencies due to the serious human health hazards these
compounds present as pollutants. The origins of these VOCs can be
from manufacturing process or wastewater treatment plants, where
the waste stream is stripped of the VOCs during aeration. Another
significant source of these pollutants is from landfill leachate.
The conventional physical/chemical treatment methods for these
gaseous pollutants are adsorption on a solid, absorption in a
solvent, incineration or catalytic conversion. An alternative to
these conventional treatment methods is the biological destruction
of the VOCs. This method has the advantages of pollution
destruction (as compared to transfer to another medium) and lower
operation and maintenance costs.
The biodegradation can be carried out in a biofilter. A
biofilter consists of a packed column containing biologically
active mass. The biologically active matter (biomass) can exist
either as a uniform biofilm on the support medium, or as a biomass
particle trapped in the void spaces between the support material.
This paper reports on the experimental study on the
biodegradation of three volatile organic compounds in an aerobic
biofilter. The three chemicals (substrates) were studied at the
following concentrations: Toluene: 520 ppm; Methylene Chloride: 180
ppm; Trichloroethylene: 2 5 ppm. The substrates were fed to the
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biofilter through the gas phase. The requisite composition of the
substrates in the gas phase was achieved by making the synthetic
gas mixtures in a cylinder and subsequent blending with air. The
biofilter was packed with GAC support material. Nutrient solution
was circulated counter to the gas through the bed. The inlet and
outlet gas streams were analyzed for the above three chemicals.
The biofilters contained active acclimated biomass. The
results showed that nearly 100% removal of the three compounds was
achieved in the biofilter.
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MFTHANOfiFNIC DEGRADATION KINETICS OF PHFNOI IC COMPOUNDS
by: E. Michael Godsy
Donald F. Goerlitz
U.S.Geological Survey
Menlo Park, California 94025
and
Dunja GrbiC-Galid
Stanford University
Stanford, California 94305
INTRODUCTION
Microbiologists have come to appreciate the power ol the quantitative approach in their
research. It is no longer enough to simply describe the organisms that occupy a given habitat;
the rates at which they carry out metabolic functions of ecological importance must be estimated.
Only when quantitative information of metabolic activities is coupled with knowledge of
organismal types can our understanding of the concerted actions of the members of a community
be considered complete.
The quantitative approach in microbial ecology involves the estimation of parameters in
equations chosen to represent the process under study, such as substrate depletion and
concomitant growth. A major factor affecting activity and growth of microorganisms in many
environments (e.g., aquifer sediment) is related to the presence of solid surfaces in those
environments. Surfaces may alter the availability of organic chemicals, change the levels of
various organic and inorganic nutrients, and/or retain microorganisms. Bacterial cells that are
attached to subsurface materials may have physiological activities quite different from cells that
are in suspension. Most of these phenomenon have not been investigated for subsurface
environments.
The generally accepted equations describing substrate utilization and concomitant bacterial
growth without decay are the ones propose by Monod, 1949 (1):
^ = maximum specific growth rate, 1/day
Ks = half-saturation constant, mg/L - numerically equal to that substrate concentration
that yields a growth rate equal to one-half \un
Y = yield coefficient, mg cells/mg substrate utilized
S = substrate concentration at time t, mg/L
X = biomass at time t, mg/L
The above relationships were developed from experiments using pure cultures of bacteria
utilizing single organic compounds. It remains to be determined if these expressions describe the
degradation of single compounds in a complex mixture of compounds in the subsurface
environment by a complex mixed microbial population.
In this study, we present evidence that the Monod equations adequately describe both the
utilization of phenolic compounds at very low environmental concentrations (oligotrophic) and the
concomitant bacterial growth.
dS HniXS
dt =Y(Ks + S)
dX Mm*S
dt = Ks + S
Where:
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MATERIALS AND METHODS
The study site is located in Escambia County within the city of Pensacola, Florida, adjacent
to an abandoned wood preserving plant (2). The wood preserving process consisted of steam
pressure treatment of pine poles with either creosote and/or pentachlorophenol (PCP). Large but
unknown quantities of wastewaters, consisting of extracted moisture from the poles, cellular
debris, creosote, PCP, and diesel fuel were discharged to surface impoundments. The
impoundments were unlined and hydraulically in direct contact with the sand-and-gravel aquifer
The aquifer consists of deltaic deposits of fine-to-coarse graveled quartz interbeded with
discontinuous silts and clay intervals.
Microcosms used for the study were prepared in 4 L sample bottles and contained
approximately 3 kg of aquifer material collected from a depth between 5 to 6 m at a site 30 m
down gradient from the contamination source. Phenolic compounds were added to 2.5 L of
mineral salts solution (3) at concentrations similar to the aquifer concentrations, 20 to 40 mg/L.
Amorphous FeS was used as a reducing agent (4) to insure methanogenic conditions. The
microcosms were prepared, stored, and sampled in an anaerobic glove box containing an 02-free
argon atmosphere at 22'C.
Samples for substrate utilization were removed from the microcosms at approximately
three-day intervals after gentle mixing. Analyses were done by reverse-phase gradient-elution
HPLC using a UV detector set at a wavelength of 280 nm.
NONLINEAR PARAMETER ESTIMATION
Substrate depletion and bacterial growth curves were fitted to the Monod equations using
nonlinear regression analysis. The method of Marquardt (5) was used for the estimation of
parameter values that minimized the sum of the squared residuals. Because the Monod
equations do not have explicit analytical solutions for substrate and biomass concentrations as a
function of time, a simultaneous solution of both equations was accomplished using a fourth-
order Runge-Kutta numerical procedure. The statistical basis for these analyses is presented by
Robinson,1985 (6), and requires that the sensitivity of the independent variable to changes in
each of the parameters be calculable. The partial derivatives of the substrate (S) with respect to
Kj and Y satisfy this requirement. These expressions are derived from the integrated Monod
equation by implicit differentiation. Unique estimates of the parameters can be obtained when
the initial substrate concentration (S0) is in the mixed-order region and proceeds through the lirst-
order region during the course of the experiment.
RESULTS
Compound 1/day mg/L Y. mg/mg
Phenol 0.104 ±0.022 2.0016.10 0.003 ±0.003
2-Methylphenol 0.040 ±0.010 0.27 ±0.35 0.003 ±0.004
3-Methylphenol 0.122 ± 0.068 0.40 ±1.11 0.002 ±0.004
4-Methylphenol 0.095 ±0.045 1.90 ±10.1 0.052 ±0.139
Parameter estimates with the 95% confidence intervals are given in the above Table. The
time interval before the onset of rapid methanogenesis varied from 28 days for 3-methylphenol to
119 days for 2-methylphenol, even though the inoculum history suggests that all of the
microorganisms in the microcosms had been exposed to all of the phenolic compounds for
considerable length of time (~ 80 years).
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DISCUSSION AND CONCLUSIONS
Laboratory microcosms containing aquifer material simulate the same biotic and abiotic
interactions that occur at the Pensacola site. Important considerations in determining the
ultimate environmental fate of contaminants are both the adsorption of substrate and biomass to
the aquifer sediment. Studies in the laboratory and at the research site have shown that
substrate adsorption to aquifer sediments of the four phenolic compounds tested was
insignificant (Retardation Factors ranged from 1.01 for phenol to 1.10 for 4-methylphenol) and
that greater than 99% of the biomass was associated with the aquifer sediment. However, for
modeling purposes, the biomass on the sediment may be treated as if it were uniformly
distributed through out the liquid volume. These considerations justify the modeling of substrate
utilization in the microcosms as batch reactions.
The bacterial substrate utilization and growth data for all of the compounds tested could be
modeled successfully using the Monod equations. The long apparent lag times for the phenolic
compounds could be attributed to extremely tow initial biomass concentration in the microcosms.
The kinetic constants for all of the compounds are very similar and given that the inocula were
acclimated to all of the phenolic compounds, it is unclear why the range of onset times was so
great. There also appears to be a lack of correlation between the onset times and the model
parameters.
Although we know of no other kinetic studies conducted under similar conditions, it appears
that the values of the parameters obtained are reasonable and consistent with values expected of
organisms from oligotrophy environments. The extremely low Y values for the phenolic
compounds suggests that these organisms have adapted to this environment by utilizing 99+% of
the available energy lor maintaining cellular integrity or that they are very inefficient at capturing
the free energy available. This phenomenon is currently under investigation.
REFERENCES
1. Monod, J. 1949. The growth of bacteria! cultures. Annual Review of Microbiology 3:371 -394.
2. Godsy, E.M., D.F. Goerlitz, and Dunja GrbiC-Galit. 1991. Methanogenic biodegradation of
creosote contaminants in natural and simulated ground water ecosystems. Ground Water
(in journal review).
3. Zeikus, J.G. 1977. The biology of melhanogenic bacteria. Bacteriological Reviews 41:514-541.
4. Brock, T.D., and K. O'Dea. 1977. Amorphous ferrous sulfide as a reducing agent for culture of
anaerobes. Applied and Environmental Microbiology 33:254-256.
5. Bard, Y. 1974. Nonlinear parameter estimation. Academic Press, Inc., New York.
6. Robinson, J.A. 1985. Nonlinear regression analysis in microbial ecology. Advances in Microbial
Ecology 8:61-114.
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Aerobic Biodegradation of Creosote
James G. Mueller1, Suzanne E. Lantz1, Ron L. Thomas1, Ellis L. Kline1,
Peter J. Chapman2, Douglas P. Middaugh2, P. Hap Pritchard2,
Richard J. Colvin^, Allan P. Rozich^ ap50% of targeted compounds) of monitored constituents.
However, removal rates were slower in surface soil slurries was
slower than that observed with subsurface soil slurries, and was
generally confined to the more readily biodegradable, lower-
molecular-weight compounds. In all cases, solid-phase
bioremediation was much less effective. The general order of
biodegradation was phenolics>low-molecular-weight
PAHs>heterocycles>high-molecular-weight PAHs=PCP. These daia
suggest that slurry-phase bioremediation strategies can be
effectively employed to treat creosote-contaminated, and possibly
PCP-contaminated, materials. The efficiency of an integrated,
multi-phasic slurry treatment process is currently being evaluated
at the pilot-scale level.
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RESULTS AND DISCUSSION
SURFACE SOIL (SS) BIOREMEDIATION
To simplify data presentation, PAHs were arbitrarily divided
into 3 groups: groups 1, 2 and 3 consist of PAHs containing 2, 3 and
4 or more fused rings, respectively. In the absence of inorganic
nutrient supplements (SS- treatment), biodegradation of phenolic,
heterocyclic and lower-molecular-weight PAH constituents of
creosote was most readily apparent (for example, see Figure 1).
Biodegradation of the more persistent chemicals was less extensive.
Biodegradation of PCP began after a 2 week lag period ultimately
resulting in the removal of approximately 70% of this chemical over
the course of the study (90 days).
When surface soils were amended on a weekly basis with
inorganic nutrients (SS+ treatment), the rate of biodegradation of
monitored chemicals was accelerated. As observed in the
unamended soils, biodegradation of phenolics, heterocyclics and
low-molecular-weight PAHs was most rapid. Moreover, the extent
of biodegradation of the more persistent chemicals (i.e., group 3
PAHs, PCP) was increased.
Biodegradation of all monitored creosote constituents was
most rapid and extensive under slurry-phase conditions (2). Within 7
days of slurry incubation, 91, 90, 45, 47 and 30% of the phenolics,
group 1 PAHs, heterocyclics, group 2 PAHs and group 3 PAHs were
biodegraded, respectively. However, PCP was not biodegraded by
indigenous microflora established in the reactor. Little change in
the total amount of chemical biodegraded was apparent with
continued incubation (30 days). This suggests that conditions for
biodegradation became limiting (eg., nutrient limitation,
accumulation of bacteriotoxic metabolites, etcetera). Alternatively,
depletion of the readily biodegradable carbon sources (i.e., group 1
PAHs, phenolics) prevented further catabolism of monitored
chemicals.
SUBSURFACE SOIL (SBS) BIOREMEDIATION
Subsurface soils recovered from a depth of 5 m beneath the
highly contaminated, capped solidified sludge material present at
the American Creosote Works Superfund site, Pensacola, Florida
contained approximately 7% (weight basis) unweathered creosote
plus PCP (3). Initial microbial population estimates showed that
this material was essentially sterile (data not shown). This was
presumably due to the high organic loading rate and the presence of
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-73-
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fly-ash (added to stabilize the above-lying sludge) which resulted in
a soil pH of 10-11. Hence, biodegradation of these chemicals during
aerobic, solid-phase bioremediation was slow to initiate*
After one to two week lag phase, however, readily biodegradable
chemicals (i.e, phenolics, heterocycles) were removed, but a
majority of the other compounds resisted biological attack. The
addition of inorganic, soluble nutrients had little effect on the rate
and extent of biodegradation.
As observed with surface soil, biodegradation occurred more
rapidly, and was more extensive, during slurry-phase treatment than
with solid-phase treatment (2). Within 7 days of slurry incubation,
95, 90, 85, 65 and 50% of the group 1 PAHs, phenolic, group 2 PAHs,
heterocyclics and group 3 PAHs were biodegraded, respectively. As
before, PCP was not degraded by the microbial community
established in the reactor.
Based on these and other data, slurry-phase biotreatment
technologies have been integrated into a multi-phasic remediation
strategy to ameliorate soil and water contaminated with creosote,
PCP and related wastes (4). The ability of this system to remove
>90% of monitored chemicals from contaminated wastes has been
demonstrated at the bench-scale level (3, 5}. In association with
the Superfund Innovative Technology Evaluation (SITE) Program,
pilot-scale performance data is currently being generated at the
American Creosote Works Superfund Site, Pensacola, Florida.
ACKNOWLEDGMENTS
Technical Assistance was provided by Beat Blattmann, Maureen
Downey, Mike Shelton and Miriam Woods (Technical Resources, Inc).
Susan Franson (U.S. EPA, EMSL, Las Vegas NV) graciously offered a
QA/QC review of these studies. Dan Thoman (U.S. EPA, ESD, Athens,
GA) obtained subsurface soil samples and performed independent
chemical analyses. Assistance from Natalie Ellington and Beverly
Houston (U.S. EPA, Region IV) is also gratefully acknowledged.
Financial support for these studies was provided by the U.S.
EPA Superfund Program (Region IV). The on-going Pilot-Scale
Technology Demonstration Project is supported by the U.S. EPA SITE
Program, Cincinnati, OH.
This work was performed as part of a Cooperative Research
and Development Agreement between the Gulf Breeze Environmental
Research Laboratory and Southern Bio Products, Inc. (Atlanta, GA) as
defined under the Federal Technology Transfer Act, 1986 (contract
no. FTTA-003).
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REFERENCES
1. Mueller, J.G., S.E. Lantz, B.O. Blattmann and P.J. Chapman. 1991.
Bench-scale evaluation of alternative biological treatment
processes for the remediation of creosote-contaminated materials:
solid-phase bioremediation. Environ. Sci. Technol. In press.
2. Mueller, J.G., S.E. Lantz, B.O. Blattmann and P.J. Chapman. 1991.
Bench-scale evaluation of alternative biological treatment
processes for the remediation of creosote-contaminated materials:
slurry-phase bioremediation. Environ. Sci. Technol. In press.
3. Mueller, J.G., S.E. Lantz, B.O. Blattmann and P.J. Chapman. 1990.
Alternative biological treatment processes for remediation of
creosote-contaminated materials: bench-scale treatability studies.
EPA/600/9-90/049. 103 p.
4. Mueller, J.G., P.J. Chapman, R. Thomas, E.L. Kline, S.E. Lantz and
P.H. Pritchard. 1990. Development of a sequential treatment system
for creosote-contaminated soil and water: bench studies.
Proceedings U.S. Environmental Protection Agency's Symposium on
Bioremediation of Hazardous Wastes: U.S. EPA's Biosystems
Technology Development Program. EPA/600/9-90/041, pages 42-45.
5. Middaugh, D.P., J.G. Mueller, R.L. Thomas, S.E. Lantz, M.J. Hemmer,
G.T. Brooks and P.J. Chapman. 1991. Detoxification of creosote- and
PCP-contaminated groundwater: chemical and biological assessment.
Arch. Environ. Contam. Toxicol. In press.
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ANAEROBIC DEGRADATION OF HIGHLY CHLORINATED
DIOXINS AND DIBENZOFl JRANS
by : Peter Adriaens and Dunja Grbic-Galic
Department of Civil Engineering
Stanford University
Stanford, CA 94305
ABSTRACT
Hudson River sediments (HR), Pensacola Soil (PS) and chlorophenol adapted
Cherokee Pond sediments (ACP) were anaerobically incubated with 50 ng/L of the
following dioxin and dibenzofuran congeners: 1,2,3,4,6,9-hexa (HexaCDD),
1,2,4,6,8,9/1,2,4,6,79-hexa isomer (HexaCDDi), and 1,2,3,4,6,7,9-hepta
(HeptaCDD) chlorinated dioxins, and 1,2,4,6,8-penta (PentaCDF) and 1,2,3,4,6,7,8-
hepta (HeptaCDF) chlorinated dibenzofurans. The initial analytical data from the first
eight weeks of incubation suggest that HR sediments exhibit a more extensive activity
towards the congeners than the PS inoculum. No data are available yet for ACP
sediments. Substrate disappearance was observed for pentaCDF (35%), hexaCDDi (<10
%), and heptaCDD (10 %) in one or more replicates of HR incubations. No detectable
levels of intermediates have been observed, except for a small amount (3-4 p.g/L) of
hexaCDD obtained from the incubation with heptaCDD. This intermediate was identified
by GC/MS, however, isomer assignment was impossible due to the extremely low
concentration levels. Although the initial results do not show extensive activity against
the congeners tested, the presence of a lower chlorinated metabolite may be the first
indication that highly chlorinated dioxin and dibenzofuran isomers could serve as
alternative electron sinks, as is observed for polychlorinated biphenyls (PCBs).
This paper has been reviewed in accordance with the U.S. Environmental Protection
Agency's peer and administrative review policies and approved for presentation and
publication.
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EXPERIMENTAL PROCEDURES
EXPERIMENTAL SET-UP AND SAMPLE PREPARATION
Three replicates of microcosms {50 mL total liquid volume, 50 ± 2 g solids) were spiked
with 100 ;iL of the respective dioxin and dibenzofuran from 50 mg/L (except for HexaCDDi, 5
mg/L) nonane stock solutions to give a final concentration of 50 ng/L (5 ^g/L for HexaCDDi).
In addition, duplicate killed (autoclaved) biological controls, live biological controls without
PCDD or PCDF, and chemical controls without inocula, have been established and were
monitored along with ihe cultures.
All bottles were manually shaken, decapped, and sampled (5 ml) with a 10-mL glass
syringe to contain both sediment or soil and aqueous phase, under a continuous stream of
nitrogen in the headspace. Two volumes of hexane/acetone (9:1) and 0.5 ^g of
octachloronaphthalene (as an internal standard) were added to each sample, which was then
shaken overnight on a wrist-action shaker. The extraction solvent was decanted from the soil,
sediment or aqueous phase, and extracted with 2 ml of concentrated H2SO4. The extract was then
back-extracted with 2 ml of a 2% CaCl2 solution (in distilled water), and dried over Na2S04.
The resulting extract was eluted over a Pasteur pipette packed with Florisil (60 mesh)/Cu-
powder (40 mesh) (1: 4 ratio), to remove excess sulfate. The sample was then concentrated to
1 ml under a constant stream of N2. To this fraction, 100 nL of dodecane was added. The sample
was then further concentrated to 100 ^L under a gentle stream of N2 and used for GC/MS and
GC-ECD analyses.
ANALYTICAL PROCEDURES
The samples were analyzed both on a Triple Stage Quadrupole TSQ 70 Finnigan MAT GC/MS,
and on a 5890A Hewlett-Packard Gas Chromatograph equipped with an electron capture detector
(ECD).
GC/MS operating conditions : Column: DB-5, 60 m, 0.32 um I.D., 0.25 |im film
thickness; column head pressure: 25 kPa; injection: on-column; injector temperature.
60 °C; initial temp.: 90 °C, hold 5 min; rate 1: 25 °C min-1; temp. 2: 200 °C, hold 15
min; rate 2: 4 °C min'1; temp. 3: 250 °C, hold 15 min.
Gas Chromatographic conditions : Column: DB-5, 30 m, 0.32 ^im I.D., 0.25 film
thickness; carrier gas: helium (linear flow velocity: 25 cm s"1); make-up gas:
argon/methane; column head pressure: 14 kPa; injection: split/splitless (10:1 ratio): injector
temperature: 250 °C; detector temperature: 275 °C; initial temperature: 250 °C, hold 5 mm:
rate 1: 2 °C min-1; final temperature: 300 °C, hold 5 min. The high initial temperature caused
the Aroclor to elute in the first 10 min, without interfering with the dioxin (25.8 min) and
dibenzofuran isomers (penta: 20 min, hepta: 24 min). The internal standard
(ociachloronaphtalene) eluted after 19 minutes.
RESULTS AND DISCUSSION
Table 1 represents the time zero samples, and Table 2 shows the results from samples taken
after 2 months. The recovery efficiencies of all isomers from the chemical controls always
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exceeded 92 %, while they varied between 20 and 45 % from the samples containing inoculum.
The low value obtained for the killed control of penta CDF incubations (Table 1) cannot be
ascribed to low recovery efficiencies, as the data are corrected for sorption with the internal
standard. Hence, they have to be explained either by incorrect spiking with the furan, or a
volatilization loss when these controls were autoclaved twice. To prevent these kinds of losses,
killed controls of the Cherokee Pond incubations were spiked with 0.5 ml of concentrated
H2S04, instead of twice autoclaved.
Table 1 : Time zero analysis of sediment and soil incubations with five dioxin and
dibenzofuran isomers. All concentrations [
-------
abundances in the molecular ion cluster agreed well with those of a hexaCDD standard
(1,2,3,4,6,9 hexa CDD) and published values (Kleopfer et al.. 1989).
1 8#
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Fig. 1: SIM trace (m/z 380 to 410) of HR incubation with heptaCDD after two months. Scans
for M* (m/z 390), M+-2+ (m/z 38B) and M++2+ (m/z 392) are shown.
Except for the presence of the m/e 327 ion, which was out of scan range, the relative abundance
percentages of the three main ions (Fig. 2)were well within the U.S.EPA acceptance criteria for
isomer identification (Alford-Stevens et al., 1986). Since background ion abundance levels
interfered strongly with the peak identification, a total extract of this replicate will be analyzed
to conclusively identify this ion cluster as a hexa CDD.
DU. «
• »». t
*«». 4
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m/z U.S EPA Acceptance
Criteria
Obtained ten
Abundances
388 > 41% and <61 %
3 9 0 1 0 0 %
392 > 71 % and < 91 %
327 Present and < 40 %
57 %
1 0 0 %
81 %
Out of scan range
Fig.2: Mass spectrum and relalive ion abundances of the hexaCDD metabolite, after background
noise substraction.
REFERENCES
1. Kleopfer, R. D., R. L. Greenall, T. S. Viswanathan, C. J. Kirchmer, A. Gier, and J. Muse.
1989. Determination of polychlorinated dibenzo-dioxins and dibenzofurans in
environmental samples using high resolution mass spectrometry. Chemosphere 18:
109-118
2. Alford-Stevens. A. L., J. W. Eichelberger, T. A. Bellar, and W. L. Budde. 1986.
Determination of chlorinated dibenzo-p-dioxins and dibenzofurans in soils and
sediments by gas chromatography/mass spectrometry. EPA Report: Physical and
Chemical Methods Branch.
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DEGRADATION OF NAPHTHALENE, PAHS, AND HETEROCYCLICS
Richard W. Eaton and Peter J. Chapman
US EPA, Environmental Research Laboratory
Gulf Breeze, FL 32561
Naphthalene is the simplest fused polycyclic aromatic hydrocarbon. Information
obtained from studies of its bacterial degradation may be used in understanding and
predicting the pathways used in the metabolism of more complex polycyclic aromatic
hydrocarbons and structurally-related heterocyclic aromatic compounds.
In spite of the relative simplicity of naphthalene, much about its bacterial
metabolism has remained unclear, particularly the steps in the metabolic pathway by
which 1,2-dihydroxynaphthalene (DHN, Fig. 1, I) is metabolized to salicylaldehyde.
This is primarily because of the chemical instability of various chemical intermediates
implicated or identified in this pathway. Thus DHN is rapidly and spontaneously
oxidized in water to 1,2-naphthoquinone and the potential ring-cleavage products, cis-
o-hydroxybenzalpyruvate (Fig. 1, IV), trans-o-hydroxybenzalpyruvate (Fig. 1, III), and 2-
hydroxychromene-2-carboxylate (Fig. 1, V) ail undergo isomerizations in water.
Davies and Evans (2) identified a product of the oxidation of DHN by cell
extracts of a naphthalene-grown Pseudomonas strain as o-hydroxybenzalpyruvate
(HBPA), initially isolated as its perchlorate, and suggested that it was probably the cis-
isomer based on its chemical properties. Cell extracts metabolized both cis- and
trans- isomers to salicylaldehyde. The cjs- isomer of HBPA was spontaneously
converted at neutral pH to its hemiketal, 2-hydroxychromene-2-carboxylate (HCCA),
which was not metabolized by cells or cell extracts and was considered to be an
artifact.
Barnsley (1) subsequently demonstrated that, instead of being an artifact, HCCA
was the initial product of the enzymatic ring cleavage of DHN. He did this by
incubating cell extracts, and, also purified dihydroxynaphthalene dioxygenase (4), with
1,2-dihydroxynaphthalene at pH 5.5. (At this pH the autoxidation of DHN to 1,2-
naphthoquinone is somewhat reduced). After two minutes the incubation mixture was
applied to and eluted from a column of Sephadex G-25 in order to separate the large
protein components of the cell extracts from lower molecular weight reaction products
which were collected and freeze-dried. In this way a small amount of a single
chemical, 2-hydroxychromene-2-carboxylate, was obtained and it was proposed that
this was the initial ring cleavage product. An HCCA-metabolizing enzyme (isomerase)
which catalyzed the conversion of that compound to trans-HBPA at pH 10 (Km = 0.2
mM) was also demonstrated.
Both of these studies were hampered by their use of 1,2-dihydroxynaphthalene
as substrate. Incubations required large amounts of enzyme and yielded product only
in quantities that were insufficient to identify rigorously.
Accordingly, a way to identify the DHN ring cleavage product was adopted that
involved cloning the genes encoding the first three enzymes of the pathway away from
genes encoding enzymes catalyzing subsequent steps. Bacteria carrying these genes
would be able to transform the stable substrate, naphthalene, to the ring-cleavage
products which could then be prepared on a large scale and identified. Such clones
would also be useful for the preparation of analogous pathway intermediates from
other hydrocarbons and heterocyclics.
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Figure 1
1,2-Dihydroxynaphthalene dioxygenase (nahC product)
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11 kb NAH7 pJasmid fragment encoding the metabolism of naphthalene to salicylate
nah genes A B F C D E
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The NAH7 plasmid from Pseudomonas putida G7 carries genes encoding the
complete degradation of naphthalene. These genes are grouped in two operons (5,6).
The operon encoding the metabolism of naphthalene to salicylate has been cloned in
the plasmid vector pMMB277 on an 11 kb EcoRI-Hindlll fragment (Fig. 2). Several
deletions and subclones have been obtained that eliminate DNA between nahC (the
1,2-dihydroxynaphthalene dioxygenase gene located 2.15 to 3.05 kb from the Hindlll
site (3)) and the Hindlll site and which inactivate the gene encoding the enzyme that
degrades the product of 1,2-dihydroxynaphthalene cleavage. One of these subclones
is a 10 kb EcoRI-Clal fragment inserted in pMMB277. Pseudomonas aeruginosa PA01
carrying this plasmid transforms naphthalene to the ring-cleavage products which
accumulate. These products have been separated by chromatography on Sephadex
G-25 and identified as trans-o-hydroxybenzalpyruvate and 2-hydroxychromene-2-
carboxylate (the hemiketaf of ds-o-hydroxybenzalpyruvate) by NMR and mass
spectrometry.
Both ds- and trans- isomers are probably formed by the spontaneous
rearomatization of an unstable ring-cleavage product (Fig. 1, II). They are both rapidly
degraded to salicylaldehyde by cell extracts of strain PA01 carrying the cloned EcoRI-
Hindlll fragment. On addition of NAD' salicylaldehyde is oxidized to salicylate.
2-Hydroxychromene-2-carboxylate could have been rapidly formed from the
accumulating ds- isomer by nucleophilic attack of the ortho-hvdroxvl oxygen on the
carbonyl carbon. This hemiketal is not metabolized at neutral pH by cell extracts of
PA01 carrying the cloned EcoRI-Hindlll fragment. In wild-type naphthalene-degrading
bacteria, the ds- isomer is probably metabolized before the formation of the hemiketal
can occur.
REFERENCES
1. Barnsley, E.A. 1980. Naphthalene metabolism by pseudomonads: the
oxidation of 1,2-dihydroxynaphthalene to 2-hydroxychromene-2-carboxylic acid
and the formation of 2'-hydroxybenzalpyruvate. Biochem. Biophys. Res.
Commun. 72:1116-1121.
2. Davies, J.L. and W.C. Evans. 1964. Oxidative metabolism of naphthalene by
soil pseudomonads. The ring-fission mechanism. Biochem. J. 91:251-261.
3. Harayama, S., and M. Rekik. 1989. Bacterial aromatic ring-cleavage enzymes
are classified into two different gene families. J. Biol. Chem. 264:15328-15333.
4. Patel, T.R. and E.A. Barnsley. 1980. Naphthalene metabolism by
pseudomonads: Purification and properties of 1,2-dihydroxynaphthalene
oxygenase. J. Bacterid. 143:668-673.
5. Yen, K.-M. and I.C. Gunsalus. 1982. Plasmid gene organization:
Naphthalene/salicylate oxidation. Proc. Natl. Acad. Sci. USA 79:874-878.
6. Yen, K.-M. and C.M. Serdar. 1988. Genetics of naphthalene catabolism in
pseudomonads. CRC Crit. Revs. Microbiol. 15:247-268.
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DEGRADATION OF HALOGENATED ALIPHATIC COMPOUNDS BY
THE AMMONIA-OXIDIZING BACTERIUM NITROSOMONAS EUROPAEA
Todd Vanelli, Peter Chapman" and Alan B. Hooper
University of Minnesota, St. Paul, MN 55108 and
'U.S. EPA, Environmental Research Laboratory,
Gulf Breeze, FL 32561
The ubiquitous soil-, marine- and freshwater-dwelling ammonia-oxidizing
nitrifying bacteria are obligate chemolithoautotrophic aerobes. They depend for growth
on the activity of the enzyme ammonia monoxygenase (AMO); 2 H+ + 2 e' + 02 + NH3
-> NH2OH + H20. The two electrons for the AMO reaction originate in the subsequent
reaction catalyzed by hydroxylamine oxidoreductase (HAO); H20 + NH2OH -> 4 e' + 5H+
+ N02-(1).
Nitrosomonas, like the methylotrophs (2), is capable of the oxidation of many
organic compounds. In addition, our laboratory at Minnesota (3,4) and the laboratory of
Arp at Corvallis (5) have observed degradation of many halogenated aliphatic compounds
as summarized in Table I.
In no case is there evidence that oxidation of an organic substrate will support
growth. Response to inhibitors of AMO (2-chloro-6-trichloromethyl pyridine
("nitrapyrin"), acetylene or aa'-dipyridyl) and the fact that the organic co-oxidized
substrate inhibits ammonia oxidation indicates that degradation is catalyzed by AMO. For
all compounds the concomitant oxidation of ammonia is required for degradation of
halogenated hydrocarbons. Depending on the substrate, hydroxylamine and/or hydrazine
(a non-biological substrate for HAO) can also serve as electron donor for degradation in
cells. Low levels of degradation are observed for a short time in the absence of added
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Table 1
Substrates*
Products
Haloalkanes:
Bromomethaneb
Formaldehyde
Chloromethanec
Formaldehyde
Dibromomethane
Dichloromethane
Trichloromethane
T etrachlorome thaned
Bromoe thane
Chloroethanec
Acetaldehyde, 2-ChloroethanoI
Fluoroethanec
Iodoethanec
1,2-Dibromoe thane
1,1,1 -Trichloroethane
1,1,2-Trichloroethane
Chloropropanec
Propionaldehyde, 3-Chloro-l-propanol,
1 -Chloro-2-propanol
1,2,3-Trichloropropane
Chlorobutanec
Butyraldehyde, 4-Chloro-l-butanol
Haloalkenes:
Chloroethylene
cis 1,2-Dibromoethylene
trans l,2-Dibromoethylened
1,1 -Dichloroethylene
cis 1,2-Dichloroethylene
trans l,2-Dichloroethylened
Trichloroethylene
Tetrachloroethylened
1,3-Dibromopropene
2,3-Dichloropropene
1,1,3-Trichloropropene
Nitrapyrin
a. All compounds reported in reference 4 except as indicated
b. See ref. 1 and 2.
c. See ref. 5 -86-
d. Tested but no degraded
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electron-donating co-substrate; it is presumed but not known that an endogenous electron
donor is involved.
Clearly Nitrosomonas is able to degrade a wide spectrum of halogenated aliphatics.
The exceptions are tetrachloroethylene and carbon tetrachloride. The observation that the
soluble methane oxidizing enzyme from methylotrophs will oxidize fluorotrichlorethylene
(6) suggests that the absence of a C-H bond need not be limiting.
The product of oxidation of some substrates by Nitrosomonas, possibly an epoxide
intermediate, forms an irreversible derivative of AMO and thus inactivates the enzyme.
Reaction with acetylene, the best example of this kind of compound, results in
derivatization of a specific membrane polypeptide (7). Nevertheless, the oxidation of
most halogenated compounds can continue for days provided that the concentration of
ammonia remains high enough.
We have recently demonstrated the ammonia-dependent degradation of 1,3
dibromopropene, 1,1,3 trichloropropene, 2-chloro-6-trichloromethyl pyridine nitrapyrin and
2,3 dichloropropene by Nitrosomonas. The rates were 4.0, 1.6, 5.8 and 29 /*moles hr'1
wet weight'1. Halogenated substrates were measured by electron capture detector after gas
chromatography. At a concentration of 30-40 ^M, the first three compounds completely
and irreversibly inhibit oxidation of ammonia. Ring ,4C-labelled nitrapyrin derivatizes all
membrane proteins equally. Thus the reactive product of oxidation of nitrapyrin appears
to be membrane soluble and long lived.
Bioremediation: In nature or in pollution treatment, actively nitrifying
Nitrosomonas would appear to have a potential role in the degradation of halogented
aliphatic compounds.
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References.
1. Hooper, A.B. (1989). Biochemistry of the Nitrifying Lithoautotrophic Bacteria.
In Autotrophic Bacteria (H.G. Schlegel and B. Bowien, Eds) pp 239-265 Sci.
Tech. Publishers, Madison, WI.
2. Bedard, C. and Knowles, R. (1989). Physiology, Biochemistry and Specific
Inhibitors of CH4, NH„+ and CO Oxidation by Methylotrophs and Nitrifiers.
Microbiol. Rev. 53: 68-64.
3. Arciero, D., Vannelli, T., Logan, M. and Hooper, A.B. (1989). Degradation of
Trichloroethylene by the Ammonia-Oxidizing Bacterium Nitrosomonas europaea.
Biochem. Biophys. Res. Commun. 159: 640-643.
4. Vannelli, T., Logan, M., Arciero, D.M. and Hooper, A.B. (1990). Degradation of
Halogenated Aliphatic Compounds by the Ammonia-Oxidizing Bacterium
Nitrosomonas europaea. Appl. Envt. Microbiol. 56: 1169-1171.
5. Rasche, M.E., Hicks, R.E., Hyman, M.R. and Arp. D.J. (1990). Oxidation of
Monohalogenated Ethanes and n-chlorinated Alkanes by Whole Cells of
Nitrosomonas europaea. J. Bacteriol. 172: 5368-5373.
6. Fox, B.G., Bomeman, J.G., Wackett, L.P. and Lipscomb, J.D. (1990). Haloalkene
Oxidation by the Soluble Methane Monoxygnease from Methylosinus
trichosporium OB3b: Mechanistic and Environmental Implications. Biochem. 29:
6419-6427.
7. Hyman, M.R. and Wood, P.M. (1985). Suicidal Inactivation and Labeling of
Ammonia Monooxygenase by Acetylene. Biochem. J. 227: 719-725.
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Degradation of Chlorinated Aromatic Compounds
Under Sulfate-Reducing Conditions
Patricia J. S. Colberg
Department of Molecular Biology, University of Wyoming, Laramie, WY
and
John E. Rogers
U. S. EPA Environmental Research Laboratory, Athens, GA
Despite recent progress made in describing microbial transformations that occur under
anaerobic conditions, our understanding of the role sulfate-reducing bacteria may play in
the remediation of environmental contaminants is still very much in its infancy. The
aromatic nucleus is the second most common naturally-occurring organic residue in the
biosphere and is the basic structural unit of innumerable anthropogenic materials that are
produced in enormous quantities all over the world. Yet the first description of an aromatic
compound being used by a sulfate reducer was not published until 1980 (Widdel, 1980).
The number of aromatic substrates known to be amenable to microbial transformations
under sulfidogenic conditions is still small, but includes a rapidly growing list of
nonhalogenated compounds. Some of these transformations are catalyzed by pure cultures
of sulfate-reducing bacteria, though none of the metabolic pathways have been elucidated.
The number of halogenated aromatic compounds reportedly susceptible to microbial
transformations under sulfate-reducing conditions is as yet limited to the five chlorinated
phenols whose structures are shown below. Results from several laboratory studies have
suggested that sulfate may inhibit the anaerobic degradation of chloroaromatic compounds
by preventing dehalogenation (Gibson and Suflita 1986, 1990; Kohring et al., 1989;
Genthner et al., 1989; Kuhn et al., 1990). The mechanism of this apparent sulfate
inhibition of dehalogenation is unclear. Some investigators have suggested that the
inhibitory effect of sulfate may be caused by competition for electron donor between the
sulfate-reducing bacteria and the organism(s) responsible for dehalogenation; that is, the
sulfidogens are able to outcompete dehalogenators for available hydrogen (Gibson and
Suflita, 1986; Suflita et al., 1988; Kuhn et al., 1990). It is interesting to note, however,
that sulfate does not inhibit dehalogenation by Desulfomonile tiedjei (DeWeerd et al., 1990;
Linkfield and Tiedje, 1990), the only known obligately anaerobic dechlorinating bacterium
which also happens to be a sulfidogen.
2-chlorophenol 3-chlorophenol 4-chlorophenol 2,4-dichlorophenol 2,6-dichlorophenol
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Based perhaps, in pan, on the rather consistently observed disparity in transformation
potential between the sulfate-reducing and methanogenic regions within their shallow
aquifer study site, Kuhn and his co-workers (1990) have speculated that the anaerobic
dehalogenation potential might be lower in environments that maintain a higher redox
potential (i.e., denitrifying and sulfidogenic), especially for less highly halogenated
aromatic compounds. Vogel et al. (1987) have proposed a similar explanation for the
anaerobic dehalogenation of aliphatic compounds.
Despite the evidence linking sulfate with inhibition of reductive dehalogenation, it is worth
noting that several of the reports of this phenomenon are from the same laboratory and used
material from the same shallow aquifer (Gibson and Suflita 1986, 1990; Kuhn et al.,
1990), so it is perhaps premature to make any generalizations. Based on the results of a
recent study in which chlorophenol degradation occurred during sulfidogenesis (Kohring et
al., 1989) and another in which haloaromatic degradation was shown to be coupled to
sulfate reduction (Haggblom and Young, 1990), it is possible that sulfate inhibition of
dehalogenation may be a site-specific characteristic.
The objectives of this research are to:
(1) develop sulfidogenic cultures or consortia that are able to dehalogenate
chlorinated benzenes, phenols, benzoates, anilines, and biphenyls
(2) determine the activity of the cultures obtained over a range of environmental
conditions
(3) evaluate the substrate specificity of the cultures towards all of the compounds
classes
(4) evaluate in microcosm experiments the ability of these cultures to enhance the
degradation of hazardous chlorinated aromatic compounds in contaminated soils
and sediments
This report will summarize some of the work in progress aimed at evaluating sewage
sludge and several freshwater sediments for sulfidogenic activity that results in microbial
transformations of a selected number of chlorinated aromatic compounds.
References Cited
DeWeerd, K. A., L. Mandelco, R. S. Tanner, C. R. Woese, and J. M Suflita. 1990.
Desulfomonile tiedjei gen. nov. and sp nov., a novel anaerobic dehalogenating,
sulfate-reducing bacterium. Arch. Microbiol. 154:23-30.
Genthner, B. R. Sharak, W. A. Price II, and P. H. Pritchard. 1989. Anaerobic
degradation of chloroaromatic compounds under a variety of enrichment conditions.
Appl. Environ. Microbiol. 55:1466-1471.
Gibson, S. A., and J. M. Suflita. 1986. Extrapolation of biodegradation results to
groundwater aquifers: reductive dehalogenation of aromatic compounds. Appl.
Environ. Microbiol. 52:681-688.
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Gibson, S. A., and J. M. Suflita. 1990. Anaerobic biodegradation of 2,4,5-
trichlorophenoxyacetic acid in samples from a methanogenic aquifer: Stimulation of
short-chain organic acids and alcohols. Appl. Environ. Microbiol. 56:1825-1832.
Haggblom, M. M., and L. Y. Young. 1990. Chlorophenol degradation coupled to sulfate
reduction. Appl. Environ. Microbiol. 56:3255-3260.
Kohring, G.-W., X. Zhang, and J. Wiegel. 1989. Anaerobic dechlorination of 2,4-
dichlorophenol in freshwater sediments in the presence of sulfate. Appl. Environ.
Microbiol. 55:2735-2737.
Kuhn, E. P., G. T. Townsend, and J. M. Suflita. 1990. Effect of sulfate and organic
carbon supplements on reductive dehalogenation of chloroanilines in anaerobic
aquifer slurries. Appl. Environ. Microbiol. 56:2630-2637.
Linkfield, T. G., and J. M. Tiedje. 1990. Characterization of the requirements and
substrates for reductive dehalogenation by strain DCB-1. J. Ind. Microbiol. 5:9-
15.
Suflita, J. M., S. A. Gibson, and R. E. Beeman. 1988. Anaerobic biotransformations of
pollutant chemicals in aquifers. J. Ind. Microbiol. 3:179-194.
Vogel, T. M., C. S. Criddle, and P. L. McCarty. 1987. Transformations of halogenated
aliphatic compounds. Environ. Sci. Technol. 21:722-736.
Widdel, F. 1980. Anaerober Abbau von Fettsauren und Benzoesaure durch neu isolierte
Arten sulfat-reduzierender Bakterien. Doctoral Dissertation, University of
Gottingen.
Acknowledgement
This project is supported by the EPA Environmental Research Laboratory, Athens, GA
through Cooperative Agreement CR-816398-01-0.
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ABSTRACT
Ring-Fission of Polycyclic Aromatic Hydrocarbons by White Rot
Fungi
Kenneth E. Hamiriel
State University of New York
College of Environmental Science & Forestry
Syracuse, NY 13210
Polycyclic aromatic hydrocarbons (PAH) are major
pollutants of both anthropogenic and natural pyrolytic
origin, occurring in soils, sediments, and airborne
particulates. The crucial step in their biodegradation is
oxidative fission of the fused aromatic ring system, an event
previously thought unique to certain bacteria. Recent
evidence necessitates a revision of this view: the lignin-
degrading fungi that cause white rot of wood have also been
shown to mineralize a wide variety of aromatic pollutants,
including certain PAH, under culture conditions that promote
the expression of ligninolytic metabolism. A key component
of the fungal ligninolytic system is thought to consist of
extracellular lignin peroxidases (LiPs), which have been
shown to catalyze the one-electron oxidation of various
lignin-related substrates. LiPs have also been shown to
oxidize certain PAH and other aromatic pollutants in vitro,
and it has been proposed that these enzymes play an important
role in fungal xenobiotic metabolism. However, it has never
been demonstrated that any PAH is oxidized by LiP in vivo, or
that the products of such a reaction are subsequently cleaved
to smaller, monocyclic compounds. To address these
questions, we have examined the fate of one PAH, anthracene,
in cultures of the ligninolytic basidiomycete Phanerochaete
chrysosporium.
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AC is the simplest PAH to be a LiP substrate. We found
that it was oxidized to a single end product by both crude
and purified preparations of the enzyme, and that this
product was indistinguishable from 9,10-anthraquinone (AQ)
when subjected to thin-layer chromatography (TLC) on silica
or gas chromatography/electron impact mass spectrometry.
Since other PAH that have been examined in detail give
mixtures of products when oxidized by LiP, we concluded that
AC was the PAH most likely to yield diagnostic metabolites in
fungal cultures, and selected it for further studies. In
fungal cultures, [14Ci-4, 4a, 9a] and 114Cphenyl ] AQ were
mineralized to the same extent, with 13.4 ± 3.9% of AQ, and
12.9 ± 1.3% of AC, oxidized to CO2 in 14 days. Moreover, the
cultures rapidly oxidized AC to AQ. The quinone was the
predominant neutral AC metabolite found when the culture
medium was analyzed by reversed-phase high performance liquid
chromatography (HPLC) or TLC on silica, and an isotope
dilution experiment done on the extracellular medium and
mycelium showed that AQ accounted for 38% of the AC
originally added after 48 h in culture. The abiotic
oxidation of AC in uninoculated cultures gave only 1%
conversion to AQ in this time. These results support a role
for LiP in AC oxidation by Phanerochaete. and show that the
pathway AC —* AQ —> CO2 is quantitatively important in AC
metabolism by the fungus.
Analysis of the acidic metabolites formed from AC and AQ
by E. chrysosporium showed that both compounds were cleaved
to phthalic acid. The identification of the ring-fission
metabolite as phthalate was based on three findings: (1) it
was indistinguishable from authentic phthalic acid by ion
exclusion HPLC, (2) it recrystallized with authentic phthalic
acid to constant 14C specific activity in an isotope dilution
experiment, and (3) after treatment with diazomethane, it was
indistinguishable from authentic dimethyl phthalate by TLC on
silica. The isotope dilution experiment showed that both AC
and AQ were accumulated as phthalic acid in 12% yield after 7
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days in culture. Phthalic acid was not a dead end metabolite
in fungal cultures: it also was mineralized, but at only
about one-third the rate that AC and AQ were. The relative
persistence of phthalate in the cultures probably explains
our success in identifying it as an intermediary metabolite,
and the bulk of AC/AQ mineralization is presumably due to
further degradation of the moiety that is cleaved from AQ to
give phthalate. Our results show that the pathway AC —» AQ —»
phthalate is a major one in AC ring-fission by Phanerochaete.
This fungal pathway clearly differs from the classical
bacterial one, which proceeds AC —> AC-cis-1,2-dihvdrodiol —»
salicylate. It is noteworthy that the principal oxidized
products to accumulate from AC, namely AQ and phthalic acid,
can both be degraded by Phanerochaete if they are given to
freshly ligninolytic cultures. This result shows that the
cessation of organopollutant mineralization activity after 2-
4 weeks that is generally observed in £. chrysosporium
cultures is not due in every case to the accumulation of
recalcitrant products that the fungus cannot further
metabolize, but is rather an artifact of laboratory culture
conditions. A principal direction for future work,
accordingly, should be the development of methods to prolong
biodegradative ability in fungal cultures.
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AEROBIC BIODEGRADATION OF POLYCHLORINATED BIPHENYLS:
GENETIC AND SOIL STUDIES
Frank J. Mondello and Bruce D. Erickson
Biological Sciences Laboratory
General Electric Corporate Research and Development
Schenectady, New York
A practical process for aerobic bioremediation of PCB containing
soil is dependent upon isolating or developing suitable organisms.
Two of the most important characteristics for such organisms are (1)
high levels of degradative activity against many PCB congeners, and
(2) the ability to survive on the soil long enough for significant
PCB degradation to occur. Recombinant DNA technology is currently
being used to develop bacterial strains with these and other
desirable properties and to study the genes involved in PCB
degradation from a strain of Pseudomonas designated LB400.
Pseudomonas sp. strain LB400 is able to degrade a wide variety
of PCBs but has many characteristics which make it unattractive for
use in a bioremediation process. One serious shortcoming is that the
organism has been reported to lose viability very rapidly on soil.
This would make it necessary to add organisms frequently to a site or
reactor, thus making a process less feasible. Previous studies have
shown that Escherichia coli stain FM4560 (a genetically modified
organism containing the LB400 bph A, B, and C genes), degrades PCBs
nearly as well as LB400 without exhibiting many of its undesirable
traits. For instance, FM4560 had better survivability than LB400 in
laboratory media containing PCBs and did not require growth on
biphenyl for high levels of degradative activity.
The abilities of FM4560 to survive and to degrade PCBs on soil
were compared with those of LB400, to determine if the recombinant
strain is potentially more useful than the naturally occurring
organism. Soil survivability was examined using PCB contaminated
material from a site in Glens Falls, New York. This material
contained approximately 550 parts per million of highly evaporated
Aroclor 1242 and therefore appeared similar in composition to Aroclor
1248. A series of 2 dram vials containing dragstrip soil were
inoculated with either FM4560 or LB400, sealed, and incubated at 23°
C without shaking. Colony formation on selective media was used to
measure cell survival. The presence of active bph genes in LB400 was
determined by growth on biphenyl, while bph gene activity in FM4560
colonies was demonstrated by their ability to produce yellow meta-
cleavage product when sprayed with an ether solution of biphenyl or
2,3-dihydroxybiphenyl.
Similar survival curves were obtained for both FM4560 and LB400.
Early time points showed significant increases in cell number
presumably resulting from growth on stored intracellular nutrients.
After 72 hours cell numbers returned to their original level and
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continued to decrease such Chat by 8 days approximately 20% of che
cells remained culturable. After 28 days of incubaCion, the number
of culturable cells was 2% of the initial value.
The presence of active bph genes in the surviving cells was
examined at each time point. For LB400 the number of viable cells
unable to grow using biphenyl remained relatively stable at
approximately 2.5% for the first 8 days and then increased to A. 2%
for the remainder of the experiment. All of the FM4560 colonies
contained active bph genes.
The ability of E. coli strain FM4560 to degrade PCBs on soil
from the Glens Falls site was compared with that of Pseudomonas sp.
strain LB400. All experiments were conducted in sealed, sterile, 2
dram glass vials containing 0.1 gram of non-sterile soil and 1.0 ml
of bacterial culture. Under most of the experimenCal conditions, PCB
degradation by strain FM4560 was significantly greater than that by
LB400.
ANALYSIS OF THE LB400 BPHA GENE
Biphenyl/PCB dioxygenase (encoded by the bphA gene), is the
enzyme primarily responsible for PCB degradation. Knowledge of the
structure, function, and regulation of bphA are therefore crucial for
the development of genetically modified bacteria with superior PCB-
degrading abilities. The nucleotide sequence of the gene(s) for
biphenyl/PCB dioxygenase from strain LB400 has been obtained. This
sequence was compared with that of toluene dioxygenase from
Pseudomonas putida strain F1, a multi-component enzyme made up of 4
distinct subunits. The genes for these subunits are co-transcribed
and arranged in the order todCl (1SPT0L large subunit), todC2 (ISPI0l
small subunit), codB (Ferrodoxin^L), and todA (ReductaseT0L) ,
Computer analysis of the bphA nucleotide sequence identified 4 open
reading frames whose DNA and protein sequences were similar to those
encoding the 4 subunits of toluene dioxygenase. For example, 67.5%
identity in the nucleotide sequence and 65.5% identity in the protein
sequence between todCl and the putative ISP large subunit of
biphenyl/PCB dioxygenase. Sequences upstream of the identified
homologous coding region showed no significant homology. These data
suggest a relationship between the dioxygenase responsible for
biphenyl/PCB degradation and those used to degrade a variety of other
aromatic hydrocarbons including toluene and benzene.
Transcription of the bphA region is being examined using the
technique of SI nuclease mapping. Two mRNA 5' ends are visible that
map to the area immediately before the first bphA coding region. An
additional transcript has been detected that maps at least 1,500 base
pairs before this coding region. The start site of this transcript
has not yet been determined. RNA 5' ends can arise either from
transcription initiation or an RNA processing event. To
differentiate between these two possibilities, DNA fragments
corresponding to the putative promoter regions were cloned into a
promoter detection vector containing the promoter-less galactokinase
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gene. Several small fragments (< 300 base pairs) that span the start
of the bphA coding region demonstrated promoter activity.
The work described in this paper has been reviewed in accordance
with the U.S. Environmental Protection Agency's peer and
administrative review policies and approved for presentation and
publication.
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GENOTOXICITY ASSAYS OF METABOLITES FROM BIOLOGICAL TREATMENT PROCESSES
by: Larry 0. Claxton
Health Effects Research Laboratory
Research Triangle Park,
North Carolina 27711
Extended Abstract
The present framework for public health evaluations at Superfund
sites and for the development of health-based performance goals is
provided in the U.S. EPA Superfund Public Health Evaluation Manual
(SPHEM) (1). Public health evaluations are one of the driving forces
for investigations that determine the need to undertake remedial action,
for feasibility studies during the cleanup phase, and for determining
the effectiveness of the cleanup procedures.
The health evaluation process as outlined in the SPHEM typically
involves five steps. In the first step, indicator compounds are
selected from among the list of compounds known to be present at the
site. The selection of these indicator compounds is based on known
toxicity, physical/chemical factors, and concentration at the site. All
evaluations after this step are based upon knowledge about these
selected indicator compounds. The second step is using knowledge about
the fate and transport of the indicator compounds to estimate exposure
concentrations. For the third step, human intake is estimated using
"standard assumptions" for daily water and air intake. The fourth step
involves a review of the toxicity of the indicator chemicals. The final
step is calculating human health risks from the exposure and toxicity
information.
These health evaluations are applicable in deciding when a site
requires remedial action, to aid in feasibility studies for cleanup
alternatives, and to evaluate the effectiveness of any cleanup
procedures used for a site. Each of these decision processes,
therefore, is dependent upon the identification and quantification of
indicator chemicals and the toxicity information for these compounds.
Several weaknesses associated with this process are identified in
the SPHEM. The following statements, for example, are made:
"...important chemical data are frequently unavailable."
"...toxicity testing has not kept pace with the need for
informations on many chemicals "
"...exposure assessment often requires many assumptions."
"...it would be unrealistic to expect that all data necessary to
determine precisely the health risks associated with every site will
be available."
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Although many of these weaknesses would result in an underestimation of
human risks associated with a site, a number of the data and knowledge
gaps may cause an overestimation of the human health risks.
For most remediation technologies (not just bioremediation), the
products from the remediation efforts are not predicted or monitored.
Instead, the process is monitored to understand to what extent the
pollutant or indicator chemical is depleted. There are a few
exceptions to this statement. For example, incinerators may be
monitored for the production of dioxins. The interaction of the known
toxic pollutants and other ancillary components and the degrading
microorganisms will not always produce non-toxic substances; however,
there is presently no provision for examining this potential.
Although present knowledge and procedures provide a usable framework
for human risk assessment associated with the remediation of Superfund
sites, those risk assessments (applied to all technologies) are highly
imprecise because of major data and knowledge gaps.
Public health evaluations for Superfund sites span a continium of
complexity, detail, level of effort, accuracy, and precision. As the
complexity of the site increases and/or the detail of information
decreases, risk assessments typically become less precise and accurate.
There is a need to fill important data and knowledge gaps concerning the
toxicology of substances found within Superfund sites, the safety of any
non-indigenous microorganisms, the toxicology of bioremediation
products, and the need for making multimedia assessments for comparing
bioremediation with other potential technologies. In addition,
substances within Superfund sites have the potential for undergoing
natural phototransformation and ecological transformation thus creating
additional products that are also potentially toxic. There is a primary
need, therefore, to have bioassay monitoring and research capabilities
that can address these complex issues in a more direct and reliable
manner.
This presentation will illustrate how in vitro bioassays that detect
mutagens and many genotoxic carcinogens can be used to evaluate the
dynamic processes that can occur within Superfund sites. For example,
sites containing high levels of the wood preservative pentachlorophenol
(PCP) could be expected to have other chlorinated phenols. Testing of
other chlorinated phenols demonstrated that some are less mutagenic and
others are more mutagenic (2). Next, the presence of one toxicant may
alter the toxicity of another pollutant. For example, PCP potentiates
the genotoxicity of 2,6-dinitrotoluene (3). During bioremediation, it
is also possible for the products of bioremediation to be genotoxic (3).
Because of its persistent nature, the herbicide 2,4,5-
trichlorophenoxyacetic acid (2,4,5-T) remains an environmental hazard.
2,4,5-trichlorophenol (2,4,5-TCP), one of the three reported metabolites
of 2,4,5-T, is 100-fold more genotoxic than 2,4,5-T (4). However, when
one tests the metabolites of 2,4,5-T produced by a bioremediation
organism (Pseudomonas cepacia AC1100), the genotoxicity of 2,4,5-T is
reduced when 2,4,5-T is the sole carbon source (4). Because 2,4,5-TCP
did not accumalate to high enough degree to increase mutagenicity,
results would support the use of bioremediation as a viable alternative
for the treatment of 2,4,5-T sites (4).
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These and other results demonstrate that the use of short-term
mutagenicity assays can be a valuable asset in evaluating treatment
alternatives including bioremediation.
REFERENCES
1. Office of Emergency and Remedial Response. Superfund Public Health
Evaluation Manual. EPA/540/1-86/060. U.S. Environmental Protection
Agency, Washington, D.C. 1986. 175 pp.
2. DeMarini, D.M., Brooks, H.G., and Parkes, D.G., Jr. Induction of
prophage lambda by chlorophenols. Environmental and Molecular
Mutagenesis 15:1-9, 1990.
3. Chadwick, R.W., George, S.E., Chang, J., Kohan, M.J., Dekker, J.P.,
Long, J.E., Duffy, M.C., and Williams, R.W. Potentiation of
2,6-dinitrotoliiene genotoxicity in Fischer 344 rats by pretreatment
with pentachlorophenol. Pesticide Biochemistry and Physiology 39:
168-181, 1991.
4. George, S.E., Whitehouse, D.A., and Claxtcr, L.D. Genotoxicity of
2,4,5-trichlorophenoxyacetic acid biodegradation products in the
Salmonella reversion and lambda prophage-induction bioassays.
Submitted to Fundamental and Applied Toxicology, 1991.
This manuscript has been reviewed by the Health Effects Research Laboratory, U.S.
Environmental Protection Agency and approved for publication. Approval cfoes not
signify that the contents necessarily reflect the views and policies of the Agency nor
does mention of trade names or commercial products constitute endorsement or
reconmendation for use.
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RESULTS OF THE BIOREMEDIATION FIELD INITIATIVE
by: Fran Kremer
U.S. Environmental Protection Agency
Cincinnati, OH
John Wilson
U.S. Environmental Protection Agency
Ada, OK
Walter Kovalick and Nancy Dean
U.S. Environmental Protection Agency
Washington, DC
BACKGROUND
As we approach the treatment of hazardous wastes, which are increasingly more diverse with
respect to the contaminants and the contaminated matrices, we will be more reliant on innovative
technologies for improved treatment efficiencies and lower costs. To meet these objectives, the U.S.
Environmental Protection Agency's (EPA) Administrator, William Reilly, has sought to develop an
agenda for the 1990s to identify strategies for increasing the use of bioremediation for the treatment
of hazardous wastes. To develop this agenda, assistance has been received from biotreatment
companies, site cleanup contractors, industry, academia, environmental organizations, and other
Federal agencies, in addition to the various offices within EPA.
One of the initial recommendations from this consortium was the need to expand our field
experience using this technology. Even though bioremediation is a viable technology to treat some
hazardous wastes, it has not been fully utilized for the many different types of wastes and site
conditions requiring remediation. It was recommended that the Agency serve as a focal point in
fostering field tests, demonstrations, and evaluations of bioremediation, using good test protocols and
documentation of results.
Based on this recommendation, the Office of Solid Waste and Emergency Response
(OSWER) and the Office of Research and Development (ORD) have instituted a Bioremediation
Field Program. This program provides assistance to the Regions and the states in conducting field
tests and carrying out evaluations of site cleanups using bioremediation. Sites considered in this
program include Superfund sites, RCRA corrective action sites and Underground Storage Tank
(UST) sites. The program is designed to:
1) more fully assess and document performance of full-scale field applications of
bioremediation,
2) provide technical assistance at various stages of site remediation, from site
characterization to full-scale implementation, and
3) regularly provide information on bioremediation projects being undertaken nationally.
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As solid full-scale performance data is needed to assess the capabilities of this technology,
evaluations of field operations are being undertaken. Sites considered for evaluation have field
biological units for treatment of wastes in situ or ex situ i.e. treatment of soils or groundwater in place
or treatment in a reactor or land treatment facility.
Technical assistance is available to the Regions and the States on treatability and field pilot
studies. This is to ensure adequate site characterizations, proper design of treatability studies, and
interpretation of results. In some cases, the EPA may conduct the treatability work. This assistance
is available through the EPA's Technical Support Centers. Presently, assistance is being provided on
a number of creosote sites and chemical facilities.
Data is being compiled on laboratory-, pilot-, and full-scale projects in order that EPA will
have a central repository of treatment information. Treatability data is being collected from the
Regions, States, other Federal agencies and the private sector. This data will be available through
the Risk Reduction Engineering Laboratory's Treatability Database Program (513-569-7503) and also
through the Alternative Treatment Technology Information Center (ATTIC) (301-816-9153). The
Treatability Database Program provides specific information on the treatability of specific chemicals
for a variety of technologies, including bioremediation. ATTIC is an on-line information retrieval
network that provides current information on innovative treatment methods for hazardous wastes.
Work is also underway on the development of a bioremediation expert system which will contain
treatability data and design information on biosystems. The Treatability Database Program and
ATTIC are presently available; the bioremediation expert system will be available in the Fall of 1991.
To date, over 130 sites have been identified across the country where bioremediation is being
planned, treatability studies are being conducted, is under full-scale operation, or has been completed.
These sites include CERCLA, RCRA and UST sites. In reviewing the types of wastes being
remediated, petroleum, creosote, and solvent wastes comprise over half of the waste types being
biologically treated.
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OIL SPILL BIOREMEDIATION PROJECT
Parmely H. Pritchard
U.S. Environmental Protection Agency
Environmental Research Laboratory
Sabine Island, Gulf Breeze, Fl 32561
INTRODUCTION
Several weeks following the Exxon Valdez oil spill in Prince William Sound,
Alaska, several million gallons of Prudhoe Bay crude oil, a well studied oil with
respect to previous cold water biodegradation studies, had contaminated almost 300
miles of rocky coast line in Prince William Sound. This confronted Exxon, the State
of Alaska, and the U.S. Coast Guard with the largest clean-up effort in U.S. history.
As a variety of clean-up options were assessed and implemented, it became clear to
EPA's Office of Research and Development and its scientists that bioremediation
was also a reasonable clean up option despite the complexity of the environmental
setting. We reasoned that the oil would become quickly colonized with oil degrading
bacteria but that their ability to degrade oil would be limited by the availability of
nitrogen and phosphorus nutrients. Artificially adding these nutrients would therefore
enhance biodegradation rates, something that has been observed many times in
laboratory studies. Thus, the Alaskan Bioremediation Project was initiated. An
approach was developed to determine if the addition of nitrogen and phosphorus-
containing fertilizers to oil-contaminated beaches would sufficiently enhance oil
biodegradation rates to permit consideration of bioremediation as a secondary
clean-up tool. A plan was conceived to conduct an initial field demonstration of this
approach and then if successful, recommendations for wider scale application would
be made to Exxon. EPA would then provide a follow up field study as a definitive
indication of the success of the large scale application.
The Field Demonstration
Field operations were begun in early May 1989. Two sites were selected,
Snug Harbor and Passage Cove. These beaches were comprised mainly of large
cobblestone overlying a mixed sand and gravel base. Both beaches had a thin layer
of oil covering the surface of the cobblestone, as well as oil mixed into the sand and
gravel under the cobbie to varying depths.
Selection of fertilizers, was based on considerations of application strategies,
logistical problems for large scale application, commercial availability (particularly if
large scale application became reasonable) and the ability to provide nitrogen and
phosphorus nutrients to the microbial communities on the surface and the
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subsurface beach material over sustained periods. Three application strategies were
adopted for testing; commercially available slow release formulations, an oleophilic
fertilizer, and water soluble fertilizer applied as a solution.
Commercial slow release fertilizer formulations were screened for the best
nutrient release rate characteristics. The strategy was to apply the best product to
the beach surface and then allow tidal action to disperse the released nutrients over
the contaminated area of the beach. The product had to remain on the beach for
several weeks while still delivering sufficient quantities of nutrients. Oleophilic
fertilizers are thought to essentially dissolve the nutrients into the oil by applying the
material directly on the oiled beach material. Nutrients sequestered in the oil phase
would presumably facilitate bacterial growth on the surface over sustained periods.
The oleophilic fertilizer Inipol EAP 22, produced by Elf Aquataine Company (Artix,
France) was selected. Fertilizer granules (about 2-3 mm in diameter), produced by
Sierra Chemicals (Milpitas, Calif), were selected as the main slow release fertilizer
formulation. These granules (Customblen) have a N:P:K ratio of 28:8:0 and slowly
release ammonia, nitrate and phosphate from inorganic ammonium nitrate and
ammonium phosphate encapsulated inside a diene-treated vegetable oil coating.
The fertilizer granules were broadcast on to the beach surface at a concentration of
90 gm/m2 using a mechanical seed spreader. Their high specific gravity , propensity
to adhere to the oil and their tendency to entrain under rocks and in interstitial
spaces, assured that they would remain on most low and moderate energy beaches
in Prince William Sound for two to three weeks.
The third type of fertilizer application involved spray irrigation with an aqueous
fertilizer solution. This approach produced the most defined, controlled and
reproducible introduction of nutrients into the oiled beach material, particularly for oil
below the beach surface. It was accomplished by dissolving commercially available
sources of ammonium nitrate (34-0-0) and triple phosphate (0-45-0) into seawater
pumped from below the beach. The resulting fertilizer solution was then sprayed
over the beach surface at low tide using a pump and lawn sprinkler heads.
The first application of the oleophilic fertilizer occurred July 8, 1989 at the Snug
Harbor site. Approximately 2-3 weeks following application, the treated beach
showed a visually pronounced reduction in the amount of oil on the surface of the
cobblestone. This produced a striking "window" against the oiled beach
background. Differences between treated and untreated portions of the beach were
dramatic. Close examination of the beach, however, showed that significant
quantities of oil remained under'the cobblestone as well as within the beach
subsurface. Over the next few weeks even this oil slowly disappeared. This
contrasted with the untreated control areas, in which there was little visual change.
Subsequent studies in the laboratory verified that Inipol was not a chemical rock
washer.
Definitive information on the role of biodegradation in this event, was
established by extracting oil from surface samples of cobble in the oleophilic
fertilizer-treated beach and analyzing the extracts by gas chromatography. The
sampling and analysis showed the this visual disappearance of the oil was
accompanied by significant decreases in total oil residues (i.e., weight of extractable
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material) and changes in hydrocarbon composition. This change in hydrocarbon
composition was largely due to biodegradation. Thus, it is reasonable to assume
that the decreases in oil residue weight on the cobblestone were caused by
biodegradation. We suspect that after a certain extent of oil bio-degradation was
achieved, the physical nature of the oil changed into a less sticky, flaky consistency
and this innocuous degraded material was then easily scoured from the rock
surfaces by tidal action.
Summary and Conclusions
Results from our initial field studies were sufficient for Exxon to consider the
use of bioremediation on a large scale as a finishing step for their clean-up effort.
We recommended that the oleophilic fertilizer, Inipol, be applied to beaches with only
surface oil and that a combination of Inipol and the fertilizer granules (Customblen)
be applied to beaches where there were both surface and subsurface oil found. The
granules provided a simple means of releasing nutrients into the beach subsurface
by tidal action and there by potentially enhancing biodegradation of subsurface oil.
Exxon began fertilizer application in early August 1989, to approximately 50
miles of beach in Prince William Sound that had been physically washed. Increasing
the biodegradation rate of oil at this point was very important because maximal
degradation could be achieved before winter conditions slowed biodegradation
processes. In many cases, the results of large scale fertilizer application were as
dramatic as our initial observations at Snug Harbor; that is, where the oil was spread
thinly over the cobble surface (as was the case on many beaches that had been
physically washed), the oil disappeared over a 20 day period. Unexpectedly, it also
appeared that even oil underneath the cobble had disappeared in a shorter time
period than observed at Snug Harbor. Although it is difficult to prove experimental-
ly, we believe that the physical cleaning process used by Exxon dispersed the oil
throughout the beach material to such an extent that the exposed oil surface area
was greatly increased, allowing greater bacterial colonization and subsequent
biodegradation.
Our Passage Cove study was initiated in late July, 1989 as the definitive
technical support site for the large scale application of Inipol and Customblen
fertilizers. These fertilizers were applied in combination to a large test beach and
samples of beach material were analyzed for changes in oil residue weight and
aliphatic hydrocarbon composition as before. These changes were compared to
those observed in an untreated control beach. In addition, a beach treated with a
seawater solution of inorganic fertilizer (applied via a sprinkler system) was examined
in the same way.
Approximately 2 to 3 weeks following initiation of this study, oil on the cobble
surfaces in the Customblen/lnipol- and fertilizer solution-treated beaches, had been
degraded to the point of producing visibly cleaner surfaces much as we had seen in
Snug Harbor. Surface oil on the control beach, however, was still very apparent
showing no visual reduction in the amount of oil. Disappearance of oil from the rock
surfaces on the beach treated with the fertilizer solution provided definitive proof that
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biodegradation (and not chemical washing) was responsible for the oil removal, as
there was no other reasonable mechanism to explain the effect of nutrient addition.
Despite sampling and interpretation complications resulting from the high
variability in oil distribution on the beaches, we have been able to show, statistically,
that oil biodegradation (as measured by changes in oil chemistry) was significantly
greater on the beach treated with the fertilizer solution than it was on the control
beach. Based on this information, we projected that after 45 days approximately 4-
5 times more oil remained on the control test beach than on the fertilizer solution-
treated test beach. This corresponded to an enhanced biodegradation rate of about
2-3 fold. Results were similar on the Inipol/Customblen-treated beach except statisti-
cally significant differences were not possible to establish. However, it appeared that
accelerated biodegradation (approximately a 2-3 fold increase in biodegradation
rates), occurred early in the test when nutrient concentrations were highest.
The long term benefit of fertilizer application was realized during examination of
the beaches in Passage Cove in November, 1989 and early June, 1990. Virtually no
oil was observed on either of the treated beaches, both at the surface and the
subsurface (12" depth). However, the untreated control still showed areas of heavy
oil contamination in the subsurface beach material. These observations provided
the final definitive demonstration of the long term success that can be expected from
bioremediation of oil contaminated beaches.
As a result, in Spring of 1990, bioremediation became an integral part of a
cleanup plan for the remaining oil-contaminated shorelines in Prince William Sound.
To follow the success of this treatment a joint bioremediation monitoring program
was conceived and implemented by scientists from Exxon, EPA, ADEC and the
University of Alaska (using logistical and resources support from Exxon). The
central success of this monitoring program paved the way for multiple reapplication
of the fertilizers during Summer 1990, a necessary step in many cases because of
large quantities of oil remaining in some areas.
The success of our field demonstration program has now set the stage for the
consideration of bioremediation as a key component in any clean-up strategy
developed for future oil spills. Its use and effectiveness will depend on the amount
of oil present in the contaminated environmental matrix; i.e., a longer time will be
required for degradation of high concentrations of oil and consequently a longer
period of fertilizer application will also be required. In addition, location of the oil (in
the absence of physical cleanup, subsurface oil may only be treatable by
bioremediation) and the acceptability of other clean-up options must be considered.
In most aquatic environments, enrichments of oil degrading microbial communities
occur relatively soon after oil contaminates shorelines. It is unlikely that natural
sources of nitrogen and phosphorus will be sufficient to give maximal degradation
rates in light of the available degradable organic carbon from the oil. Thus, the
application of fertilizers should enhance degradation and eventually remove the oil.
Although oxygen may become limiting in certain situations (e.g., fine grain sandy
beaches) the high porosity and large tidal fluxes characteristic of Prince William
Sound beaches precluded this as a limitation.
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Acknowledgements
The success of this project would not have been possible without the able
assistance and collaboration of the following EPA scientists and their relevant
expertise; Dr. John Rogers (microbial analyses), Drs. Al Venosa and John Glaser
(fertilizer evaluation), Dr. Fran Kremer (field site coordination), Mr. Dan Heggem
(quality assurance), Drs. Jim Clark and Rick Coffin (ecological assessment), Dr.
Larry Claxton (genotoxicity) and Dr. John Haines (fertilizer application). We wish to
also acknowledge the special contributions from John Baker, Steve McCutcheon,
Dick Valentinetti, Dennis Millar, and Steve Safferman. Without the resources and
support of Exxon this project would not have been possible; we acknowledge the
valuable scientific interaction with Drs. Steve Hinton, Roger Prince and Russ
Chianelli at Exxon). We gratefully acknowledge the guidance, encouragement, and
wisdom offered by Dr. Eric Bretthauer and John Skinner, director and deputy
director of the EPA Office of Research and Development, respectively.
References
(1) Atlas, R.; Busdosh, M. 3rd International Biodearedation Symposium 1976. 79-
86.
(2) Fedorak, P.; Westlake, D. Can. J. Microbiol. 1981, 27, 432-443.
(3) Horowitz, A.; Atlas, R. Appl. Environ. Microbiol. 1977, 33, 647-653.
(4) Atlas, R. et al. J. Fish. Res. Board Can. 1978, 35 585-590.
(5) Cook, F.; Westlake, D. Task Force on Northern Oil Development; 1974 Report
No. 74-1. #R72-12774
(6) Lee, M.D. et al. CRC Critical Rev. Environ. Control 1988, 18, 29-89.
(7) Atlas, R.M. Microbiol. Rev. 1981, 45, 180-209.
(8) Atlas, R.M. Ed; "Petroleum Microbiology". Macmillian: New York, 1984, 692
pp.
(9) National Academy of Sciences. "Oil in the Sea: Inputs, Fates and Effects".
1985, National Academy of Sciences, Washington, D.C., 348 pp.
(10) Leahy, J.G.; Colwell, R.R. Microbiol. Rev. 1990, 54, 305-315.
(11) Bartha, R. Microbiol. Ecol. 1986, 12, 155-172.
(12) Morgan, P.; Watkinson, R.J. CRC Critical Rev. Biotech. 1989, 8, 305-333.
(13) Office of Research and Development "Alaskan Oil Spill Bioremediation Project"
Environmental Protection Agency, Washington, D.C. EPA/600/8-89/073, 1989,
pp. 1-16.
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(14) Office of Research and Development "Alaskan Oil Spill Bioremediation Project,
Update" Environmental Protection Agency, Washington, D.C. EPA/600/8-
89/073, 1990, pp. 1-20.
(15) Pritchard, P. et al. Interim Report: Oil Spill Bioremediation Project 1990, 223
PP-
(16) Kremer, F. et al. Proceedings Air and Waste Management Meeting 1990, 90-
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(17) Rogers. J. et al. Proceedings Air and Waste Management Meeting 1990. 90-
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(18) Venosa, A. et al. Proceedings Air and Waste Management Meeting 1990, 90-
22.3.
(19) Clark, J. et al. Proceedings Air Waste Management Meeting 1990, 90-22.4.
(20) Office of Research and Development "Biosystem Technology Development
Program" U.S. Environmental Protection Agency, 1990, EPA/600/9-90/041, 58
pp. 985.
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