&EPA
            United States
            Environmental Protection
            Agency
            Office of Research and
            Development
            Washington DC 20460
EPA/540/R-95/532
September 1995
Bioremediation of
Hazardous Wastes
Research, Development,
and Field Evaluations

-------
                                                 EPA/540/R-95/532
                                                  September 1995
BIOREMEDIATION OF HAZARDOUS WASTES:
Research, Development, and Field Evaluations
         Biosystems Technology Development Program
             Office of Research and Development
            U.S. Environmental Protection Agency
            U.S. Environmental Protection Agency
      Ada, OK; Athens, GA; Cincinnati, OH; Gulf Breeze, FL;
              and Research Triangle Park, NC

-------
                                     Disclaimer
The information in this document has been funded wholly or in part by the U.S. Environmental
Protection Agency (EPA) and has been reviewed in accordance with EPAs peer and administrative
review policies and approved for presentation and publication. Mention of trade names or commer-
cial products does not constitute endorsement or recommendation for use.

-------
                                             Contents


                                                                                               Page


Executive Summary	  ix


Introduction	  xi


Section 1:  Bioremediation  Field Initiative	  1

Intrinsic Bioremediation of Trichloroethylene at the St. Joseph Aquifer/Lake Michigan Interface:
A Role for Iron and Sulfate Reduction
       Jack Lendvay, Mike McCormick, Peter Adriaens, University of Michigan, Ann Arbor, Ml	  3

Modeling Intrinsic Remediation as Ground-Water Discharges to a Lake:
The Trichloroethylene Plume at St. Joseph, Michigan
       Sean  M. Dean, Nikolaos D. Katopodes, University of Michigan, Ann Arbor, Ml	  6

The U.S. Environmental Protection Agency's Development of Bioventing
       Gregory Sayles, U.S. EPA, Cincinnati, OH	  8

Section 2:  Field Research	  11

A Review of Intrinsic Bioremediation of Trichloroethylene in Ground Water at Picatinny Arsenal,
New Jersey, and St. Joseph, Michigan
       John T  Wilson, Don Kampbell, James Weaver, Barbara Wilson, U.S. EPA, Ada, OK;
       Tom Imbrigiotta, Ted Ehlke, U.S. Geological Survey, Trenton, NJ	  13

Intrinsic Bioremediation of a Gasoline Plume: Comparison of Field and Laboratory Results
       Morton A. Barlaz,  Melody J. Hunt,  Sreenivas Kota, Robert C.  Borden, North  Carolina State
       University, Raleigh, NC	  17

Toxicity Effects on Methanogenic Degradation of Phenol in Ground Water
       Barbara A. Bekins, E. Michael Godsy,  Ean Warren, U.S. Geological Survey,  Menlo Park, CA	  20

A Multiphase,  Multicomponent Numerical Model of Bioventing With Nonequilibrium Mass Exchange
       Linda  M. Abriola, John R. Lang,  Klaus M. Rathfelder, University of Michigan, Ann Arbor, Ml	  22

Aromatic Hydrocarbon Biotransformation Under Mixed Oxygen/Nitrate Electron  Acceptor Conditions
       Liza P. Wilson, Peter C. D'Adamo,  Edward J. Bouwer, The Johns Hopkins University,
       Baltimore, MD	  24

-------
                                       Contents (continued)

                                                                                               Page

Nutrient Transport in a Sandy Beach
       Brian A. Wrenn, Makram T. Suidan, B. Loye Eberhardt, Gregory J. Wilson, University of
       Cincinnati, Cincinnati, OH; Kevin L. Strohmeier, Environmental Technologies and Solutions,
       Inc., Covington, KY; Albert D. Venosa, U.S. EPA, Cincinnati, OH	  26

Bioremediation of Crude Oil Intentionally Released on the Shoreline of Fowler Beach, Delaware
       Albert D. Venosa, John  R. Haines, U.S. EPA, Cincinnati, OH; Makram T. Suidan, Brian A.
       Wrenn, B. Loye Eberhardt,  Miryam Kadkhodayan, Edith Holder, University of Cincinnati,
       Cincinnati, OH; Kevin L. Strohmeier, Environmental Technologies and Solutions, Inc.,
       Covington, KY; Dennis King, Kingstat Consulting, Fairfield, OH;  Bennet Anderson, Delaware
       Department of Natural Resources and Environmental Control, Dover, DE	  29

Dynamics of Oil Degradation in  Coastal Environments: Effect of Bioremediation Products and Some
Environmental Parameters
       Marirosa Molina, Rochelle Araujo, U.S. EPA, Athens, GA; Jennifer R. Bond, DYNCORP,
       Athens,  GA	  32

Progress Toward Verification of  Intrinsic Cobioremediation of Chlorinated Aliphatics
       Mark Henry, Michigan Department of Natural Resources, Oscoda,  Ml	  35

Development and Capabilities of the National Center for Integrated Bioremediation Research and
Development (NCIBRD)
       Mark Henry, Michigan Department of Natural Resources, Oscoda,  Ml	  36

Phytoremediation of Petroleum-Contaminated Soil: Laboratory, Greenhouse, and Field Studies
       M. Katherine Banks, A.  Paul Schwab, Kansas State University,  Manhattan, KS	  37

Intrinsic Bioremediation of Fuel  Contamination in Ground Water at a Field Site
       Don H. Campbell, Robert S. Kerr Environmental Research Laboratory, Ada, OK;
       T.H. Wiedemeier, Parsons Engineering Science, Inc., Denver, CO;  J.E.  Hansen, U.S. Air Force
       Center for Environmental Excellence, Brooks Air Force Base, TX	  38
Section 3:  Performance Evaluation	  39


Detoxification of Model Compounds and  Complex Waste Mixtures Using Indigenous and Enriched
Microbial Cultures
       K.C. Donnelly, Jeannine L. Capizzi, Ling-Yu He, Henry J. Huebner, Texas A&M University,
       College Station, TX	  41

Assessing the Genotoxicity of Complex Waste Mixtures
       Larry D. Claxton, Virginia S. Houk, Sarah H. Warren, Thomas J. Hughes, and Susan E.
       George, U.S. EPA, Research Triangle Park, NC	  43


                                                 iv

-------
                                       Contents (continued)

                                                                                                Page

Section 4:  Pilot-Scale Research	  45

In Situ Bioremediation of Trichloroethylene With Burkholderia Cepacia PR1: Analysis of Parameters
for Establishing a Treatment Zone
       Richard A. Snyder, University of West Florida, Pensacola, FL; M. James Hendry, John R.
       Lawrence, Environment Canada, Saskatoon, Canada	  47

Characterization of Trichloroethylene-Degrading Bacteria From  an Aerobic Biofilter
       Alec Breen, Todd Ward, Ginger Reinemeyer, John Loper, Rakesh Govind, University of
       Cincinnati, Cincinnati, OH; John  Haines, U.S. EPA, Cincinnati, OH	  51

Anaerobic/Aerobic Degradation of Aliphatic Chlorinated Hydrocarbons in an Encapsulated
Biomass Biofilter
       Rakesh Govind, P.S.R.V. Prasad, University of Cincinnati, Cincinnati,  OH; Dolloff F.  Bishop,
       U.S. EPA, Cincinnati, OH	  54

Operation and Optimization of Granular Air Biofilters
       Francis Lee Smith, George A. Serial, Makram T. Suidan, Amit Pandit, Pratim Biswas, University
       of Cincinnati,  Cincinnati,  OH; Richard C. Brenner,  U.S. EPA, Cincinnati, OH	  57

Abiotic Fate  Mechanisms in Soil  Slurry Bioreators
       John A. Glaser, Paul T. McCauley, U.S. EPA, Cincinnati, OH; Majid A. Dosani, Jennifer S. Platt,
       E. Radha Krishnan, IT Corporation, Cincinnati, OH	  61

Design and Testing of an Experimental In-Vessel Composting System
       Carl  L. Potter, John A. Glaser, U.S. EPA, Cincinnati, OH; Majid A. Dosani, Srinivas Krishnan,
       Timothy A. Deets, E. Radha Krishnan,  IT Corporation, Cincinnati, OH	  64

Integrated Systems To Remediate Soil Contaminated With Wood Treating Wastes
       Makram T. Suidan, Amid P. Khodadoust, Gregory J. Wilson, Karen  M. Miller,
       University of Cincinnati, Cincinnati, OH; Carolyn M. Acheson, Richard C. Brenner, U.S. EPA,
       Cincinnati, OH	  66

Biological Treatment of Contaminated Soils Using Redox Control
       Margaret J. Kupferle, Tiehong L. Huang, Yonggui Shan, Maoxiu Wang, Guanrong You,
       University of Cincinnati, Cincinnati, OH; Gregory D. Sayles, Carolyn M. Acheson,  U.S. EPA,
       Cincinnati, OH	  68

Development of a Sulfate-Reducing  Bioprocess To Remove Heavy Metals From Contaminated
Water and Soil
       Munish Gupta, Makram T. Suidan, University of Cincinnati, Cincinnati, OH; Gregory D. Sayles,
       Carolyn M. Acheson, U.S. EPA, Cincinnati, OH	  71

Development of Techniques for the Bioremediation of Chromium-Contaminated Soil and Ground Water
       Michael J. Mclnerney, Nydia  Leon, Veronica E. Worrell, John D. Coates, University of
       Oklahoma, Norman, OK	  73

-------
                                        Contents (continued)

                                                                                                 Page

Bioremediation of Chlorinated Pesticide-Contaminated Sites Using Compost
       James C. Young, Jean-Marc Bollag, Raymond W. Regan, Pennsylvania State University,
       University Park, PA	   75

Reductive Electrolytic Dechlorination
       John W.  Norton, Jr., Makram T. Suidan, University  of Cincinnati, Cincinnati, OH; Carolyn M.
       Acheson, Albert D. Venosa, U.S.  EPA, Cincinnati, OH	   76

Biological Ex Situ Treatment of Soil Contaminated With Polynuclear Aromatic Hydrocarbons
       Carl L. Potter,  U.S. EPA, Cincinnati, OH; Roy C. Haught, IT Corporation, Cincinnati, OH	   77

Effectiveness of Gas-Phase Bioremediation Stimulating Agents (BSAs) for Unsaturated
Zone In Situ Bioremediation
       James G. Uber, Ronghui Liang, R.  Scott Smith, University of Cincinnati, Cincinnati, OH;
       Paul T. McCauley, U.S. EPA, Cincinnati, OH	   78

Section 5: Process Research	   81

Monitoring Crude Oil Mineralization  in Salt Marshes: Use of Stable Carbon Isotope Ratios
       Andrew W. Jackson, John H. Pardue, Louisiana State University, Baton Rouge, LA	   83

Mercury and Arsenic Biotransformation
       Ronald S. Oremland, U.S. Geological Survey, Menlo Park, CA	   86

Monod Degradation Kinetics of Quinoline in Natural and Microbially  Enriched Methanogenic Microcosms
       E. Michael  Godsy,  Ean Warren, Barbara A. Bekins, U.S. Geological Survey, Menlo  Park, CA	   87

Stimulating the Biotransformation of Polychlorinated Biphenyls
       John F. Quensen, III,  Stephen A.  Boyd, James M. Tiedje, Michigan State University, East
       Lansing,  Ml	   89

Bioaugmentation for In Situ Co-metabolic Biodegradation of Trichloroethylene in Ground Water
       Junko Munakata Marr, Perry L. McCarty, Stanford University, Stanford,  CA; V. Grace Matheson,
       Larry J. Forney, James M. Tiedje, Michigan State University, East Lansing, Ml; Stephen
       Francesconi, Malcolm S. Shields, University of West Florida, Pensacola,  FL;
       PH. Pritchard, U.S. EPA, Gulf Breeze, FL	   93

Biodegradation of Chlorinated Solvents
       Larry Wackett, Lisa Newman, Sergey Selifonov, University of Minnesota, St. Paul, MN;
       Peter Chapman, Michael Shelton, U.S. EPA, Gulf Breeze, FL	   96

Biological and Nutritional Factors Affecting Reductive Dechlorination of Chlorinated Organic Chemicals
       Dingyi Ye, National Research Council, Athens, GA; W. Jack Jones, U.S. EPA, Athens, GA	   98

Predicting Heavy Metal Inhibition of the In Situ Reductive Dechlorination of Organics at the Petro
Processor's Superfund Site
       John H.  Pardue, Louisiana State  University, Baton  Rouge, LA	  102

                                                  vi

-------
                                        Contents (continued)

                                                                                                 Page

Effect of Primary Substrate on the Reduction of 2,4-Dinitrotoluene
       Jiayang Cheng, Makram T. Suidan, University of Cincinnati, Cincinnati, OH; Albert D. Venosa,
       U.S. EPA, Cincinnati, OH	 104

Surfactants in Sediment Slurries:  Partitioning Behavior and Effects on Apparent Polychlorinated
Biphenyl Solubilization
       Jae-Woo Park, John F. Quensen, III, Stephen A. Boyd, Michigan State University, East
       Lansing, Ml	 108

Partial Characterization of an Anaerobic, Aryl, and Alkyl Dehalogenating Microorganism
       Xiaoming Zhang, National Research Council, Athens, GA; W. Jack Jones, John E. Rogers,
       U.S. EPA, Athens, GA	 109

Microbial Degradation of Petroleum Hydrocarbons in  Unsaturated Soils: The Mechanistic Importance of
Water Potential and the Exopolymer Matrix
       Patricia A.  Holden, James R. Hunt, Mary K. Firestone,  University of California, Berkeley, CA	 111

Metabolic Indicators of Anaerobic In  Situ Bioremediation of Gasoline-Contaminated Aquifers
       Harry R. Beller, Martin Reinhard, Alfred M. Spormann, Stanford University, Stanford, CA	 113

Contaminant Dissolution  and Biodegradation in Soils  Containing Nonaqueous-Phase Organics
       Larry E.  Erickson, L.T. Fan, J. Patrick McDonald, George X. Yang, Satish K. Santharam,
       Kansas State University, Manhattan, KS	 114

Protein Expression of Mycobacteria That Metabolize Polycyclic Aromatic Hydrocarbons
       David E. Wennerstrom, University  of Arkansas for Medical Sciences, Little Rock, AR;
       Carl E. Cerniglia, National Center for Toxicological Research, Jefferson, AR	 115
Section 6:  Hazardous Substance Research Centers	  117

Co-metabolic Biodegradation Kinetics of Trichloroethylene in Unsaturated Soils
       Karen L. Skubal, Peter Adriaens, University of Michigan, Ann Arbor, Ml	  119

The Effect of Water Potential on Biodegradation Kinetics and Population Dynamics
       Astrid Hillers, Peter Adriaens, University of Michigan, Ann Arbor, Ml	  120

Anaerobic-Aerobic Bioventing Development
       Gregory Sayles, U.S. EPA, Cincinnati, OH; Munish Gupta, Makram T. Suidan, University of
       Cincinnati, Cincinnati, OH	  121

Development of Co-metabolic Bioventing: Laboratory Tests
       Gregory Sayles, U.S. EPA, Cincinnati, OH; Alan Zaffiro, Jennifer Platt, IT Corporation,
       Cincinnati, OH	  123

                                                  vii

-------
                                       Contents (continued)

                                                                                                Page

Evaluating the Environmental Safety of Using Commercial Oil Spill Bioremediation Agents
       Jeffrey L. Kavanaugh, University of West Florida, Pensacola, FL; C. Richard Gripe, Carol B.
       Daniels, U.S. EPA, Gulf Breeze, FL; Rochelle Araujo, U.S. EPA, Athens, GA; Joe E. Lepo,
       University of West Florida, Pensacola, FL	  125

UNIFAC Phase Equilibrium Modeling To Assess  the Bioavailability of Multicomponent
Nonaqueous-Phase Liquids Containing Polycyclic Aromatic Hydrocarbons
       Catherine A. Peters, Princeton University, Princeton,  NJ	  126

Field Evaluation of Pneumatic Fracturing Enhanced Bioremediation
       Sankar N. Venkatraman, David S. Kosson, Rutgers University,  Piscataway, NJ;
       Thomas M. Boland, John R. Schuring, New Jersey Institute of  Technology, Newark, NJ	  127

Solids Suspension Characteristics Related  to Slurry Biotreatment Performances
       J.-W  Jim Tzeng, Paul T. McCauley, John A. Glaser,  U.S. EPA,  Cincinnati, OH	  128
                                                 VIM

-------
                                       Executive Summary
The  U.S.  Environmental  Protection Agency's (EPA's)
Office of Research and Development (ORD) hosted the
eighth annual Symposium on Bioremediation of Hazard-
ous Wastes: Research, Development, and Field Evalu-
ations in Rye Brook, New York, August 8-10,1995. More
than 250 people attended,  including leading  bioreme-
diation researchers, field personnel from federal, state,
and local agencies, and representatives from industry
and academia. Three speakers opened the symposium
with introductions and background information on biore-
mediation  research.

Fran Kremer, Coordinator of the Bioremediation Field
Initiative and Symposium  Chairperson, began by intro-
ducing several members of  the scientific steering com-
mittee  of  the Biosystems  Technology Development
Program (BTDP).  The BTDP draws on ORD scientists
who possess unique skills and expertise in biodegrada-
tion,  toxicology, engineering, modeling, biological and
analytical chemistry, and molecular biology.

George Pavlou,  Deputy  Director of EPA's  Region 2
Emergency and Remedial Response Division in  New
York City, presented a summary of recent  advances in
bioremediation technology.  Mr. Pavlou  noted  that new
techniques have increased the availability of in situ con-
taminants to biological degradation, and they have also
increased the range of contaminants that can be treated
biologically. EPA scientists  have  learned that microor-
ganisms are not only able to transform simple hydrocar-
bons but also such toxic and resistant contaminants as
chlorinated aromatics and heavy metal salts. Mr. Pavlou
provided a regional perspective on these developments,
describing how bioremediation has been put  to use in
Region 2.

Timothy Oppelt, Director of  the National Risk  Manage-
ment  Research  Laboratory in  Cincinnati, Ohio, dis-
cussed  the  future of bioremediation on  the  national
level. He explained that bioremediation has been essen-
tial to the  development of cost-effective cleanup tech-
nology and that it  has already been used at more than
450 sites.  This technology could potentially save  hun-
dreds of millions of dollars in future cleanup costs. While
bioremediation is now predominantly used for petroleum
decontamination, it will be applied to a wider range of
sites in the future. Mr. Oppelt concluded by describing
the reorganization of research under ORD and by warn-
ing that funding shortages may hinder the development
of bioremediation technology.

The 33 papers and 22 posters presented at the confer-
ence highlighted recent program achievements and re-
search projects aimed  at bringing bioremediation into
more widespread use. Taken as a whole, these topic areas
represent a comprehensive approach to bioremediation
of hazardous waste sites. The presentations were or-
ganized into five key research and program areas:

• Bioremediation Field Initiative: This initiative was in-
  stituted in 1990 to collect and disseminate perform-
  ance data  on bioremediation techniques from field
  application experiences. The Agency assists regions
  and states in conducting  field tests  and in carrying
  out independent evaluations of site  cleanups  using
  bioremediation.  Through this initiative, tests are un-
  der way at Superfund sites, Resource Conservation
  and Recovery Act corrective action facilities, and un-
  derground  storage  tank  sites.  Three  papers  pre-
  sented at the symposium were devoted to this key
  program area.

• Field Research: Once a bioremediation approach has
  proven effective in a laboratory or pilot-scale treata-
  bility study, it must be monitored and evaluated at a
  field site. The objective of  this  level of research is
  to  demonstrate  that  the  particular  bioremediation
  process performs as expected in the field. For most
  bioremediation technologies, certain  key factors con-
  cerning applicability, such  as cost effectiveness, can-
  not be thoroughly evaluated  until the  approach is
  scaled up and field tested. Ten papers and two post-
  ers provided information on recent field research.

• Performance Evaluation: Performance evaluation in-
  volves assessing  the extent and rate of cleanup for
  particular bioremediation processes as well as moni-
  toring the environmental fate and effects of contami-
  nants and  their  biological byproducts.  Two papers
  and one poster  addressed this area.

• Pilot-Scale Research: Pilot-scale research provides
  information on the operation and control of bioreme-
  diation technologies  and  the management of proc-
  ess-related residuals  and  emissions to enable the
  full-scale application  of a  technology. Given the ex-
  panding base of experience with various bioremedia-
  tion  methods, the need for pilot-scale research is
                                                   IX

-------
  increasing. Ten papers and three posters were pre-
  sented concerning research based on microcosms of
  field sites.

• Process Research: Process research involves isolat-
  ing  and  identifying  microorganisms that carry out
  biodegradation processes as well as developing tech-
  niques for modeling and  monitoring such processes.
  This research is fundamental to the development of
  new biosystems for treatment of environmental pol-
  lutants in surface waters, sediments, soils, and sub-
  surface materials.  Nine papers addressed this critical
  area, focusing on  the role of  metals and chlorinated
  organics  in bioremediation.   In addition, six poster
  presentations discussed  process research.
• Hazardous Substances Research Centers: In addition
  to presentations on research being carried out under
  the BTDP, the symposium included nine poster pres-
  entations from the EPA Hazardous Substance  Re-
  search Centers (HSRC). The scientists and engineers
  involved in HSRC conduct EPA research sponsored
  by the following  centers: the  Northeast Hazardous
  Substance Research Center (Regions 1  and 2), the
  Great Lakes and Mid-Atlantic Hazardous Substance
  Research  Center (Regions 3  and 5),  the  South/
  Southwest Hazardous  Substance  Research  Center
  (Regions 4 and 6), the Great Plains and Rocky Moun-
  tain  Hazardous Substance Research Center (Re-
  gions 7 and 8), and the Western Region Hazardous
  Substance Research Center (Regions 9  and  10).

-------
                                           Introduction
Bioremediation is one of the most promising technologi-
cal approaches to the problem of hazardous waste. This
process relies on  microorganisms such as bacteria or
fungi to transform hazardous chemicals into  less toxic
or nontoxic substances. There are several reasons why
such biological transformation is often  more attractive
than direct chemical or physical treatment. Microorgan-
isms typically:

• Directly degrade  contaminants rather than merely
  transferring them from one medium to another.

• Employ metabolic degradation pathways that can ter-
  minate with benign waste products (e.g., carbon di-
  oxide and water).

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

• Can  be used in situ to minimize disturbance of the
  cleanup site.

For these reasons,  microorganisms  can be effective,
economical, and nondisruptive tools for eliminating haz-
ardous chemicals. Until recent years, however, the use
of bioremediation was limited by the lack of a thorough
understanding of  biodegradation processes, their ap-
propriate applications, their control and enhancement in
environmental matrices,  and the engineering  tech-
niques required for broad application  of the technology.

Because the U.S. Environmental Protection Agency (EPA)
believes that bioremediation offers an attractive alterna-
tive to conventional methods of hazardous waste cleanup,
it has developed  a  strategic plan for  its  acceptance
and use by the technical and regulatory communities.
The Agency's strategic plan is centered  on  site-directed
bioremediation research to expedite  the development
and use of relevant technology. EPAs Office of Research
and Development (ORD) developed an integrated Bio-
remediation  Research Program to advance the under-
standing, development, and application of bioremedia-
tion solutions to hazardous waste problems threatening
human health and the environment.

Related bioremediation studies are being carried out at
five  EPA  Hazardous  Substance  Research  Centers
(HSRCs) under the direction of ORD's National Center
for  Extramural  Research  and  Quality  Assurance
(NCERQA). EPA was authorized to establish these cen-
ters by provisions in the 1986 amendments to the Su-
perfund law  calling for research  into  all aspects of the
"manufacture, use, transportation,  disposal, and  man-
agement of hazardous substances."

EPAs bioremediation research efforts have produced
significant results in the laboratory, at the pilot scale, and
in the field. The many accomplishments include aquifer
restoration, soil cleanup, process characterization, and
technology transfer. Research also focuses on extend-
ing the range of substances that can be  treated with
biological agents. This symposium was held to present
and discuss  recent developments in bioremediation re-
search undertaken during 1994 and 1995 under the
Biosystems Technology Development Program.

In this document, abstracts of paper and poster presen-
tations from the symposium are organized within five
key research and program areas:

• Bioremediation Field Initiative

• Field research

• Performance  evaluation

• Pilot-scale research

• Process research

The last section of this document includes abstracts of
presentations on bioremediation research performed as
part of the HSRC program.
                                                  XI

-------
                                      Section 1
                        Bioremediation  Field Initiative
The  Bioremediation Field Initiative is one  of the major components  of  EPA's  Bioremediation
Research Program. The Initiative was undertaken in 1990 to expand the nation's field experience
in  bioremediation techniques. The Initiative's goals are to  more fully assess  and document the
performance of full-scale bioremediation applications, to create a database of current field data on
the treatability of contaminants, and to assist regional and state site managers using or considering
bioremediation. The Initiative is currently tracking bioremediation  activities at more than 400 sites
under government and private-sector jurisdiction, in both the United States and Canada.  Perform-
ance evaluations are currently being conducted at nine sites, three of which were reported on  at
this symposium.

Data were presented from work at the St. Joseph, Michigan, Superfund  site on the use of iron as
an electron acceptor and the potential for natural attenuation  of chlorinated solvents. A presentation
was given on the modeling of the natural attenuation of solvents at the ground-water and  lake
interface. Studies also were carried out on the design and  field applications of bioventing in the
bioremediation of jet fuel spills.

-------
       Intrinsic Bioremediation of Trichloroethylene at the St. Joseph Aquifer/
            Lake Michigan Interface: A Role for Iron and Sulfate Reduction
                        Jack Lendvay, Mike McCormick, and Peter Adriaens
                            University of Michigan, Ann Arbor, Michigan
Introduction
The anaerobic aquifer at the St. Joseph, Michigan, Na-
tional Priorities List (NPL) site was contaminated with
trichloroethylene (TCE), which has been shown to have
dechlorinated to cis- and trans-dichloroethylene (DCE),
vinyl chloride (VC), ethylene, and ethane. These  prod-
ucts occur as a result of natural attenuation processes,
presumably under methanogenic conditions (1).
The flux of all alkyl halides into Lake Michigan is of major
public concern because of the suspected carcinogenic-
ity of VC. As the  plume moves toward the aerobic sur-
face  water,  the  dominant  redox conditions can be
expected to change because of wave action and vertical
seepage, which promote the interchange of oxygen-rich
lake water and anaerobic ground water. This presenta-
tion provides preliminary results of laboratory investiga-
tions geared  toward determining the  prevailing redox
processes and the potential  for natural attenuation of
chlorinated solvents at the interface.

Background
Three of the most important redox processes in the
natural anaerobic environment are the coupling of the
oxidation of organic matter to iron (Fe)  (III) reduction, to
sulfate reduction, and to methanogenesis. These  three
processes are considered mutually exclusive which, in
the anaerobic subsurface environment, results in the
development of spatially or  temporally distinct redox
zones. It has been demonstrated in aquatic sediment
and aquifer samples that Fe(lll)-reducing bacteria can
outcompete sulfate reducers, as well as methanogens,
for organic matter (2, 3).
Mineral-bound Fe(lll) has been shown to contribute sig-
nificantly to the total oxidation capacity of both pristine
and contaminated aquifers,  as  it often represents the
most abundant anaerobic terminal electron acceptor (4,
5). The speciation of iron  in aquifer  solids  is greatly
influenced by microbial  processes, particularly under
redox conditions favoring sulfate- and Fe(lll)-reduction.
Depending on the temporal or spatial succession in the
development of subsurface redox conditions, Fe(ll) pro-
duced by iron reducers precipitates as iron sulfides once
sulfate-reducing conditions develop,  or as iron oxides
such as magnetite (Fe3O4)  in the  absence of sulfide.
Alternatively, biogenically produced sulfide may precipi-
tate as FeS(1_x) (mackinawite) after reductive dissolution
of Fe(lll)  minerals. The iron sulfide and iron oxide  min-
erals thus formed may then contribute significantly to the
reduction capacity of aquifer solids and, in turn, play a
major role in the fate of organic contaminants.  The
presence  of these  precipitated minerals is direct evi-
dence for past or present iron- or sulfate-reducing  con-
ditions.

Materials and Methods
A three-stage iron analysis was performed on sediments
collected  from the same depths in sampling wells 55AB
(upstream) and 55AD  (near shore) to evaluate the oc-
currence  of oxidized and reduced iron minerals, and to
calculate  inorganic reducing equivalents present in aqui-
fer materials.
• Bioavailable iron (Fe(lll)): Microbial Fe(lll)-reduction
  has been shown to predominantly use amorphous
  oxyhydroxides and goethite as terminal electron ac-
  ceptors. Quantitation of these minerals can be ap-
  proximated   by  extracting  sediments  according to
  Lovley  and  Phillips (6).
• Ferrous monosulfides and amorphous iron oxides
  (Fe(ll)): Reduced  Fe(ll) minerals resulting from mi-
  crobial  iron  and sulfate reduction, predominantly fer-
  rous sulfides and iron oxides, can be quantified using
  a 24-hr extraction with 0.5 M hydrochloric acid (HCI)
  (7). As this extraction  removes bioavailable iron as
  well, FeS can be approximated by substration.
• Siderite, crystalline iron oxide, and magnetite (Fe(ll)):
  This extraction  represents the precipitated  ferrous

-------
  iron fraction in the absence of sulfate reduction. Wet
  sediment samples were extracted with 5 M HCI for a
  21-day period, and analyzed according to Heron and
  Christensen (5).

This information was then used in conjunction with avail-
able data on ground-water chemistry (pH,  redox, car-
bonate,  sulfate, sulfide)  and compiled in a  chemical
equilibrium  model (MINEQL+) to  predict speciation of
iron. Precipitation  of ferrous iron solids was based on
stability constants from the literature.

Results

The analysis of samples from wells 55AB and 55AD
indicated the presence  of  similar concentrations of
bioavailable  Fe(lll)  (5  mmol  equiv./kg)  (Table  1).
Whereas precipitated iron oxides and ferrous monosul-
fides increased dramatically toward the  shore, however,
solids representative of the absence of sulfate reduction
significantly decreased  (12  versus 5  mmol  equiv./kg
aquifer  material, after subtraction of bioavailable iron).
This observation,  based on iron  extraction  data, was
confirmed by chemical modeling, which  predicted a pre-
dominant occurrence of siderite in the sampling point
farthest from the lakeshore and increasing  occurrence
of iron sulfide and mackinawite closer to shore and into
the lake. Iron- and sulfate-reducing  activity was rela-
tively easily stimulated  in these sediments.

Ongoing dechlorination experiments under sulfate- and
iron-reducing conditions, using TCE and VC, have re-
sulted in the production of t-DCE from TCE under iron-
reducing conditions. Whetherthe production of the trans
rather than the cis isomer is indicative of an abiotic
dechlorination mechanism (by Fe(ll)) remains to be elu-
cidated. VC did not dechlorinate under sulfate-reducing
conditions during the 2 months monitored to date.

Information available from previous  field studies in col-
laboration with the Robert S. Kerr Laboratories (Dr. John
Wilson)  suggested  that neither  methanotrophic nor
methanogenic activity predominates in the contaminant
plume underthe shoreline and Lake Michigan, based on
redox potential  and  oxygen  and  methane  measure-
ments. Redox potentials and an increase in soluble iron
concentrations between  both shore samples and below
the Lake  Michigan bottom, however, suggest that iron
reduction may be the dominant process in  regions near
plume emergence (8).


References

1. Wilson, J.T., J.W. Weaver, and  D.H.  Kampbell.  1994. Intrinsic
  bioremediation of TCE  in  ground water at an NPL Site in St.
  Joseph, Michigan. In: U.S.  EPA Symposium on Intrinsic Bioreme-
  diation of Ground Water. EPA/540/R-94/515. Washington, DC.

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

3. Chapelle, F.H., and D.R. Lovley.  1992. Competitive exclusion of
  sulfate reduction by Fe(lll)-bacteria: A mechanism for producing
  discrete zones of high-iron ground water. Ground Water 30:29-36.

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

5. Heron, G., and T.H. Christensen. 1995.  Impact of sediment-bound
  iron on redox buffering in a landfill leachate-polluted aquifer (Vejen,
  Denmark). Environ. Sci. Technol.  29:187-192.
Table 1.  Determination of Iron and Sulfate Equivalents in St. Joseph Sediments
Sample Location and
Analysis
Well 55AB
1-hr/0.5M HCI
1 -day/0.5 M HCI
1-hr/0.25 M HCI and 0.25 M
NH2OH • HCI
21 -day/5.0 M HCI

Ion chromatography
Well 55AD
1-hr/0.5 M HCI
1 -day/0.5 M HCI
1-hr/0.25 M HCI and 0.25 M
NH2OH • HCI
21 -day/5.0 M HCI

Ion chromatography
Chemical Species
Analyzed
Iron Extractions
Limited Fe(ll)
Limited Fe(ll)
Limited Fe(ll) and
Bioavailable Fe(lll)
Limited Fe(ll)
Sulfate Extractions
Soluble sulfate
Iron Extractions
Limited Fe(ll)
Limited Fe(ll)
Limited Fe(ll) and
Bioavailable Fe(lll)
Limited Fe(ll)
Sulfate Extractions
Soluble sulfate
Chemical
Concentration
(mg/kg soil)
219.2 ±15.8
21 0.6 ±30.6
269.6 ± 29.7
881 .8 ± 66.2
(mg/L)
25.8 ± 3.5
(mg/kg soil)
333.4 ± 11.5
445.5 ± 16.9
270.8 ± 8.2
735.9 ± 48.7
(mg/L)
29.3 ± 4.2
Redox Equivalents
(mmol equiv./kg soil)
3.9 ± 0.3
3.8 ± 0.6
4.7 ± 0.5
15.8 ±1.2
(mmol equiv./L)
2.2 ± 0.3
(mmol equiv./kg soil)
6.0 ± 0.2
8.0 ± 0.3
4.9 ± 0.2
13.2 ±0.9
(mmol equiv./L)
2.4 ±0.4

-------
Lovley, D.R., and E.J.P. Phillips. 1987. Rapid assay for microbially    8. Adriaens,  P., J.V. Lendvay, N. Katopodes, S. Dean, J.T. Wilson,
reducible ferric iron in aquatic sediments. Appl. Environ. Microbiol.       and D. Kampbell. 1995. Intrinsic bioremediation of chlorinated sol-
53:1,536-1,540.                                                     vents at the St. Joseph, Michigan Aquifer/Lake Michigan Interface.
                                                                   EPA/600/R-95/012. In: Proceedings of the 21st Annual RREL Re-
Heron, G., C. Crouzet, A.C.M. Bourg, and T.H. Christensen. 1994.       search Symposium, Abstract.
Speciation of Fe(ll) and Fe(lll)  in contaminated aquifer sediments
using chemical  extraction techniques. Environ.  Sci.  Technol.
28:1,698-1,705.

-------
      Modeling Intrinsic Remediation as Ground-Water Discharges to a Lake:
                  The Trichloroethylene Plume at St. Joseph,  Michigan
                             Sean M. Dean and Nikolaos D. Katopodes
           Department of Civil and Environmental Engineering, University of Michigan,
                                       Ann Arbor, Michigan
Introduction

Contamination of ground water by chlorinated solvents
is widespread and has been in the forefront of public and
regulatory concern for  the last decade.  As a result,
considerable research efforts, both in the laboratory and
in the field, have addressed the potential for using bio-
logical processes to degrade these pollutants via either
in situ or  onsite bioremediation technologies. Natural
attenuation has been observed to be responsible for
removal or partial transformation of both chlorinated and
nonchlorinated organic contaminants. At several sites,
naturally occurring  reductive dechlorination has  been
found to be responsible for the anaerobic transformation
of trichloroethylene (TCE) to lesser chlorinated interme-
diates, such as c-  and t-dichloroethylene (DCE) and
vinyl chloride (VC),  and to ethylene (1,  2). Because
aquifers become oxygenated near groundwater and
lake interfaces, concerns have been raised with respect
to surface water contamination by VC. In the current
study, a  modeling approach addresses the fate of  TCE
and its lesser chlorinated transformation products at this
anaerobic/aerobic interface. The  modeling effort de-
scribes and predicts the fate of the chlorinated solvents
at and near the interface, taking into account ground-
water flow rates and microbial degradation rates, as well
as the oxygenating effects of wave action and lake water
intrusion near the shore line.

Numerical  Model

The numerical model for the simulation of the hydrody-
namic, physicochemical, and  biological processes that
take place at the  lake-aquifer interface is validated
based on  specific data  from an application site at St.
Joseph,  Michigan. Although  the  hydrodynamic proc-
esses are truly three-dimensional,  most of the phenom-
ena of interest, such as migration of TCE, DCE, and VC
into the  lake and transfer of dissolved oxygen into the
aquifer from water infiltrating through the surf region,
can be modeled by a two-dimensional model on the
vertical plane.

Model Components

Due to the significant time-scale difference between the
near-shore circulation and wave runup and breaking in
the lake compared with the flow in the porous media,
two separate models  are  constructed for the corre-
sponding hydrodynamic phenomena. The resulting flow
fields are then integrated in a single mass transport and
contaminant fate model. All three  components of the
model are two-dimensional, covering a vertical  plane
extending from a location inland where uniform flow and
mass  flux are observed in the aquifer to a distance
inside the lake where most near-shore current activity
has diminished. The various modules are verified by
analytical solutions and  intermodel comparisons.

Ground-Water Module

In the porous media, a finite-element module for variably
saturated flow has been constructed on a vertical plane.
This module uses pressure heads calculated by the lake
module as a boundary condition at the lake-aquifer in-
terface. For the St. Joseph site, this module  has been
used in conjunction with  two separate grids to  take
advantage of naturally defined boundary conditions.

A coarse grid has been developed extending from Lake
Michigan to a ground-water divide approximately 900 m
inland. Zero  flux boundaries are  prescribed at  the
ground-water divide and an underlying clay layer. A
seasonally varying flux boundary condition at the ground
surface reflects recharge from rainfall.  To focus on the
lake-aquifer interface region, a refined grid has been
developed extending from Lake Michigan to a point
approximately  100 m inland. One  year's output  from
running this module on  the coarse  grid defines a sea-

-------
sonally varying inland boundary condition for the refined
grid.

Lake Module

The near-shore/free-surface flow simulation is based on
the numerical solution  of the Navier-Stokes equations
by  means of the finite-element  method. For turbulent
flow, a widely accepted two-equation closure model is
employed together with certain approximations near the
bed and free-surface  boundaries.  At  high Reynolds
numbers, an upwind formulation known as the Petrov-
Galerkin method of weighted residuals is introduced for
the suppression of nonlinear instabilities. The model can
predict the vertical structure of the flow from the seep-
age face between the aquifer and the  lake to the  free
surface. Wave action is incorporated, and special atten-
tion is focused on wave runup and breaking. The beach
is assumed  to be a porous bed so that water from the
surf and break region is allowed to infiltrate and reach
the aquifer.

For the complete formulation of  the problem, bed  per-
meability resulting in seepage through the surf region
would  be computed by simultaneous  solution of the
free-surface flow problem in the lake with the associated
unsaturated  flow problem in the subsurface domain. In
this model,  bed seepage is introduced as a boundary
condition. This eliminates the difficulty of having to deal
simultaneously with two time scales without affecting the
robustness of the model. The bed seepage is averaged
over time to provide an interface boundary condition to
the ground-water flow module.

Contaminant Transport Module

The contaminants are assumed to be well mixed later-
ally. The contaminant fate and transport module uses a
finite-element model to solve the two-dimensional trans-
port equation based on the flow fields computed by the
ground-water and lake modules. This accounts for con-
taminant transport due to advection  and dispersion in
the aqueous phase and for interphase mass transfer
due to sorption and volatilization.

Microbial transformations are incorporated using modi-
fied Monod  kinetics to describe  a source/sink term in
transport equation. Rate constants have been estimated
by  a parallel experimental  effort focusing on  microbial
interactions. The microbial biomass  is  assumed to  be
immobile below a limiting concentration at which slough-
ing occurs.


Conclusion

A two-dimensional finite element has been developed to
simulate the transport  and biodegradation of chlorinated
solvents at and near ground-water and lake interfaces.
Example simulations  consider the effects that factors
such as heterogeneities  in the porous media, uncertain-
ties in  parameter estimation,  and  varying  recharge
through the  beach have on the location and concentra-
tion of a plume of chlorinated solvents.


References

1.  Freedman,  D.L., and J.M. Gossett.  1989. Biological reductive
   dechlorination of tetrachloroethylene and trichloroethylene to eth-
   ylene under methanogenic conditions. Appl.  Environ. Microbiol.
   55(9):2,144-2,151.

2.  Vogel, T.M.,  and  P.L. McCarty. 1985.  Biotransformation of
   tetrachloroethylene  to trichloroethylene, dichloroethylene, vinyl
   chloride, and carbon dioxide under methanogenic conditions. Appl.
   Environ. Microbiol. 49(5):1,080-1,083.

-------
                        The U.S. Environmental Protection Agency's
                                  Development ofBioventing
                                          Gregory D. Sayles
                         National Risk Management Research Laboratory,
                      U.S. Environmental Protection Agency, Cincinnati, Ohio
Research conducted in the mid to late 1980s by the U.S.
Air Force (1, 2), researchers in the Netherlands (3-6),
Texas Research Institute (7, 8), Battelle Memorial Insti-
tute (2, 9, 10,  11), Utah State University (11), and the
U.S. Environmental  Protection Agency  (EPA) (12),
among others, suggested  that delivering air  to  the
vadose zone to promote biodegradation could  be a low-
cost means of cleaning  fuel-contaminated vadose zone
soils. This approach was motivated  by the attempt to
solve two different remediation development problems:

• Soil vacuum extraction for treatment of contaminated
  vadose zones involved costly off-gas treatment and
  only removed the volatile fraction of the contamination.

• Oxygen delivery to the vadose zone to promote aero-
  bic biodegradation  using  the approaches attempted
  in promoting biodegradation in ground water (namely,
  delivering oxygen saturated water or aqueous solu-
  tions of hydrogen peroxide or nitrate to the contami-
  nated area)  was neither efficient nor cost-effective.

A  process was needed that could  deliver oxygen  by
introducing air into the vadose at a rate that minimized
volatilization of the contamination. Several groups simul-
taneously developed what is  now known as bioventing.

EPA recognized the potential cost savings of such a
technology over traditional remediation approaches and
began an aggressive bioventing development program
in  1990. The mission of the program, in essence, was
to  develop bioventing so that it could be applied at as
many contaminated sites as possible. To date, EPAs
program has  demonstrated or is currently developing
the use of bioventing for the following situations:

• For operation with air injection.

• In cold climates.

• With soil warming.
• For jet fuel/aviation fuel.

• For nonfuel contaminants such as acetone, toluene,
  polycyclic aromatic hydrocarbons (PAHs), and trichlo-
  roethylene (TCE).

Table 1 provides a list of EPAs involvement in bioventing
research and development. The cumulative knowledge
of EPA, the Air Force,  and  Battelle Memorial  Institute
regarding bioventing of fuel-contaminated sites was dis-
tilled in Principles and Practices Manual for Bioventing,
to be released  in late 1995.

The next frontier for aerobic bioventing is the application
of the process to sites contaminated with  chlorinated
solvents. EPA is currently involved in two laboratory and
field   projects  to  develop "co-metabolic bioventing."
Co-metabolic bioventing is the promotion of the aerobic
biodegradation of chlorinated solvents, such as TCE, in
the vadose zone by delivering oxygen and, if necessary,
a volatile co-metabolite to the contaminated site. EPA
projects will consider two scenarios: 1) the co-metabolite
is a co-contaminant of the chlorinated solvent, and thus
only  air must  be  delivered;  and 2) the  co-metabolite
must be delivered  with the air stream and must therefore
be volatile.

In summary, EPA and its collaborators, primarily the Air
Force, the  U.S. Coast  Guard, and Battelle, has been
successful in developing bioventing into an inexpensive,
robust process, applicable to the cleanup of many con-
taminated sites.
References

 1. Miller, R.N. 1990. A field-scale investigation of enhanced petro-
   leum hydrocarbon biodegradation in the vadose zone combining
   soil venting as an oxygen source with moisture and nutrient ad-
   ditions. Ph.D. dissertation. Utah State University, Logan, UT.

-------
Table 1.   Summary of EPA Bioventing Research and Development
Project/
Location
                   Dates
                                  Contaminants    Scale
                                                                Approach/Results
                                                                                                                   References
Coast Guard        1990-1991      Aviation fuel      Lab, pilot
Station, Michigan
Hill AFB, Utah
Eielson AFB,
Alaska
Reilly Tar SF
Site, MN

Greenwood
Chemical SF
Site, Virginia
Dover AFB,
Delaware
1990-1994     Jet fuel (JP-4)     Full
1991-1994     Jet fuel (JP-4)     Lab, pilot
                   1992-
                                  PAHs
                                                   Pilot
1993-1995     Acetone,          Pilot
              toluene, others
                   1995-
Principles and      1995
Practices Manual   release
                                  TCE
                                  Fuels
                                Lab, pilot
                                Two
                                volumes
First test of air-injection bioventing. Showed              13-16
that air injection near the water table could
induce biodegradation in the vadose zone and
in ground water.

First full-scale air-injection bioventing. Showed             17
that low-rate air injection could supply
biodegradation oxygen demand and produce
no measurable surface emissions of volatile
organic compounds.

Showed that bioventing in cold climates is                 18
feasible and that simple soil warming
techniques can increase the rate of
biodegradation when bioventing, thereby
decreasing the time required for remediation.

Attempting to show that PAHs at wood treating             20
sites can be remediated with bioventing.

Attempting to show that bioventing of sites               N/A
contaminated with non-fuel,  aerobically
biodegradable organics and in low permeability
soils is feasible.

Attempting to show that TCE can be treated              N/A
with bioventing if the necessary co-metabolite
is present, either as a co-contaminant or
delivered as a volatile organic in the injected
air stream.

Volume 1: Principles of Bioventing                       N/A
Volume 2: The Practice of Bioventing
 2. Miller, R.N.,  C.C. Vogel, and R.E. Hinchee.  1991. A field-scale
    investigation of petroleum hydrocarbon  biodegradation in  the
    vadose zone enhanced by soil venting at Tyndall AFB, Florida.
    In: Hinchee,  R.E., and R.F. Olfenbuttel, eds.  In situ bioreclama-
    tion. Stoneham, MA: Butterworth-Heinemann. pp. 283-302.

 3. Staatsuitgeverij. 1986. Proceedings of a  Workshop  (March  21-
    21).  Bodembeschermingsreeeks  No.   9.  Biotechnologische
    Bodemsanering. Rapportnr. 851105002. ISBN 90-12-054133, Or-
    dernr. 250-154-59; Staatsuitgeverij Den Haag, The Netherlands.
    pp. 31-33.

 4. van Eyk, J., and C. Vreeken. 1988. Venting-mediated removal of
    petrol from subsurface soil strata as a result of stimulated evapo-
    ration and enhanced biodegradation.  Med. Fac. Landbouww.
    Riiksuniv. Gent 53(4b):1,873-1,884.

 5. van Eyk, J., and C. Vreeken. 1989. Model of petroleum minerali-
    zation response  to soil aeration to aid in site-specific, in  situ
    biological remediation. In: Jousma et al., eds. Groundwater con-
    tamination:  Use of models in  decision-making.  Proceedings of
    the  International Conference on Ground-water Contamination.
    Boston, MA/London, England:  Kluwer. pp. 365-371.

 6. van Eyk, J., and C. Vreeken. 1989. Venting-mediated removal of
    diesel oil from subsurface  soil strata  as a result of stimulated
    evaporation and enhanced biodegradation. In: Hazardous waste
    and contaminated sites, Envirotech, Vienna. Vol. 2,  Session 3.
    ISBN 389432-009-5. Essen, Germany: Westarp Wiss. pp. 475-485.

 7. Texas Research Institute. 1980. Laboratory-scale gasoline spill
    and venting experiment.  Interim Report No. 7743-5:JST American
    Petroleum Institute.
                                                8. Texas Research Institute. 1984. Forced venting to remove gaso-
                                                  line vapor from a large-scale  model aquifer.  Final Report No.
                                                  8210I-F:TAV American Petroleum Institute.

                                                9. Hinchee, R.E., and  M. Arthur.  1991. Bench-scale studies of the
                                                  soil aeration process for bioremediation of petroleum hydrocar-
                                                  bons. J. Appl. Biochem. Biotech. 28/29:901-906.

                                               10. Hinchee, R.E., and S.K. Ong. 1992. A rapid in situ respiration test
                                                  for measuring  aerobic biodegradation rates of hydrocarbons in
                                                  soil. J. Air Waste Mgmt. Assoc. 42(10):1,305-1,312.

                                               11. Dupont, R.R., WJ. Doucette, and R.E.  Hinchee. 1991. Assess-
                                                  ment  of in situ bioremediation potential and the application of
                                                  bioventing at a fuel-contaminated site. In: Hinchee, R.E., and R.F.
                                                  Olfenbuttel, eds. In situ bioreclamation: Applications and investi-
                                                  gations  for  hydrocarbon and  contaminated  site  remediation.
                                                  Stoneham, MA: Butterworth-Heinemann. pp. 262-282.

                                               12. Wilson, J.T., and C.H.  Ward. 1986. Opportunities for bioremedia-
                                                  tion of aquifers contaminated  with petroleum hydrocarbons. J.
                                                  Ind. Microbiol.  27:109-116.

                                               13. Ostendorf, D.W, and  D.H. Kampbell. 1990.  Boioremediated soil
                                                  venting of light hydrocarbons. Haz. Waste Haz. Mat.  1(4):319-334.

                                               14. Kampbell, D.H., and J.T Wilson. 1991. Bioventing to treat fuel
                                                  spills  from underground storage tanks. J. Haz. Mat. 28:75-80.

                                               15. Kampbell, D.H., J.T. Wilson, and C.J. Griffin. 1992. Performance
                                                  of bioventing at Traverse City,  Michigan. In: U.S. EPA. Sympo-
                                                  sium on Bioremediation of Hazardous Wastes. EPA/600/R-92/126.
                                                  pp. 61-64.

-------
16.  Kampbell,  D.H., C.J. Griffin, and F.A. Blaha. 1993. Comparison
    of bioventing and air sparging for in situ bioremediation of fuels.
    In:  U.S. EPA. Symposium on  Bioremediation of Hazardous
    Wastes:  Research,  Development,  and  Field  Evaluations.
    EPA/600/R-93/054.  pp. 61-67.

17.  Sayles, G.D., R.C.  Brenner, R.E. Hinchee, and R. Elliott. 1994.
    Bioventing  of jet fuel spills II: Bioventing in a deep vadose zone
    at Hill AFB, Utah. In: U.S.  EPA. Symposium on Bioremediation
    of Hazardous Wastes: Research, Development and Field Appli-
    cations. EPA/600/R-94/075. pp. 22-28.

18.  Sayles, G.D., R.C. Brenner, R.E. Hinchee, A. Leeson, C.M. Vogel,
    and R.N. Miller. 1994. Bioventing of jet fuel spills I: Bioventing in
    a cold climate with soil warming at Eielson AFB, Alaska. In: U.S.
    EPA. Symposium on Bioremediation of Hazardous Wastes: Re-
    search, Development and Field Applications. EPA/600/R-94/075.
    pp. 15-21.
19. Leeson, A., R.E. Hinchee, J.  Kittel, G. Sayles, C. Vogel, and R.
    Miller. 1993. Optimizing bioventing in  shallow vadose zones in
    cold climates. Hydrological Sci. J. 38(4).

20. McCauley, P.T., R.C. Brenner, F.V. Kremer, B.C. Alleman, and
    D.C.  Beckwith. 1994. Bioventing soils contaminated with  wood
    preservatives.  In:  U.S. EPA.  Symposium on Bioremediation of
    Hazardous Wastes:  Research, Development and Field Applica-
    tions. EPA/600/R-94/075. pp.  40-45.
                                                              10

-------
                                      Section 2
                                  Field  Research
Field research is essential for evaluating the performance of full-scale bioremediation precesses
and for conducting accelerated testing on technologies that are appropriate for scaled-up applica-
tion. For example, problems associated with the use of bacteria used in the laboratory include
optimizing the activity of the organism under site conditions and defining the risks associated with
introducing a non-native microorganism to the site. The objective of this level of research  is to
demonstrate  that  the particular bioremediation process performs as expected  in the field. Re-
searchers at the symposium provided information on several  ongoing field experiments.

Field studies conducted at St. Joseph, Michigan, and  Rocky Point,  North Carolina, sought to
determine the extent of intrinsic bioremediation in subsurface contaminant plumes.  EPA found
extensive dechlorination of trichloroethylene contamination by endogenous microorganisms at St.
Joseph. At Rocky  Point, researchers found considerable  agreement between laboratory models of
hydrocarbon biodegradation and field observations.

At Pensacola,  Florida, researchers obtained data from a  site  contaminated with  creosote to
construct a model  of toxic inhibition of bioremediation. Anothergroup of scientists constructed a set
of models to better understand the processes of soil vapor extraction and bioventing.

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

Finally, a group of presentations dealt with the use of bioremediation to clean petroleum contami-
nation in beaches and wetlands. Researchers performed studies to determine how often nutrients
should be applied  to beaches to support bacterial growth, as well as directly studied the process of
bacterial petroleum degradation on beaches. Another study  examined what environmental influ-
ences might affect the biodegradation of petroleum in wetland areas.

There were two field research poster presentations. The first of these described the National Center
for Integrated Bioremediation Research and Development (NCIBRD) and  its work at Wurtsmith Air
Force Base  in  Michigan. The second presentation described use of plant growth  to enhance
bioremediation in the field.
                                           11

-------
   A Review of Intrinsic Bioremediation of Trichloroethylene in Ground Water at
               Picatinny Arsenal, New Jersey, and St. Joseph,  Michigan
               John T. Wilson, Don Kampbell, James Weaver, and Barbara Wilson
                     U.S. Environmental Protection Agency, Ada, Oklahoma

                                  Tom Imbrigiotta and Ted Ehlke
                          U.S. Geological Survey, Trenton, New Jersey
Reductive  dechlorination occurs frequently in large
trichloroethylene (TCE)  plumes.  TCE  is transformed
largely to cis-dichloroethylene (cis-DCE), then to vinyl
chloride, and finally to compounds that do not contain
organic chlorine. This abstract evaluates the rate and
extent of natural reductive dechlorination of TCE in two
large plumes with similar properties.

Description of the Plumes

Both plumes originated in a release of liquid TCE. The
plume at St Joseph, Michigan, originates from an indus-
trial park, while the plume at the Picatinny Arsenal, New
Jersey, originates in a release from a degreasing vat at
a plating shop. Cross sections of the plumes are de-
picted in Figures 1 and 2. Both plumes have high con-
centrations of TCE in the core of the plume (over 25,000
u,g/L), are devoid of oxygen or nitrate, contain low con-
centrations of iron (II) and methane (generally less than
10 mg/L),  and have relatively low concentrations  of
sulfate (generally less than 15 mg/L). Both plumes have
concentrations of dissolved organic carbon that are ele-
vated over background.  The ground  water in both
plumes is cold (near 10°C). The water is hard, with pH
near neutrality.

Both plumes discharge to surface water. The interstitial
seepage velocities of the plumes are very similar. The
seepage velocity of the TCE plume at St. Joseph (cor-
rected for retardation) is near 0.1 m/day, while the ve-
locity of the plume on the Picatinny Arsenal varies from
0.3 to 1.0 m/day. For purposes of calculation, 0.3 m/day
is used in this abstract.

Monitoring

The plume at St.  Joseph was characterized by four
transects that extended across the plume, perpendicular
to  ground-water flow. At each point in each transect,
water was sampled in 1.5-m vertical intervals extending
from the water table to a clay layer at the bottom of the
aquifer.  Each transect contains at least 20  sampling
points. Table 1 compares the average concentration of
TCE, cis-DCE, and  vinyl  chloride in each transect,  as
well  as  the  highest concentration encountered. The
most distant transect was sampled from the sediments
of Lake Michigan. The plume was encountered approxi-
mately 1.5 m  below the sediment surface, 100 m from
the shore line.

The plume of TCE at the Picatinny Arsenal is monitored
by a series of well clusters installed along the centerline
of the plume. Table 1 presents data from the monitoring
well in a cluster that had the highest concentration of
TCE. The data were collected in 1989.

Extent of Attenuation

Dechlorination in the plume at St. Joseph is extensive.
Vinyl chloride and cis-DCE accumulated near the spill,
then were degraded as the plume moved downgradient
(Table 1). Dechlorination in the plume at Picatinny Arse-
nal was  also extensive. Comparing the location of the
highest  concentration with the  point of discharge,
dechlorination destroyed  approximately 90 percent of
the TCE. Vinyl chloride and cis-DCE did not accumulate
to an appreciable extent. Because the plume at Picat-
inny  Arsenal discharged  to  surface water  before
dechlorination was complete,  the  U.S. Army installed
and continues to operate a pump-and-treat system  on
the plume.

Comparison of Attenuation due to
Dilution and Dechlorination

The plume at St. Joseph has high concentrations of TCE
at its core, while the concentration of chloride in the
aquifer is low. This  makes  it possible to estimate the
                                                13

-------
                      2001
                    SCALE
      VERTICAL EXAGGERATION 1:10
   NORTH
PARKING LOT
                       CLAY	
                                                        •CLAY-
                                                                        ...--CLAY.
                                                           ,,,"-	"
                                                                                  CLAY
Figure 1.  Cross section of the plume at St. Joseph, Michigan, as it leaves the industrial park and enters the sediments under Lake
         Michigan. Concentrations are in ug/L total chloroethenes.
                                                                        - 690     EXPLANATION
             670 -
             660 -
             650 -
             640 -
               200
                                               - 650



                                               - 640


200  400  800   800  1,000 1.200  1,400 1,600 1,800 2.000

   DISTANCE; FROM SOURCE, IN FEET

      VERTICAL EXAGGERATION  x21
          DATUM IS SEA LEVEL
                                                      Concentration of
                                                      trlohloroethylene, in
                                                      micrograms per liter

                                                       [:^;>;::| <100

                                                       \//\ 100-1,000

                                                       |s^j 1,000-10,000

                                                       |'"3J >10,000

                                                        • Sampling point
Figure 2.  Cross section of the plume at Picatinny Arsenal, New Jersey, as it moves from its source near Building 24 and discharges
         at Green Pond Brook.
contribution of dilution by comparing the accumulation
of chloride from reductive dechlorination to attenuation
of chloroethenes. Table 2 portrays the accumulation of
chloride and reduction of total organic chlorine along the
flow path. Table 2 compares water from the most con-
                               centrated sample in each transect. Based on KOC rela-
                               tionships and the fraction of organic carbon in the aqui-
                               fer, approximately 60 percent of the TCE in the aquifer
                               should be in solution. TCE was  largely depleted, and
                               sorption of cis-DCE  and  vinyl chloride in the  aquifer
                                                     14

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

                                                                            Average cone. (ug/L)
                                                                            Highest cone. (y.g/L)
Location
St. Joseph






Picatinny


Distance From Source
(m)
130
390

550

855

240
320
460
Time in Aquifer
(y)
3.2
9.7

12.5

17.9

2.2
2.9
4.2
TCE
6,500
63,000
520
8,700
15
56
<1
1.4
25,000
10,000
1,400
cis-DCE
8,100
723,000
830
9,800
18
870
<1
0.8
220
35
310
Vinyl
Chloride
930
4,400
450
7,660
106
205
<1
0.5
4
1
6
Table 2.  Comparison of the Relative Attenuation of TCE, cis-DCE, and Vinyl Chloride With the Attenuation of Chloride in the
        Plume at St. Joseph, Michigan (1, 2)

                                                                          Highest concentration
Distance From
Source (m)
Background
130
390
550
855
Chloride Ion
(mg/L)
14
55
109
71
57
Organic Chlorine
(mg/L)

104
15
0.8
<0.1
TCE

4,000
8,700
11
1.4
c-DCE

128,000
9,800
828
0.8
Vinyl Chloride

4,400
1,660
205
0.5
should be  minimal. We will assume that the organic
chlorine in ground water represents the pool of chlorine
available for dechlorination to chloride.

Near the source, the concentration of chloride plus po-
tential biogenic chloride minus background chloride was
145 mg/L. Only 38 percent of this quantity was actually
chloride. Total organic and inorganic chlorine attenuated
with  distance  downgradient. By the  time  the plume
reached the lake, the  concentration  of total chlorine
(minus background) was 43 mg/L, which is significantly
higher than background. Apparently the plume was at-
tenuated three- to four-fold due to dilution. Total attenu-
ation of chloroethenes was at least 100,000-fold.

Kinetics  of  Reductive Dechlorination in
Ground Water

Table 3 compares first-order rate constants calculated
between transects in the plume at  St. Joseph and be-
tween monitoring wells in the plume at Picatinny Arse-
nal. Field-scale estimates of rates are also compared
with attenuation in  microcosms constructed from mate-
rial collected  along the flow path at Picatinny Arsenal.
There is surprising agreement in the rates of dechlori-
nation of TCE within the same plume, between plumes,
and between microcosm studies and  field-scale  esti-
mates. Nine separate estimates vary less than an order
of magnitude. The  rates of degradation of vinyl chloride
and cis-DCE  were  comparable to the rates of degrada-
tion of TCE (Table  3).
The rates of attenuation in the two plumes are as slow
as humans experience time. In particular, they are slow
compared with the time usually devoted to site charac-
terization. In plumes with a long residence time, on the
order of decades,  however, they have significance for
protection of waters that receive the  plumes.
                                                   15

-------
Table 3.  Rates of Reductive Dechlorination of TCE, cis-DCE, and Vinyl Chloride in Ground Water (residence time refers to time in
         the segment of the plume being described, or incubation time of microcosms) (1-5)

                                                                                        Apparent Loss Coefficient (1/yr)
Location
Field Scale Estimates
St. Joseph


Picatinny




Laboratory Microcosm
Picatinny


Distance From
Source (m)

1 30 to 390
390 to 550
550 to 855
240 to 460
320 to 460
0 to 460
240 to 320
0 to 250
Studies
240
320
460
Time From
Source (yr)

3.2 to 9.7
9.7 to 12.5
12.5 to 17.9
2.2 to 4.2
2.9 to 4.2
0.0 to 4.2
2.2 to 2.9
0.0 to 2.3

2.2
2.9
4.2
Residence
Time (yr)

6.5
2.8
5.4
2.0
1.3
4.2
0.7


0.5
0.5
0.5
TCE

0.38
1.3
0.93
1.4
1.2
1.0



0.64
0.42
0.21
cis-DCE

0.50
0.83
3.1
Produced
Produced

1.6
0.5

0.52
9.4
3.1
Vinyl
Chloride

0.18
0.88
2.2
Produced
Produced







References

1.  Semprini, L, P.K. Kitanidis, D.H. Kampbell,  and J.T. Wilson. An-
   aerobic transformation of chlorinated aliphatic hydrocarbons in a
   sand aquifer based on spatial chemical distributions.  Water Re-
   sources  Res. In press.

2.  Wilson,  J.T., J.W Weaver, and D.H. Kampbell. 1994. Intrinsic
   bioremediation  of TCE in ground  water at  an NPL  site  in  St.
   Joseph,  Michigan. In: Symposium  on Intrinsic  Bioremediation of
   Ground Water.  EPA/540/R-95/515.  pp. 154-160.

3.  Martin, M.,  and I.E. Imbrigiotta. 1994. Contamination of ground
   water with trichloroethylene at the Building 24 site at Picatinny
   Arsenal, New Jersey. In: Symposium on Intrinsic Bioremediation
   of Ground Water. EPA/540/R-95/515. pp. 143-153.
4.  Ehlke, T.A., I.E. Imbrigiotta, B.H. Wilson, and J.T. Wilson. 1991.
   Biotransformation of cis-1,2-dichloroethylene in  aquifer material
   from Picatinny Arsenal, Morris County, New Jersey. In: U.S. Geo-
   logical Survey Toxic Substances Hydrology Program—Proceed-
   ings of the Technical Meeting. Water-Resources Investigations
   Report 91-4034. pp. 689-697.
5.  Wilson, B.H., T.A. Ehlke, T.E. Imbrigiotta, and J.T. Wilson. 1991.
   Reductive dechlorination of trichloroethylene in anoxic aquifer ma-
   terial from Picatinny Arsenal, New Jersey. In: U.S. Geological  Sur-
   vey Toxic Substances Hydrology  Program—Proceedings  of the
   Technical Meeting.  Water-Resources  Investigations Report  91-
   4034. pp. 704-707.
                                                              16

-------
       Intrinsic Bioremediation of a Gasoline Plume: Comparison of Field and
                                      Laboratory Results
             Morton A. Barlaz, Melody J. Hunt, Sreenivas Kota, and Robert C. Borden
                     North Carolina State University, Raleigh, North Carolina
Introduction

Assessing the potential for natural bioremediation in the
subsurface  is complicated  by site-specific conditions
and the methods used to estimate biodegradation rates.
Controlled laboratory experiments often are necessary
to verify biological loss of a compound and to assess
factors that  influence biodegradation. The effect of re-
moving samples from such a stable environment and
placing them in laboratory microcosms, however, is not
understood. In situ columns have been used to measure
biodegradation on a limited basis, and little is known
about their  reliability. In this paper, we use laboratory
microcosms and in situ column experiments to estimate
intrinsic biodegradation rates of benzene, toluene, ethyl-
benzene, and xylene (BTEX) isomers in the subsurface.

Site Background

This research was conducted  at a petroleum-contami-
nated aquifer in the southeastern coastal plain  near
Rocky Point, North Carolina. The plume  is characterized
by negligible dissolved oxygen and redox potentials of
-100 to -200 mV due to intrinsic biodegradation of BTEX
(1). The dominant electron acceptors within the plume
are sulfate and iron (1). The  midpoint  of the plume is
characterized by high  dissolved iron (Fe)(ll) (greater
than 40 mg/L) and low SO4-2 concentrations (less than
4 mg/L). Toluene and o-xylene  are nearly depleted (less
than 20 u,g/L), whereas high quantities of benzene,
ethylbenzene, and  m-,p-xylene remain (greater than
500 ug/L).

Experimental Methods

Laboratory Microcosms

Multiple  replicate microcosms  with no headspace  were
constructed  in an anaerobic chamber under aseptic con-
ditions  using blended aquifer sediment and  ground
water recovered under anaerobic conditions. Microcosm
preparation was designed to  simulate ambient condi-
tions to the maximum  extent possible. Microcosms
were spiked with approximately  10,000  u,g/L BTEX
(2,000  u,g/L of each compound) and  incubated in an-
aerobic containers stored at the ambient ground-water
temperature, 16°C. Because it cannot be distinguished
from m-xylene by the analytical  procedure used, p-
xylene  was not added.  BTEX loss was monitored by
destructively sampling three live and three abiotic micro-
cosms  at monthly intervals for 300 days.  A final time
point was taken 100 days later (after 400 days).

In Situ Columns
The  in situ columns  were similar to a system used
previously (2).  Each column consisted  of a 1-m long
chamber where sediment and ground water were iso-
lated from the surrounding aquifer. Two sets of columns,
Group  A and Group B,  were  installed at the midpoint
area of the plume. Each set contained three individual
columns: two live and one abiotic control. After  installa-
tion,  the  columns were filled with anaerobic  ground
water containing BTEX, which had been recovered from
nearby wells.  The contaminant concentrations in
water added to Group A were 1,300, 40,  1,800, 700,
and  15 u,g/L  for benzene,  toluene, ethylbenzene,
m-,p-xylene, and o-xylene, respectively. The approxi-
mate concentrations of compounds added to Group B
columns were 200, 200, 600,  1,600, and 800 u,g/L for
benzene,  toluene, ethylbenzene, m-,p-xylene,  and
o-xylene, respectively. The abiotic control columns were
prepared by adjusting the pH to less than 2 with hydro-
chloric  acid (HCI). Tracer tests were conducted on all
columns before each experiment began to ensure that
they were properly installed.

Results
A distinct order of compound disappearance was meas-
ured in the laboratory incubations: m-xylene degrada-
tion began with  no lag period, followed  by toluene,
                                                17

-------
    10000
                                                         1000 F
       ""•— Abiotic Ben/,ene
       —Q— Benzene
 Toluene
• Ethylbenzene
                                            400
- m-Xylene
~ o-Xylene
Figure 1.  BTEX biodegradation in laboratory microcosms.

o-xylene, benzene, and ethylbenzene (Figure 1). The
rate of m-xylene loss slowed once toluene loss be-
gan; once toluene and o-xylene were below 20 u,g/L
(120 days), the rate of m-xylene loss  increased. The
aquifer material was obtained in an area of the plume
where  toluene and o-xylene concentrations were very
low (less than 50 u,g/L) but significant quantities of m-,p-
xylene remained (greater than  1,000 u,g/L). Thus, the
microbial population appeared to have an initial prefer-
ence for m-xylene,  but switched to toluene and o-xylene
after a 22-day acclimation  period. Benzene began  to
biodegrade once m-xylene was depleted and was at or
below  10 u,g/L in all  microcosms at the final sampling
(403 days). First-order decay rates (K) were determined
during the time of loss for each compound (Table 1).

All  live and  abiotic in situ columns exhibited an initial
concentration  decrease of several  hundred  micro-
grams/per liter between the injection water and the first
sample taken from  the chamber. This initial loss is attrib-
uted to sorption. After the sorption loss, the initial com-
pound concentrations were less than 500 u,g/L in most
                                                       I
                                                        u
                                                        §
                                                       U
                              100
                                                           10
                                                               ll
                                         Live Column -1
                                         Live Column - 2
                                         Abiotic Column

                                        j	I	
                                                                   50
                                              100
                                      150

                                     Days
                                200    250
300
Figure 2.  m-,p-Xylene biodegradation in Group A in situ
         columns.
                           columns. The  concentrations of hydrocarbons  in the
                           abiotic  columns remained fairly  constant or declined
                           slowly after the initial decrease,  indicating biological
                           activity  or short circuiting did not occur in the control
                           columns.

                           In Group A, benzene and m-,p-xylene exhibited signifi-
                           cantly higher losses in the live columns relative to the
                           abiotic columns. The concentration of m-,p-xylene de-
                           creased in the live columns after an initial lag of 85 to
                           121  days (Figure 2). Benzene concentrations remained
                           constant in both live columns for 155 days, after which
                           time  decreases attributed to biological activity were
                           measured (data not shown). Initial toluene and o-xylene
                           concentrations were too  low (less than  50 u,g/L) to ac-
                           curately measure  concentration  changes. Figure  2
                           shows the measured m-,p-xylene loss in the Group  A
                           columns and illustrates the timeframe used to calculate
                           the decay rates. Results from the two live columns were
                           pooled  to  estimate the  live decay  rate. In Group B,
                           significant biological  loss of toluene occurred with no
                           apparent lag  time. The  short sampling period of 75
                           days was not adequate to measure losses of the other
                           compounds.
Table 1.   Comparison of Microcosm and In Situ Column
         Biodegradation Rates3
Compound
Benzene
Toluene
m,p-xylene
o-xylene
Ethylbenzene
Laboratory Rate
(percent -day"1)
(time interval in
days)
2.37 (184 to 403)
4.46 (22 to 120)
2.04 (0 to 1 84)
5.59 (37 to 120)
0.19 (0 to 403)
In Situ Column Rate
(percent -day"1)
(time interval in
days)
0.41 (121 to 251)
1.15(13to75)
1.43(121 to 251)
NS
NS
 Rate calculated is the difference between the live and abiotic loss
 rates assuming first-order model.
NS = The difference between the live and abiotic loss rate is not
significant at the 95-percent confidence level.
                           Comparison of In Situ Columns and
                           Laboratory Microcosms

                           Biological loss for three of the five BTEX compounds
                           occurred over similar periods in the laboratory and in situ
                           experiments. In both cases,  toluene  degradation was
                           followed by  m-,p-xylene and benzene. This order  is
                           consistent with previous field investigations (1). Ethyl-
                           benzene loss was minimal in the laboratory microcosms
                           during the 400 days of incubation, and no ethylbenzene
                           degradation  was measured during the  7 months of  in
                           situ monitoring.  Loss of o-xylene was not observed  in
                           the Group B columns, but fairly rapid depletion concur-
                           rent with toluene loss was measured  in the laboratory.
                           The initial concentration of o-xylene (less than 500u,g/L)
                                                   18

-------
was possibly too low to stimulate in situ degradation, or
the 75-day monitoring period could have been too short.

Although the monthly sampling frequency was consis-
tent for both types of measurements, the length of moni-
toring  was shorter in the in situ  columns due to  the
limited sample volume available. Thus, direct compari-
son of decay rates between the two types of measure-
ments is difficult. Given these limitations, the measured
rates are comparable in both columns and microcosms.
The slightly lower decay rates measured in the in situ
columns may be due to the  lower  initial concentrations
used in these experiments. Biological decay was dem-
onstrated in the controlled column and microcosm ex-
periments. Use of in situ columns could provide a prac-
tical link between laboratory evaluations and full-scale
field studies.


References

1. Borden, R.C.,  C.A. Gomez, and M.T. Becker. 1995. Geochemical
  indicators of natural bioremediation. Ground Water 82(2):180-89.

2. Gillham, R.W., R.C. Starr, and D.J. Miller. 1990. A device for in situ
  determination of geochemical transport parameters; two biochemi-
  cal reactions. Ground Water 28(6):858-862.
                                                    19

-------
      Toxicity Effects on Methanogenic Degradation of Phenol in Ground Water
                       Barbara A. Bekins, E. Michael Godsy, and Ean Warren
                           U.S. Geological Survey, Menlo Park, California
Introduction

At an abandoned creosote works located near Pensa-
cola, Florida, the shallow ground water is contaminated
with phenolic compounds, heterocyclic compounds, and
polyaromatic hydrocarbons. Based on the use of unlined
disposal ponds during the 80 years of operations at the
plant, the contaminants have probably been present in
the ground water for several decades. A methanogenic
consortium in the aquifer is degrading some of the com-
pounds, and the  concentrations of the degradable frac-
tion drop to less than 1 percent of the source values by
100 to 150 m downgradient from the nonaqueous phase
(1). In spite of the long exposure time to a continuous
source, the results of acridine orange direct counts and
most  probable number determinations indicate a low,
uniform microbial density (1). The continued existence
of low microbial numbers suggests that some factor is
limiting growth. Several possibilities, such  as microbial
transport,  nutrient limitation, predation, and  toxicity,
have  been examined.  Of these,  toxicity appears to be
the most promising explanation. The toxicity of creosote
compounds to various organisms has been studied for
a long time (2). Very little work has been done, however,
on the effects of creosote on methanogenic consortia
known to be especially sensitive to toxic compounds (3).
The results of our work indicate that the dimethylphenols
and methylphenols present in the ground water at this
site inhibit the degradation of phenol. Furthermore, in-
corporating these inhibition  effects into a one-dimen-
sional  model  of the aquifer predicts a steady-state
degradation profile and zero net growth of the active
methanogenic consortium.

Toxicity Assay Results

A serum bottle assay similar to that described by Owens
et al. (3) was performed to determine which of the
compounds present  in the ground water might  be toxic
to the methanogenic  consortium. These compounds
were grouped by class and added at three concentration
levels equivalent to 1.5, 1.0,  and 0.5 times the highest
measured field value. The classes of compounds tested
were 1) indene, benzothiophene, 2-methylnaphthalene,
biphenyl, flourene, and 2-naphthol; 2) 2-,  3-, and  4-
methylphenol; 3) 2,4- and 3,5-dimethylphenols; 4) qui-
noline  and  isoquinoline;  5) 2(1H)-quinolinone  and
1(2H)-isoquinolinone; and 6) all of the preceding com-
pounds combined. Duplicate 100-mL serum bottles con-
taining the target compounds, enriched  methanogenic
culture derived from the aquifer, mineral salts, and phe-
nol as the growth substrate (at the highest concentration
observed in  the field) were prepared in an  anaerobic
glove box and capped. The volume of gas produced in
each bottle was monitored by allowing a wetted glass
syringe inserted through the septum to equilibrate with
atmospheric pressure. Figure 1 shows the gas produc-
tion in  the bottles containing  Mixtures 2, 3, 4, and 6 at
the 1.5 times concentration as well as a control contain-
ing only  phenol. Methylphenols and  dimethylphenols
showed a substantial toxicity effect, whereas nitrogen
heterocycles had a smaller effect and  polyaromatic hy-
drocarbons had no measurable effect (data not shown).
         --»-- Methyphenols
         •-••••••• Dimethylphenols
         -••*— Nitrogen heterocycles
         —x— All combined
         —•— Control
Figure 1.  Gas production by the aquifer microbes with phenol
         as the growth substrate and various inhibitor concen-
         trations equal to  1.5 times the  maximum value ob-
         served in the aquifer. The results are not shown for
         quinolinone, which was similar to quinoline, or for
         polyaromatic hydrocarbons,  which were similar to
         the positive control.
                                                  20

-------
In the bottles with concentrations equal to those in the
field,  dimethylphenols  had  a substantial  effect,  and
methylphenols had  a smaller effect.  In the  0.5 times
concentration bottles, methylphenols and dimethylphe-
nols had a slight effect. Preliminary results showing the
buildup of fatty acids suggest that the intermediate steps
in the degradation process are being inhibited.

Model Results
The monitoring  of ground-water concentrations of the
degradable compounds for more than  12 years shows
that the concentration profiles are constant in time. The
existence of a steady-state degradation profile  of each
substrate together with a low, uniform microbial density
indicates that the microbial numbers do not change with
time.  In theory, the functional form of the Monod growth
expression  cannot be  balanced by a  constant decay
rate. To address this problem, toxicity effects are incor-
porated into the following equations for one-dimensional
substrate transport  with  degradation  and  microbial
growth:
              -V^-T?
                                                                                      0.07
S
Ksfl

sc] _ s2
*~K \+S+~K
KC I ft/


(Eq. 1)
dB
 dt ''
B
                                              (Eq. 2)
where R is the retardation factor, S is phenol concentra-
tion, v is the flow  velocity, D is dispersion, \im is the
maximum growth rate, Y is the yield, B is biomass, 9 is
porosity, Ks is the half saturation constant, K, and KC are
haldane  and competitive inhibition constants, Sc is the
concentration of the inhibiting compound, and  kj is the
biomass decay or maintenance  rate.  When only the
toxicity of phenol  is incorporated using the Haldane
inhibition model, the  predicted growth is about  50 per-
cent lower but still much higher than the only published
decay rate (4). In addition, the equations do not produce
a steady-state solution.  Incorporating the  effect of the
dimethylphenol toxicity produces a steady-state solution
for phenol and microbial concentrations that matches
the character of the data (Figure 2).  The values of the
parameters used in the solution are given in the figure
                                 100      200

                                Distance (m)
                                       100      200

                                       Distance (m)
                     Figure 2.
         Solution to Equation 1 for phenol after 2,000, 6,000,
         and 10,000 days (left), and coupled solutions to Equa-
         tion2 for microbe concentrations at the same times.
         The values of the parameters used were R =  1.01,
         S(0) = 26 mg/L, v = 1.0 mid, D = 1.0 m2/d,|im = 0.111,
         Y = 0.013, Emit = 0.005 mg/L(1.6 x 10s per 100g),6 =
         0.38, Ks = 1.33, KI = 250 mg/L, Kc = 0.52 mg/L, Sc = 23
         exp(x2/(2 (47)2)) mg/L (an empirical fit to the observed
         dimethylphenol concentrations), and kd =0.0326 d"1
caption. The model results show that the aquifer con-
centrations take  about 6 years to evolve to a steady
state, while the microbial population takes about 25 years.
The population of aquifer microorganisms oscillates
as  they  adjust their distribution to account for two
competing effects: 1) the maximum concentration of phe-
nol nearthe source should lead to maximum growth and
substrate utilization there, and 2) the maximum concen-
trations of dimethylphenols near the  source lead  to
maximum inhibition  of growth and substrate utilization.
The result is a tradeoff between a location  where the
growth substrate concentrations are higher versus one
farther from the source where the inhibitor  concentra-
tions are lower. The final microbial concentrations  stabi-
lize at about an order of magnitude higher 50 m from the
source than immediately adjacent to it.

References

1. Godsy E.M., D.F. Goerlitz, and D. Grbic-Galic. 1992. Methano-
  genic biodegradation of creosote contaminants in natural and
  simulated ground-water ecosystems. Ground Water 30:232-242.
2. Mayfield, P.B. The toxic elements of high-temperature coal  tar
  creosote. 1951. Proc. Am. Wood-Preserver's Assoc. 47:62-85.
3. Owen, W.F., D.C. Stuckey,  J.B.  Healy, L.Y. Young, and PL.
  McCarty. 1979. Bioassay for monitoring biochemical methane po-
  tential and anaerobic toxicity. Water Research Res. 13:485-492.
4. Bekins,  B.A.,  E.M. Godsy,  and D.F. Goerlitz. 1993. Modeling
  steady-state methanogenic degradation of phenols  in groundwa-
  ter. J. Contam. Hydrol. 14:279-294.
                                                     21

-------
         A Multiphase, Multicomponent Numerical Model of Bioventing With
                              Nonequilibrium Mass Exchange
                     Linda M. Abriola, John R. Lang, and Klaus M. Rathfelder
          Department of Civil and Environmental Engineering, University of Michigan,
                                       Ann Arbor, Michigan
Introduction

Soil vapor extraction  (SVE) and bioventing (BV) are
common  remediation  practices for unsaturated  soils
contaminated with volatile organic compounds (VOCs).
These methods  have  been demonstrated to be effec-
tive at comparatively low costs. The efficiency of these
techniques is known  to be restricted by soil charac-
teristics; by mass transfer limitations between phases,
including liquid/solid, liquid/gas, and liquid/microorganism
phases;  by the  availability of oxygen; and  by system
design  and operation  parameters (1). Assessment of
SVE/BV systems is often hindered by the complex inter-
play of physical, chemical, and biological  processes.
Consequently, design  and  operation of these systems
are typically based  on engineering  experience and/or
simple design equations. Numerical  models  of SVE/BV
systems can be valuable tools for the investigation of
the effects of various processes on system performance
and for optimal system design. In this work, a numerical
model is presented that has been specifically developed
to incorporate the complete range of processes occur-
ring at the field  scale  and  to include interphase mass
transfer rate limitations.

Model Formulation

Three fluid phases are modeled: gas, aqueous, and a
nonaqueous phase liquid (NAPL). The gas and aqueous
phases may flow simultaneously in response to applied
pumping/injection or density gradients. The  movement
of these phases is described by standard macroscopi-
cally averaged flow equations (2). The NAPL phase is
assumed to be at an  immobile residual  saturation.
Changes in  NAPL  saturation,  therefore, result solely
from interphase  mass  transfer.

The NAPL may be a mixture of an unrestricted number
of organic components. The gas phase  is assumed to
be composed of nitrogen  and  oxygen (the two major
constituents of air), water vapor, volatile components of
the NAPL, and a single limiting nutrient. The aqueous
phase is composed of water, oxygen, soluble compo-
nents of the NAPL, and the limiting nutrient. Sorption to
the soil particles is  restricted to  components of the
NAPL. The migration of each component in each phase
is  described  by  standard  macroscopically averaged
transport equations (2).

Quantification of the biotransformation processes fol-
lows the conceptual approach of Chen et al. (3). Biode-
gradation is assumed to occur only within the aqueous
phase by an indigenous, spatially heterogeneous, mixed
microbial population that is present as attached micro-
colonies. There is no biomass transport or detachment
or sloughing  of the microcolonies, and biomass growth
does not affect permeability. Monod-type kinetic expres-
sions are employed to  describe biophase utilization of
substrates, oxygen, and a limiting nutrient, as well as
growth of the microbial population. Additionally, a  mini-
mum biophase concentration reflecting the indigenous
population  is maintained when growth is restricted due
to  oxygen, substrate, or nutrient limitations.

A linear driving force expression is used to model non-
equilibrium interphase exchange. Interphase partition-
ing processes included in the model are: volatilization
and  dissolution  of components  from the  NAPL;
gas/aqueous exchange of oxygen, water vapor, and the
components  of the NAPL; sorption of the NAPL compo-
nents to the  soil particles through the aqueous phase;
and  rate-limited  uptake by  the biophase of oxygen,
substrate (components of the NAPL), and the limiting
nutrient.

Numerical Solution

The  flow and transport equations are solved in two
space dimensions (vertical cross  section or radial ge-
ometry) using a standard Galerkin finite element method
                                                22

-------
with linear triangular elements. A set-iterative scheme is
used for computational efficiency. The sets of coupled
flow, transport, and biodegradation equations, as well as
multiple  equations within  sets,  are decoupled and
solved sequentially. Decoupling is accomplished by lag-
ging, either by one iteration or one time step, the cou-
pling terms  which are phase density and interphase
mass exchange. Iteration within and  between equation
sets is performed to account for nonlinearities and en-
sure solution accuracy. Numerical solutions of the flow,
transport, and biodegradation equations have been  in-
dependently verified with analytical solutions and inter-
model  comparisons. A detailed description of the model
and example simulations are presented in Lang et al. (4).

Demonstration of SVE and BV Simulations

Hypothetical field-scale SVE and BV systems are pre-
sented to demonstrate model capabilities. The modeled
scenario involves the remediation  of a residual NAPL
distributed  within  a  layered soil system.  Here  the
nonuniform initial NAPL distribution was generated with
a multiphase flow model. NAPL contamination is present
in both the unsaturated zone and capillary fringe.

SVE operations are examined by simulating relatively
large pumping rates to an extraction well positioned in
the center of the  residual  NAPL zone. Removal effi-
ciency in the test simulations is shown to be sensitive to
mass transfer rates,  permeability contrasts, the initial
NAPL distribution, and, to a lesser extent, pumping rate
and well screen position.

A BV operation is also modeled by simulating small gas
injection rates at a well located  in  the center of the
contamination zone. Removal efficiency is shown to be
sensitive to flow rate, interphase mass transfer, biode-
gradation rates, and NAPL distribution.
An  example of the  complex interplay  between chemical
and biological processes in BV systems is demonstrated
in simulation  results shown  in Figure 1.  Here contami-
nant removal is compared  for BV systems run at a
comparatively low flow (0.1  pore volumes/day) and high
flow (1 pore volume/day). Total mass removed (Figure
1a) is greater at high  flow. Due to nonequilibrium inter-
                                phase partitioning occurring at the high flow, however,
                                the difference in mass removed is less than the propor-
                                tionate difference in flow rate. The rate of contaminant
                                interphase partitioning also affects the quantity of mass
                                removed  by  biodegradation  (Figure 1b),  which  is far
                                greater at high  flow. At low flow,  interphase partitioning
                                of the contaminant is approximately at equilibrium, re-
                                sulting in downgradient aqueous concentrations that are
                                greater than  an  inhibitory threshold.  Consequently,
                                biodegradation in the low-flow scenario is restricted to
                                the  region  upgradient of the  NAPL  contamination
                                zone. Nonequilibrium  partitioning at high flow rates
                                produces downgradient aqueous concentrations be-
                                low  the  inhibitory threshold, resulting in enhanced
                                biodegradation.  This  greater degradation in  the
                                high-flow system produces a reduction in the  con-
                                taminant mass in the gas phase arriving at a down-
                                gradient extraction point (Figure 1c).

                                Conclusions

                                A numerical model of SVE/BV systems has been devel-
                                oped that incorporates the complete compositional and
                                biological processes representative of field conditions.
                                Example simulations demonstrate the  model capabili-
                                ties  and illustrate the  complex  interplay  of chemical,
                                physical, and biological processes occurring  in SVE/BV
                                systems.

                                References

                                1. Rathfelder, K., J.L. Lang, and L.M. Abriola. 1995. Soil vapor ex-
                                   traction and  bioventing: Applications, limitations, and future re-
                                   search  directions.  Reviews of Geophysics  IUGG Quadrennial
                                   Report. In press.
                                2. Abriola, L.M. 1989. Modeling multiphase migration of organic
                                   chemicals in groundwater systems—A review and  assessment.
                                   Environmental Health Perspectives 83:117-143.
                                3. Chen, Y.-M., L.M. Abriola, P.J.J. Alverez, P.J. Anid, and T.M. Vogel.
                                   1992. Modeling transport and biodegradation of benzene and tolu-
                                   ene in  sandy aquifer  material: Comparisons with  experimental
                                   measurements. Water Resources Res. 28(7):1,833-1,847.
                                4. Lang, J.L., K.M. Rathfelder, and L.M. Abriola. 1995. A multiphase,
                                   multicomponent numerical model of  bioventing with non-equilib-
                                   rium mass exchange. In: Proceedings of the Bioremediation Sym-
                                   posium, San Diego, CA. April.
          15
          10
              (a) Total Mass Removed
                       (b) Biodegradation
(c)  Gas Phase Recovery
                 10
20    30    40
time (day)
                                                          — high flow
                                                           - low flow
                                                   20    30
                                                   time (day)
   10
20   30
time (day)
40   50
Figure 1.  Contaminant removal versus time for hypothetical BV scenarios at comparatively low and high flow rates.
                                                   23

-------
  Aromatic Hydrocarbon Biotransformation Under Mixed Oxygen/Nitrate Electron
                                      Acceptor Conditions
                     Liza P. Wilson, Peter C. D'Adamo, and Edward J. Bouwer
    Department of Geography and Environmental Engineering, The Johns Hopkins University,
                                        Baltimore, Maryland
Introduction

Biodegradation of contaminants associated with sedi-
ments  and ground water under mixed oxygen/nitrate
electron acceptor conditions may prove to be more suc-
cessful and feasible than remediation under strict aero-
bic or anaerobic conditions. In particular, the low level of
oxygen may allow subsurface microorganisms to attack
the aromatic rings of many organic compounds (using
oxygenases),  with nitrate serving as the electron ac-
ceptor to complete the degradation. Providing nitrate to
the subsurface is less expensive than maintaining aero-
bic conditions, and, as nitrate is highly soluble, it is
easier to maintain a residual concentration in  ground
water.

A laboratory investigation is being conducted to provide
a better understanding of the effect of dual oxygen/ni-
trate electron acceptor conditions on the biodegradation
of monocyclic and polycyclic aromatic hydrocarbon mix-
tures in aqueous solution. The specific objectives of the
research are 1) to quantify the stoichiometry and kinetics
of biodegradation of a mixture of aromatic hydrocarbons
under microaerophilic conditions (defined as less than
or equal to 2  mg/L O2), and 2) to  assess the  relative
efficacy of bioremediation under microaerophilic condi-
tions compared with strict aerobic or denitrification con-
ditions  in the  laboratory, using  batch microcosms and
aquifer sediment columns.

Microaerophilic Biodegradation by an
Enrichment of Aquifer Bacteria

The microaerophilic biodegradation of a mixture of aro-
matic compounds was  investigated  by varying com-
bined concentrations of  oxygen and  nitrate. Batch
microcosms were prepared using a liquid enrichment of
aquifer bacteria as inocula; a mixture of benzene, tolu-
ene, ethylbenzene, m-xylene, naphthalene, and phen-
anthrene as substrate; and oxygen and nitrate as elec-
tron acceptors.

The results of this study indicated that the level of oxy-
gen had a significant affect on the extent of biodegrada-
tion of most of the aromatic hydrocarbons. Analysis of
the consumption of electron acceptors indicated that
both nitrate and oxygen acted as electron acceptors
during biodegradation of the mixture of aromatic hydro-
carbons. Denitrification may be inhibited by oxygen lev-
els above  1  mg/L (1).  In  this  study, toluene and
naphthalene biodegradation was favored at microaero-
philic oxygen levels between 1.5 and 2 mg/L. The data
(measurements of oxygen and nitrate  not shown here)
suggested that oxygen and nitrate were used sequen-
tially to  biodegrade naphthalene and toluene, respec-
tively (i.e., denitrification was inhibited until oxygen was
depleted)(Table  1).

At lower levels of oxygen  (0.5 to 1 mg/L), toluene and
ethylbenzene biodegradation was favored. The mecha-
nism for biodegradation of toluene and ethylbenzene at
very low oxygen levels (less than or equal to 1 mg/L)

Table 1.  Aromatic Hydrocarbons Degraded  Under the
        Various Combinations of Oxygen and Nitrate
        Investigated (degradation is removal greater than or
        equal to  10 percent relative to killed controls)
                   Oxygen (mg/L)

Nitrate
(mg/L)
10
50
150
400
0


T
T
T
T
0.


T,
T,
T,
T,
,5


E
E
E
E
1 1


T,
T, E T,
T, E T,
T, E T,
.5


N
N
N
N
2


B,
T, N B,
T, N B,
T, N B,



T,
T,
T,
T,



E,
E,
E,
E,
8


m-X,
m-X,
m-X,
m-X,



N,
N,
N,
N,



P
P
P
P
B = benzene, T = toluene, E = ethylbenzene, m-X = m-xylene,
N = naphthalene, and P = phenanthrene
                                                 24

-------
                          Days

Figure 1.  Conversion of benzene to intermediates, cells, and
         carbon dioxide.
may be quite different than at higher levels (greater than
or equal to  1.5  mg/L). The data indicated that at low
levels (less than or equal to 1  mg/L) of oxygen, nitrate
played a  role in biodegradation of both toluene and
ethylbenzene. Evidence  for simultaneous utilization  of
nitrate and oxygen  has been documented (2). Oxygen
levels below 1 mg/L may not  inhibit denitrification and
may actually be beneficial by increasing cell numbers
(3, 4).
Benzene  was recalcitrant  under denitrifying  and mi-
croaerophilic conditions. Extensive benzene mineraliza-
tion, however, was  observed under aerobic conditions
(Figure 1). Although the  majority of benzene biodegra-
dation occurred in the presence of oxygen, partial trans-
formation  of the  parent compound to intermediates and
carbon dioxide was observed in the absence of oxygen.

Microaerophilic Biodegradation  in the
Presence of Sediments
A laboratory study  of the stoichiometry  of microaero-
philic biodegradation in the presence  of sediments  is
being conducted.  Aquifer  sediments  may  affect the
stoichiometry of microaerophilic biodegradation by ex-
erting an additional oxygen  demand from natural or-
ganic matter or reduced metals, and  by providing sur-
faces for microbial attachment that are not present in
sediment-free microcosms. In this study,  aquifer sedi-
ments were used as inocula in microcosms instead of a
liquid enrichment of aquifer bacteria. Initial results indi-
cate that toluene biodegradation under denitrifying con-
ditions occurs after a significant lag time in microcosms
containing sediment as  inocula when compared with
denitrification in liquid enrichment microcosms. This de-
lay in denitrifying activity is likely due to the development
of a sufficient denitrifying population. No biodegradation
of aromatic compounds was observed under microaero-
philic conditions (microaerophilic oxygen consumed by
sediment demands). Biodegradation under aerobic con-
ditions (7 mg/L), however, exceeded what was observed
in  the liquid enrichment microcosms. Studies of sedi-
ment microcosms and  columns are ongoing.

Acknowledgment

This research was made possible by the generous sup-
port  of  the  U.S.  Environmental Protection Agency
Robert  S. Kerr Environmental  Research Laboratory
(Project CR-821907).

References
1.  Kuhn,  E.P., J. Zeyer, P. Eicher, and  R.P. Schwarzenbach. 1988.
   Anaerobic degradation of alkylated benzenes in denitrifying labo-
   ratory columns. Appl. Environ. Microbiol. 54(2):490-496.
2.  Ottow, J.C.G., and W. Fabig. 1985. Influence of oxygen aeration
   on denitrification  and redox level in  different  batch cultures. In:
   Caldwell, D.E., J.A. Brierly, and C.L.  Brierly, eds. Planetary ecol-
   ogy. New York: Van Nostrad Reinhold. pp. 427-440.
3.  Hutchins, S.R. 1991. Optimizing BTEX biodegradation conditions.
   Environ. Toxicol. Chem. 10:1,437-1,448.
4.  Su, J., and D. Kafkewitz. 1994. Utilization of toluene and xylenes
   by a nitrate reducing strain of Pseudomonas maltophilia under low
   oxygen and anoxic conditions. FEMS Microbiol. Ecol.  15:249-258.
                                                    25

-------
                             Nutrient Transport in a Sandy Beach
           Brian A. Wrenn, Makram T. Suidan, B. Loye Eberhardt, and Gregory J. Wilson
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                         Kevin L. Strohmeier
               Environmental Technologies & Solutions, Inc., Covington, Kentucky

                                          Albert D. Venosa
 Risk Reduction  Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio
Introduction

Bioremediation of beaches that are contaminated with
crude oil is expected to be limited by the availability of
nutrients. Addition  of nutrients, such  as nitrogen and
phosphorous, has been shown to stimulate the rate of
crude oil biodegradation in the beach environment (1,
2). For bioremediation to be effective,  the nutrient con-
centration in contact with the oil-contaminated beach
material must be high enough to allow hydrocarbon-de-
grading bacteria to grow at their maximum rates.

Oil often strands in the intertidal zone (3). When this
occurs, washout will probably dominate the nutrient resi-
dence  time in the bioremediation zone. The  periodic
flooding that occurs when the tide  rises is one of several
important nutrient transport mechanisms. Wave action
can also affect nutrient transport in the intertidal region
of beaches (4), as can  the flow of fresh ground water
from inland (5, 6). The objective of this research was to
characterize the nutrient transport  rates for a low-energy
sandy beach. The results were used to determine the
nutrient application  rate that  is required for effective
bioremediation in this type of environment.

Methods

The  tracer study  was conducted on  a long,  uniform
stretch of sandy beach south of Slaughter Beach, Dela-
ware, near the southern  end of Delaware  Bay. The
beach is composed largely of coarse sand with some
gravel, which lies on top of an impermeable peat layer.
The beach  has a slope of approximately 11 percent in
the upper intertidal zone.

Eight replicate plots (5 m x 10 m) were established in
the  upper intertidal zone. The tops of the plots were
placed approximately at the spring high-tide line. The
plots were divided into two blocks of four plots each. A
water soluble tracer, either lithium or sodium nitrate, was
applied to one plot in each block at each of four different
stages in the  tidal cycle (Figure  1): spring tide (full
moon), falling midtide (i.e., as the tide proceeded from
spring to neap), neap tide (last quarter moon), and rising
midtide (i.e., as the tide proceeded from neap to spring).
(Spring tide is the point in the lunar tidal cycle at which
the difference  between high  and low tide is greatest.
Neap tide is the point at which the difference is small-
est.) Lithium nitrate was used as the  tracer  for the
             O
D
    -5
 2 -10
   -15
                        neap tide
              falling mid-tide       rising mid-tide
         spring tide                  	
     -72    0    72    144   216   288   360    432
                       time (hrs)
Figure 1. The actual and expected tidal elevations are shown,
        with the start times for the four tracer transport ex-
        periments. The elevations of the tops and bottoms of
        the plots are also shown. All elevations are relative
        to a benchmark that was placed behind the beach in
        a dune area well above the maximum high tide line.
                                                  26

-------
spring- and neap-tide experiments, and sodium nitrate
was used for both midtidal experiments. The tracer was
dissolved in fresh water (the initial concentration was
20 g/L) and applied to the plots with a sprinkler sys-
tem. Tracer was always  applied at low tide, and  the
first samples were collected shortly afterward.

The spring- and  neap-tide plots contained  multiport
wells that were  used to collect water samples from
discrete depths within  the beach. These wells were
placed along a transect through the middle of the plots
at 2.5-m intervals, beginning 2.5 m above the top of the
plots and continuing  to 5 m below the bottom of the
plots. The wells had sample ports every 6 in. The top
port of each well  was installed 3 in.  below the beach
surface. Sand  samples were  collected from randomly
selected  positions within  five  subsections that were
marked off in each plot. The locations of these subsec-
tions corresponded to the positions of the wells in  the
spring and neap-tide plots.

Sand samples were collected  as 10-in. cores using a
5-in. auger. Samples that were analyzed for lithium were
kept as two separate 5-in. cores: an upper sample (0 to
5 in.) and a lower sample  (5 to  10 in.). The cores taken
for nitrate analysis, on the other hand, were composited
into a  single 10-in. (0 to  10 in.) sample. Lithium was
extracted into 1 M ammonium acetate and analyzed by
atomic absorption spectrophotometry (7). Nitrate, which
was extracted and analyzed in the field, was analyzed
by the cadmium reduction autoanalyzer method follow-
ing extraction into 2 M potassium chloride (KCI) (8).

Results and Discussion

The average lithium concentrations in  the two 5-in. core
samples for the  spring- and  neap-tide tracer experi-
ments are shown in Figure 2. The concentrations meas-
ured in  samples collected from the five subsections of
each plot were pooled to  obtain the plot averages that
are shown in this figure. The plot averages for each of
the replicate plots are shown independently in this  fig-
ure. The average lithium concentration was reduced to
zero very rapidly  following  the spring-tide application
(Figure 2A), but it was washed out more slowly after the
neap-tide application  (Figure  2B). Washout  of  nitrate
following the two midcycle applications behaved simi-
larly; nitrate was washed out quickly following the falling
midtidal application, and it persisted for a relatively long
time following the rising midtidal application  (data  not
shown).  Although the  initial  lithium concentrations
were higher in the samples from the upper 5-in., it
disappeared more rapidly in this region than it did from the
5- to 10-in. region.

It is clear from Figure 1 that one of the  major differences
among the four tracer experiments is the extent to which
the plots were  covered with water at high tide. Wave
action also contributed to plot  coverage, and the total
       90

       80

       70

       60

       50

       40

       30

       20

       10

      100
       90

       80
       70
0-5 inches
  5-10 inches
                            B
              0 - 5 inches
                         time (hrs)


Figure 2.  Lithium concentrations in sand samples collected in
         the bioremediation zone following tracer application
         at spring tide (A) and neap tide (B). The concentra-
         tions for the upper (0 to 5 in.) and lower (5 to 10 in.)
         samples are plotted separately for each experiment.

coverage is the sum of both effects. The cumulative
effects of plot coverage on nutrient retention are shown
in  Figure 3,  in  which the ratio of the remaining sand
nitrate concentration to its initial concentration is plotted
as a function of the maximum extent of plot coverage
that had occurred before collecting each sample (e.g.,  if
       1.00 1
       0.75 -
O
S     0.50 ;
O

       0.25
       0.00 '  '  '  ' '
          0.00
    0.25
              0.50
0.75
                                 1.00
                     maximum plot coverage
Figure 3.  Relationship between the fraction of nitrate remaining
         and the maximum extent to which water had covered
         the plots before the samples were collected. Data
         from all four tracer experiments are included in this
         plot.
                                                    27

-------
the first high tide following tracer application covered 75
percent of the plot with water and subsequent high tides
covered only 50 percent, the maximum coverage for all
of the data  collected after the first high tide  is reported
as 75 percent). Data from  all four tracer experiments,
calculated as described for the lithium data plotted in
Figure 2, were used to construct this plot.  Figure  3
shows a strong correlation  between the  maximum ex-
tent of plot coverage and the remaining nitrate concen-
tration,  suggesting  that  nutrient  retention  in   the
bioremediation zone of a sandy beach can be predicted
based solely on the extent of water coverage.

The simplest explanation for the results shown in Figure
3 is that the tracers become diluted by mixing  with bulk
seawater when the plots are covered by the rising  tide,
and they are washed away when the tide recedes.  This
explanation is consistent with a model for nutrient trans-
port in  a  beach in Prince William  Sound, Alaska  (9).
Pore-water  data, however, show that the tracer move-
ment is  predominantly downward into the beach (Figure
4).  These data are  plot averages, calculated as  de-
scribed  above, for all pore-water samples collected  from
each depth within the plot. Samples collected from wells
outside  of the plots were not used to compute these
averages.

Figure 4 shows that nutrient is  probably  removed  from
the bioremediation zone by advective flow through the
porous  matrix of  the beach, not by  mixing with  bulk
seawater.

Conclusions

Effective  bioremediation requires  a sufficient supply of
the growth-limiting substrate to  be available to the  bac-
teria responsible for biodegradation. For bioremediation
of oil-contaminated beaches, nutrients such as nitrogen
Figure 4.
Average lithium concentrations in pore water col-
lected at discrete depths below the beach surface for
three time points during the spring tide tracer experi-
ment. All samples were collected at high tide.
and  phosphorus are expected to be most important (1,
2), and washout is expected to be the dominant nutrient
removal mechanism. In many cases, it will be desirable
to optimize nutrient application rates to minimize costs
and  to reduce the opportunity for  eutrophication of
adjacent bodies of water. Our data  suggest a simple
method for  determining the  frequency  required for
nutrient application.

Nutrient retention  in the bioremediation  zone of the in-
tertidal region of a sandy beach varies with the lunar
tidal cycle. Our tracers were washed out of the bioreme-
diation zone very quickly when they were applied during
and  shortly after  the  spring  tide (when the high  tide
reached  its  maximum elevation), but  they persisted
through several tidal cycles when applied around neap
tide, when the lowest high tides occurred. Our  data
suggest that the differences in nutrient retention time are
related to the maximum extent to which our experimen-
tal plots were covered by water at high  tide. Total  cov-
erage appears to be more important than coverage due
to the tide alone. Therefore, wave activity and  tidal
elevation  must both  be considered  to  determine the
appropriate fertilization frequency. When water-soluble
fertilizers  are used, visual  inspection of the extent to
which  the contaminated  area is covered  by water
during high tide is a reliable alternative to  expensive
and  time-consuming  chemical analyses for nutrient
concentration.

References

1. Lee,  K., G.H. Tremblay, and  E.M. Levy.  1993. Bioremediation:
  Application of slow release fertilizers on low-energy shorelines. In:
  Proceedings of the 1993 Oil Spill Conference, American Petroleum
  Institute, Washington, DC. pp. 449-454.
2. Pritchard,  PH., and C.F. Costa. 1991. EPA's Alaska oil spill  biore-
  mediation  project. Environ. Sci. Technol. 25:372-379.
3. Payne, J.R., and C.R. Phillips. 1985. Petroleum spills in the marine
  environment: The chemistry and formation of water-in-oil  emul-
  sions and  tar balls. Chelsea, Ml: Lewis Publishers.
4. Brown, A.C., and A. McLachlan. 1990. Ecology of sandy shores.
  New York, NY: Elsevier.
5. Glover, R.E.  1959. The pattern of fresh-water flow in  a coastal
  aquifer. J.  Geophys. Res. 64:457-459.
6. Nielsen, P. 1990. Tidal dynamics of the water table in beaches.
  Water Resources Res. 26:2,127-2,134.
7. American  Public Health Association. 1989. Direct air-acetylene
  flame method 3111B. In: Standard methods for the examination of
  water and wastewater, 17th ed., Washington, DC.
8. Keeney, D.R., and D.W Nelson. 1982. Nitrogen—inorganic forms.
  In: Page, A.L., R.H. Miller, and  D.R. Keeney, eds. Methods of soil
  analysis,  Part 2. Chemical and microbiological properties.  Madi-
  son, Wl: American Society of Agronomy, Inc. pp. 643-698.
9. Wise, W.R., O. Guven, F.J. Molz, and S.C. McCutcheon. Nutrient
  retention time in a high permeability, oil-fouled beach. J. Environ.
  Engin. In press.
                                                     28

-------
              Bioremediation of Crude Oil Intentionally Released on the
                           Shoreline of Fowler Beach, Delaware
                              Albert D. Venosa and John R. Haines
                     U.S. Environmental Protection Agency, Cincinnati, Ohio

  Makram T. Suidan, Brian A. Wrenn, B. Loye Eberhart, Miry am Kadkhodayan, and Edith Holder
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                       Kevin L. Strohmeier
             Environmental Technologies and Solutions, Inc., Covington, Kentucky

                                           Dennis King
                               Kingstat Consulting, Fairfield, Ohio

                                        Bennet Anderson
    Delaware Department of Natural Resources and Environmental Control, Dover, Delaware
Introduction

A major factor contributing to the equivocal findings of
past field  studies (1-5) was that conclusions were usu-
ally based on comparisons between one large treatment
plot and one large control plot. The problem with this
type of experiment is that no replicate plots  are estab-
lished to  provide a basis  for estimating experimental
error. The collection of numerous subsamples from one
treatment plot and one control plot, termed pseudorep-
lication (6), is statistically invalid for drawing  inferences
on treatment effects because no experimental error can
be computed. An experiment lacking  replication is an
uncontrolled experiment because it does not control for
among-replicate variability inherent in the experimental
material, introduced by the experimenter, or arising from
chance occurrences. To  eliminate uncontrollable and
unknown  environmental factors that could skew results
in one direction, several replicate plots must be set up
in  random fashion on the beach  surface. The experi-
mental approach described  herein was carefully de-
signed to allow for valid  and  statistically  authentic
comparisons between treatments.

The goals of the study were 1) to obtain sufficient sta-
tistical and scientifically credible evidence to determine
whether bioremediation with inorganic mineral nutrients
or microbial inoculation enhances the removal of crude
oil contaminating mixed sand and gravel beaches, and
2) to compute the rate at which such enhancement takes
place.

Materials and Methods

The plan was to maximize the effectiveness of bioreme-
diation by maintaining a certain level of nitrogen, in the
form of nitrate (agricultural grade sodium nitrate), and
phosphorus, in the form  of tripolyphosphate  (sodium
tripolyphosphate), in contact with the degrading popula-
tions so that they would be able to grow at their maximal
rates at all times. In a previous study (7), we had shown
that the minimum nitrate-N concentration needed by oil
degraders to grow on hydrocarbons at an  accelerated
rate under semicontinuous flow conditions was approxi-
mately 1.5 mg/L. It was also known that, if the incoming
tide completely submerged the plot, the levels of nitrate
in the interstitial pore-water diluted to undetectable limits
(8). Thus, to  maximize bioremediation  during spring
tides, we reasoned that  nutrients would  have  to be
applied every day.  To achieve the target 1.5 mg/L inter-
stitial  pore-water concentrations, we assumed a  100-
fold safety factor to account for dilution. The amount of
nitrate-N needed under these circumstances was thus
calculated to be about  55 g/m2, applied once daily to
each plot.

The approach used to assess treatment effects in the
field  study was  a  randomized complete block (RGB)
                                                29

-------
design with repeated measures.  Five  areas of beach
were  selected based on the homogeneity of geomor-
phology within each area. Each area ("block") was large
enough to accommodate four experimental units or test
plots. The blocks were situated in a  row on the beach
parallel to the shoreline. Three treatments were tested
on oiled  plots: no-nutrient control, water-soluble  nutri-
ents,  and water-soluble nutrients  supplemented with a
natural microbial inoculum from the site. The inoculum
was grown  by isolating a mixed culture from the site and
adding it to a 55-gal drum containing Delaware  Bay
seawater, the same Bonny Light crude oil, and the same
nutrient mix used on the beach. A fourth treatment, an
unoiled and  untreated plot, served  as a background
control for microbiological characterization and baseline
bioassays.  The  four treatments  were  randomized in
each of the five blocks so that whatever inferences could
be ascertained from the data would be applicable to the
entire beach, not just the test plots.

Results and Discussion

The mean interstitial  nitrate-N concentrations measured
over the course of the 14-week investigation were:
0.8 + 0.3 mg/L in the unamended control plots, 6.3 +
2.7 mg/L in the nutrient-treated plots, and 3.5 + 1.7 mg/L in
the inoculum-treated plots. These results indicate that
background nutrient  levels on Fowler Beach were high
enough to sustain nearly maximum oil degrader growth.

Figure 1 is a summary of the hopane-normalized alkane
(Figure 1) and aromatic (Figure 2) oil  components re-
maining after the first 8 weeks of the study. In regards
to the alkanes, statistical analysis of variance (ANOVA)
showed that the difference between  both treated  plots
and the control plots  were highly significant at Weeks 2,
4, and 8 (p  < 0.01) but not at Week 6.  Differences
between the nutrient-treated and inoculum-treated plots
were  not significant at any time. Clearly, substantial
natural biodegradation was taking place on the control
                       4        6

                       time, weeks
                                                 10
                       4        6

                       time, weeks
                                                 10
Figure 2.
Decline in hopane-normalized total PAHs during the
first 8 weeks.
Figure 1.  Decline in hopane-normalized total  alkanes during
         the first 8 weeks.
plots without addition of nutrients. This observation  is
consistent with observations made on the background
nutrient levels  existent  on Fowler Beach.  Addition  of
nutrients significantly enhanced the natural biodegrada-
tion rates  of the alkane fraction,  but not to the extent
expected on a less eutrophic beach.

With respect to the aromatic components, results of the
ANOVA revealed no significant differences  among any
of the  treatments at Weeks 0, 2, 4,  and 6,  although
substantial biodegradation of  the polycyclic  aromatic
hydrocarbons (PAHs) occurred on all plots.  At Week  8,
statistically significant differences between  the treated
and untreated plots were evident. Most of the disappear-
ance occurred among the two-  and three-ring PAHs and
the  lower  alkyl-substituted   homologues  (data  not
shown). The four-ring PAHs began to show  evidence  of
biodegradation  during  the eighth week  of the study.
These  results suggest that biostimulation may not al-
ways be necessary to  promote bioremediation  if suffi-
cient nutrients are naturally present at a spill site in high
enough concentrations  to effect  natural cleanup. The
evidence suggests that  nutrient application  to maintain
a residual nitrate concentration in the interstitial waters
at high enough levels to sustain maximum biodegrada-
tive metabolism resulted in a significant enhancement  of
alkane and, to a lesser extent,  aromatic biodegradation
over natural attenuation. Bioaugmentation (i.e., supple-
mentation  with  a bacterial  inoculum indigenous to the
area), however, did not appear to result in further en-
hancement.


References

1.  Pritchard, P.M., and C.F. Costa. 1991. EPA's Alaska oil spill biore-
   mediation project. Environ. Sci. Technol. 25:372-379.

2.  Sveum, P., and A. Ladousse. 1989. Biodegradation of oil in the
   Arctic: Enhancement by oil-soluble fertilizer application. In: Pro-
   ceedings of the  1989 International Oil Spill Conference. Washing-
   ton, DC: American Petroleum Institute.
                                                   30

-------
3.  Sveum, P. and Ladousse, A. 1989. "Biodegration of Oil in the Artie:
   Enhancement by Oil-Soluble  Fertilizer Application." In: Proceed-
   ings of the 1989 International Oil Spill Conference. Washington,
   DC: American Petroleum Institute.

4.  Rosenburg, E., R. Legmann, A. Kushmaro, R. Taube, R. Adler, and
   E.Z. Ron.  1992.  Petroleum bioremediation—a multiphase prob-
   lem. Biodegradation 3:337-350.

5.  Lee,  K., G.H. Tremblay, and  E.M.  Levy. 1993. Bioremediation:
   Application of slow-release fertilizers on low-energy shorelines. In:
   Proceedings of the 1993 International Oil Spill Conference. Wash-
   ington, DC: American  Petroleum  Institute.

6.  Hurlbert, S.H. 1984. Pseudoreplication and the design of ecologi-
   cal field experiments.  Ecol. Monographs  54(2):187-211.
7.  Venosa, A.D., J.R.  Haines,  M.T.  Suidan,  B.A. Wrenn, K.L.
   Strohmeier, B.L. Eberhart, E.L. Holder, and X. Wang. 1994. Re-
   search leading to the bioremediation of oil-contaminated beaches.
   In: Symposium  on Bioremediation of Hazardous Wastes:  Re-
   search, Development, and Field Evaluations, June 28-30, 1994,
   San Francisco, CA. EPA/600/R-94/075. pp. 103-108.

8.  Venosa, A.D., M.T.  Suidan,  B.A.  Wrenn,  J.R.  Haines, K.L.
   Strohmeier, E.L. Holder, and B.L. Eberhart. 1994. Nutrient appli-
   cation strategies for oil spill bioremediation in the field. In: Twen-
   tieth Annual RREL Research  Symposium, March 15-17, 1994,
   Cincinnati, OH. EPA/600/R-94/011. pp. 139-143.
                                                              31

-------
  Dynamics of Oil Degradation in Coastal Environments: Effect of Bioremediation
                      Products and Some Environmental Parameters
                               Marirosa Molina and Rochelle Araujo
                     U.S. Environmental  Protection Agency, Athens, Georgia

                                         Jennifer R. Bond
                                    DYNCORP, Athens, Georgia
Introduction

Oil extraction, refining, and transshipment are often lo-
cated in coastal regions, putting wetland systems at risk
of exposure to spilled oil and oil products. The inacces-
sibility of wetlands and the fragile nature of those eco-
systems preclude mechanical cleanup  of oil, making
bioremediation a preferred option. Moreover, the high
level of indigenous microbial activity suggests a poten-
tial for biodegradation, especially if fertilizer additions
can relieve environmental nutrient limitations.

Bioremediation strategies that have been proposed for
oil spills in wetlands include fertilization, solubilization of
oil, and bioaugmentation with oil-degrading  bacteria.
Although bioaugmentation has been demonstrated to be
effective in engineered systems, the ability of introduced
organisms to establish themselves in the complex web
of microbial relationships that characterize wetlands is
questionable.  Moreover, supplies of oxygen and nutri-
ents may be insufficient for concurrent degradation of oil
and natural substrates. In this research, we propose a
system for assessing  the efficacy of bioremediation in
wetlands and for testing the effectiveness of several
bioremediation products. We also present data on the
dynamics of the bacterial populations and the relative
rates of degradation of natural substrates and oil in
wetlands.

Methods and Materials

Sediment microcosms were constructed using glass col-
umns (10-cm internal diameter, 20-cm length) fitted with
fritted glass supports and filled with homogenized marsh
sediments  from Sapelo Island,  Georgia. Seawater or
artificial seawaterwas adjusted to a brackish salinity of
20%o and  exchanged on a  tidal basis.  An artificially
weathered Alaska North Slope (ANS)  crude  oil was
added at low tide to cover the sediment surface to a
depth of 0.5 mm.

Bioremediation products consisted of bacterial prepara-
tions enriched in oil degraders, nutrients, surfactants, or
combinations thereof. Products were applied in a man-
ner consistent with the manufacturer's recommenda-
tions. In  addition to commercial products, inorganic
nutrients were added to microcosms to test the potential
response of  indigenous bacteria. Ground Spartina al-
terniflora, a salt marsh grass common to the coast of
Georgia, was used as an alternate natural organic sub-
strate. The grass (4.5  g) was  added to the sediment
surface in an amount equivalent to a 1-yr standing stock
of aboveground biomass.

After 3 months,  the residual hydrocarbons were  ex-
tracted and  analyzed  by gas chromatography/mass
spectrometry (GC/MS).  Efficacy was assessed on  the
basis of reduced concentrations of specific components
of oil, including straight-chained and branched alkanes
and aromatics, expressed as ratios to conserved inter-
nal markers. Microbial populations were estimated using
modified heterotrophic  and oil  most probable number
(MPN) techniques (1). Heterotrophic bacteria were also
quantified using standard plate counts.

Deoxyribonucleic acid  (DMA) samples were extracted
from each microcosm using a modification of the method
of Tsa  and Olson (2).  Target  DMA was filtered onto
maximum-strength Nytran membranes. Sediment DMA
was loaded onto the membranes in triplicate. The mem-
branes were  hybridized with a Pseudomonas23S rRNA
oligoprobe (Group I). Detection was carried out using
Rad-Free Lumi-Phos 530 substrate sheet and exposure
to x-ray film for 3 hr at room temperature. The signal was
quantified using densitometry.
                                                 32

-------
Results and Discussion

The alkane fraction of oil was degraded in the presence
or absence of bioremediation  products, as indicated by
the absence of C13 and C14 compounds and the greatly
diminished peaks for C13 through C33, including the
branched alkanes, pristane, and  phytane. Addition  of
inorganic nutrients to sediments containing only indige-
nous bacteria resulted in greater depletion of the alkane
fraction than did additions of products  composed  of
surfactants or bacterial enrichments. Moreover, in the
presence of surfactants, the extent and range of degra-
dation was less than that in the control treatment, sug-
gesting inhibition of microbial activity.

The aromatic fractions of the oil  (Figure 1) were de-
graded to a lesser extent in all treatments than were the
alkane fractions (data  not shown),  although the rank
order of the treatments was the same. Both the range
of compounds degraded  and  the  extent to which they
were degraded were significantly greater in  the treat-
ment containing nutrients than in the  presence of the
commercial  products.

The abundance of oil-degrading bacteria (Figure 2) was
not consistent with the extent  of oil  degradation ob-
served. In the absence of bioremediation products, the
number of oil degraders (1.51 x 103 bacteria/g sediment)
after 3 months of experimentation was not appreciably
different from that observed in the initial  Sapelo Island
sediment (1.38 x 103). Even at such low numbers, how-
ever, the  microbial community was capable of degrading
the alkane fraction of oil, especially when fertilized with
inorganic nutrients. No  enhanced  degradation resulted
from the addition  of a product purported to  be en-
riched in  oil-degrading bacteria, regardless of the rela-
tively high numbers of total heterotrophs and potential
oil-degrading bacteria that the treatment yielded.
                      Sediment layer
 Ratio to C-2 Chrysenes
D NUTRIENTS
  OILDEG.
D CONTROL
D SURFACT.
  521 OIL
      O O
                              O O O
Figure 1.  Residual aromatics in surface layer of microcosms
         after 3 months in the presence of bioremediation
         agents.
                                                  5.33
                       Middle
                      Bottom
                                          3.25
                                                4.92
                            4.86

                            4.86


                            14.92
                                    2468
                                      Log bacteria/g sediment
                                              10
                          D Nutrients  D Added Oil Degraders • Control
                          B Surfactants E3 Sapelo Is. Sed.
                    Figure 2.  Distribution of oil degraders in wetlands microcosms
                             after 3 months of exposure to bioremediation agents.
The inhibition of oil degradation by the addition of sur-
factant was consistent with  the microbial  numbers,
which show that, in the presence of the surfactant, the
abundance of oil degraders was less than in the treat-
ments containing nutrients or added oil degraders. The
heterotrophic bacteria,  however, were not affected by
the addition of surfactants, as indicated by the numbers
in  this treatment being comparable with the numbers
obtained  in the presence of added nutrients (data not
shown). The surfactant product could have been used
as a carbon  or nutrient  source by the heterotrophic
bacteria,  thus  increasing their numbers.  Furthermore,
the surfactant-based product  may  be inhibitory to the
indigenous oil degraders, which were only enhanced in
the presence of inorganic nutrients.

Enumeration of oil-degrading bacteria gives no informa-
tion about the source of the active bacteria because the
oil-degrading  bacteria may have been of sediment or
product  origin. DMA hybridizations  with 16S rRNA
oligoprobes indicated that Pseudomonas Group I  was
significantly reduced in the presence of the microbial
product. This suggests that addition of the bacterial
preparation suppressed some  indigenous populations,
including perhaps the indigenous oil-degrading bacteria.
None of the  other  treatments produced  numbers of
                                                   33

-------
Pseudomonads that were significantly different from
those observed in the initial  sediment; data on other
sediment microbial groups will be forthcoming.

Under nutrient-amended conditions, both alkane  and
aromatic fractions were degraded relative to the 521 oil;
however,  additions of Spartina alterniflora detritus de-
creased the extent of degradation for both fractions. The
results for aromatic constituents are shown  in Figure 3.
Given that microbial activities  in wetlands ecosystems
are usually limited by oxygen and nutrients, oil degrada-
tion  may  compete with  the degradation of natural or-
ganic substrates for these substances. In addition, the
hemicellulose fraction  of  lignocellulose constitutes a
readily degradable carbon source  that may compete
with the  utilization of petroleum hydrocarbons by the
indigenous microbial community.

In the treatments containing no oil, the number of oil
degraders was constant over the 3-month  period  (ap-
proximately 6.31 x 10s bacteria/g of sediment),  which
indicates  that addition of nutrients only does not in-
crease the population of degraders. The two treatments
containing oil produced an increase of oil  degraders
during the second month of experiment (Figure 4), which
may correspond to the initial, rapid  degradation  of the
alkane fraction. By the third month, the oil degraders had
decreased, possibly reflecting the slower rate of degra-
dation of aromatic compounds.

The numbers of heterotrophic bacteria were higher in
the presence of Spartina and oil together than in any of
the other treatments. Although the  numbers of oil de-
graders in both treatments containing oil were very simi-
lar, the  extent  of oil degradation was  significantly
different. The diminished oil degradation in the presence
of oil and Spartina suggests that  the  combination of
substrates may antagonize the activity of indigenous oil
degraders, either directly or by competition for nutrients
and/or oxygen. Given the abundance of Spartina alterni-
flora in coastal regions of the United  States, the interac-
tions  between oil-degrading  and  other heterotrophic
bacteria and the impact of natural substrates should not
be overlooked in the design of remediation strategies.

References
1.  Brown, E.J., and J.F. Braddock. 1990. Sheen screen, a miniatur-
   ized most-probable-number method for enumeration of oil-degrad-
   ing microorganisms. Appl. Environ. Microbiol. 56:3,895-3,896.
2.  Tsai, Y, and B.H. Olson. 1991. Rapid method  for direct extraction
   of DMA from soil and sediment. Appl. Environ. Microbiol. 57:1,070-
   1,074.
Ratio to C-2 Chrysenes
                      DOil
                      D Oil + Spartina
                      H 521 oil control
   ^ -C -C p .c — :3 ^ ;3 i— *--•-• -i-1 ^ ^ y — Q_D_-C

   "fl-fl-  v^9llllsSSS6o
     ""^       fificiCLn-n^^ccc
               0 ° ° ^ ^ Q- Q- s % £ £
• z
i •*

     OO Q O
                               T|- C\l CO
                               666
Figure 3.  Residual aromatics in surface layer of microcosms
         after 3 months in the presence of a natural organic
         substrate.
                  Sediment + Oil
              B Oil degraders EH Heterotrophic bacteria
               Sediment + Spartina + Oil
              H Oil degraders D Heterotrophic bacteria
                                                        Figure 4.  Heterotrophic bacteria and oil degraders in surface
                                                                 layer of microcosms after 3 months of exposure to
                                                                 oil.
                                                     34

-------
            Progress  Toward Verification of Intrinsic Cobioremediation of
                                     Chlorinated Aliphatics
                                            Mark Henry
                  Michigan Department of Natural Resources, Oscoda, Michigan
A plume consisting of chlorinated aliphatic and aromatic
hydrocarbons mixed with JP-4 (benzene, toluene, ethyl-
benzene, and the xylenes;  BTEX) from a former Air
Force fire-fighting training area shows evidence of con-
tinual natural bioremediation. This site is  being charac-
terized  and monitored by  the National  Center for
Integrated Bioremediation  Research and Development
(NCIBRD) at Wurtsmith Air  Force Base, a decommis-
sioned installation located  in lower northeast Michigan.

Wurtsmith is bounded by the Au Sable Riverto the south
and west and by Van Etten Lake to the north and east.
The  base sits on a 20-m bed of homogeneous glacial
alluvial sand and gravel aquifer underlain by a thick clay
aquitard. The average  ground-water depth in the study
area is 6 m. The hydraulic  conductivity  has been re-
ported to be 3e-3 cm/sec (v =  2.6 m/day) at the site,
which has resulted  in a narrow  50 m x 300+ m plume.
The plume is monitored on a  quarterly basis through the
use of dedicated bladder pumps installed  in 37 monitor-
ing wells at the site. Local ground-water elevations are
continually  recorded by a datalogging network of re-
corder wells.

The  site has been characterized through 2 years of
quarterly sampling of the well and pieziometer network,
as well as direct analysis of continuous cores (gathered
by resident Geoprobe sampling equipment) across the
plume. This information is supplemented by a  weekly
monitoring of the vertical temperature profile of the site
and periodic soil gas profiles.
In general, the site has a  large amount of residual
fuel/solvent residing near the interface of the  water
column and capillary fringe, extending at least 125 m
from the source. Soil gas measurements in the vadose
zone near the source indicate that the interstices con-
tain  approximately 65 percent methane,  30 percent
carbon dioxide, ppb level hydrogen sulfide and nitrogen,
and virtually no oxygen. The ground water beneath the
free product has almost no dissolved oxygen  and  has
depressed redox potential, increased electrolytic con-
ductivity, depressed  pH levels, and increased concen-
trations of reduced iron. BTEX levels steadily decrease
over the length of the  plume. While perchloroethylene
(PCE) and trichloroethylene (TCE) levels are significant
(greaterthan 1,500 mg/kg) in the solids, only trace levels
are found  in the ground water. As the dissolved plume
moves downgradient, the predominant chlorinated spe-
cies  are  cis-1,2-dichloroethylene and  vinyl chloride.
On the fringe  of the contamination, BTEX  metabolites
such as m,p-toluic acid and  salicylic acid have been
identified.
The disappearance  of TCE, PCE, and  BTEX and the
appearance of  bacterial metabolites  of these com-
pounds over the length of the plume suggest that these
contaminants are being bioattenuated within the same
plume. Changes in redox potential, temperature, and pH
support this assumption. It remains to be seen whether
or not these processes are interrelated, and are perhaps
influenced by  bacterially mediated iron  or manganese
reduction.
                                                 35

-------
                   Phytoremediation of Petroleum-Contaminated Soil:
                        Laboratory, Greenhouse, and Field Studies
                                        M. Katharine Banks
          Department of Civil Engineering, Kansas State University, Manhattan, Kansas

                                          A. Paul Schwab
              Department of Agronomy, Kansas State University, Manhattan, Kansas
Common environmental problems associated with the
pumping  and refining of crude oil are the disposal  of
petroleum sludge and pipeline leaks. Contaminants are
often treated by incorporation into the soil. If the soil is
frequently tilled  and  fertilized, soil microorganisms will
be stimulated and organic contaminants biodegraded.
Unfortunately, the biodegradation rate of more recalci-
trant and potentially toxic contaminants, such as the
polycyclic aromatic hydrocarbons (PAHs), is rapid at first
but declines  quickly. Biodegradation  of these  com-
pounds is limited by their strong adsorption potential and
low solubility.
Recent research suggests that vegetation may play an
important role in the biodegradation of toxic organic
chemicals in soil. The establishment of vegetation on
hazardous waste sites may be an economical, effective,
low-maintenance approach to waste remediation and
stabilization. The use of plants for remediation may be
especially appropriate for soils contaminated by organic
chemicals to depths of  less than 2  m.  The beneficial
effects of vegetation on the biodegradation of hazardous
organics  are two-fold: organic  contaminants may be
taken up by the plant and accumulated, metabolized, or
volatilized; and the rhizosphere microflora may acceler-
ate biodegradation of the contaminants.
Completed greenhouse studies indicate that vegetative
remediation is a feasible method for cleanup of surface
soil contaminated with petroleum products. Afield dem-
onstration is necessary,  however, to exhibit  this new
technology to the industrial community. In this project,
several petroleum-contaminated field sites have been
chosen in collaboration with three industrial  partners.
These sites have been  thoroughly characterized for
chemical  properties, physical properties, and initial con-
taminant concentrations. A variety of plant species have
been established on the sites, including warm and cool
season grasses and  legumes.  Soil analyses for the
target compounds overtime indicate that the interaction
between plants and rhizosphere microflora significantly
enhances remediation of the contaminated soils. Con-
tinued monitoring will allow us to assess the efficiency
and applicability of this remediation approach.
                                                 37

-------
                                     Section 3
                           Performance Evaluation
In an effort to evaluate the performance of various bioremediation technologies, researchers assess
the extent and rate of cleanup for particular bioremediation methods. They also study the environ-
mental fate and effects  of compounds  and their byproducts  because remediation efforts at a
contaminated site can produce intermediate compounds that can themselves be hazardous. Thus,
another important aspect of performance evaluation projects involves assessing the risk of potential
health effects and identifying bioremediation approaches that best protect public health.

To this end, EPA's National Health and  Environmental  Effects Research Laboratory (NHEERL)
developed an integrated program to address 1) the toxicity of known hazardous waste site contami-
nants, their natural breakdown products,  and their bioremediation products; 2) the development of
methods to screen microorganisms for potential adverse health effects; 3) the potential for adverse
effects when  chemical/chemical chemical/microorganism interactions occur; and 4) the develop-
ment of methods to better extrapolate toxicological bioassay results to the understanding of potential
human toxicity.

Two  performance evaluation  papers were presented at the symposium studying the toxicity of
hazardous waste mixtures before and after bioremediation. One group of microorganisms (Pseudo-
monas aeruginosa and Phanerochaete chrysosporium) were found to be unable to significantly
degrade or eliminate the  mutagenic activity of a mixture of several aromatic chemicals.
                                          39

-------
      Detoxification of Model Compounds and Complex Waste Mixtures Using
                       Indigenous and Enriched Microbial Cultures
              K.C. Donnelly, Jeannine L. Capizzi, Ling-Yu He, and Henry J. Huebner
                          Texas A&M University, College Station, Texas
Introduction
Biological degradation of organic compounds may be
considered an economical tool for remediating hazard-
ous waste contaminated environments. While some en-
vironments may be too severely contaminated for initial
in  situ  treatment to  be  effective,  most contaminated
media will use some form of biological degradation in
the final treatment phases. Traditional studies monitor
the success of bioremediation by the loss of a parent
compound or class of chemicals. Prior studies, however,
have observed both an increase and decrease in toxicity
as a result of degradation (1-3). This study uses both
genotoxicity bioassays and chemical analysis to monitor
the efficacy of biological degradation of model com-
pounds and a complex chemical  mixture in soils.

Methods
A Weswood sandy loam soil was allotted in 7-kg portions
into 40 stainless steel boxes. Four soil treatment cate-
gories included: 1) no amendment; 2) amendment with
three  model chemicals,  i.e.,  benzo(a)pyrene  (BAP),
2,4,6-trinitrotoluene   (TNT),  and  pentachlorophenol
(PCP); 3) amendment with a wood preserving bottom
sediment waste (WPW); and 4)  amendment with both
the model chemicals and WPW. Additionally, each of the
four soil treatment categories were divided into  three-
box subunits and inoculated with either an indigenous
culture,  a Pseudomonas aeruginosa culture,  or  a
Phanerochaete chrysosporium culture. Thus, a total of
40  boxes  were  prepared, including  the soil  with no
chemical amendment and the soil with indigenous, bac-
teria, or fungi  inocula (Nl, NB, NF); a wood preserving
waste amended soil with each of the three inocula (Wl,
WB, WF); soil amended with the model chemicals and
each  of the  microbial  inocula (Ml, MB, MF); soil
amended with both the model chemicals and wood pre-
serving waste and each inolula (Cl, CB, CF); and steril-
ized soil receiving each  of the chemical  amendments
(St-N, St-W, St-M, St-C). The soil boxes were stored in
a greenhouse under constant temperature and humidity,
and monitored to maintain moisture content. Samples
were collected on Days 0, 90, 180, and  360 posttreat-
ment. The solvent was extracted with methylene chlo-
ride and methanol, then evaluated for genotoxicity in the
Sa/mone//a/microsome  assay.  Benzo(a)pyrene  and
pentachlorophenol were quantified using gas chroma-
tography (GC), whereas TNT was quantified using high-
performance liquid chromatography (HPLC).

Results
The data presented in Table 1 describe the quantity of
solvent-extractable  organics recovered  from  the  soil
treatments using methylene chloride and  methanol. The
data indicate that smaller quantities  of  organics were
recovered  from the  treatments with indigenous organ-
isms on 0 and 30-day sampling periods.  The soil treat-
ment with the wood  preserving waste, model chemicals
(Cl), and indigenous organisms was not appreciably
changed during the 360-day incubation. The waste and
bacteria treatment (WB) yielded an average of 32 mg/g
solvent-extractable organics on Day 0 and 20 mg/g on
Day 360. Overall, none of the treatments  amended with
wood-preserving waste displayed an appreciable reduc-
tion over the initial 360 days of treatment.
The extract of the unamended Weswood soil induced an
average negative mutagenic response at Days 0 and 90.
In general, treatments with the model chemicals, wood
preserving waste, or combined treatments induced an
average positive mutagenic response at  all time points.
With metabolic activation, the methylene chloride  ex-
tract of the MB treatment induced an average of 167 net
TA98 revertants on Day 0 and 122 net revertants on Day
90.  A slight increase in the mutagenic  response was
observed in the majority of extracts of samples collected
on Day 180. The methylene  chloride extracts of  the
waste amended  soils (Wl, WB, and WF) collected on
Day 180 induced responses with activation that ranged
from 46 net revertants (Wl) to 63 net revertants per mg
                                                41

-------
Table 1.  Solvent Extractable Organics Recovered From Soil Treatments (mg residue/g soil)

                                                    Time (days after treatment)

Treatment3             0
                                      30
                                                      60
                                                                      90
                                                                                       180
 Wl = wood preserving waste with indigenous bacteria.
 WB = wood preserving waste with P. aeruginosa.
 WF = wood preserving waste with P. chrysosporium.
 Cl = combined model chemicals and wood preserving waste with indigenous bacteria.
 CB = combined model chemicals and wood preserving waste with P. aeruginosa.
 CF = combined model chemicals and wood preserving waste with P. chrysosporium.
 StC = sterilized soil with combined model chemicals and wood preserving waste.
b Values are the mean of three replicates ± standard deviation.
                                                                                                       360
Wl
WB
WF
Cl
CB
CF
StC
19.80 ±3.2b
32.25 ±1.1
21 .04 ±0.9
19.1 7 ±3.5
26.45 ±1.6
23.36 ± 2.0
27.05
23.18 ± 1.6
27.25 ± 2.5
28.35 ± 2.0
29.26 ± 1 .7
25.42 ± 0.9
31.51 ±0.6
30.31
29.45 ± 2.2
26. 10 ±2.8
30.84 ± 4.6
28. 16 ±3.2
26.35 ± 1 .0
29.90 ± 1 .9
32.88
24.18 ±0.8
23.81 ± 1.2
23.79 ± 2.0
26.67 ± 2.7
23.54 ± 0.4
26.1 7 ±0.2
36.60
20.28 ±1.8
25.24 ± 3.8
21.19± 1.7
24.06 ± 0.7
23.97 ± 1 .8
21.71 ±2.6
26.95
18.25 ±0.8
19.78 ±1.6
18.14± 1.0
20.21 ± 0.4
19.62 ±1.1
19.57 ±0.9
28.78
of residue (WB).  The methylene chloride extracts of
the combined waste and  model  chemical  soils (Cl,
CB, and CF) collected on Day 180 induced responses
that  ranged from  71 net  revertants (CB) to 76  net
revertants per mg  of residue (CF and Cl). Most of the
methanol extracts induced a stronger  response than
was  observed in the methylene chloride extract.  At
least one solvent-extractable fraction from each of the
soils  collected on Day  180 induced  a positive mu-
tagenic response.

Discussion

Approximately 1 year after the application of both wood
preserving waste and model chemicals to  a  Weswood
soil, the level of solvent-extractable organics was not
appreciably reduced. Although  chemical analysis sug-
gested reductions in the model chemicals in some of the
treatments, contaminant concentrations  were all detect-
able after 360 days. The data do suggest that the pres-
ence  of oils and hydrophobic chemicals in the waste
amended soil may have limited the availability of chemi-
cals for microbial degradation. Changes in mutagenicity
over the initial year of the study indicate that the bioavail-
ability of chemicals was increased in some treatments.
These results indicate that 360 days  of treatment was
insufficient to eliminate the mutagenic  activity of soils
amended with either the model chemicals or the wood
preserving waste. Data collection will continue for  an
additional 720 days to monitor the chemical and toxico-
logical changes associated with each treatment.

References

1. Aprill, W, R.C. Sims, J.L. Sims, and J.E. Matthews. 1990. Assess-
  ing detoxification and degradation of wood preserving and petro-
  leum wastes in  contaminated soil. Waste  Manag. and Res.
  8:45-65.
2. George, S.E., D.A. Whitehouse, and  L.D. Claxton. 1992. Gentoxic-
  ity of 2,3,4-trichlorophenoxyacetic acid biodegradation products in
  the  Salmonella reversion and lamba prophage-induction bioas-
  says. Environ. Toxicol. Chem. 116:733-740.
3. Donnelly, K.C., P. Davol, K.W. Brown, M. Estiri, and J.C. Thomas.
  1987.  Mutagenic activity of two soils amended with  a wood-
  preserving waste. Environ. Sci. Technol. 21:57-64.
                                                     42

-------
                Assessing the Genotoxicity of Complex Waste Mixtures
  Larry D. Claxton, Virginia S. Houk, Sarah H. Warren, Thomas J. Hughes, and Susan E. George
  Environmental Carcinogenesis Division, National Health and Environmental Effects Research
   Laboratory, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina
Introduction

Some chemicals associated with environmental spills
and hazardous waste sites can induce permanent ge-
netic alterations in all organisms. These changes, called
mutations, can have deleterious effects on individuals
and their descendants. In human populations, mutations
are known  to  increase the incidence of cancer and
genetic diseases and may play a role in numerous other
diseases. In nonhuman biota, mutations may alter the
balance of  an  ecosystem or change the  virulence of
pathogens.

In spite of  20  years of research concerning environ-
mental mutagenesis, current knowledge concerning the
identity  and effects of most environmental  mutagens
(including genotoxic carcinogens) is limited, and data on
their effects on other living organisms and the ecosys-
tem are practically nonexistent.  Because their mode of
action theoretically does not depend on a threshold limit,
genotoxicants  are a  class  of  toxic  substances that
should be examined when found in even low concentra-
tions in the  environment.

The major objective of this  paper is to review the U.S.
Environmental  Protection Agency's (EPA's) research in
the area of  genetic toxicology as it relates to bioreme-
diation. Readers can referto a completed manuscript (1)
for more in-depth information, examples of studies and
data, and a  discussion of the regulatory context.

Research Targeted by the Risk
Assessment Paradigm

The complexity of environmental situations complicates
the monitoring, evaluation, and  risk assessment of haz-
ardous waste sites and spills. The composition of pollut-
ant mixtures may consist  of a few or thousands of
individual compounds, and remediation efforts produce
additional components that add to the complexity of
evaluating  risk.  EPA  research targets efforts that
enhance the risk assessment process. Research, there-
fore, can be placed into the four categories associated
with the risk assessment process: hazard identification,
exposure assessment, dose-response assessment, and
risk assessment modeling and  calculation methods.

Hazard Identification Research

Hazard identification research strengthens the risk as-
sessment process by developing methods that are more
reliable  and more cost effective, and,  when possible,
give some  sense of relative toxicity when identifying
toxic agents in the environment. This research involves
the development of new assay methods (e.g., the spiral
mutagenicity assay), the enhancement of existing as-
says (e.g., the modification of the prophage assay), the
integration of bioassay and analytical chemistry meth-
ods, and the modification of methods for use with com-
plex environmental mixtures  rather than with single
environmental chemicals.

Exposure Assessment Research

Generally, a hazardous substance must come into con-
tact with a specific biological component (e.g., deoxyri-
bonucleic acid)  before  a toxic  response can occur.
Measuring or estimating  the potential  level of contact
between a toxicant and its reactive target (or a surrogate
for that target) is exposure assessment. Exposure as-
sessment is done on a population and not an individual
basis. Quantitative toxicological assays and analytical
chemistry methods have  been developed to determine
relative amounts of toxicity potential existing in different
environmental (e.g., treated versus untreated) sites, to
determine the relative  bioavailability of environmental
toxicants and their metabolites, and to follow the change
in  environmental concentrations of toxicants over time.
In  collaboration with other EPA laboratories, the National
Health and Environmental Effects Research Labora-
tory (NHEERL)  has demonstrated the usefulness of
bioassays to enhance and calibrate exposure assess-
ment methods.
                                                43

-------
Dose-Response Assessment

Dose-response assessments quantitate the potency of
an environmental agent(s) for a specific outcome.  Al-
though the NHEERL bioremediation program does not
support whole animal carcinogenicity and mutagenicity
studies, the  Environmental  Carcinogenesis Division
(ECD)  of NHEERL does  develop, in support of other
programs, information  relevant for bioremediation haz-
ard and risk assessments. In addition, ECD has devel-
oped a statistical software system for analyzing, in a
quantitative manner, short-term mutagenicity tests.

Risk Assessment Modeling Research

When  information is limited or only a general charac-
terization is required, a qualitative assessment can be
made;  however, there is often a need to be as quantita-
tive as  possible in risk assessments. ECD, therefore, not
only develops quantitative data  but  also  examines
mechanisms and developing models to help scientists
understand to what extent the toxic effect is relevant to
human populations. Mechanistic research, for example,
may demonstrate that rodents respond to a toxin in a
manner not relevant to humans, alleviating the need for
a quantitative human risk assessment. Likewise, mecha-
nistic research may demonstrate whether or not syner-
gism  should  be  considered  when  evaluating  a
multichemical site.

The purpose of the presentation given at this conference
is to catalog NHEERL bioremediation research, to show
its relevance, and to provide examples of how the infor-
mation obtained will be useful in future assessments.


Reference

1. Claxton, L.D., V.S. Houk, and S.E. George. 1995.  Integration of
  complex mixture toxicity and microbiological analyses for envi-
  ronmental  remediation research. In:  de  Serres, F.J., and  A.
  Bloom, eds. Ecotoxicology and human health. Boca Raton, FL:
  Lewis Publishers. In press.
                                                  44

-------
                                      Section 4
                              Pilot-Scale Research
By studying bioremediation processes under actual site conditions on a small scale, researchers
can gather critical information on issues such as operation, control, and management of residuals
and emissions before moving to full-scale research. This is a critical intermediate step in which the
success of laboratory experiments are further tested in an expanded but controlled setting.

Pilot-scale evaluations covered many different tools for bioremediation, including biofilters, compost
piles, and slurry bioreactors. One paper studied the optimization of biofilters for use in removing
volatile organic contaminants from the air, while another sought to establish the optimal operating
conditions for compost media. Finally, a third paper found that many organic contaminants become
concentrated in the foam produced  in slurry bioreactors and suggested that this effect could be
utilized to isolate contaminants from the rest of the slurry.

Two papers focused on the use of combinations of aerobic and anaerobic conditions to degrade
recalcitrant chlorinated wastes. Work was presented on biofiltration with gel beads that provide an
oxic environment at their surface and an anoxic environment at their centers. The authors found
that such filtration was highly effective at degrading trichloroethylene (TCE). Asystem was evaluated
for land treatment that involved switching between  anaerobic and aerobic conditions.

Research continued on the integration of physical and  biological processes to clean up organic
wastes. One paper showed how wood preserving wastes can be first removed from their soil matrix
by washing and distillation processes and then degraded in a bioreactor.

Two presentations covered research into optimizing the  biodegradation of TCE. The first was the
result of  the study  of a variety of parameters that might affect the success of field-scale TCE
treatment. The next characterized several potentially useful bacteria isolated from a biofiltration
device.

Two papers investigated the use of microorganisms to clean up inorganic wastes. One technique
mentioned involves the biological reduction of sulfates into sulfides. This technique helps to remove
metal cations from solution. The other technique involves the biological reduction  of metal cations
into less harmful oxidation states. The authors of the final paper tested the ability of microorganisms
to reduce chromium (VI).

Poster presentations described several pilot-scale research projects. The poster projects within this
category  involved the testing of a compost bioreactor, a biofilm-electrode reactor, and a laminar-type
flow reactor.
                                           45

-------
    In Situ Bioremediation of Trichloroethylene With Burkholderia Cepacia PR1:
               Analysis of Parameters for Establishing a Treatment Zone
                                        Richard A. Snyder
      Center for Environmental Diagnostics and Bioremediation, University of West Florida,
                                        Pensacola, Florida

                             M. James Hendry and John R. Lawrence
        National Hydrology Research Institute,  Environment Canada, Saskatoon, Canada
Introduction

The use  of chloroethenes, including trichloroethylene
(TCE), has led to an extensive contamination of ground-
water resources in the United States. In situ bioremedia-
tion of this  contaminant  is being pursued with the
aerobic microorganism Burkholderia cepacia (formerly
Pseudomonas  cepacia)  G4 (1).  Mutant strains G4
PR12s and PR131 constitutively produce a toluene ortho-
monooxygenase (TOM) that mineralizes TCE. The gene
for this enzyme is located on the degradative plasmid
TOM  of B. cepacia (2). The constitutive expression of
this gene  is the result of secondary transposition follow-
ing  Tn5  insertion  mutagenesis, which  also confers
kanamycin (km)  resistance  to these  bacteria.  PR12s
contains a single entire Tn5 in the chromosome and an
IS50R of  Tn5 in the plasmid. PR131 bears an entire Tn5
and an IS50R in the TOM31c plasmid. The IS50 elements
are at nearly the same  locations and are thought to be
responsible for the constitutive expression of torn.

This  project  involves ground-water flow control with
sealed sheet piling to funnel ground water through  a
narrow gate  area in which treatment technologies can
be applied. Parameters for the establishment of a bio-
logical treatment zone with PR1 have been investigated
in the laboratory. Transport and survival  characteristics
of the bacterium have been examined in ground water
and sediment from a targeted release site (the aquifer
under the Canadian Armed Forces Base, Borden, On-
tario). Monitoring  techniques have been developed for
tracking PR1 populations in the treatment zone and
determining the extent of dispersal and survival beyond
the treatment zone.

The functional integration  of non-native bacteria in natu-
ral microbial  communities and maintenance of bacterial
populations above normal environmental  background
concentrations provides a challenge for both microbial
ecologists and applied  microbiologists. Monitoring of
population dynamics and trophic interactions is critical
for successful  bioaugmentation  applications. Risk as-
sessment associated with uncontained biotechnological
introductions also requires careful monitoring of the
survival and dispersal of released microorganisms and
altered genes. Although releases of non-native or re-
combinant bacteria have not been  reported to result in
adverse environmental effects to date, there is a respon-
sibility to  ensure that released microorganisms will be
constrained by the selective pressures of the  target
environment.

Tracking

Selective media provide a first step in tracking the or-
ganism. Phenol, o-cresol, and phthalic acid combined
with kanamycin (km) were tested for growth of PR1 and
isolates from the Borden aquifer. Growth of PR1 was
optimal on 20 mM phthalate medium. In aquifer sedi-
ment slurries, numbers of native bacteria isolated on this
medium range from 0 to less than 5 x 105 colony forming
units  (CPU)  and  direct epifluorescence  microscopy
counts of bacterivorous protists  are less than 8 x 102
ml"1  in unamended incubations.

A monoclonal antibody (mab) specific to G4 lipopolysac-
charide (IPS)  (3) has  been used both in direct im-
munofluorescent counts and to confirm CPU of PR1 on
phthalate plates. Details of the production and testing of
this mab can be found in Winkler et al. (3).

Polymerase chain reaction (PCR) primers targeting the
Tn5  insertion  junctions  have  been  developed  and
tested. The  primers  were  designed  based on the
                                                47

-------
assumption that the insertion points would be unique for
PR1. The PCR products resulting from this primer set
are of slightly different size for PR12s and PR131, reflect-
ing the slightly shifted location of the  IS50R in the plas-
mid in these two strains. Attempts to  extract indigenous
bacterial deoxyribonucleic  acid  (DMA)  from  Borden
aquifer sediments and to amplify product from these
extractions with the PR1 primers have failed. We have
demonstrated that PCR amplification  with these primers
will work with whole, intact  cells added directly to the
PCR reaction. The current limit of detection for positive
amplification from cell suspensions (equivalent to pore
water samples) is 1 x 103 cells/mL"1,  but from sediment
slurries the detection  limit drops an order of magnitude.
These results indicate that this assay  will be most useful
for large volumes of  pore water samples  collected on
Sterivex filters (Millipore Corp.) for extraction of DMA (4).

Potential host range for the TOM plasmid from PR131 in
24 random isolates from Borden aquifer sediments (R2A
medium; 5) was assessed by direct filter matings. Over-
night cultures of PR131 grown in  lactate medium and
aquifer isolates in R2A medium were pelleted and resus-
pended  in R2A medium to an optical density of 1.0 @
600 nm. One milliliter of donor and recipient were mixed
and filtered on a 0.2  u,m pore-size polycarbonate filter.
Filters were  placed on R2A  plates and incubated over-
night at 30°C. Filters  were then removed and vortexed
in R2A broth and suspensions plated on selective media
predetermined by antibiotic screening. Out of  24 iso-
lates,  42 percent were positive for TOM31c transfer by
PCR detection, and 80 percent of these were positive
for torn activity as indicated by a positive transformation
of trifluoromethyl phenol (TFMP) (1) and mineralization
of TCE in standing cell assays. These data indicate  a
wide potential host range within the target environment
for TOM, highlighting  the need to ensure tracking capa-
bility for  both the organism to  be introduced  and its
associated genetic elements.  These  native bacterial
strains bearing  functional TOM plasmids may also be
better adapted for bioremediation application in  the tar-
get environment than PR1.

Analysis of PR1  Transport in Borden
Aquifer Material

Transport characteristics of PR1  in  Borden sand  are
important in determining the  retardation and dispersal of
cells within a treatment zone and in the  downstream
aquifer.  Cell retardation is important to allow contami-
nated water to flow past the inoculated populations of
PR1. A series of experiments was carried out at three
scales:  3.8-cm,  10-cm, and 40-cm length columns
packed  with sterile  and nonsterile  Borden sand and
commercially available  silica  sand. Artificial  ground
water (AGW) was used as the solute. All columns were
packed understanding water to minimize air entrapment
and tamped with a glass rod to  attain a uniform bulk
density of 1.8 g/cm"3 and a porosity of 0.4. Ground-water
velocity was set to the approximate velocity of water in
the target  site.  Bacterial suspensions were  pulsed
through for 1 void volume, and the effluents were col-
lected  in  sterile vials with a chromatography fraction
collector.  Numbers of PR1 were determined by plating
on selective media. Chloride (Cl) ion as a conservative
tracerwas measured with Ag/AgCI electrodes calibrated
against Cl ion analysis by ion chromatography.

Both the  Cl tracer and bacterial breakthrough curves
displayed similar patterns between columns,  indicting
good replication. All bacterial breakthrough curves ex-
hibited a  notable pulse of bacteria corresponding to the
breakthrough of the Cl tracer. Reduced peak concentra-
tions of bacteria relative to breakthrough concentrations
of Cl (greaterthan 99 percent) indicate irreversible sorp-
tion of PR1 onto the geologic media. Well-defined tailing
was also  observed overthe duration of the experiments,
indicating reversible sorption of PR1 to the geologic
media. Peak heights and tailing  were three orders of
magnitude lower in  Borden material than silica sand,
possibly  due to clays  and  iron coatings  on Borden
sands. Breakthrough of PR1 was not affected by sterili-
zation of the sediment orthe addition of a cotransported
bacterium. These results were integrated with predation
loss rates of PR1 added  to Borden sediment  in slurry
microcosms to  develop a predictive  model with both
physical and biological parameters for the transport and
fate of the organism within the Borden aquifer.

Survival of PR1 in Aquifer Microcosms

Much of  the target environment consists of anaerobic
saturated sediments. Survival of the organism beyond
an oxygenated target zone will likely depend on its ability
to withstand conditions of little  or no  oxygen. Plate
counts of a suspension of cells with Nitrogen (N2) gas
flushed through the head space indicate little effect on
PR131 viability through 6 days. Afterthis point, culturable
numbers  drop precipitously but maintained a low popu-
lation level through 25 days of anaerobic conditions.

No  PR1-specific viruses could be isolated from the tar-
get environment.  Samples of aquifer sediment were
incubated with growing PR1 cells as an enrichment, and
supernatants were tested  on PR1 overlay plates to look
for cleared viral plaques.

Survival of PR1  above 1 x 107 cells/mL"1 is of interest in
establishing  needed inoculation densities for  effective
bioremediation  (6). Survival below this concentration
and long-term integration of the organism into the native
microbial community was of interest for risk assess-
ment. A series of experiments has  been conducted in
ground water, sediment slurries, and flow columns to
examine  population dynamics of PR1  and  the bac-
terivorous protists from  the Borden aquifer that are the
primary vector for loss of PR1. Use  of a monoclonal
                                                  48

-------
antibody (mab) to PR1  IPS (3) allows enumeration of
PR1 after PR1  numbers have  been  reduced to  the
background  level  of phthalate utilizers in the  system.
The loss of PR1 cells and corresponding tracking data
by mab indicates  that with increasing inoculation den-
sity, PR1 populations are sustained for longer periods
prior to rapid loss. This delay in the loss of cell numbers
can be attributed to exceeding the maximum response
of the bacterivorous protists in the system.

Data on the time of sustained PR1 populations and  the
loss rate of PR1 cells after this delay were compiled for
all sediment  slurry incubations. The delay in rapid loss
of PR1 was  incorporated into a regression analysis of
time of sustained  PR1 populations above 1 x 107 as a
function of inoculation density. A half-life  estimation by
regression analysis incorporates loss rate data after the
period of sustained population numbers as a function of
inoculation density. This former analysis provides an
estimated  lifetime  for a pulse of PR1  into  the treatment
zone.  The latter analysis provides an extinction coeffi-
cient for cells in the downstream aquifer.

An  aquifer sediment  column has been established to
test the response  of PR1 within the flow  regime  of the
target system. A commercial spring water (GMW) was
used as the diluent. A chromatography column was fitted
with cut gas chromatography (GC)  vials closed with
Teflon septa to provide sampling ports. Teflon tubing and
fittings were used throughout the setup.  Tygon tubing
connected to a constant temperature recirculating bath
was used to jacket the column and maintain 15°C. Flow
was controlled at the column outflow to the flow rate in
the target environment (2 cm/day"1). Cells and TCE were
added by syringe  pump. An overflow ensured constant
supply of diluent, cell suspension, or TCE solution to the
top of the sediment. Pore-water samples were taken by
syringe  (2:300 ul each), and used for  plate counts,
direct counts of bacteria and protists, and  TCE analysis
(alternate sampling periods).

With an inoculum of 1 x 108 cells/mL"1 for 1 void volume,
a population of greater than 107 CFU/mL"1 pore  water
was maintained for 5 days at in the  top  portion  of the
column. Column data expressed as depth profiles show-
ing a  roughly linear decrease in  PR131 numbers with
distance  traveled through the column from  Days 1
through 5, and  the combined effect of predation and
elution decreased numbers in the upper  portion  of the
column at Days 8 and  10. By Day 15, the pulse has been
eliminated at the  upper and lower portions of the col-
umn, leaving residual cells in the central portion. Bac-
terivorous protists increased  in  proportion  to  the
numbers of PR1 added. These organisms  form resistant
cysts on sediment surfaces when food  is  not available.
Decreases observed over time for protist numbers  are
likely due mainly to encystment of these organisms after
depletion  of PR131.  This results  in  a  reservoir  of
bacterivores capable of responding to subsequent addi-
tions of bacteria.

Integration of PR1 Into Stable Microbial
Consortia

Persistence  of a non-native bacterium introduced into
an environment is dependent on the ability of that organ-
ism to find refuges from  predation and compete with
native bacteria. One such refuge may be in biofilms. To
examine the ability of PR1  to integrate into biofilms, PR1
was  introduced into existing biofilms developed from
Borden aquifer material  and into  a defined microbial
community (including  protists) grown with input of TCE
and dibutyl  phthalate. PR1  was found not to  stick to
pre-existing  biofilms of Borden  aquifer origin, but suc-
cessfully  integrated into  the degradative  community
growing in the presence  of TCE and dibutyl phthalate.
The addition of these substances apparently provided a
competitive advantage to PR1 in this system. One year
post-inoculation,  PR1  was located in biofilms from this
system by fluorescent monoclonal antibody and scan-
ning confocal laser microscopy. PR1 was found through-
out the biofilms as scattered cells and microcolonies.
This information suggests that in the presence of TCE it
may  be possible to maintain and grow PR1 within a
treatment zone in the target aquifer.

Modeling  of Transport and Fate of PR1 in
Borden Aquifer

Using the transport data and loss rates due to predation,
preliminary modeling  exercises have  been conducted.
These initial  approximations assume grazing losses will
be the same for pore  water and attached  bacteria, and
do not address  excystment/encystment  processes  of
the protists, but do  incorporate  protist grazing as a
dynamic response  to  variable bacterial density. In  all
simulations,  predation reduces the bacterial concentra-
tions in the  effluent from a modeled transport column.
With   an  increased predation constant,  peak  break-
through numbers are reduced,  but tailing is affected
more significantly. Tailing, which can be attributed to the
process of reversible sorption, is of concern for transport
of bacteria  to greater distances  in geologic media.
Thus, reductions in tailing by the  protist response re-
duces the concern of offsite migration of the introduced
microorganism.

Conclusions

Data collected from these laboratory studies indicate
that biological interactions, particularly with bacterivor-
ous protists,  will limit the survival of PR1 introduced into
the system and may provide for a natural containment
of the bacterium. Maintenance of PR1  numbers  high
enough for mineralization activity will likely require re-
peated additions of sufficient cells to exceed  protist
                                                  49

-------
maximum response. Preliminary data on TCE additions
suggest that TCE may impede protist activity until it is
mineralized, providing a gradient of protist response in
proportion to TCE removal. Ultimately, the utility of labo-
ratory analyses will be gauged against data  from the
field release.

Acknowledgments
The authors thank M.S. Shields and S.C. Francisconi for
advise  and assistance in the research. This work was
supported  by  the  U.S.   Environmental  Protection
Agency's Gulf Breeze  Environmental Research Labora-
tory through cooperative research agreement 822568 to
RAS.
References

1.  Shields, M.S., S.O. Montgomery, S.M. Cuskey, P.J. Chapman, and
   P.M. Pritchard. 1991.  Mutants of Pseudomonas cepacia G4 defec-
   tive in catabolism of aromatic  compounds and trichloroethylene.
   Appl. Environ. Microbiol. 57:1,935-1,941.

2.  Shields, M.S., Reagin, M.J., Gerger, R.R., Cambell, R., Somerville,
   C.C. 1995. TOM: A new aeromatic degradative plasmid from Burk-
   holderia (Pseudomonas) cepacia G4.  Appl. Environ. Microbiol.
   61:1,352-1,356.

3.  Winkler, J., K.N. Timmis, and R.A. Snyder. 1995. Tracking  the
   response of Burkholderia cepacia G4 5223 PR1 in aquifer micro-
   cosms. Appl. Environ. Microbiol. 61:448-455.

4.  Somerville, C.C., IT. Knight, W.L.  Straube, and R.R. Colwell. 1989.
   Simple, rapid  method for direct  isolation  of nucleic acids from
   aquatic environments. Appl.  Environ. Microbiol. 55:548-554.

5.  Reasoner, D.J., and E.E. Geldreich. 1985. Anew medium for the
   enumeration and subculture of bacteria from potable water. Appl.
   Environ. Microbiol 49:1-7.

6.  Krumme,  M.L., K.N.  Timmis, and D.F. Dwyer. 1993. Degradation
   of trichloroethylene by Pseudomonas cepacia G4 and the consti-
   tutive mutant strain G4  5223 PR1  in aquifer microcosms. Appl.
   Environ. Microbiol. 59:2,746-2,749.
                                                        50

-------
               Characterization of Trichloroethylene-Degrading Bacteria
                                   From an Aerobic Bio filter
          Alec Breen, Todd Ward, Ginger Reinemeyer, John Loper, and Rakesh Govind
                             University of Cincinnati, Cincinnati, Ohio

                                           John Haines
                     U.S. Environmental Protection Agency, Cincinnati, Ohio
Introduction

The  microbial community  colonizing  a vapor-phase
biofilter was examined to determine the population(s)
capable of trichloroethylene  (TCE) degradation. The
community had been exposed to  low  levels  of TCE
continuously for 24 months and maintained degradation
in the absence of a canonical co-metabolite. Although
low levels of autotrophic ammonia-oxidizing  bacteria
were present, nitrapyrin inhibitor studies suggested that
alternative bacteria were responsible for TCE oxidation.
In addition, replacement of ammonia with nitrate did not
affect TCE degradation. Incubation of biofilter biomass
in a toluene or benzene atmosphere resulted in a turbid
culture within 2 to 3 days. In light of these observations,
aromatic hydrocarbon-oxidizing bacteria were  pursued
as putative candidates mediating TCE degradation.

A significant fraction of the  culturable heterotrophs
(greater than 80  percent) were capable of growth on
toluene or benzene. This study describes the naturally
occurring TCE-degradative populations that became es-
tablished over time in the  biofilter.  Individual  isolates
were tested for TCE-degradative capacity under several
growth conditions. The pure cultures  tested were all
capable of co-metabolic TCE degradation. These organ-
isms had persisted in the biofilter regardless of condi-
tions that would exert a negative selective pressure due
to generation of the TCE-derived epoxide during aerobic
TCE degradation. Several  isolates were selected for
further study. Initial sampling of the biofilter yielded three
isolates: Rhodococcus sp.  TA1, Pseudomonas putida
TA2, and Nocardia sp. AR1. Both the Rhodococcus and
the P. putida could be repeatedly isolated from the biofil-
ter.  Two other organisms,  P. putida DC1  and  Burk-
holderia cepaciaGRZ, were isolated more recently. This
consortium appears relatively resistant  to the toxic ef-
fects of TCE oxidation at the concentrations used in the
biofilter.

Background

Halogenated aliphatic compounds are a major class of
industrially important chemicals that have become sig-
nificant environmental contaminants with mutagenic and
carcinogenic potential. A widespread ground-water con-
taminant, TCE can undergo co-metabolic oxidation by
a variety of physiologically diverse bacteria (1). Co-
metabolic TCE degradation by aromatic hydrocarbon
utilizing bacteria was originally reported by Nelson et al.
(2). Since that time, efforts to employ these organisms
to ameliorate  TCE contamination problems in situ and
in reactors have been conducted. This study examines
the toluene-degrading bacteria surviving  in a vapor-
phase biofilter. Although  only a mineral salts  medium
was supplied to the biofilter, toluene oxidizers survived
and TCE degradation was maintained at a level of 20
percent of a 21-ppmv gas stream.

Experimental  Methods

Biofilter and Sampling

The biofilter consisted of ceramic  plates in a stainless
steel casing. The initial inoculum was a municipal sludge
sample that was acclimated to a volatile organic com-
pound (VOC) mixture (benzene, toluene, ethylbenzene,
and TCE) fora period of 3 months. At this point, all VOCs
except TCE were removed. The biofilter was operated
at a gas-flow rate of 520 mL/min, and had an empty-bed
residence time of 1.9 min. TCE inlet concentration was
21 ppmv. A mineral  salts solution was  applied to the
biofilter at a flow rate  of 357 ml/day. Biofilter sampling
was conducted by opening the biofilter and scraping
biomass off the ceramic matrix. VOC-degrading  bacteria
                                                 51

-------
were  isolated  by  incubation  of biofilter material in a
mineral salts medium, with the appropriate VOC sup-
plied in the vapor phase.

TCE Mineralization Assays

Degradation  experiments using 14C-TCE were  con-
ducted in 20.0  ml vials, with teflon-lined silicone septa
closures allowing  injection into the vial. An  inner vial
containing 0.4  N NaOH served as a CO2 trap. Sterile
control vials were subtracted from experimental values
when determining conversion of TCE to CO2 or soluble
products. All data  represent a mean value of triplicate
vials.

Biochemical and Genetic Characterization of
Biofilter Bacteria

Aromatic hydrocarbon utilizing bacteria were isolated as
follows.  Biofilter biomass  was inoculated into mineral
salts medium, and the medium was exposed to toluene
or benzene vapor as a sole  carbon source  for 48 hr.
Cultures were  then plated onto mineral salts medium
and grown in a toluene or benzene atmosphere for 10
days. Organisms  that appeared   were picked  and
streaked onto mineral salts plates and grown again with
the carbon source in the vapor phase. The isolates were
then checked for purity and TCE mineralization capabil-
ity.  Isolates were  sent  to Microbial ID, Inc.  (Newark,
Delaware), for  fatty acid methyl ester (FAME) analysis.
16S rDNA sequencing was carried out by amplifying the
27 to 321  base pair region of the  16S rDNA gene by
polymerase chain reaction (PCR). The primers used are
forward  '5AGAGTTTGATCCTGGCTCAG-3'  (positions
27-46) and reverse '5AGTCTGGACCGTGTCTCAGT-3'
(positions  321-301). Both forward  and reverse  DNA
strands  were  sequenced. Gene probes for charac-
terized toluene/TCE co-metabolic oxygenases were ob-
tained from other researchers. The todABC probe was
obtained from  Dr.  D.T Gibson,  and the tbu probe was
obtained from Dr. A. Byrne and  Dr.  R.H. Olsen (3, 4).

Results  and Discussion

The predominant bacteria  in the biofilter were shown to
be degrading TCE  by toluene/benzene oxygenase co-
metabolic route. The biofilter community had not  been
exposed to these compounds for over 24 months, yet
these organisms persisted and were shown  to be the
key population  mineralizing TCE. These isolates are of
interest because they arose spontaneously from a  natu-
rally occurring  population  and were maintained in the
continual presence of TCE. It might be expected that
TCE-oxidizing organisms would be selected against in
such a system (5). Mixed cultures mineralized 14C-TCE
(data not shown) and exhibited very little diversity when
plated out onto a mineral salts medium and grown with
toluene or benzene vapor as a  sole carbon source.
These  plates generally had  only one or two colony
morphotypes shown to be a P. putida (designated TA2)
and a Rhodococcus (designated TA1). A Nocardia sp.
(designated AR1) was isolated during an early sampling
time but was not reisolated. Two additional organisms,
B. cepacia GR3 and P. putida DC1, were more recently
isolated.

The predominant toluene degrader  isolated from the
biofilter was TA2. Growth on glucose,  succinate, ortryp-
tophan completely inhibited TCE mineralization by TA2.
A gene probe for the alpha subunit  of the tbu toluene
monooxygenase strongly hybridizes with TA2 (Figure 1).
Figure 1. Slot blot analysis of biofilter DNA extracts probed
        with the tbu monooxygenase alpha component:
        A) P. putida DC1, B) B. cepacia GR3, C) P. putida
        TA2, D) Nocardia sp.  AR1, and E) Rhodococcus
        sp. TA1.

This probe will also hybridize to two  other toluene
monooxygenases, tmo and torn (data from our labora-
tory  and unpublished information from  M.S. Shields,
University of West Florida,  and R.H. Olsen, University
of Michigan; therefore, the mode of toluene oxidation is
not definitively established.  Slot blot  analysis of biofilter
isolates probed with  tbu is shown in  Figure 1.
Rhodococcus sp.I'M did not hybridize strongly to either
the tbu or tod toluene oxygenase probes, indicating the
uniqueness of its toluene oxygenase. In addition to TA1
and TA2, three other TCE-co-metabolizing organisms
from the reactor were investigated, and their properties
are listed in Table 1. All organisms are being evaluated
to determine  basal levels of TCE  catabolic activity as
toluene is depleted. The effect of acclimation to alterna-
tive substrates on  TCE degradation is  also  being
examined.
                                                 52

-------
Table 1.   Properties and Growth Substrates of TCE-Co-metabolizing Biofilter Isolates

Organism      Oxidase       lndolea      Benzene      Ethylbenzene      Phenol       o-Cresol       m-Cresol       p-Cresol

TA1            -              -           + + +          + + +               + + +         + + +          + + +           +

TA2            +              +           + + +          + + +               -            +             +


DC1            +              -           + + +          + + +

GR3           +              +           + + +          + + +               -            +
a Conversion of indole to indigo.

References                                                 3- Byrne, A.M., J.  Kukor, and R.H. Olsen. 1995. Sequence of the
                                                                    gene cluster encoding toluene-3-monooxygenase  from Pseudo-
                                                                    monas pickettii PKO1. Gene.  In press.

1.  Ensley, B.D. 1991.  Biochemical diversity of trichloroethylene me-     4. Wackett, L.P., and S.R.  Householder.  1989. Toxicity of trichlo-
   tabolism. Arm. Rev. Microbiol. 45:283-299.                            roethylene  to Pseudomonas  putida F1  is mediated by toluene
                                                                    dioxygenase. Appl. Environ. Microbiol. 55:2,723-2,725.

2.  Nelson, M.K.J., S.O. Montgomery,  E.J. O'Neil, and P.M. Pritchard.     S.Alvarez-Cohen,  L, and  PL. McCarty.  1991.  A  co-metabolic
   1986. Aerobic metabolism of trichloroethylene by a  bacterial  iso-        biotransformation model for halogenated aliphatic compounds ex-
   late. Appl. Environ.  Microbiol.  52:383-384.                            hibiting product toxicity. Environ. Sci. Tech. 25:1,381-1,387.
                                                             53

-------
    Anaerobic/Aerobic Degradation of Aliphatic Chlorinated Hydrocarbons in an
                               Encapsulated Biomass Biofilter
                                Rakesh Govind and P.S.R.V. Prasad
         Department of Chemical Engineering, University of Cincinnati, Cincinnati, Ohio

                                         Dolloff F. Bishop
    National Risk Management Research Laboratory, U.S. Environmental Protection Agency,
                                         Cincinnati, Ohio
Introduction

During the past decade, public awareness and concern
about the quantity and diversity of persistent (recalci-
trant to degradation) synthetic chemicals produced by
industry have increased. Release of these chemicals
into the environment is inevitable, and hence there is a
strong need to control,  direct, and improve the  proc-
esses for degradation of these chemicals. Large differ-
ences in the rates and mechanisms of biodegradation
of various compounds under oxic and anoxic conditions
exist. Consequently,  sequential anoxic and  oxic condi-
tions, enabling cooperation of anaerobic and  aerobic
bacteria, are desirable for rapid complete mineralization
of many polyhalogenated compounds (1-6).

Encapsulation of Biomass in Hydrogels

The goal of this research is to use hydrogel-encapsu-
lated bacteria for simultaneous creation of oxic and
anoxic zones inside the hydrogel bead. Further, the oxic
and anoxic zones created inside the hydrogel bead can
be successfully used to mineralize chlorinated  com-
pounds such as trichloroethylene  (TCE)  and perchlo-
roethylene (PCE). Hydrogel beads with encapsulated
bacteria can be used to mineralize chlorinated  com-
pounds present in air, ground water, and soil, and can
also  be used to promote ecological competitiveness of
laboratory-grown cultures that are specially adapted for
biodegradation of specific environmental pollutants.

Preparation of New Gel Material (7)

The gelation procedure used is as follows: Silica sol,
obtained  commercially as LUDOX colloidal silica SM
grade, is mixed with 1 to 3 percent sodium alginate in
the following proportion: silica sol (90 wt  percent to
99wtpercent): 1 toSpercentsodiumalginatesolution
(1 wt percent to 10 wt percent). The mixture, after ad-
justment of pH between 7 and 8 by 5N HCI, is mixed
with  active aerobic and anaerobic biomass  cells, with
the initial cell wt percent varying from 2.0 to 10.0. The
mixture is stirred, then poured into a petri dish to a depth
of 5 mm. Calcium chloride solution (0.1 molar) is poured
on top of the mixture in the petri dish. The silica sol and
sodium  alginate  mixture immediately gel due to the
diffusion  of calcium forming calcium alginate on the
outer surface, and the gel is allowed to cure from 10 min
to 24 hr. During the curing process, the pH of the silica
sol decreases, thereby forming silica gel with  pockets of
calcium  alginate  inside the silica gel. The survival of
active cells is maximized by using a combination of silica
sol and sodium alginate. Once the silica sol has formed
silica gel, stainless-steel wire mesh cylinders 2.5 mm in
diameter and 5.0 mm long  (open  at both  ends) are
pushed  into  the  gel layer, thereby  enclosing  the gel
inside the wire  mesh. The use of the stainless steel
mesh gives the  silica gel/calcium alginate bead struc-
tural  strength so that the beads can be packed in a bed
without compaction.

Experimental Studies Conducted

A 40-mL bioreactor (1.9 cm inner diameter)  consisting
of a jacketed cylinder was constructed from borosilicate
glass. The reactor was packed randomly  with the gel
beads. Air was passed at a controlled rate through the
bioreactor, and nutrient solution was trickled down from
the top of the bioreactor counter current to the  air flow.
The  air  was  contaminated with chlorinated  pollutants,
such as TCE or PCE, using a syringe pump that injected
the liquid contaminant into the air line through a septum.
The  concentration of the contaminant in the  air stream
                                                54

-------
was varied within the following range: toluene 0 to
100 ppmv; TCE 0 to 25 ppmv; and PCE 0 to 25 ppmv.
The reactor temperature was maintained at 25°C by circu-
lating water from a constant temperature bath through the
jacket  of the bioreactor.  Nutrient solution was trickled
down the bioreactor at a  flowrate of 1 liter per day, and
the nutrient composition was as follows: KhfePCX, (85 mg/L),
K2HPO4 (217.5 mg/L),  Na2HPO4-2H2O (334 mg/L), NH4CI
(25 mg/L),  MgSO4-7H2O (22.5 mg/L), CaCfe (27.5 mg/L),
FeCI3-6H2O (0.25  mg/L),  MnSO4-H2O  (0.0399  mg/L),
H3BO3  (0.0572  mg/L),   ZnSO4-7H2O  (0.0428  mg/L),
(NH4)6Mo7O24 (0.0347 mg/L), FeCI3-EDTA(0.1 mg/L), yeast
extract (0.15 mg/L), and formate (50 mg/L). Results obtained
are shown in Table 1.
Table 1.   Percent degradation of TCE in the bioreactor at
         various air flowrates. Inlet TCE concentration was
         25 ppmv, and nutrient flowrate was 1 liter/day.
   Air Flowrate
     (mL/min)
Percent Degradation
     of TCE
       35

       40

       50

       60

       65
       100.0

       67.2

       40.7

       22.1

       10.8
Carbon and chlorine balances were made by monitoring
the increase in carbon dioxide in the exit air,  and in-
crease in chloride ion concentration in the exit nutrients
was analyzed by an  ion chromatograph.  The chlorine
balance was developed at steady-state conditions within
an error band of 15 percent of the calculated increase
in chloride  ion concentration.

The proposed degradation  pathway was shown to be a
partial dehalogenation in the anoxic zone followed by
oxic biodegradation of the anoxic degradation products
in the outer aerobic zone of the gel bead. The anoxic
zone was  created due to  oxygen consumption in the
aerobic zone  by  the  oxic  degradation of the partially
dehalogenated products as they  diffused out from the
anoxic zone.

A mathematical model was developed to describe the
diffusion of TCE and  oxygen, and consumption of oxy-
gen  due to aerobic degradation of the dehalogenated
products. At the outer surface of the gel bead (denoted
by dimensionless position of 1.0) the oxygen concentra-
tion  is about 8 mg/L due to presence of air outside the
bead. As oxygen diffuses inside the gel bead, it is con-
sumed due to aerobic degradation of the dehalogenated
products diffusing outwards. At some point in the interior
of the gel bead, oxygen is completely consumed produc-
ing an anoxic zone in the interior portion of the gel bead.
It is in  this anoxic zone that dehalogenation of TCE
                     occurs. The  formate  in the nutrient medium is  rapidly
                     absorbed by the  gel  bead and  provides  the  organic
                     carbon source needed for partial dehalogenation of TCE
                     in the anoxic zone by anaerobic microbiota. Other po-
                     tential carbon  sources for anaerobic microorganisms
                     are acetate and other carboxylic acids.

                     Experiments  also  were   conducted  with  perchlo-
                     roethylene (PCE)  at an inlet concentration  of 25 ppmv.
                     Results obtained  are shown  in Table 2. Chloride ion
                     balances  were obtained at steady-state to prove that
                     complete  mineralization of PCE had occurred. Each
                     experiment had to be conducted for over 5 days to
                     achieve a stable exit concentration of chloride ion in the
                     exit nutrients. No other by products were observed in the
                     exit gas phase at the above operating conditions.

                     Table 2. Percent degradation of PCE in the bioreactor at
                             various air flowrates. Inlet PCE concentration was
                             25 ppmv, and nutrient flowrate was 1 liter/day.
Air Flowrate
(mL/min)
10
15
20
30
50
Percent Degradation
of PCE
100.0
86.7
72.4
41.8
12.8
                     Conclusions

                     The hydrogel-encapsulated biomass reactor is capable
                     of biodegrading trichloroethylene (TCE)  and perchlo-
                     roethylene (PCE) through an anaerobic/aerobic degra-
                     dation mechanism.  Experimental results indicate that
                     the degradation of TCE and PCE is complete, and the
                     empty-bed  gas phase residence time for  complete
                     removal is  less than a few minutes. Further studies
                     are ongoing to quantitate the transport parameters
                     and apply the process for treatment of TCE or PCE
                     in ground water.

                     References

                     1. Abramowitz, D.A. 1990. Aerobic and anaerobic biodegradation of
                        PCBs: A review. Crit. Rev. Biotechnol. 10:241-251.
                     2. Beunink, G., and H.J. Rehm.  1988. Synchronous anaerobic and
                        aerobic degradation of DDT by an immobilized mixed culture sys-
                        tem. Appl. Environ. Microbiol.  151:95-100.

                     3. Beunink, J., and H.J. Rehm. 1990. Coupled reductive and oxida-
                        tive degradation of 4-chloro-2-nitrophenol by  a co-immobilized
                        mixed culture system. Appl. Microbiol. Biotechnol. 29:72-80.

                     4. Fathepure,  B.Z., and T.M. Vogel. 1991. Complete degradation of
                        polychlorinated hydrocarbons by a two-stage biofilm reactor. Appl.
                        Environ. Microbiol. 57:3,418-3,422.
                     5. Fogel, S., R.L. Lancione, and A.E. Sewall. 1982. Enhanced biode-
                        gradation of Methoxychlor in soil under sequential environmental
                        conditions.  Appl. Environ. Microbiol. 44:113-120.
                                                    55

-------
6.  Kastner, M.  1991. Reductive dechlorination of tri- and tetrachlo-    7. Bishop, D.F., and R. Govind.  1995.  New hydrogel material for
   roethylenes depends on transition from aerobic to anaerobic con-       degradation of persistent pollutants in immobilized film bioreactors.
   ditions. Appl. Environ. Microbiol. 57:2,039-2,046.                       U.S. Patent Application.
                                                                56

-------
                  Operation and Optimization of Granular Air Biofilters
     Francis Lee Smith, George A. Serial, Makram T. Suidan, Amit Pandit, and Pratim Biswas
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                        Richard C. Brenner
Risk Reduction Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati,  Ohio
Introduction

Since enactment of the 1990 amendments to the Clean
Air Act, the control and removal of volatile organic com-
pounds  (VOCs) from contaminated air streams have
become major public concerns (1). Consequently, con-
siderable interest has evolved in developing more eco-
nomical technologies  for cleaning contaminated air
streams, especially large, dilute air streams. Biofiltration
has emerged as a practical air pollution control (ARC)
technology for VOC  removal (2-4). In fact, biofiltration
can be a cost-effective alternative to the more traditional
technologies, such as carbon absorption and incinera-
tion, for removal of low levels of VOCs in  large air
streams (5, 6). Such cost-effectiveness stems from  a
combination  of low energy requirements and  microbial
oxidation of the VOCs at ambient conditions.

Our biofiltration research has focused on expanding the
range of application of biofiltration  technology to  the
treatment  of  high VOC loadings at consistently  high
removal efficiencies. The preliminary period of our re-
search was dedicated to the pilot-scale comparison of
three  different types  of biological attachment media:  a
patented peat mixture and two synthetic inorganic me-
dia, one channelized  and the other pelletized. The biofil-
ters containing the latter two media were operated as
trickle bed air biofilters (TBABs), called such because
the  media received a steady application of water. After
18  months of testing, the  pelletized medium  (Celite
6-mm R-635 Bio-Catalyst Carrier) was demonstrated to
be significantly better than the other two for handling
high VOC loadings (7-9). Subsequent work to evaluate
the  performance and behavior of biofilters using  the
R-635 pelletized medium produced two significant find-
ings:  first, that an increase in the biofilter operating
temperature permits a significantly higher practical VOC
loading  (i.e., a significantly smaller required media vol-
ume), and second, that biofilter performance decreases
substantially with  the  buildup of back  pressure due
to the accumulation of biomass within the media bed
(10,11).

Working exclusively with this  pelletized medium,  our
continued  research focused  on the development of
strategies for long-term operation with high VOC load-
ings at  sustained  high-removal  efficiencies. This re-
search effort demonstrated that this objective could be
achieved using a biomass removal and control strategy
employing  periodic backwashing of the media  with
water. Backwashing (the upflow washing of the fluidized
media with water) gently removes excess biomass from
the media, circumventing the problems noted earlier for
this medium. A second finding of this research was that
NO3-N as the sole nitrogen  source was superior to
NH3-N.  The use of NO3-N resulted in  lower volatile
suspended  solids  (VSS):  chemical  oxygen demand
(COD) and VSS:N ratios. In other words, for a given set
of operating conditions, less biomass is produced and
less nitrogen is consumed. Finally, it was also observed
that both the recovery of the VOC  removal efficiency
with time after backwashing (unsteady state) and the
VOC removal efficiency with depth (at near steady state)
were superior when using NO3-N.

This paper discusses the continuing research being per-
formed for the development of biofiltration as an effi-
cient,  reliable, and cost-effective VOC ARC technology.
The objectives of this effort were to investigate the re-
moval efficiencies of TBABs under high toluene loadings
and low residence times, and to evaluate the associated
development and control of excess biomass with time.
The biofilter operational period between backwashings
was evaluated to determine its effect on the stability of
biofilter performance. Backwashing variables, including
backwashing frequency  and  backwashing  duration,
were evaluated.
                                                 57

-------
Methodology

The biofilter apparatus used in this study consisted of
four independent, parallel biofilter trains, each contain-
ing 3.75 ft of pelletized Celite 6-mm R-635  biological
attachment medium. A detailed schematic, equipment
description, and typical system operation are given else-
where (8). Each biofilter had a circular cross section with
a 5.75-in. internal diameter (ID). The air feed was mass-
flow controlled, and the VOC (liquid toluene) was me-
tered by syringe pumps into the air feed stream. Each
biofilter was fed a  buffered nutrient feed solution con-
taining all necessary macro- and micronutrients with a
sodium bicarbonate buffer, described elsewhere (8). For
each biofilter, the sole nitrogen nutrient source was NCV
N. The flow directions of the air and  nutrients were
downward. All biofilters were  insulated and inde-
pendently temperature controlled at 32.2°C.

Results

Each biofilter was  loaded with clean, sterilized pellets
and  seeded with  backwashing water from a  similar,
previous run. For each biofilter, a significant dip in per-
formance occurred from Days 17 to 26 due to an error
in preparing the nutrient solution that resulted in feeding
insufficient NO3-N. After the target VOC loading was
achieved, at a COD:NO3-N ratio of 50:1, it was main-
tained for the duration of the run.

Three different backwashing strategies were tested  on
all biofilters sequentially. After restart following back-
washing, effluent samples were collected to determine
the recovery of the VOC removal efficiency with time.
On days when no backwashing occurred, samples were
collected along the  length of the bed to determine the
VOC removal efficiency with respect to depth.

Biofilter A

This biofilter was started up at 50 ppmv toluene influent
concentration,  1.33  min empty bed  residence  time
(EBRT), and 21 mmol NO3-N per day. On Day 17, the
biofilter was backwashed  for the first  time.  Detailed
schematic  and backwashing descriptions  are given
elsewhere (12). The procedure used was to recycle  70
L of 32.2°C tap water through the bed, bottom to top, at
a rate of 57 L/min to induce full media fluidization at a
bed  expansion of about 40 percent. At the end of the
backwashing period, the media was flushed at the same
rate  with another 50 L of clean, 32.2°C tap water.  For
this first backwashing strategy, the period and frequency
were 1  hr twice per  week. The target influent VOC
concentration (500 ppmv toluene) and loading (6.2  kg
COD/m3 day) were  reached on Day 53. On Day 129, the
second  backwashing strategy was started using a pe-
riod and frequency  of 2 hr twice per week. On Day 171,
the third and final  backwashing  strategy was started
using a period and frequency of 1 hr every 2 days. The
performance of Biofilter A is shown in Figure 1.

Biofilter B

This biofilter was started up at 50 ppmv toluene influent
concentration, 0.67 min EBRT, and 21 mmol NO3-N per
day. On Day 17, the biofilter was backwashed for the
first time using the  first backwashing strategy of 1 hr
twice per week. The target influent VOC concentration
(250 ppmv toluene) and loading (6.2 kg COD/m3 day)
were  reached on Day 35. On  Day  129, the second
backwashing strategy of 2 hrtwice perweekwas begun,
and on Day 171, the third strategy of 1 hr every 2 days
was begun. The performance  of Biofilter B is shown in
Figure 2.

Biofilter C

This biofilter was started up at 50 ppmv toluene influent
concentration, 1.0 min EBRT, and 21  mmol NO3-N per
day. On Day 17, the biofilter was backwashed for the
first time using the  first backwashing strategy of 1 hr
twice per week. The target influent VOC concentration
(250 ppmv toluene) and loading (4.1  kg COD/m3 day)
   800
                                             100
                          6.2 kg COD/m

                            •  Influent Cone

                            T  Effluent Cone.
                            o  Percent Remova
                    100      150

                   Sequential Date, days



Figure 1. Performance of Biofilter A with backwashing.
                                                  58

-------
   700
   600
   500
 1 400
   300
   200
EBRT: 0.67 minutes

6 2 kg COD/m3.day

  •   Influent Cone

  »   Effluent Cone.
  a   Percent Removal
                                              100
                                              80
                                              60
                                              40
                                              20  .a
             50      100     150

                   Sequential Date, days
                                    200
                                            250
Figure 2.  Performance of Biofilter B with backwashing.

were  reached on  Day 53. On Day 129, the second
backwashing strategy of 2 hrtwice perweekwas begun,
and on Day 171, the third strategy of 1 hr every 2 days
was begun. The performance of Biofilter C is shown in
Figure 3.

Biofilter D

This biofilterwas started up at 50 ppmv toluene influent
concentration, 2.0  min  EBRT, and 21 mmol NO3-N per
day. On Day 17, the biofilterwas backwashed for the
first time using the first backwashing strategy of 1  hr
twice  per week. The target influent VOC concentration
(500 ppmv toluene) and loading (4.1 kg COD/m3 day)
were  reached on  Day 53. On Day 129, the second
backwashing strategy of 2 hrtwice perweekwas begun,
and on Day 171, the third strategy of 1 hr every 2 days
was begun. The performance of Biofilter D is shown in
Figure 4.

Conclusions

The performances of the four biofilters with respect to
the three backwashing strategies were similar,  although
clearly affected by both the loading and the residence
time. The effectiveness of the three strategies increased
from the first through the third strategy. This shows that
both backwashing duration and frequency are very im-
portant  parameters for control of the  biofilters' VOC
removal efficiency. The third and best strategy, however,
actually had less total backwashing time per week. At
the higher loading, the greater than 90 percent removal
efficiencies of both biofilters were unexpectedly high for
the third backwashing strategy but below the sustained
99.9 percent achieved by the lower loaded  biofilters. It
can also be seen that for a given  loading, the perform-
ance at the lower EBRTs is more sensitive to the back-
washing strategy employed. Both  of these effects were
anticipated;  what was not anticipated  was that this
pelletized medium would perform  so well for any back-
washing strategy at a loading of  6.2 kg COD/m3 day.
These findings, as well as biofilter recovery of perform-
ance after backwashing, will be presented.

Acknowledgment

This research  was  supported by Cooperative Agree-
ment CR-821029 with the U.S. Environmental Protec-
tion Agency.
                                                         800
                             700
                                                         60°
                             400
U>
8
                           t-1  200
                             100 -
                              0


                             4.0

                            . 3.5

                           § 3.0
                           cd
                           s
                            EBRT: 1.0 minutes

                            4.1kg COD/m3.day

                              •  Influent Cone,
                              T  F.ffluent Cone.
                              a  Percent Removal
         r
         riL
                                                 w~%*-»
                                                                      80
                                                                                                  60
                                                                      40
                                                                         I
                                                I
                             1.5
                             0.5

                             0,0
                                       50
                                              100     150

                                            Sequential Date, days
                         Figure 3.  Performance of Biofilter C with backwashing.
                                                   59

-------
     800
  cf  500
  o

  1
  g  400
  a
  o
  ^  300
  u
  _3

  °  200


     100


      0


     4.0

     3.5

   8  3.0


  I"
   & 2.0
     0.5

     0.0
EBRT: 2.0 minutes

4.1 kgCOD/m3.day

  •   Influent Cone.
  T   Effluent Cone.
  D   Percent Removal
                      100


                      80


                      60


                      40  £
                          >*

                      20  |
                          e
                         100      150

                        Sequential Date, days
                                           200
                                                    250
Figure 4.  Performance of Biofilter D with backwashing.

References
 1.  Lee,  B.  1991.  Highlights of the Clean  Air Act Amendments of
    1990. J. Air Waste Manag. Assoc. 41(1):16-31.

 2.  Leson,  G., and A.M. Winer.  1991. Biofiltration: An innovative air
    pollution control technology for VOC  Emissions. J. Air Waste
    Manag. Assoc. 41(8):1,045-1,054.
 3. Leson, G., F. Tabatabal, and A.M. Winer. 1992. Control of haz-
    ardous and toxic air emissions by biofiltration.  Paper presented
    at the Annual Meeting and Exhibition of the Air & Waste Manage-
    ment Association, Kansas City, MO, June 21-26.

 4. Ottengraf, S.P.P.  1986. Exhaust gas purification. Rehn, H.J., and
    G.  Reed, eds. In Biotechnology, Vol. 8. Weinham, Germany: VCH
    Verlagsgesellschaft.

 5. Ottengraf, S.P.P. 1986. Biological elimination of volatile xenobiotic
    compounds in biofilters. Bioprocess Eng.  1:61-69.

 6. Severin,  B.F., J.  Shi, and  T.  Hayes. 1993.  Destruction of gas
    industry VOCs in a  biofilter.  Paper presented  at the IGT sixth
    International  Symposium on Gas, Oil, and Environmental  Tech-
    nology. Colorado  Springs, CO, November 29 -  December 1.

 7. Smith, F.L., G.A.  Serial, P.J. Smith, M.T. Suidan, P.  Biswas, and
    R.C. Brenner. 1993.  Preliminary evaluation of attachment media
    for gas phase biofilters. Paper presented at the U.S.  EPA Sym-
    posium on Bioremediation of Hazardous Wastes: Research, De-
    velopment, and Field Evaluation. Dallas, TX, May 4-6.

 8. Serial, G.A.,  F.L.  Smith, P.J. Smith, M.T. Suidan, P.  Biswas, and
    R.C. Brenner. 1993.  Development of aerobic biofilter design cri-
    teria for treating  VOCs. Paper no. 93-TP-52A.04.  Presented  at
    the 86th Annual Meeting and Exhibition of Air & Waste Manage-
    ment Association. Denver, CO, June  13-18.

 9. Serial, G.A.,  F.L.  Smith, P.J. Smith, M.T. Suidan, P.  Biswas, and
    R.C. Brenner. 1993.  Evaluation of biofilter media for treatment of
    air streams containing VOCs.  In: Proceedings of the Water En-
    vironment Federation 66th Annual Conference and Exposition,
    Facility Operations Symposia,  Volume X. pp. 429-439.

10. Serial, G.A.,  F.L.  Smith, M.T. Suidan, P. Biswas, and R.C.  Bren-
    ner. 1994. Evaluation of the performance of trickle bed biofilters—
    Impact of periodic removal of accumulated  biomass.  Paper no.
    94-RA115A.05. Presented at the 87th Annual Meeting and Exhi-
    bition  of Air & Waste Management Association. Cincinnati, OH,
    June 19-24.

11. Smith, F.L., G.A. Serial, M.T. Suidan, P. Biswas, and R.C. Brenner.
    1994. Pilot-scale evaluation of alternative biofilter attachment me-
    dia for the treatment of VOCs. Paper presented at the U.S. EPA
    Symposium on Bioremediation of Hazardous Wastes:  Research,
    Development, and Field Evaluations. San Francisco, CA, June
    28-30.

12. Smith,  F.L., M.T. Suidan, G.A.  Serial, P. Biswas, and R.C.  Bren-
    ner. 1994. Trickle  bed biofilter performance: Biomass control and
    N-nutrient effects. Paper no. AC946004. Presented at the Water
    Environment Federation 67th Annual Conference and Exposition,
    Facility Operations Symposia.  Chicago, IL, October 15-19.
                                                              60

-------
                   Abiotic Fate Mechanisms in Soil Slurry Bioreactors
                               John A. Glaser and Paul T. McCauley
    National Risk Management Research Laboratory, U.S. Environmental Protection Agency,
                                          Cincinnati, Ohio

                     Majid A. Dosani, Jennifer S. Platt, and E. Radha Krishnan
                                  IT Corporation, Cincinnati, Ohio
Introduction

Soil slurry treatment of contaminated  soil has been
shown to offer a viable technology for soil bioremedia-
tion. This technology, however, has not sufficiently pro-
gressed to be a durable,  reliable,  and cost-effective
treatment option (1).

The use of aggressive mixing energy to provide condi-
tions for improved contact between  soil contamination
and microorganisms capable of degrading the contami-
nation is the hallmark of slurry treatment technology. A
more  complete description of pollutant mass transfer
during the treatment phase is required that includes
treatment  fate mechanisms attributable to biotic and
abiotic processes. Losses attributable to abiotic means
can be overlooked in field application of the technology,
because  limited  questions  can  be successfully  ad-
dressed at field scale. Discussions  with  U.S. Environ-
mental Protection Agency (EPA) regional personnel and
inspection of active field-scale soil slurry bioreactor op-
erations have identified operational  problems, such as
foaming, that could result in possible abiotic loss (2).

Field  bioslurry operations  have adopted various  ap-
proaches to reduce foaming: 1) addition of defoaming
agents, 2) reduction of the rotational  speed of the agita-
tor, and 3) reduction of gas flow through the bioreactor
system. The foaming phenomenon is generally consid-
ered a nuisance, rather than a potential beneficial re-
moval mechanism. Where  pollutants have a specific
gravity less than water once desorbed from the slurried
soil, the pollutants  would rise to  the surface, as in a
flotation  process. One of our working hypotheses was
that foam formation could  be  related to this pollutant
release process. If this analysis has  merit, it is possible
that the operational strategy used in the field is counter-
productive, because a separated contaminant phase is
re-entrained with partially cleaned soil material.

We have conducted two bench-scale slurry bioreactor
treatability studies at the EPA Testing & Evaluation Fa-
cility in Cincinnati, Ohio, which were designed to assess
operating factors leading to foam formation, and to iden-
tify the most advantageous means to deal with foaming.
The initial study was previously presented as a general
treatability study for treatment of creosote contamination
in a soil (3). During this previous study, foaming became
a major problem for operation. Use of a defoamer con-
trolled foaming conditions, as did reduction of the mixer
rotational speed and gas flow in the  more extreme
cases. Subsequent studies  devoted specifically to in-
vestigating the causes and  conditions of foaming and
using  a different batch of soil from the same site as the
earlier study showed little foaming at the beginning of
the study.

Methodology and  Experimental Designs

Foam Study (First Study)

A soil from St. Louis Park, Minnesota, contaminated with
polynuclear aromatic hydrocarbons (PAHs) was used to
assess the importance of foaming conditions to  the
performance of bench-scale  slurry reactors. The design
of the bench-scale experimental bioslurry reactor has
been reported (3). Operational slurry volume was 6 L,
representing  75 percent of the total reactor volume.

To evaluate the conditions and causes of foam forma-
tion, a subsequent study was designed.  This investi-
gation used  the monitoring  conditions  specified
for the treatability study,  and was conducted us-
ing six bench-scale slurry reactors.  Each  reactor
was loaded at 30 percent solids, with an initial volume
                                                 61

-------
of 6 L.  The study design employed two  reactors that
were permitted to develop foam, two reactors in which
foam formation was suppressed through  the use of a
defoamer, and two reactors in which formaldehyde was
used to suppress biological activity. The study duration
was 1 week, with foam sampling based on the cumula-
tive production throughout the study period.

Foam and Scale Study

Experimental variables selected for the foam and scale
study were 30 percent soil solids loading, two treatment
conditions  of condensed foam (removed),  and  foam
retention within the reactors. Each condition was repli-
cated with a single replicated control based on the foam
retention condition. The foam-retained conditions were
maintained through the use of a dispersant (Westvaco,
Reax 100M). The soil was classified to a minus 3/16 in.
dimension; no provisions were developed for prelimi-
nary  dispersion  of the soil solids  or  sand  exclusion.
Scale was collected from the reactor walls after the
mobile soil solids were removed. The scale was strongly
fixed to  the reactor wall, and some effort was  required
to chip the scale away from the reactor.

General Reactor Operation Condition

The control reactors were operated under abiotic condi-
tions to  serve as bioinactive control reactors. Formalde-
hyde was  used as a  biocide in these  reactors and
maintained at 2 percent residual concentration.

The following monitoring and operating conditions were
held constant for the reactors:

• Dissolved oxygen: greater than 2  mg/L

• pH: range of 6 to 9

• Ambient temperature: recorded daily

• Treatment duration:  10 weeks

• Nutrients: C:N:P ratio =  100:10:1

• Antifoam: as needed to control foam

Results

In a previous  treatability study,  a  high solids control
reactor showed the greatest amount of foaming (3). The
amount  of foaming was surprising because foam forma-
tion was expected to  be  related to the  formation of
biosurfactants by the microbiota. Addition of formalde-
hyde to  the control reactor was the only other explana-
tion forthe foam formation observed. Other reactors had
foaming problems, but this reactor was very noticeable
by contrast. Higher solids loading was also observed to
contribute to foam formation.

In contrast to the earlier studies in which  foaming was
observed, the foaming study showed little foam forma-
                      mg/kg
 2-& 3-PAHs
    t-PAHs
   c-PAHs







1474
2311
3786



1016
7950
11905
19855
5367








Foam Cone
 Factor

    5.4


    5.2
                                             5.2
                                             5.3
                        100    1000    10000   100000
       Thousands
                 Dlnit.Conc. DFoam Cone.
                                          7 DAY STUDY
Figure 1.  Foam composition and concentration factor (Study 1).

tion. A small amount of foaming occurred on the first day
of operation. Figure 1 shows the increased concentra-
tion of the t-PAH analytes in the foam, which is five times
greater than the concentration found in the initial sus-
pended soil slurry. A second attempt to evaluate foam
formation is shown in  Figure  2. These data show a
decrease in the foam concentration factor, and are prob-
ably more realistic than the first foam study results. The
second study was also designed to evaluate the depo-
sition  of t-PAH analytes as part of a scale formed in the
reactor. Figures 3 and 4 show the  results  of t-PAH
deposition  in the scale  under  conditions  in which the
foam  was condensed (removed from the reactor) and
retained within the reactor. From inspection of the  re-
sults,  it is clear that the higher molecular weight compo-
nents of the t-PAHs were deposited in  the scale in
quantities 10 to 90 percent above the initial suspended
slurry concentration. A mass  balance analysis will  be
presented that puts the importance of these abiotic fate
mechanisms into perspective.

Differences in physical characteristics of the soil and
operation of the bioslurry reactor between the two stud-
ies may have contributed to decreased foam  formation
in  the foam study. Although soil for the same site was
used for both the treatability study and foam study, the
batch of soil used forthe foam study was coarser (less
than 1/4  in.) which may have resulted in lower PAH
2-& 3-PAHs
4-to 6-PAHs
t-PAHs
c-PAHs
36
mg/kg








1848


1147
2995




812
1467
2806
4273




2077

OQ 300a 2600 20OO 15OO 1OOO 500 1OO 1O»

c
D nit. Cone. dFoam Cone


Foam Cone
Factor
0.8
2.4
1.4
2.6
10000
14 DAY STUDY
Figure 2.  Foam composition and concentration factor (Study 2).
                                                  62

-------
                         mg/kg
 2-& 3-PAHs
 4-to 6-PAHs








1848


1147
2995




812
1208
2008
3215





1568









                                             Scale Cone
                                               Factor
                                                   0.6
                                                                                    mg/kg
                                                   1.7
                                                   1.1
                                                   1.9
         I 3000 2SOO 2OOO 15OO 10OO 5OO 100
                     Dlnil.Conc. D Scale Cone.
                                             10OOO

                                             14 DAY STUDY
Figure 3. Scale concentrations for foam condensed reactors
         (Study 2).

concentrations in the foaming study and to decreased
foam formation.  Furthermore, the air flow rate for re-
actors in the foam study (1 ft3/min) was approximately
20  percent  of that used in the  treatability study  (5
ft3/min), which also may have contributed to decreased
foam formation.

Conclusions

Foam  formation continues to be an  unpredictable and
poorly  understood  event associated  with slurry  treat-
ment. The results of ourstudies are based on the bench-
scale  reactor and may  exaggerate  the  abiotic fate
mechanisms due to high surface-to-volume ratio consid-
erations. The concentration effects associated with foam
formation indicate that foam removal may  be desirable





1717


1013
2730





730
2337
1312

3649
976

                                             Scale Cone
                                              Factor

                                                 1.4
                                                                                                            1.3
                                                                                                            1.3
                                                                  3OOO 25OO  2000 15OO 10OO 5OO  ID
                                 100    10OO    10OOO

                    Dlnit.Conc. CDScale Cone.     14 DAY STUDY
Figure 4.  Scale  concentrations  for  foam retained  reactors
         (Study 2).

to optimize slurry reactor performance. Future studies
will endeavor to evaluate foam separation as part of the
slurry process.
References
1.  U.S. EPA.  1990. Engineering bulletin:  Slurry  biodegradation.
   EPA/540/2-90/016. Cincinnati, OH.

2.  Jerger, D.E., S.A. Erickson, and  R.D. Rigger. 1994. Full-scale
   slurry phase biological treatment of wood-preserving wastes at a
   Superfund site. Draft manuscript.

3.  Glaser, J.A., M.A. Dosani, P.T. McCauley, J.S. Platt, E.J. Opatken,
   and E.R. Krishnan. 1994. Soil slurry bioreactors: Bench scale stud-
   ies. In: U.S. EPA Twentieth Annual RREL Research Symposium:
   Abstract Proceedings. EPA 600/R-94-011. Cincinnati, OH. p.  127.
                                                       63

-------
        Design and Testing of an Experimental In-Vessel Composting System
                                 Carl L. Potter and John A. Glaser
                      U.S. Environmental Protection Agency, Cincinnati, Ohio

          Majid A. Dosani, Srinivas Krishnan, Timothy A. Deets, and E. Radha Krishnan
                                  IT Corporation, Cincinnati, Ohio
Introduction

The goal of this compost research is to  evaluate the
potential use of compost systems in remediation of soils
contaminated with hazardous chemicals. We have de-
veloped bench-scale composters to evaluate  factors
controlling compost  treatment at large scale. We are
currently studying the ability of compost microorganisms
to biodegrade polynuclear  aromatic  hydrocarbons
(PAHs) in in-vessel reactors located at the U.S. Environ-
mental Protection Agency's Test & Evaluation  (T&E)
Facility in Cincinnati, Ohio.

Composting  differs from other ex situ soil treatment
systems in that bulking agents are added to the compost
mixture to increase  porosity and serve as sources of
easily assimilated carbon for biomass growth. Aerobic
metabolism generates heat, resulting in significant tem-
perature increases that bring about changes in the mi-
crobial ecology of the compost mixture.

Optimal conditions for composting may vary depending
on many factors, but generally aerobic conditions with
45° to 55°C (mesophilic temperature range), 40 to 60
percent moisture, and a carbon-to-nitrogen ratio of 20:1
to 30:1 have been considered  best. Mesophilic com-
posting in the range  of 35°C to 50°C might prove to be
the most effective at destroying  certain wastes. Main-
taining temperature below 50°C, however, may not al-
ways  be cost effective if cooling requires too much
energy.

In an  active compost pile, temperature can easily ex-
ceed 55°C, and temperatures above 70°C have been
reported. When the temperature exceeds 55°C, called
the thermophilic stage, most bacteria are killed.  Organ-
isms capable of sporulation, such as some bacteria (2)
and fungi (3, 4), will sporulate and remain dormant until
aerobic activity slows; the temperature falls back into the
mesophilic range when they re-emerge.
Reactor Design

Ten 55-gal, insulated stainless steel compost reactors
have been fabricated to provide the closely monitored
and controlled conditions required for treatability stud-
ies. These fully enclosed, computer-monitored, bench-
scale reactors hold about 1/4 yd3 total compost mixture.

The reactor units stand upright with air flowing vertically
up through the compost mixture for 23 hours per day.
Enclosed units permit on-line analysis of oxygen, carbon
dioxide, and methane at inlet and exit locations. A data-
logging system accumulates data and transmits them to
the PC-based  central data system  that monitors and
controls each reactor. XAD traps in the exit line of each
composter  permit trapping  of  volatile organic com-
pounds  (VOCs) for analysis.

The bottom of each reactor contains a conical collection
system for periodic sampling of any leachate leaving the
reaction mixture. The space above the leachate collec-
tion system holds 2 in. of gravel. Mass balance studies
on soil contaminants are possible by direct sampling of
the reaction  mixture at different depths through bung
holes  in the lid,  together with  capture of VOCs and
leachate leaving the  reactor.

Periodic determination of compost moisture content in
each reactor unit permits adjustment of total  moisture
content  in the compost  matrix to 40 to 50  percent.
Moisture condensers inside compost units promote re-
tention of moisture within the reactor. Otherwise, with
typical airflows, each unit could lose significant amounts
of water daily. If moisture falls below 40 percent, a water
distribution system inside the reactor may be used to add
water to the reaction mixture without opening the reactor.

The cylindrical reactor design permits mixing of reactor
contents by rolling each unit on a drum roller at desired
intervals. Mixing  can be used to  break up anaerobic
pockets and to avoid  packing of the compost mixture. All
                                                 64

-------
reactors are mixed simultaneously by placing them on
rollers over a modified conveyor  belt that forces the
reactors to turn  in unison. Baffles inside  the reactors
promote mixing during  rolling.

Insulation between the reactor core and outer shell re-
duces heat loss from the reactor during aerobic activity.
Heating coils provide the option of warming the reactor
to accelerate composting during startup. Each compos-
ter houses five thermocouples  connected to  a central
computer for on-line temperature measurements. Ther-
mocouples reside at four equally spaced locations within
the compost mixture, and a fifth  thermocouple tracks
ambient temperature outside the reactor.  If the  mean
temperature of the  middle two reactor thermocouples
exceeds  a  predetermined high value,  the computer
switches that unit to high air flow (60 L/min) to cool the
reaction mixture. After the high-temperature unit cools
to a specified  low temperature, the computer switches
the unit back to low air flow (5 L/min) to reduce further
heat loss from the reaction mixture.

Methods

Current studies focus on defining  acceptable operating
conditions and process characteristics to establish suit-
able  parameters for treatment effectiveness. Parame-
ters  of interest include aeration,  moisture dynamics,
heat production, and physical and chemical properties
of the compost mixture. Growth of microorganisms and
disappearance of parent compounds serve  as indicators
of parameter suitability.

A 24-day treatability study,  using field soil from the Reilly
Tar Pit Superfund site near Minneapolis, Minnesota, was
conducted to evaluate  performance of the compost re-
actor system. The soil was contaminated with creosote and
contained 22 PAHs that were measured during the study.

The study design included five replicated treatment con-
ditions involving different ratios of corn cobs to soil and
different airflow rates in 10 reactors.  Soil/bulking agent
compositions evaluated in this study were 50:50 (four
reactors) and 30:70 (two reactors) ratios of corn cobs to
soil (50 percent soil and 70  percent soil, respectively).
Selected airflow rates  were 5 and 10 L/min.

Results and Discussion

Temperatures  in  reactors  with  50  percent  soil  and
moisture content of about 50 percent or less climbed to
the upper mesophilic  and  lower  thermophilic ranges.
Temperatures  in reactors with  moisture  content above
53 percent failed to  increase much  above 30°C. This
might indicate that higher moisture content restricted air
flow through the compost mixture, resulting in  insuffi-
cient aerobic activity to attain high temperatures.  Reac-
tors with air flows of 5 and 10 L/min exhibited similar
temperatures within the compost mixture.

Reactors with 70 percent soil in the compost remained
relatively cool throughout the entire run, never reaching
the mid to upper mesophilic temperature  range. These
reactors tended to maintain higher moisture content
throughout the study. Fewer corn cobs to absorb excess
moisture in the mixture may have resulted in flooding of
the pore space, blocking of air flow through the mixture,
and reduced drying.

Total heterotrophic populations increased from a range of
107-108 to 109-7.6 x 1010 (60- to 300-fold increases) in
reactors during the first 24 hr of composting.  Heterotroph
counts ranging from 1.6 x 109 to 1.4 x 1010 remained after
24 days in reactors with 50 percent soil, but had returned
to around 2 x 108 in reactors with 70 percent soil.

Small PAHs  (two to three rings) were reduced  by aver-
ages of 50 and  30 percent in compost mixtures of 50
and 70 percent soil,  respectively, after 24 days. Large
PAHs (four to six rings) were not decreased under any
treatment condition after 24 days. Continued evaluation
of the compost mixture will  provide more information on
the long-term ability of composting to destroy large PAHs.

Future investigations  will include application  to  pen-
tachlorophenol  and other soil contaminants yet to be
specified. Evaluation  of  pollutant  mass  balance  and
biotransformation products is an important aspect of
future research.

To judge the  abilities of  microorganisms to  degrade
hazardous wastes in soil under various composting con-
ditions, emphasis will  be  placed on diagnosing popula-
tion  changes throughout  treatment and  identifying
microbial species responsible forbiodegradation of con-
taminants. Early microbiological studies  have focused
on enumerating  total microorganisms and determining
the presence of PAH degraders. Future studies will fo-
cus on characterizing changes in biological activity  dur-
ing the four stages of composting, and on identifying the
microbial species responsible for significant biodegrada-
tion of PAHs during each  composting stage. Reappear-
ance of fungi and other mesophiles (e.g., Actinomycetes)
during the cooling stage is also of interest.

References
1. Nakasaki, K., M. Sasaki, M. Shoda, and H. Kubota. 1985. Change
  in microbial  numbers during thermophilic composting of sewage
  sludge with reference to CO2 evolution rate. Appl. Environ. Micro-
  biol. 49(1):37-41.
2. Strom, P.P. 1985. Identification of thermophilic bacteria in solid-
  waste composting. Appl. Environ. Microbiol. 50(4):906-913.
3. Fogarty, A.M., and O.H. Tuovinen. 1991. Microbiological degrada-
  tion of pesticides in yard  waste composting. Microbiol.  Rev.
  June:225-233.
                                                   65

-------
 Integrated Systems To Remediate Soil Contaminated With Wood Treating Wastes
         Makram T. Suidan, Amid P. Khodadoust, Gregory J. Wilson, and Karen M. Miller
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                          Carolyn M. Acheson and Richard C. Brenner
    National Risk Management Research Laboratory, U.S. Environmental Protection Agency,
                                         Cincinnati, Ohio
Introduction

Approximately 15 percent of Superfund Records of De-
cision (RODs) are directed towards sites contaminated
with wood treating wastes (1). Several types of pollut-
ants characterize these sites, including pentachlorophe-
nol (PCP), creosote, polycyclic aromatic hydrocarbons
(PAHs), other hydrocarbons, and heavy metals such as
copper, chromium, arsenic, and zinc (2). A process (Fig-
ure 1) that integrates soil washing with sequential an-
aerobic and aerobic biotreatment is being developed to
       Wash Solution Recycle
                    Fresh Wash
                      Feed
                         Soil Feed
                                     Water
                                     Feed
                             Outlet Liquid
Figure 1.  Integrated soil treatment process.
cost-effectively remediate soil contaminated with these
wood treating wastes. Soil washing facilitates degrada-
tion by mobilizing the target compounds and expanding
the range of feasible remediation technologies (3). To
reduce costs and the volume of PCP-bearing liquid, the
soil wash liquid is concentrated via distillation, and the
recovered ethanol and water is recycled to the first soil
washing unit. The remainder of the wash solution is
initially bioremediated  in an anaerobic environment.
Mineralization of the target  compounds is completed
aerobically (4). Process development began by inde-
pendently evaluating soil washing and target compound
bioremediation. PCP-contaminated soils were the initial
focus, but this work is currently being  extended to in-
clude  soils  contaminated with both  PCP and PAHs.
Based on preliminary results, the integrated process will
meet the target cleanup level in 73 to 55 percent of the
RODs directed towards PCP remediation, resulting in
soil with a residual PCP level of 8 to 13 mg/kg, respec-
tively (2).

Soil Washing/Solvent  Extraction Studies

An equimass (50 percent) mixture of ethanol and water
(5) was found to be the optimal solution to remove PCP
from a variety of  spiked soils in  a bench-scale  soil
washing process. This soil  washing  method removes
PCP at levels comparable with those achieved through
the analytical techniques of sonication and soxhlet ex-
traction. Starting with initial spike levels of 85 to 100
mg/kg, 70 to 100 percent of the PCP added to the soil
was removed by washing,  depending  on  soil particle
size, contamination  age, and soil washing format. PCP
is  extracted  from soil  in a  30-min  contact time. The
availability of residual PCP  on  soils of 20  x 40, 100 x
140, and greater than 200  U.S.  meshsize has been
evaluated  through  a  serial procedure:  soils were
                                                66

-------
washed with 50  percent  ethanol solution, rinsed with
water, and finally treated by soxhlet or sonication extrac-
tion  using  methanol/methylene chloride. Less than  4
percent of the residual PCP (less than 0.6 mg/kg) was
removed from the soil by the final sonication  or soxhlet
extraction,  demonstrating the limited availability of the
residual PCP. The solvent washing of soil with mixtures
of water and ethanol is also being investigated for PAH-
contaminated soils, using four compounds on the U.S.
Environmental Protection Agency's list of priority pollut-
ants as model compounds: naphthalene, acenaphthene,
pyrene, and benzo(b)fluoranthene (6). A more ethanol-
rich mixture may be required to effectively mobilize PAHs
from soil.

A sequential soil washing train is being optimized for
PCP-contaminated soils in which ex  situ soil washing  is
performed with the 50 percent ethanol solution in three
batch-wash stages. After washing the soil for 30 min  in
each stage, the washed soil is recovered from the soil-
solvent slurry via vacuum filtration of the slurry, and  a
fresh batch of solvent is added to the soil in the next
stage. Preliminary design data indicate that a series
of three 20-mL solvent washes will clean 5 g of soil
(1:12 soil:solvent ratio) as effectively as a single ex-
traction of  100 ml cleans 1 g of soil (1:100 soil:solvent
ratio). Additional  optimization will further decrease the
soil:solvent ratio.

Biological Treatment Studies

In  the integrated  process, the distillate bottoms will be
fed to an  anaerobic fluidized-bed  granular activated
carbon (GAG) reactor. Two of these  reactors were con-
structed and operated for over 40  months, evaluating
variables such as PCP loading and  reactor empty bed
contact time (EBCT) (7). The reactor volume is 10 Lwith
a 1-L recycle loop. Based  on this evaluation, the follow-
ing optimal operating variables were identified: EBCT,
2.3 hr; ethanol loading, 33.3 g/day (loading  rate 6.3  g
chemical oxygen demand/L day); and  PCP loading,
4.8 g/day (loading rate 0.55 g/L day). When the GAG
reactor operated  at an EBCT of 2.3 hr, on a molar basis,
greater than 99.97  percent  of the  influent  PCP was
dechlorinated to monochlorophenol (MCP). In addition,
data from the extraction of the reactor GAG during the
operating period indicated negligible accumulation of
PCP on the surface of the GAG. An aerobic fluidized-bed
GAG reactor will polish the effluent from the anaerobic
GAG reactor to attain complete mineralization of PCP.
Operation of the  aerobic reactor has recently been
initiated.

An additional anaerobic fluidized-bed GAG reactor has
been constructed to evaluate the biotreatment of chemi-
cally synthesized solutions of the four PAHs and PCP in
ethanol. Greater than 99  percent transformation of the
influent  PCP concentration  of 100 mg/L has been
achieved in the reactor, while operating the reactor with
an EBCT of 9.3 hr. The reactor effluent data represent
greater than  99  percent removal for naphthalene, ace-
naphthene,  and pyrene,  and greater than 90 percent
removal for benzo(b)fluoranthene in the reactor.


References

1. U.S. EPA. 1994. Innovative treatment technologies: Annual status
  report, 6th ed. EPA/542/R-94/005. Cincinnati, OH.

2. U.S. EPA. 1992. Contaminants and remedial options at wood pre-
  serving sites. EPA/600/R-92/182. Cincinnati, OH.

3. U.S.  EPA. 1990. Soil washing  treatment. Engineering bulletin.
  EPA/540/2-90/017. Cincinnati, OH.

4. Khodadoust, A.P., J.A. Wagner, M.T. Suidan, and S.I. Safferman.
  1994.  Solvent washing of PCP contaminated soils with anaerobic
  treatment of wash fluids. Water Environ. Res. 66:692.

5. U.S. EPA. 1986.  Microbial decomposition of chlorinated aromatic
  compounds. EPA/600/2-86/090. Cincinnati, OH.

6. Keith,  L.H.,  and WA. Telliard. 1979. Priority Pollutants I: A per-
  spective view. Environ. Sci. Technol. 13:416.

7. Wilson, G.J., J.A. Wagner, A.P. Khodadoust, M.T. Suidan, and R.C.
  Brenner. 1994. The evaluation of empty bed contact time on the
  biodegradation of pentachlorophenol using an anaerobic GAC
  fluidized-bed. In:  Proceedings of the National Conference on En-
  vironmental Engineering: Critical  Issues in Water and Wastewater
  Treatment. American Society of Civil Engineers, p. 624.
                                                    67

-------
          Biological Treatment of Contaminated Soils Using Redox Control
                                       Margaret J. Kupferle
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                        Gregory D. Sayles
    National Risk Management Research Laboratory, U.S. Environmental Protection Agency,
                                         Cincinnati, Ohio

               Tiehong L. Huang, Yonggui Shan, Maoxiu Wang, and Guanrong You
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                       Carolyn  M. Acheson
    National Risk Management Research Laboratory, U.S. Environmental Protection Agency,
                                         Cincinnati, Ohio
Introduction

Land  treatment is a well-understood,  cost-effective
means of conducting aerobic biological treatment of
soils contaminated with aerobically biodegradable com-
pounds, such as petroleum. Common contaminants in
soil also include highly chlorinated organics that are not
readily biodegraded aerobically,  such as pentachlo-
rophenol (PCP), polychlorinated biphenyls (PCBs), and
1,1,1-trichloro-2, 2-bis(p-chlorophenyl)ethane (DDT).
These compounds may, however,  be efficiently de-
graded using a sequential anaerobic/aerobic treatment
strategy. A cost-effective process to treat soils con-
taminated with these highly chlorinated contaminants
is needed. A modified type of prepared-bed land treat-
ment that incorporates variable redox states (i.e., an-
aerobic and aerobic phases) is being evaluated in this
project. The  first  pilot-scale  study, using PCP-con-
taminated soil from the American Wood Products site
in Lake City, Florida, is  in progress at the U.S. Envi-
ronmental Protection Agency's Test and  Evaluation
(T&E) Facility in Cincinnati, Ohio.

Methodology

The pilot-scale study is being conducted in soil pans that
simulate a prepared-bed  land treatment unit with perme-
ate  collection. Each Plexiglas enclosure contains four
soil pans suspended in a controlled-temperature water
bath maintained at 20 + 2°C. Each soil pan (13 in. x
13 in.) is loaded with contaminated soil to a depth of
8 in. over a graded gravel underdrain system, which is
separated from the soil layer by a coarse mesh stainless
steel screen.  During the anaerobic phase, permeate
recycle is optional from the underdrain to the top of the
pan at flowrates of 5 to 20 mL/min. Anaerobic conditions
are maintained in the soil by flooding the pans with clean
creek water from the site. Aerobic conditions are pro-
duced and maintained in the soil by draining the water
and tilling the soil in a manner consistent with landfarm-
ing techniques.

Initial Anaerobic Phase

Based on a review of the literature, which is dominated
by studies evaluating spiked soils, three variables were
selected for study in the initial anaerobic phase: 1) per-
meate recycle vs. no recycle; 2) addition of a supple-
mental organic source (ethanol or anaerobic sewage
sludge); and 3) soil PCP contamination level. The ex-
perimental design was a three-factor analysis of vari-
ance with replication.  The  amounts of soil,  ethanol,
anaerobic digester sludge (32.8 g dry solids/L, 60 per-
cent volatile solids), and site water initially added to each
of the 24 pans are summarized  in Table 1. In situ oxida-
tion-reduction potential (ORP)  probes were placed in
one pan representative of each of the 12  treatments.
Four probes were buried in  each pan, two in opposite
corners approximately 1 in. from the soil/gravel interface
and two in the remaining corners  approximately 1 in.
from the soil/water interface  near the top of the reactor.
                                                68

-------
Recycle flow rates were maintained at 8 to 10 mL/min.
The pH of the water flooding the pans was measured
in situ each week. The soil in the pans was sampled and
analyzed for PCP and its less chlorinated phenolic me-
tabolites, hydrocarbons, and  percent  moisture  on a
monthly basis. Hydrocarbons were present in the site
soil in significant quantities  (1,000 to 2,500 mg/kg  dry
soil) because diesel fuel serves as a carrier for PCP in
wood  treating operations. The water flooding the pans
was analyzed  for PCP and less chlorinated  phenolic
metabolites each time a soil sample was collected.

Aerobic Phase

After 6 months, all of the pans except four of the sludge-
amended pans (2D, 3A, 4B, and  5C) were converted to
aerobic conditions. Afterthe free waterwas drained from
the soil, the soil was tilled  three times a week for 4
weeks, until it dried to less than  10 percent total  mois-
ture. When the conversion phase was completed,  some
of the pans were continuously supplied with air at a  low
flow rate (in addition to weekly tilling of all pans) and/or
amended with  poultry  manure  (see Table 1).  Each
month, the soil in the pans is sampled and analyzed for
PCP and its less chlorinated phenolic metabolites, as
well as hydrocarbons. Moisture content and water addi-
tion volumes are measured weekly to maintain moisture
content in a constant range.

Results

Anaerobic conditions were established in the soil pans
after the first week. The measured ORPs ranged  be-
tween -150 to -500  mV (versus Ag/AgCI reference elec-
trode). No apparent correlation was found with respect
to probe depth, soil type, or treatments. The soil sample
PCP concentration  data showed no significant amount
of PCP removal in any of the treatments after 6 months.
Changes in  PCP concentration  in the flood water of
several of the pans were noted, however. After2 months,
the PCP concentration in the flood water of the soil pans
containing sludge dropped  from  15 to 55 mg/L to less
than 0.5 mg/L. After 4 months, the PCP concentration
also  dropped to less than 0.5 mg/L in the two replicate
pans with  low-PCP soil treated without recycle or sup-
plemental  organic source. The less chlorinated phenolic
metabolites were not detected as intermediates in the
flood water from any of these  pans. No degradation  of
hydrocarbons was noted in any of the pans in the initial
anaerobic phase, but degradation has occurred in the
aerobic phase. The presence of hydrocarbons may have
interfered  with the bioavailability of  PCP in the initial
anaerobic phase. To test  this  hypothesis,  aerobically
treated soil will be reconverted to anaerobic conditions
once the hydrocarbon concentration has been reduced.
Another possibility is that  appropriate anaerobic PCP
degraders are not present  in sufficient quantities in the
soil  pans. A bench-scale  study using  soil from the
same source as the  pilot-scale study  has been initi-
ated to investigate the effect of amendment with PCP-
acclimated culture.
Conclusions

Adaptation of the pilot-scale land-treatment units to an-
aerobic operation has been evaluated. Flooding the soil
with water successfully creates a low redox (anaerobic)
state. The in situ ORP probes constructed for the project
work well. Monthly sampling intervals and the analytical
techniques  used adequately characterize system be-
havior.  The tilling strategy used in the conversion from
anaerobic to aerobic operation was successful.  The
presence of significant amounts of hydrocarbon co-
contamination  may have affected  PCP  degradation,
suggesting that a more appropriate treatment sequence
may be aerobic-anaerobic-aerobic. This observation re-
inforces the importance of technology evaluation  with
soils characteristic of those found at actual sites.
                                                  69

-------
Table 1.  Soil Pan Operation Summary
Pan3
6B
1C
2B
6Cf
1Af
3B
3C
2Cf
4C
5Cf
4Bf
6D
2A
1Df
5A
4Df
3Df
5B
6A
5Df
1B
2Df
3Af
4A

Soil
Lowd
36
36
36
36
36
36
36
36
36
36
36
36
-
-
-
-
-
-
-
-
-
-
-


(kg as is)
High6
-
-
-
-
-
-
-
-
-
-
-
-
35
35
35
35
35
35
35
35
35
35
35
35
Initial Anaerobic Phase Treatments'3
Ethanol Sludge Water
(mL) (L) (L)
16
16
16
16
17.2 - 16
17.2 - 16
17.2 - 16
17.2 - 16
5.16 11
5.16 11
5.16 11
5.16 11
16
16
16
16
17.2 - 16
17.2 - 18h
17.2 - 16
17.2 - 16
4.88 12
4.88 12
4.88 10.5h
4.88 12
Aerobic Phase
Treatments0
Recycle
(mL/min)
-
-
8-10
8-10
-
-
8-10
8-10
8-1 Og
-
8-1 Og
-
-
-
8-10
8-10
-
-
8-10
8-10
-
-
8-1 Og
8-1 Og
Air
(mL/min)
-
10
-
10
-
10
-
10
-
Not
Not
10
-
10
-
10
-
10
-
10
-
Not
Not
10
Manure
(g/pan)
-
750
750
-
-
750
750
-
-
converted
converted
-
-
750
750
-
-
750
750
-
-
converted
converted

  Pan location for treatments (Pans A-D in Boxes 1-6) randomly assigned for statistical purposes.
b Initial anaerobic phase from August 11, 1994, to March 2, 1995.
c Aerobic phase from March 2, 1995, to July 7, 1995.
d Soil from 5 ft depth at site containing approximately 250  mg PCP per kg dry soil.
e Soil from 12 ft depth at site  containing approximately 650 mg PCP per kg dry soil.
f  In situ ORP probes added to pan during  initial anaerobic phase.
g Recycle was set at 9 mL/min initially but was discontinued  after the first week due to  extremely low flow (less than 1 mL/min) through
  sludge-amended soils.
h Amount of water required to maintain a constant depth of 2 in. above soil surface varied somewhat.
                                                             70

-------
  Development of a Sulfate-Reducing Bioprocess To Remove Heavy Metals From
                                Contaminated Water and Soil
                               Munish Gupta and Makram T. Suidan
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                           Gregory D. Sayles and Carolyn M. Acheson
                     U.S. Environmental Protection Agency,  Cincinnati, Ohio
Introduction
Acid mine drainage is characterized by low pH (1.5 to
3.5) and  high concentrations of sulfate and dissolved
heavy metals. Bacterial sulfate reduction has been iden-
tified as a potentially cost-effective process for removing
metals from mine drainage (1, 2). Sulfate-reducing bac-
teria convert sulfate to sulfide using an organic carbon
source as the electron donor. The sulfide precipitates
the various metals present in the wastewater, yielding a
very low concentration of dissolved metals in the efflu-
ent. In this study, acetic acid was used as the carbon
source for two reasons: it is relatively inexpensive, and
being an  acid, it can effectively leach out metals from
contaminated soils such as mine tailings.

Reactor Selection
To effectively treat metal-contaminated wastewater,  a
reactor must establish an anaerobic environment to sup-
port sulfate reduction, resulting in metal precipitation as
metal sulfides, and must provide an efficient clarifier to
remove metal precipitates from the effluent. Because
sludge (metal precipitates and biomass solids) would
accumulate and eventually clog  the reactor, ease of
sludge removal or cleaning is an important considera-
tion in selecting a reactor.  Two reactors were evaluated:
an upflow anaerobic filter packed with plastic Pall rings
and an anaerobic sludge  blanket reactor. To clean the
reactor, sludge can be removed from the bottom of the
sludge blanket reactor. The same  technique  can be
used for the filter; however, it may be more difficult due
to packing material. The cleaning of the filter, therefore,
was an additional aspect of research.
Reactor Operation and Performance
Two filters (A and B) and one sludge  blanket reactor
were operated at a temperature of 30°C and at a pH of
7.2, optimal  pH for  sulfate-reducing microorganisms.
The feed concentration  of metals (shown in  Table 1)
used in this study were among the highest concentra-
tions observed  at mines in Montana  and  Colorado.
After an initial acclimation, Filter A was fed the metals
listed in Table 1, while Filter B and the sludge blanket
reactor were fed iron at a concentration equal to the
sum of the molar concentrations of all the metals fed
to Reactor A. Table 1 shows characteristic effluent
Table 1.  Influent and Effluent Concentrations (mg/L) for Filter A

                              Effluent Concentration
Influent
Constituent
Iron
Zinc
Manganese
Copper
Cadmium
Lead
Arsenic
Acetate
Sulfate
Total sulfide
Influent
Concentration
840
650
280
130
2.3
2.1
1.5
3000
5000
0
Filtered
0.09
0.14
5.4
0.02
0.02
0.005
0.01
11.0
800
31
Total
0.497
0.310
5.60
0.022
0.019
0.005
0.01
—
—
—
                                                71

-------
concentrations  for Filter A,  operating at a hydraulic
detention time of 5 days.

During this period, the level  of sludge rose above the
packed bed in Filter A. To investigate whether sludge
withdrawal from the bottom  would control the sludge
height in the filter, 1 L of sludge was  removed. Sludge
withdrawal lowered the sludge level and the filter con-
tinued to  operate efficiently,  with  less than 1 percent
change in effluent conditions. A similar situation in Filter
B was also corrected in the same fashion. Sludge with-
drawal from the bottom can, therefore, control the accu-
mulation of sludge and prevent clogging of the filters.

The sludge blanket reactor did not perform very well as
a clarifier. Although the soluble iron concentration in the
effluent was less than 0.25 mg/L, the total concentration
was  as high as  25 mg/L  and varied  between 18 and
22 mg/L. This reactor had a high concentration of total
suspended solids in the effluent compared with  Reac-
tors A and B. It was concluded that this type of reactor
was not effective in clarification and was unable to meet
the requirements. Therefore, the operation of the sludge
blanket reactor was discontinued.
Conclusions

Compared with an anaerobic sludge blanket reactor, an
upflow anaerobic filter packed with Pall rings was found
to be a very efficient reactor for the treatment of water
contaminated with heavy metals. The filter, unlike the
sludge blanket  reactor, worked very well as a clarifier,
and  all metals  except  manganese were reduced to a
concentration close to drinking-level standards. Sludge
withdrawal from the bottom  of the filter can be used to
remove accumulating sludge,  and, therefore, the filter
can be operated continuously.  Ongoing work will evalu-
ate the performance of the filters as a function of hydrau-
lic retention  time, lower temperatures, and pH. Sludge
removal frequency will also be optimized.

References

1.  Dvorak, D.H., R.S. Hedin, H.M. Edenborn, and RE. Mclntire. 1992.
   Treatment of metal-contaminated water using sulfate reduction:
   Results from pilot-scale reactors. Biotech. & Bioeng. 40:609-616.
2.  Kuyucak, N., D. Lyew, P. St. Germain,  and K.G. Wheeland. 1991.
   In  situ bacterial  treatment of AMD in open pits. Presented at the
   Second International Conference  on  the Abatement of Acidic
   Drainage, Montreal, Canada, September 16-18.
                                                    72

-------
  Development of Techniques for the Bioremediation of Chromium-Contaminated
                                     Soil and Ground Water
            Michael J. Mclnerney, Nydia Leon, Veronica E. Worrell, and John D. Coates
                            University of Oklahoma, Norman, Oklahoma
The potential for biotic Cr(VI) reduction in samples from
a  Cr(VI)-contaminated aquifer (Elizabeth City,  North
Carolina) was evaluated by inoculating aquifer material
into anaerobically prepared mineral salts medium that
did not contain chemical reductant. In  inoculated micro-
cosms, the Cr(VI) concentration decreased after 5 days
incubation at 25°C, and almost all of the Cr(VI) was gone
after 25 days (Figure 1). Little or no  change in Cr(VI)
       0  5
10 15  20 25 30 35 40 45  50  55
       Time (days)
   - uninoculated
   - autaclaved+HgCI2
      - autoclaved
      - 2Xautoclaved+HgCI2 -
 2X autoclaved
- nonsterile
Figure 1.  Biological reduction of Cr(VI) with aquifer material.

concentration was observed in uninoculated controls, or
in sterile controls prepared by autoclaving, boiling, orthe
addition  of HgCI2  or  chloramphenicol.  Hydrogen  re-
duced the lag time before Cr(VI) reduction occurred but
did not markedly affect the rate of Cr(VI) reduction (Fig-
ure 2). The addition of other exogenous electron donors
such as glucose, acetate, formate, or benzoate did  not
affect the rate or lag time associated with Cr(VI) reduc-
tion in microcosms compared  with controls that lacked
an exogenous electron donor. The  addition of phenol,
lactate, and ethanol to microcosms inhibited  Cr(VI) re-
duction. Subsequent addition  of Cr(VI) to microcosms
with benzoate as the electron  donor decreased the  lag
time and  increased the rate of Cr(VI) reduced compared
with that  observed initially.
                                                                       10     15
                                                                       Time (days)
-•-
sterile
-*-
no donor
-B~
acetate
-•-
hydrogen
Figure 2.  Effect of electron donors on Cr(VI) reduction with
         aquifer material.

The effect of sodium sulfate, sodium nitrate, amorphous
ferric hydroxide  (each at 10 mM) on Cr(VI) reduction
was tested with benzoate-amended microcosms. The
presence of sulfate and nitrate inhibited the reduction of
Cr(VI) compared with microcosms that did not receive
any of the three additional electron acceptors. Sulfide
levels remained unchanged during the course of the
experiments. In bottles with nitrate, nitrite accumu-
lated after 8 days and decreased after 16 days. Ferric-
hydroxide-supplemented microcosms reduced Cr(VI)
to a much greater extent than unsupplemented control
cultures; ferrous iron production coincided with Cr(VI)
reduction.

Two facultative bacteria that can  reduce  Cr(VI) were
isolated from Elizabeth City aquifer material,  and one
bacteria was isolated from an aquifer underlying a land-
fill in Norman, Oklahoma, using a mineral salts medium
with 5 mM  benzoate, 500 u,M Cr(VI). All three isolates
are gram-negative,  motile rods that grow singly, in pairs,
and in branched chains. On agar medium, the isolates
formed shiny, smooth,  pink colonies  and  produced a
diffusible green pigment. Upon  initial isolation, Cr(VI)
was rapidly reduced (Figure 3);  however, the  rate and
                                                  73

-------
                                                       Table 1.  Electron Donors That Support Cr(VI) Reduction by
                                                               Strain NLB
                 10
15   20   25
  Time (days)
            + benzoate
    no donor
killed cells
Figure 3.  Cr(VI) reduction by strain NLB.

extent of Cr(VI)  reduction  decreased  with  repeated
transfer of the culture in benzoate-Cr(VI) medium. The
use of Cr(VI)  was dependent on the presence of an
electron donor and an active  inoculum.  In addition to
benzoate, other substrates supported Cr(VI)  reduction
(Table 1). Increases  in cell numbers were  observed
when the electron donor and Cr(VI) were both pre-
sent.  In the  absence  of Cr(VI) or electron donor, little
or no increase in cell  number was observed:  less than
6x 10s cells/ml.

Four  observations supported  the conclusion that the
decrease in  Cr(VI) concentration was a biologically me-
diated reduction process:  1) Cr(VI) concentrations de-
creased faster and to a  greater extent in nonsterile
versus sterile microcosms; 2) phenol, ethanol, and  lac-
Additions
Fumarate
Purine/Pyrimidine mix3
p-Toluic acid
Lactate
Phenoxyacetate
Malate
Benzoate
Phenol
Ethanol
No addition
Cr(VI)
Reduced
(|iM)
414
337
184
169
166
80
68
59
32
0
Increase in Cell
Numbers
([cells/mL] x 106)
0
31.8
30.1
62.8
56.3
3.8
142.9
19.3
7.8
0
                                  Mix contains 0.5 mM each: adenine, guanine, thymine, uracil.

                                 tate inhibited Cr(VI)  reduction in microcosms; 3) re-
                                 peated additions of Cr(VI) to microcosms decreased the
                                 lag time and stimulated the rate of Cr(VI) reduction; 4)
                                 bacteria were isolated and capable of using Cr(VI) as an
                                 electron acceptor. Iron  hydroxide stimulated Cr(VI) re-
                                 duction in microcosms, most likely by an indirect mecha-
                                 nism involving the production of ferrous iron. The extent
                                 and rate of Cr(VI) reduction by aquifer microcosms was
                                 not affected when exogenous electron donors, with the
                                 exception of hydrogen, were added.  This indicates that
                                 the aquifer material had sufficient levels of endogenous
                                 electron donors to support Cr(VI) reduction.
                                                   74

-------
    Bioremediation of Chlorinated Pesticide-Contaminated Sites Using Compost
                   James C. Young, Jean-Marc Bollag, and Raymond W. Regan
                   Pennsylvania State University, University Park, Pennsylvania
Many sites throughout the United States are contami-
nated with chlorinated  pesticides. Of particular interest
to this project are those sites contaminated with chlor-
dane and toxaphene. One objective is to determine the
feasibility of using  compost as a culture medium for
mediating the biodegradation of these pesticides. A sec-
ond objective is to determine major pathways of chlor-
dane  and  toxaphene  biodegradation that lead  to
mineralization. These objectives are  particularly chal-
lenging because chlordane and toxaphene each consist
of several chlorinated cyclic hydrocarbons that individu-
ally may follow different biodegradation pathways, or
may be only partially dechlorinated.

Biodegradation of chlordane and toxaphene and other
chlorinated pesticides is expected to require an organic
co-substrate as a carbon source for the growth of accli-
mated microorganisms that enzymatically are capable
of dechlorinating the  pesticides through reductive or
oxidative reactions.  Co-substrates considered for use in
field applications include milk solids, sugar, blood meal,
sewage solids,  methane, or,  in  the  current project,
compost.


The test program includes the development and op-
eration of a pilot-scale compost reactor that contains
a mixture of 10 percent municipal-sludge compost,
10  percent  spent-mushroom  compost,  40 percent
grass, and 40 percent alfalfa hay to provide an environ-
ment suitable for the culture of chlordane- and toxaphene-
degrading microorganisms. This compost is used to amend
various contaminated-soil matrices followed by analysis of
the fate of the pesticide. Residual  pesticides are moni-
tored using gas  chromatography, thin  layer chromatog-
raphy, and mass spectroscopy. Test parameters include
soil type, compost-soil ratio, moisture level, oxidation-
reduction potential, pH, presence of sulfates and ni-
trates, and the effect of supplemental soluble organic
co-substrates.
                                                 75

-------
                            Reductive Electrolytic Dechlorination
                            John W. Norton, Jr., and Makram T. Suidan
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                             Carolyn M. Acheson and Albert D. Venosa
   National Risk Management Engineering Laboratory, U.S. Environmental Protection Agency,
                                          Cincinnati, Ohio
Reductive dehalogenation is the only known mechanism
to biologically degrade some highly chlorinated organic
compounds, including  pentachlorophenol (PCP)  (1),
and occurs primarily in anaerobic environments (2). A
biofilm-electrode reactor  (BER) was  constructed to
evaluate  PCP  dechlorination as a function of ethanol
concentration and the presence of an electrical current.
The  BER was  operated  as follows: current, 20  mA;
hydraulic retention time, 0.38 days;  PCP feed, 5 mg/L;
ethanol feed range, 0 to 100 mg/L. The best observed
dechlorination occurred when 5 mg/L PCP and 25 mg/L
ethanol were fed to the reactor. The effluent under these
conditions contained 0.013 mg/L PCP,  0.26 percent of
the feed concentration. At lower ethanol levels, PCP was
not as effectively dechlorinated. The trichlorophenols (TCP)
and dichlorophenols (DCP) displayed a two-  to three-
fold increase in effluent concentration as the substrate
ethanol was decreased, particularly at concentrations  less
than 10 mg/L. The monochlorophenols (MCP), however,
reached a maximum of 0.014 mM at ethanol concentra-
tions of 10 to 25 mg/L.  The  total  dechlorination  de-
creased significantly when the ethanol was  removed
from the feed, indicating that the  ethanol  stabilized
dechlorination.

After characterizing the ethanol requirements in the  sys-
tem, the role of the current in the dehalogenation of the
PCP was evaluated by turning  off the current. Electrical
current was shown to play a necessary  role in dechlori-
nation, although it is unknown whether this role was the
result of the hydrogen generation or the low reducing
potential surface formed on the cathode of the anode-
cathode cell. Two trials of reactor operation without cur-
rent were conducted. Each current removal trial caused
a reduction in PCP dehalogenation, demonstrated by
the successive appearance of the  higher chlorinated
phenols in the effluent. The two trials displayed very
different temporal behavior, however. The first removal
resulted in a quick  rise in effluent PCP concentrations,
increasing one order of magnitude in a few hours. Fol-
lowing the second current removal,  a much slower ap-
pearance  of  chlorinated  phenols was  observed; the
effluent PCP concentration increased one order of mag-
nitude in approximately 6 days. During each trial, after
the current was reapplied, the system recovered. Each
trial showed  a recovery pattern similar to the failure
preceding  it, the first trial showing a  quick recovery and
the second trial showing a much slower recovery. The
causes of the different behaviors have not been  charac-
terized. We are presently evaluating the role of electrical
current by  varying the current while keeping the ethanol
concentration constant at 10 mg/L.

References

1. Mohn, W, and J. Tiedje. 1992. Microbial reductive dehalogenation.
  Microb. Rev. 56:482-507.
2. Suflita, J., A.  Horowitz, D. Shelton, and J. Tiedje. 1982. Dehalo-
  genation:  A novel pathway for the anaerobic biodegradation of
  haloaromatic compounds. Science 218:1,115-1,117.
                                                  76

-------
                 Biological Ex Situ Treatment of PAH-Contaminated Soil
                                            Carl L. Potter
                      U.S. Environmental Protection Agency, Cincinnati, Ohio

                                           Roy C. Haught
                                  IT Corporation, Cincinnati, Ohio
The goal of this  project is to evaluate the potential of
biological ex  situ soil  treatment systems  (biopiles) to
remediate soils contaminated with hazardous chemi-
cals. A laminar-type flow pilot-scale reactor with a vol-
ume of  3  yd3  has been constructed  at  the U.S.
Environmental Protection Agency's Test  & Evaluation
(T&E) Facility in  Cincinnati, Ohio. Laminar-type  flow
from one side of the reactor to the other may provide
even aeration to  all areas of the reactor while avoiding
the use of pipes  inside the reactor. This design greatly
facilitates loading and  unloading of the reactor and is
readily scalable to larger systems.

Passing smoke through the reactor for visual observa-
tion of flow indicated uniform, laminar-type flow through
the empty reactor. Further testing involved filling  the
reactor with vermiculite or a synthetic soil, flushing with
argon, and then passing airthrough the reactorto evalu-
ate air flow through this uniform solid matrix. Oxygen
probes, located at 27 positions within the reactor,  indi-
cated rapid and  uniform air saturation of the system.
Analysis  of gas  flow  through  an  empty  reactor and
through uniform matrices allowed evaluation of reactor
performance without confounding effects of soil inhomo-
geneities that may lead to nonuniform aeration of the
reactor space.

The  reactor uses pulsed air flow through the  pile  to
permit maximum distribution of air within the soil. Air is
driven into the soil during pulse action, then allowed to
diffuse in all directions during the rest interval.

Soil contaminated with polynuclear aromatic hydrocar-
bons (PAHs) from the Reilly Tar Pit Superfund site in St.
Louis Park, Minnesota, has been  brought to the T&E
Facility for research on soil aeration  and effectiveness
of this ex situ reactor design for biological treatment of
contaminated soils. Micronutrients were adjusted  to
100:20:1 phosphorus:carbon:nitrogen, and 0.5 percent
by weight cow manure was  added to the soil. A10-week
treatability study is under way to evaluate disappear-
ance  of parent PAHs and microorganism population
changes in this reactor system.
                                                  77

-------
     Effectiveness of Gas-Phase Bio re mediation Stimulating Agents (BSAs) for
                         Unsaturated Zone In Situ Bioremediation
                        James G. Uber, Ronghui Liang, and R. Scott Smith
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                         Paul T. McCauley
Water and Hazardous Waste Treatment Research Division, National Risk Management Research
               Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio
Background

Successful  in situ  bioremediation  in the unsaturated
zone requires that water, oxygen, nutrients, primary sub-
strate, and perhaps co-metabolites be available to  the
microorganism via physical transport mechanisms. Any
of these substances may be called a bioremediation
stimulating agent (BSA), given that a shortage of any
one may adversely affect the performance of an in situ
bioremediation system. Other  potential BSAs include
substances (e.g., surfactants) that are not ordinarily re-
quired for microbial growth but that may enhance sub-
strate or nutrient bioavailability.

Much work has focused on engineering approaches to
deliver BSAs at field scale.  Little  research has been
conducted, however, to evaluate which in situ  delivery
approaches are best for transporting BSAs to microor-
ganisms. Given the complexity of two-phase (gas/water)
or three-phase (gas/water/nonaqueous-phase liquid)
fluid and contaminant transport in the unsaturated zone,
considerable uncertainty exists about the ultimate distri-
bution of BSAs in contaminated soils. Further, microbial
growth processes affect fluid and contaminant transport
not only through biochemical reactions but also through
a spatial-temporal influence on fluid permeability. (Plug-
ging  of pore spaces by microorganisms can reduce
wetting fluid  permeability by  greater than 99 percent.)
The present study will identify in situ BSA delivery strate-
gies that are most likely to achieve a uniform BSA spatial
distribution and, therefore, most likely to improve biore-
mediation  field  performance. As  a byproduct  of this
work, the project aims to identify and measure the fun-
damental physical and microbial processes that affect
bioremediation performance enhancement through BSA
delivery methods.
Because of strong capillary forces that affect the distri-
bution and  movement of wetting fluids in unsaturated
soils, gas-phase BSAs are more likely to achieve uni-
form in situ  spatial distribution. It is, in fact, well known
that movement of water in the unsaturated zone often
occurs in discrete fingers that occupy a small fraction of
the total pore space. Relatively little is known, however,
about the characteristics of in situ gas-phase BSA trans-
port, including physical factors that may lead to complex
and undesirable flow patterns (e.g., interactions of water
saturation and air permeability), chemical transport fac-
tors that may limit  gas-phase BSA spatial distribution
(e.g., BSA solubility), and dynamic microbial factors that
may affect BSA transport in the field (e.g., microbial BSA
utilization rates and plugging that leads to heterogene-
ous effects  on air/water permeability). Because of the
promise of gas-phase BSAs  and  the  significant un-
knowns  regarding their effectiveness, this project will
focus on effective gas-phase addition of nutrients, co-
metabolites, oxygen, and moisture.

Objectives

The objectives of this work are to:

• Evaluate  the effectiveness  of field systems  for gas-
  phase delivery of BSAs to the unsaturated zone for
  enhancing in situ  bioremediation performance. These
  BSAs include nutrients  (organophosphates), co-
  metabolites, surfactants or solvents, and water vapor.

• Identify and measure the physical and microbial fac-
  tors affecting the bioavailability of gas-phase  BSAs
  in the  unsaturated zone, including uneven spatial dis-
  tribution of BSAs at the pore- and core-scales and
  complex changes in unsaturated zone air permeabil-
  ity caused by microbial-growth dynamics.
                                                 78

-------
• Develop visible light tomography (VLT) systems that
  allow visualization of in situ unsaturated zone physi-
  cal and microbial processes for controlled evaluation
  of alternative BSAs and delivery systems.

• Use  the VLT systems to evaluate alternative  BSAs
  for remediation of aged contaminated soils in control-
  led but realistic environments.

Accomplishments

We have designed and constructed two different labo-
ratory systems for observing dynamic fluid distribution
in the unsaturated zone under simulated BSA delivery.
A three-dimensional column system has been  designed
to collect data on fluid migration through discrete fingers
in disturbed and undisturbed soil cores, and will be used
to  measure the limitations on BSA delivery caused
by fingering under a variety of soil conditions and
fluid application rates. This work will be completed
by December 1995. The columns are 30 cm in diameter
and comprised of stackable 10-cm sections separated
by 1-mm spacers. The columns rest on a base that
allows  manipulation of the bottom pressure  boundary
condition, and  the side boundary condition is manipu-
lated through 1-mm gaps  between rings. Water is ap-
plied uniformly to the top surface of the soil columns via
a carefully designed air-atomizing nozzle. After the fluid
flow is  developed, a dye mixture marks the locations of
any preferential flow  pathways.  The pathways will be
exposed at the surfaces of each  10-cm ring, and the
complete three-dimensional character of each pathway
will be recorded. Different color dyes will  be used to
investigate  the  persistence of individual fingers  when
fluid application is cycled, allowing the soil to drain to
varying water contents between application.
A two-dimensional,  vertical, thin-slab  visible  light to-
mography (VLT) system has been designed to visualize
and measure the interactions between  gas-phase BSA
and liquid-phase flow in a controlled environment. The
system will also serve as a bioremediation simulator to
measure the effectiveness of various gas-phase BSAs,
and to visualize dynamic microbial-growth  processes
under simulated in situ bioremediation  conditions (with
and without BSA addition). The vertically oriented cham-
ber dimensions are 1 m x 2 m x 1 cm. The top boundary
will be either open or closed to the atmosphere, and will
be capable of having controlled amounts of liquid added
uniformly over the slab length.  Side boundaries will be
either closed or open to the atmosphere, thereby provid-
ing  the  ability to  control gas-phase BSA injection or
extraction  (simulating the operation of BSA injection or
extraction  wells). The bottom  boundary will be  either
open to the atmosphere or, via  a manifold, will simulate
water table conditions.

The advantage of the  thin-slab system is the ability to
visualize the complex flow and  microbial processes oc-
curring in the unsaturated zone under simulated in situ
conditions. A bank of high frequency fluorescent lamps
will illuminate the system from  the back. Because light
transmission  is related to water saturation, the  water
distribution can be easily visualized without the use of
dyes. Fluorescent gases will be investigated for visuali-
zation of the gas movement, as will color-marked gas-
and liquid-phase pH indicator solutions. Data will be
recorded via a CCD  camera  and  a data  acquisition
system so that actual fluid flow and microbial processes
can be recorded and visualized. The CCD camera sys-
tem will  collect data  at a spatial resolution on the order
of the pore scale (approximately 0.5 to  1.0 mm).
                                                  79

-------
                                      Section 5
                                Process Research
Process research involves isolating and identifying microorganisms that carry out biodegradation
processes and the environmental factors affecting these processes. It also deals with the develop-
ment of techniques for modeling and monitoring biodegradation. Through this research, scientists
establish the building blocks  of  new biosystems for treatment of pollutants  in surface  waters,
sediments, soils, and subsurface  materials. Thorough evaluation is critical at this level of research
because a firm scientific foundation can facilitate the scaling up of a promising technology. Process
research is being conducted on many environmental pollutants.

Two research projects quantified the extent of biodegradation of organic compounds using carbon
isotopes. The methods employed in this research can be used both in the laboratory, where target
compounds are isotopically labeled, and in the field, where contaminants contain different concen-
trations of isotopes than the surrounding environment. The methods were applied in the measure-
ment of petroleum degradation and the demethylation of organometallic compounds.

Another project determined the  kinetic rate constants of anaerobic degradation  of quinoline by
methanogenic bacteria. This system is potentially of great practical significance because quinoline
is found in wastes at wood preserving plants.

Several papers focused on the use of microorganisms to degrade alkyl halides and polychlorinated
aromatics. Research included the characterization  of a  bacterial enzyme  (toluene 2-monooxy-
genase) that can degrade trichloroethylene and of  a strain of Pseudomonas cepacia G4 that
expresses the enzyme constitutively. Further research covered the effects of varying environmental
conditions on the bioremediation of chlorinated compounds.

Other process research projects studied heavy metal inhibition of the bioremediation of  polychlori-
nated aromatics and the effects of different primary substrates on the reduction of 2,4-dinitrotoluene.
Primary substrates are used to sustain the growth of microorganisms when the target contaminant
cannot act as a food source.

Several poster  presentations  at  the symposium  involved process research.  The presentations
covered methods for monitoring  bioremediation, the use of surfactants in sediments, and  charac-
terization of an anaerobic dehalogenating microorganism. Another presentation  dealt with the
potential of Mycobacterium to  mineralize polycyclic aromatic hydrocarbons.
                                           81

-------
                  Monitoring Crude Oil Mineralization in Salt Marshes:
                            Use of Stable Carbon Isotope Ratios
                                       Andrew W. Jackson
         Department of Civil and Environmental Engineering, Louisiana State University,
                                     Baton Rouge, Louisiana

                                          John H. Pardue
     Wetland Biogeochemistry Institute, Louisiana State University, Baton Rouge, Louisiana
Introduction
The ability to monitor mineralization of hydrocarbons is
of prime importance in a successful remediation strat-
egy. Hydrocarbon mineralization must be ensured, be-
cause hydrocarbons can be  sorbed, transformed, or
buried, or otherwise be undetected but still pose threats
to the existing system ecology. One successful tech-
nique  has been monitoring changes in oil composition
relative to a stable, nondegradable compound (1,  2).
Two disadvantages  to this method exist, however: 1) its
inability to demonstrate mineralization instead of trans-
formation, and 2) its inability  to measure absolute oil
degradation, because only "resolved" compounds are
quantified.
A promising new technique for the detection and quan-
tification of hydrocarbon mineralization is the use of
stable carbon isotope  ratios (3). Carbon dioxide gas
ratios vary from of 12C to 13C depending on the source
of the gas. Crude oils are more depleted in 13C, and thus
the mineralization of oil produces CO2 with lower 8-13C
values. Oil has a 8-13C value of-29 to -32 (0/00) depend-
ing on the source of the oil. Salt marshes are predomi-
nantly colonized by C3  plants, and CO2 evolved from
these soils has 8-13C ratios of-14.4 to -17.7 (0/00) (4).
If biodegradation is  occurring  in a contaminated salt
marsh, the  8-13C value of the produced CO2 should
decrease due to the presence of 13C-depleted CO2 from
the crude oil. If this occurs, it would be possible to qualify
and  quantify hydrocarbon degradation by  measuring
total  CO2 production and changes in the 13C signature
of CO2 produced from the marsh.
Theoretical

The rate of CO2 produced from each carbon source can
be easily computed  using three  equations describing
CO2 production and the 8-13C signature:

                   R0+Ri=R,             (Eq. 1)

                   RO   R,  „              (Eq. 2)
                                          (Eq. 3)
where R0 and R| are the rates of CO2 production from
the crude oil and indigenous carbon  sources respec-
tively, and R, is the total rate of CO2 production. S0 and
S| are the 8-13C signatures of the crude oil and indige-
nous carbon, and S, is the measured 8-13C signature of
the produced CO2. S0,S|, St, and R, are experimentally
determined. R0 and Rm can  then be determined from
Equations 2 and 3. This assumes that CO2 is generated
from only these two carbon pools and that there is no
addition of atmospheric CO2.

Results and Discussion

Kinetic Experiments

Microcosm studies showed rapid and nearly complete
(greater than 90 percent reduction in the hopane ratio)
degradation of parent alkanes in the fertilized treatments
under completely mixed, aerated conditions. In the un-
fertilized treatment, less than a 10 percent reduction was
observed in the hopane ratio of the alkanes (Figure 1 A).
Polycyclic  aromatic  hydrocarbon  (PAH) degradation
                                                83

-------
      900
  tn
  g
 '
 or
 0)
 (0
 Q.
 o
 X
 
                                                        I  I
                                                        CL


-g- 3 -
O
-*?
£ 2"
X—.
1 1 -
9
S 0
0 °

Fertilized (1)
Fertilized (2)

BH Unfertilized (1)
^B Unfertilized (2)

I
m \
_ 1 1 !

1 a
Jjj
||








1

                                                                     2345

                                                                         Time (Weeks)
                                                Figure 3.  Rates of CO2-C mineralized from crude oil in fertilized
                                                         and unfertilized salt marsh soils (calculated from iso-
                                                         tope dilution equations).
                                                    84

-------
threne, however, appear to be stable over the period of
this experiment.

The CO2 production rates and the 8-13C ratios measured
were used to calculate the CO2 produced from crude oil
(Figure 3). No mineralization of crude oil was detected
until Week 2, and the majority of mineralization appears
to begin  at Week 5.  The fertilized treatments appeared
to show  higher mineralization rates before the unfertil-
ized and to mineralize at a more even prolonged rate.
The  unfertilized treatments have a more intense  rate
of mineralization  but  for  only  one sampling date.
Amendments of fertilizer inconclusively increased deg-
radation, as evidenced by  hopane ratios of specific oil
components.

The importance of the 8-13C data is the data's ability to
calculate mineralization rates directly. They measure the
final product, while monitoring hopane ratios only meas-
ures the disappearance of the parent compound, not
mineralization. These experiments support the ability to
use  8-13C ratios in conjunction with CO2 production  to
qualitatively and quantitatively monitor crude oil degra-
dation.


References

1. Bragg, J.R., R.C. Prince, E.J. Harner, and R.M. Atlas. 1994. Nature
  368:413-418.

2. Bragg, J.R., R.C. Prince,  E.J. Harner, and R.M. Atlas. 1993.  In:
  Proceedings of the 1993  International Oil Spill Conference, pp.
  435-447.

3. Aggarwal, P.K., and R.E.  Hinchee. 1991. Environ. Sci. Technol.
  25:1,178-1,180.

4. Chmura, G.L., R.A. Socki, and R. Abernethy.  1987.  Oecologia
  74:264-271.
                                                     85

-------
                          Mercury and Arsenic Biotransformation
                                       Ronald S. Oremland
                          U.S. Geological Survey, Menlo Park, California
This presentation  will cover our recent findings with
regard to bacterial processes affecting 1) methylmer-
cury demethylation and 2) the dissimilatory reduction of
arsenic (As) (V).

Methylmercury Oxidative Degradation
Potentials in Contaminated and Pristine
Sediments of the Carson  River, Nevada

Sediments from mercury-contaminated and uncontami-
nated  reaches  of the Carson  River, Nevada,  were
assayed forsulfate-reduction, methanogenesis, denitri-
fication, and monomethylmercury (MeHg) degradation.
Demethylation of 14C-MeHg was detected at all sites, as
indicated by the formation of 14CO2 and 14CH4. Oxidative
demethylation was indicated  by the formation of 14CO2
and was  present  at significant  levels in all samples.
Oxidized/Reduced demethylation product (ORDP)  ra-
tios (e.g., 14CO2/14CH4) generally ranged from 4.0 in
surface layers to as low as 0.5 at depth. Production of
14CO2 was most pronounced at sediments surfaces that
were zones of active denitrification and sulfate-reduc-
tion, but was also  significant within zones of methano-
genesis. In a core taken from an uncontaminated site
having more oxidized, coarse-grained sediments, sul-
fate-reduction and methanogenic  activities were very
low, and 14CO2 accounted for 98 percent of the product
formed from  14C-MeHg. No relationship was apparent
between the degree of mercury contamination of the
sediments and the occurrence of oxidative demethyla-
tion. Sediments from Fort Churchill, the most contami-
nated  site,  however, were  most active  in terms of
demethylation potentials. Inhibition of sulfate reduction
with  molybdate  resulted  in  significantly  depressed
ORDP ratios, but overall demethylation rates were com-
parable  between  inhibited and  uninhibited samples.
Addition  of  sulfate  to  sediment slurries stimulated
production  of  14CO2 from  14C-MeHg,  while  2-bro-
moethane- sulfonic acid blocked production of 14CH4.
These results reveal the importance of sulfate-reducing
and methanogenic bacteria in oxidative demethylation
of MeHg in anoxic environments.


The Dissimilatory Reduction of As(V) to
As(lll) in Anoxic Sediments and as an
Electron Acceptor for Growth of Strain
SES-3

Anoxic sediment slurries amended with millimolar levels
of As(V) achieved a complete reduction of this oxyanion
to As(lll) upon incubation.  As reduction was enhanced
when slurries were provided with the electron donors H2,
lactate, or glucose, although no effect was achieved with
acetate or succinate. Aerobically incubated slurries did
not  reduce As(V),  nor did  formalin-killed or autoclaved
controls. Even though acetate did not stimulate As re-
duction, the oxidation of 2-14C-acetate to 14CO2 in an-
oxic slurries could be coupled with  the abundance of
As(V). The selenium (Se) (VI) respiring anaerobe strain
SES-3 was found to be capable of achieving growth by
carrying out the dissimilatory reduction  of As(V) to
As(lll). Although growth parameters were meager (e.g.,
Ym = 0.53 g cells/mole lactate; maximal cell density =
9.2 x 107 cells/ml), the ability to reduce As(V) to As(lll)
was constitutive and occurred rapidly in either selenate-
or nitrate-grown cells. These  results suggest that the
reduction of As(V) to As(lll) in nature may be achieved
by bacteria-like strain SES-3 carrying out dissimilatory
As(V) reduction.
                                                 86

-------
   Monod Degradation Kinetics of Quinoline in Natural and Microbially Enriched
                                  Methanogenic Microcosms
                      E. Michael Godsy, Ean Warren, and Barbara A. Bekins
                          U.S. Geological Survey,  Menlo Park, California
Introduction

The methanogenic biodegradation of nitrogen-contain-
ing heterocyclic compounds found in wastes from petro-
leum refineries, coke operations, coal tar production,
and wood preservation has not been studied in  detail.
Quinoline, the largest single component in creosote (1),
is first oxidized to  2(1 H)-quinolinone, which is then de-
graded to CH4 and CO2. In this study, the Monod no-
growth kinetic constants for the oxidation of quinoline
and the Monod kinetic constants for the methanogene-
sis of 2(1 H)-quinolinone were determined under natural
and microbially enriched methanogenic conditions using
nonlinear regression analysis (2). In microcosms simu-
lating natural aquifer conditions, it was necessary to
model the oxidation and subsequent methanogenesis
independently (1). In microbially enriched microcosms,
the  two must be  coupled to include the biomass in-
crease from the methanogenesis of 2(1H)-quinolinone
as below:
dQ
 dt ''
       Kn+Q
                     Qn
dQn
    _        _
 dt   Kn+Q   Y(Ks+Qn)
dX
~df~~
     U,max*a
      Ks+Qn
where
   Q =
Umax
   Y =
      quinoline, mg-L"1
      no-growth oxidation constant, day"1
      active biomass, mg-L"1
      one-half saturation no-growth coefficient,
      mg-L"1
      maximum specific growth rate, mg-L"1-day"1
      2(1H)-quinolinone,  mg-L"1
      one-half saturation coefficient, mg-L"1
      growth yield, mg biomass-mg substrate"1
Materials and Methods

The study site is located adjacent to an  abandoned
wood preserving plant within the city limits of Pensacola,
Florida (1). The wood preserving process consisted of
steam pressure treatment of pine poles  with creosote
and/or pentachlorophenol.  For more  than 80 years,
large but unknown quantities of waste waters (consist-
ing of extracted moisture from the poles, cellular debris,
creosote, pentachlorophenol, and diesel fuel from the
treatment processes) were discharged to  nearby sur-
face impoundments. These impoundments were unlined
and in direct hydraulic contact with the underlying sand-
and-gravel aquifer. Contamination of the ground water
resulted from the accretion  of wastes from these im-
poundments. Methanogenesis in the aquifer was simu-
lated using microcosms containing approximately 3 kg
of freshly collected anaerobic aquifer material in a 4-L
glass serum  bottle with 2.5 L of prereduced anaerobi-
cally sterilized mineral salts solution. Approximately
40 mg/L of quinoline was added, simulating a concen-
tration similarto that found in the aquifer (1). The micro-
cosms were  prepared,  incubated, and sampled in  an
anaerobic glove box containing an O2-free atmosphere
maintained at 22°C to 24°C. Microbially enriched micro-
cosms were prepared as above but were batch fed for
three cycles by removing 50 percent of the liquid volume
and replacing that volume with fresh mineral salts con-
taining enough quinoline to bring the final concentration
back to 40 mg/L.  After the last feeding cycle, the liquid
culture was removed from the sediment,  resulting in
liquid-only culture which was batch fed for six more
cycles.

Substrate concentrations were determined at approx-
imately  4-day  intervals by  high-performance  liquid
chromatography.  Total  biomass  concentrations were
determined at approximately 20-day intervals by total
protein using the  Coomassie brilliant blue staining pro-
cedure of Galli (3).
                                                 87

-------
Results and Conclusions

The oxidation of quinoline, a reaction that is endergonic
(4), is modeled using derived Monod no-growth  kinetic
constants.  This oxidation is uncoupled with the  degra-
dation  of 2(1H)-quinolinone in  microcosms simulating
natural  conditions,  as shown  by  the  complete  and
stoichiometric oxidation to 2(1H)-quinolinone before the
onset of methanogenesis. In microbially enriched micro-
cosms,  however, the oxidation of quinoline is linked  to
the degradation of 2(1H)-quinolinone;  the increase  in
biomass from methanogenesis must be included in the
equations describing the  oxidation of quinoline (Figure
1).  The kinetic values derived from the microbially en-
riched,  all-liquid microcosm experiments were not sig-
nificantly different from those values from sand-filled
natural  microcosms (Table  1). The  Monod kinetic  con-
stants for both the oxidation and subsequent methano-
genesis are  representative   of  values  describing
substrate utilization in  an oligotrophic and somewhat
hostile environment (4).

It is still unclear,  however, what number of microbial
populations are involved and to what extent each of the
populations influences  the steps in  the  biodegradation
of quinoline. This  uncertainty can be seen by the  high
concentration of biomass capable  of the oxidation  of
quinoline in  natural microcosms, suggesting  that the
ability to oxidize  quinoline is not  unique to just this
consortium but may be common to many of the individ-
ual  members of the creosote-degrading consortia. The
enrichment procedure has altered the microbial popula-
tion of the  natural microcosms by potentially removing
all of the microorganisms that can oxidize quinoline but
are not directly involved in the methanogenesis of 2(1 H)-
quinolinone.  These results suggest  that as long  as the
culture is derived from the contaminated aquifer, enrich-
Table 1.  Monod Kinetic Constants ±95 Percent Confidence
        Interval for Parameters Determined by Nonlinear
        Regression for Both Natural and Microbially
        Enriched Microcosms
Kinetic Constant
An, day1
/<„, mg-L-1
Umax, day1
KS, mg-L'1
Y, mg-mg-1
Starting biomass, mg-L"1
Oxidation
Starting biomass, mg-L"1
Methanogenesis
Natural
Microcosms
0.31 ±0.06
2.0 ± 1.4
0.09 ± 0.06
11.4±0.6
0.03
17.3
0.003
Microbially
Enriched
Microcosms
0.29 ± 0.02
7.58 ± 4.0
0.14 ±0.07
33.1 ± 11.5
0.07
1.39
1.39
ment does not alter the kinetics of quinoline oxidation
and subsequent methanogenesis of 2(1H)-quinolinone.
The  size  of  the  various  active  microbial populations,
however, must be known before fate-and-transport mod-
eling can be  attempted.


References

1. Godsy, E.M.,  D.F. Goer-lite, and D. GrbiCEGaliCE 1 992. Anaerobic
  biodegradation of creosote contaminants in natural and simulated
  ground water ecosystems. Ground Water 30:232-242.

2. Monod, J. 1 949. The growth of bacterial cultures. Ann. Rev. Micro-
  biol. 3:371 -394

3. Galli, R. 1987. Biodegradation of dichloromettiane in waste water
  using fluidized bed bioreactor. Appl. Microbiol. Biotechnol. 27:206-
  21 3.

4. Godsy E.M.  1993. Mettianogenic biodegradation of creosote-de-
  rived contaminants in natural and simulated ground water ecosys-
  tems. Ph.D. dissertation. Stanford University, Stanford, CA. p. 155.
                       Quinolme
                                    * Quinoline
                                    D 2(lH)-Quiiiolinnne
                                    A Biomass
                                    	Model
                       40        60
                        Time (days)


Figure 1.  Quinoline oxidation and 2(1H)-quinolinone methano-
         genesis  in microbially  enriched  laboratory micro-
         cosms.
                                                     88

-------
           Stimulating the Biotransformation of Polychlorinated Biphenyls
                   John F. Quensen, III, Stephen A. Boyd, and James M. Tiedje
                        Michigan State University, East Lansing, Michigan
Introduction

The  discovery that poly chlorinated  biphenyls (PCBs)
can be reductively dechlorinated by microorganisms un-
der anaerobic conditions has stimulated interest in the
development of a sequential anaerobic/aerobic biotreat-
ment process for their destruction.  While the aerobic
degradation of PCBs  is generally limited to congeners
with four or fewer chlorines, the anaerobic process can
dechlorinate  more highly substituted congeners, pro-
ducing products  that are aerobically degradable. In-
deed, all products from the anaerobic dechlorination of
Aroclor  1254  (1) have been shown to  be aerobically
degradable by one or more strains of aerobic bacteria
(2).  Also, the  high  proportion  of monochlorinated
biphenyls that can accumulate as a result of anaerobic
PCB dechlorination may serve to induce PCB-degrad-
ing enzymes in  aerobic microorganisms (3).  More
highly chlorinated congeners can be aerobically co-
metabolized but are not inducing substrates (4).

A greater  understanding of the factors  controlling the
anaerobic dechlorination of PCBs is necessary before a
successful sequential  anaerobic/aerobic biotreatment
process can be developed for PCBs. In particular, it is
important to determine how to stimulate more rapid and
complete dechlorination in areas where the natural rate
and/or extent  of dechlorination  is limited. The general
approach we have taken is to identify the most probable
site-specific factors  limiting in situ PCB dechlorination,
then to  apply treatments to alleviate the limitation(s).
During the past year of this project, we have focused on
enhancing the dechlorination of PCBs present in River
Raisin (Michigan) and  Silver  Lake (Massachusetts)
sediments.

River Raisin Sediment Experiment

In a previous project, we found that little in situ dechlori-
nation of the  PCBs present in River Raisin sediment
collected near  Monroe, Michigan, had occurred. PCB-
dechlorinating microorganisms were found to exist in the
sediment,  however. The sediment supported dechlori-
nation in laboratory assays when spiked with additional
PCBs and inoculated with  PCB-dechlorinating microor-
ganisms (meaning inhibitory compounds were not pre-
sent), and the PCBs already present in the sediments
were bioavailable because they were dechlorinated un-
der the conditions of our treatability assay. In fact, indi-
vidual  congeners  in the  contaminated  sediment
decreased 30 to 70 percent in 24 weeks at rates nearly
identical to rates for the same congeners freshly spiked
into noncontaminated sediments.

The treatability assays were conducted using air-dried
River Raisin sediments. The sediments were slurried
with  an  equal weight of non-PCB-contaminated sedi-
ments and reduced anaerobic mineral medium (RAMM).
The  slurry was then  inoculated with  microorganisms
eluted from Hudson River sediment to ensure that PCB-
dechlorinating microorganisms were present, and with
2',3,4-trichlorobiphenyl (34-2-CB) in  a small volume of
acetone. The 34-2-CB was added because the addition
of a single PCB congener (or other halogenated aro-
matic compound) can sometimes "prime" the dechlori-
nation of  PCBs  already  present in a  contaminated
sediment (5). Non-PCB-contaminated Red Cedar River
sediments were added to  provide a source of unidenti-
fied nutrients. The RAMM included essential mineral
salts and a chemical reductant (Na2S) to lower the initial
redox potential. We  have  conducted separate experi-
ments with slurries made from both air-dried and always
wet River Raisin sediments to help determine what as-
pect of our treatability assay fosters the dechlorination
of the PCBs  present in these sediments.

Materials and Methods

Air-Dried  River Raisin Sediments

With slurries made from  the air-dried sediments,  the
factors considered were 1) addition of 34-2-CB, 2) addi-
tion of the mineral salts in  RAMM, 3) addition of Na2S,
and 4) addition of the non-PCB-contaminated sediment.
                                                 89

-------
The  River Raisin sediment was added to Balch tubes
(1  g  per tube). An additional 1 g of non-PCB-contami-
nated Red Cedar River sediment was added to the
appropriate treatments. Inocula for each treatment were
prepared  by  eluting PCB-dechlorinating  microorgan-
isms from Hudson River sediments with a medium ap-
propriate to the treatment (i.e., with or without mineral
salts, with or without reductant), and  7 ml of an inocu-
lum was added to each tube using an anaerobic tech-
nique. 34-2-CB in a small volume of acetone was added
to  one treatment, while the rest received the same vol-
ume of acetone. The tubes were sealed with Teflon-lined
rubber stoppers and  aluminum crimps. Autoclaved treat-
ments served as controls. A tube was sacrificed for each
sample. Triplicate samples were taken at 8-week inter-
vals, extracted,  and  analyzed for PCBs using capillary
gas chromatography with electron  capture detection.

Wet  River Raisin Sediments

The same four factors described above plus the neces-
sity of inoculating with Hudson River microorganisms
were considered in an experiment with River Raisin
sediments that  had been  kept  wet since the time of
collection. Portions of the sediment were mixed with the
appropriate medium (i.e., with or without salts, reduc-
tant,  or inoculum)  in  a small Erlenmeyer flask on  a
magnetic stirrer in an  anaerobic chamber. Red Cedar
River sediments, acetone with or without 34-2-CB, and
PCB-dechlorinating  microorganisms  eluted from Hud-
son River sediments were also added  as appropriate to
each treatment. Portions (7 ml) of the slurries were then
dispensed to  Balch  tubes,  and the tubes were  sealed
with Teflon-lined rubber stoppers and aluminum crimps.
The sampling and analytical procedure was the same
as the experiment  with slurries made  from air-dried
sediment.

Results and Discussion

Decreases in the concentrations of certain PCB conge-
ners  in the live samples relative to the autoclaved con-
trols  were used to  compare  the  effectiveness of the
various treatments. These  congeners (245-25-CB and
235-24-CB in chromatographic peak 42, and 34-34-CB
and 236-34-CB in peak 49) were chosen because each
peak represents more  than 2  mole percent of the con-
geners initially present, and  because they could  not
have been formed  in  significant  quantities from  the
dechlorination of other PCBs present.

In  the experiment with slurries made from the air-dried
sediment,  approximately 50 percent  of the congeners
present in  each indicator peak were dechlorinated in the
treatment  receiving 34-2-CB (Figure 1). No dechlorina-
tion was apparent in any of the other treatments.

In  the experiment with slurries made from wet sedi-
ments, no inoculation was required for PCB dechlorina-
                 Decrease of 245-25-CB/235-24-CB
   cd u
  > c
  o o
    '
    .S 0.5
           o—o No amendment
           3—E 34-2-CB
           a—£> Trace sails
           v-^? Sodium Kulfi.de
           o—o Red Cedar sediment
                      8           16

                  Incubalioii Time (Weeks)




                 Decreasn of 236-34-CB/34-34-CU
   > a
   0-0
           o—o No amendment
           E-E 34-2-CB
           z—a Trace salts
           v-^f Sodium sulfide
           •3-^ Red Cedar sediment
Figure 1.
                      8           16

                  Incubation Time (Weeks)
Decrease of indicator PCB congeners due to dechlor-
ination in slurries prepared from air-dried River Rai-
sin sediments.
tion, and some dechlorination occurred in all treatments.
Thus, merely  making a slurry from the River Raisin
sediments appears to have stimulated some dechlorina-
tion of the PCBs present in them. The most extensive
dechlorination, however, occurred in the treatment re-
ceiving 34-2-CB, showing that dechlorination could be
enhanced by "priming." Somewhat surprisingly, the ad-
dition  of Red Cedar River sediments inhibited dechlori-
nation (Figure 2); perhaps the additional organic matter
served as a  sorptive sink for some of the PCBs.

Silver  Lake Sediment Experiments

Although there is evidence that the PCBs present in
Silver Lake sediments have undergone in situ dechlori-
nation, these sediments do not support PCB dechlorina-
tion in laboratory experiments.  These sediments have
high concentrations of several  metals, especially zinc,
copper, lead, and chromium. We suspect that the metals
are present mainly in a reduced state in situ and become
partially oxidized and therefore more toxic to dechlori-
nating microorganisms during  the sediment handling
                                                  90

-------
                 Decrease of a45-25-CB/235-24-CD
      1.0-
 *-. O
 01
 > a
 o o

 I-
 K o
   a
   o
   ^
  a,
  e—e No amendment
  3-3 34-2-CB
  A—a Trace salts
  ^-^ Sodium sulfide
  Q—& Red Cedar sediment
                                   16
                  Incubation Time  (Weeks)
                 Decrease of 236-34-CB/34-34-CB
   G
 tn O
 s-, c
  C
 o o
   D.
   o
   Du
  &-^D No amendment
  B-B 34-2-CB
  &—A Trace salts
  v-^7 Sodium sulfide
  o—e Red Cedar sediment
Figure 2.
          Incubation Time (Weeks)


Decrease of indicator PCB congeners due to dechlor-
ination in slurries prepared from wet River Raisin
sediments.
required to set up dechlorination experiments. We also
have previously shown that high concentrations of zinc
can inhibit PCB dechlorination even after highly reduced
conditions are restored. The experiments reported here
were designed to stimulate dechlorination  by reducing
the bioavailability of toxic metals  through  chelation  or
precipitation in both a  model system and in Silver Lake
sediments.

Materials and Methods

General Procedure

Anaerobic sediment slurries containing PCBs were in-
oculated with  a  PCB-dechlorinating microbial  consor-
tium eluted  from   PCB-contaminated  Hudson  River
sediments. Treatments consisted of the addition of met-
al salts and/or amendments to precipitate or chelate
metals. Autoclaved slurries served as negative or sterile
controls, while untreated slurries served as positive con-
trols. Triplicate samples were sacrificed at 4-week inter-
vals, solvent extracted,  and analyzed for PCBs  using
capillary gas chromatography with electron capture de-
tection. The course of PCB dechlorination was followed
by calculating the average meta plus para chlorines for
each treatment versus incubation time. No dechlorina-
tion from the ortho positions was evident. Dechlorination
patterns were evaluated by assessing changes in spe-
cific congener concentrations overtime.

Model System

Anaerobic  slurries of non-PCB-contaminated Hudson
River sediment were spiked with Aroclor 1242 (500 |j,g/g
sediment) and inoculated  with PCB-dechlorinating mi-
croorganisms eluted from PCB-contaminated Hudson
River sediments. We consistently observed dechlorina-
tion of the Aroclor 1242 in  such preparations. Zinc (Zn)
or Lead (Pb) (as  chloride  salts) was added at solution
concentrations of 500 |j,g/mL to  induce metal toxicity.
Amendments of FeSO4,  ethylene diamine triacetic acid
(EDTA), and citrate were added individually to samples
before incubation to test their effectiveness in alleviating
the toxicity of Zn and Pb.

Silver Lake Sediment Slurries

Anaerobic slurries of Silver Lake sediments were spiked
with 34-2-CB and inoculated with Hudson River micro-
organisms. The 34-2-CB was added so that we  could
monitor the dechlorination  of a freshly added PCB con-
gener  in addition to  the PCBs already present in the
sediment. Experimental treatments consisted of the ad-
dition of FeSO4,  EDTA, and citrate, as in  the  model
system described above.

Results and Discussion

In the model  system, ZnCI  prevented Aroclor  1242
dechlorination while  PbCI decreased the extent  of
dechlorination. EDTA, citrate, and FeSO4 amendments
all reversed the inhibitory effect of PbCI while EDTA and
FeSO4 eliminated the inhibition by ZnCI. In fact, FeSO4-
amended treatments exhibited more  extensive dechlori-
nation than the unamended positive controls (i.e., those
without PbCI or ZnCI additions). Apparently, the FeSO4
greatly stimulated dechlorination from para positions. In
all non-FeSO4-amended slurries  exhibiting dechlorina-
tion, dechlorination  occurred primarily from  the  meta
positions to yield ortho and  para substituted products
(pattern M). But in FeSO4-amended treatments, the ma-
jor products were 2-CB,  2-2-CB,  and 26-CB, indicating
that dechlorination occurred from both meta and para
positions (pattern  C). We have often  noted that the para
dechlorination activity present in Hudson  River sedi-
ments is lost during storage of the sediments. It appears
that addition of FeSO4 somehow "rescues" this dechlori-
nation  activity.
                                                   91

-------
In the Silver Lake sediment slurries, the added 34-2-CB
was dechlorinated in citrate-  and FeSO4-amended slur-
ries,  but not in  EDTA-amended slurries. There was no
indication of further dechlorination of the PCBs already
present in the sediments.

References

1.  Quensen, J.F., III, S.A. Boyd, and J.M. Tiedje. 1990. Dechlorination
   of four commercial polychlorinated biphenyl mixtures (Aroclors) by
   anaerobic microorganisms from sediments. Appl. Environ. Micro-
   biol. 56:2,360-2,369.
2.  Bedard, D.L., R.E. Wagner, M.J. Brennan, M.L. Haberl, and J.F.
   Brown, Jr. 1987. Extensive degradation of Aroclors  and environ-
   mentally transformed polychlorinated biphenyls by Alcalignes eu-
   trophus H850. Appl. Environ. Microbiol. 53:1,094-1,102.
3.  Masse, R., F. Messier, L. Peloquin, C. Ayotte, and M. Sylvestre.
   1984. Microbial biodegradation of 4-chlorobiphenyl, a model com-
   pound of chlorinated biphenyls. Appl. Environ. Microbiol. 41:947-
   951.
4.  Furukawa,  K., F. Matusumura, and K. Tonomura. 1978. Alcali-
   genes and Acinetobacter strains capable of degrading polychlori-
   nated biphenyls. Agric. Biol. Chem. 42:543-548.

5.  Bedard, D.L., H.M. VanDort,  R.J. May, K.A. DeWeerd, J.M. Prin-
   cipe, and L.A. Smullen. 1992. Stimulation of dechlorination of Aro-
   clor 1260 in Woods Pond sediment. In: General Electric Company
   research and development program for the destruction of PCBs,
   11th progress report. Schenectady, NY: General Electric Corpora-
   tion Research and Development, pp. 269-280.
                                                           92

-------
             Bioaugmentation for In Situ Co-metabolic Biodegradation of
                            Trichloroethylene in Ground Water
                           Junko Munakata Marr and Perry L. McCarty
                            Stanford University, Stanford, California

                    V. Grace Matheson, Larry J. Forney, and James M. Tiedje
                       Michigan State University, East Lansing, Michigan

                          Stephen Francesconi and  Malcolm S. Shields
                          University of West Florida, Pensacola, Florida

                                          P.M. Pritchard
                   U.S. Environmental Protection Agency, Gulf Breeze, Florida
Introduction

Trichloroethylene (TCE), a common ground-water con-
taminant, has been found to be fortuitously degraded
(co-metabolized) by organisms grown on a variety of
substrates (1, 2). Addition of such substrates can lead
to  two  significant problems in ground-water aquifers.
First, organisms stimulated by substrate addition may
be unable to degrade TCE. Second, the most promising
compounds for inducing TCE degradation, phenol and
toluene, are  themselves hazardous  substances and
therefore cause regulatory concern. To address the first
issue, aquifers may be bioaugmented with wild-type
strains  known to be effective at TCE degradation. Per-
haps ideally,  both problems can  be overcome through
the use of mutant strains known both to degrade TCE
efficiently and to grow on a nontoxic substrate. Labora-
tory studies were conducted to  investigate these two
alternatives.

Laboratory Studies

Bacterial Cultures

The  wild-type strain  evaluated  was Pseudomonas
cepacia G4 (G4), a strain isolated from a holding pond
at  an industrial waste treatment facility in Pensacola,
Florida  (2). This organism co-metabolizes TCE using
toluene ortho-monooxygenase (TOM), which is induced
by phenol or toluene  (3). The  mutant  used was P.
cepacia G4 PR130i (PR1), a chemically induced mutant
of G4 that constitutively expresses TOM while grown on
substrates such as lactate. A more complete description
of PR1  is presented at this meeting.
So/7 Microcosms

Small-column  microcosms (17 cm3) were constructed
using aquifer material from a test area at Moffett Federal
Air Station. Column fluids were exchanged every 2 to 3
days by pumping 10 ml of solutions held in gas-tight
glass barrel syringes through the column influent port
with a syringe pump. At the start of each fluid exchange
period,  1 ml of bacterial culture was added  to the mi-
crocosms followed by 9 ml of oxygenated ground water
containing  about 200 |j,g/L TCE  and/or  primary sub-
strates. Microcosm effluent samples were  collected dur-
ing each exchange for analysis.


Detection of Bacteria

A deoxyribonucleic acid (DMA) probe specific for both
strains of G4 was constructed using polymerase chain
reaction (PCR) to amplify segments of G4  DMA between
repetitive extragenic palindromic (REP) sequences. The
REP-PCR  reaction can be performed directly on envi-
ronmental samples and therefore  does not require ex-
traction of  DMA before amplification. The method was
tested using ground-water and sediment  samples con-
taining indigenous bacterial populations with and with-
out added G4, with parallel plate counts on R2A agar.
                                               93

-------
Results

Soil Microcosms

The  6.5-mg/L, phenol-fed,  nonbioaugmented  column
followed a pattern similar to that observed in the field (4)
and consumed approximately 60 |j,g/L TCE relative to a
nonfed control. Columns augmented with  induced G4
without a  primary substrate achieved similar levels of
TCE degradation. With the addition of 15 mg/L lactate,
degradation increased to 100 |j,g/LTCE in G4-amended
columns,  but no  such  increase  occurred  in  a  PR1-
amended column. In these lactate-fed  columns, the G4
was pregrown on phenol while the PR1 was pregrown
on lactate. When columns amended  with either G4 or
PR1  were fed 6.5 mg/L  phenol,  130 |j,g/L  TCE  was
degraded. The results are summarized in Figure 1.

Detection of Bacteria

The  REP-PCR  probe was able  to detect 10 colony
forming units (CPUs) of G4 against a background of 10^
nontarget  CPUs contained in 1 |j,L of template  (Figure
2). The probe's sensitivity compares favorably  to other
PCR-based detection methods (5).
                    Log10 CFU nontarget mixture
                             +10CFUG4
Figure 2. Sensitivity of G4 REP-PCR products using the strain
        G4 GF13 probe; agarose gel electrophoresis of REP-
        PCR reactions and hybridization of the Southern blot
        to GF13.

Application of REP-PCR to aqueous effluent samples
from the soil microcosms produced mixed results (Fig-
ure 3). G4 was not detected in the control or phenol-only
microcosms  and gave a strong signal in  phenol- and
lactate-fed microcosms augmented with G4, as antici-
        o
        s
        o
        c
        o
        o
        LU
        P
        a)
        3
        O
        0)

        I
              200 -
8


1

Control-



> 1
1 1
1 1
W 0)
1 1-
^ CO
CD E>
A °
£ i


1
i— « — i
CD
_£

o « & i •§.
— — B ffl °-

fluent
fluent
1
1
•o
a>
a>
0
_co
S

-------
                            G4  123488789 1ft
Figure 3.  Detection of G4 in column effluents, ninth column
         exchange (Day 20); agarose gel electrophoresis of
         REP-PCR reactions of column effluents and  probe
         GF13 hybridization of the Southern blot.

pated. In microcosms containing unfed G4 or phenol-fed
PR1, however,  responses were weak or absent. The
reason for the latter results has not yet been determined.

Conclusion

Bioaugmentation with G4 or PR1 and phenol feed pro-
vides a means for enhancing native activity toward TCE.
Addition of phenol to aquifers could be avoided by sim-
ply adding G4 previously induced forthe TCE-degrading
enzyme.  Lactate enhanced  activity  of preinduced  G4
toward TCE. In PR1-augmented systems, however, lac-
tate  did  not support the same level of activity toward
TCE  as did phenol.  G4 and PR1  were  identified in
constructed samples with a sensitivity of 1  in 104. De-
tection in aqueous column samples gave some unex-
pected results,  the causes  for  which  remain  to  be
elucidated.


References

1.  Wilson, J.T., and  B.H. Wilson. 1985. Biotransformation of trichlo-
   roethylene in soil. Appl. Environ. Microbiol. 49:242-243.

2.  Nelson, M.J.K., S.O. Montgomery, E.J. O'Neill, and P.M. Pritchard.
   1986. Aerobic metabolism of trichloroethylene by a bacterial iso-
   late. Appl. Environ. Microbiol. 55:383-384.

3.  Shields, M.S., S.O. Montgomery, S.M. Cuskey, P.J. Chapman, and
   PH. Pritchard. 1991. Mutants of Pseudomonas cepacia G4 defec-
   tive in catabolism of aromatic compounds and trichloroethylene.
   Appl. Environ. Microbiol. 57:1,935-1,941.

4.  Hopkins, G.D., J.  Munakata, L. Semprini, and PL. McCarty. 1993.
   Trichloroethylene  concentration effects on pilot field-scale in situ
   groundwater bioremediation by phenol-oxidizing microorganisms.
   Environ. Sci. Technol. 27:2,542-2,547.

5.  Thiem,  S.M.,  M.L. Krumme, R.L. Smith, and J.M. Tiedje. 1994.
   Use of molecular techniques to evaluate the survival of a micro-
   organism injected  into  an aquifer.  Appl.  Environ.  Microbiol.
   60:1,059-1,067.
                                                       95

-------
                          Biodegradation of Chlorinated Solvents
                        Larry Wackett, Lisa Newman, and Sergey Selifonov
                           University of Minnesota, St. Paul, Minnesota

                               Peter Chapman and Michael Shelton
                    U.S. Environmental Protection Agency, Gulf Breeze, Florida
General Scope of Research

Research is being conducted on the bacterial metabo-
lism  of chlorinated aliphatic compounds, with a focus
on oxidative mechanisms of biodegradation. Pseudo-
monas cepaciaG4 oxidizes trichloroethylene (TCE) and
related chlorinated alkenes with relatively little  loss of
activity over time  (1), and the molecular basis of this
observation is being elucidated. In vivo experiments are
delineating the substrate range and concentration limits
of P.  cepacia G4 for chlorinated solvents. In vitro experi-
ments are defining the properties of toluene 2-monooxy-
genase, the enzyme catalyzing the oxidation of TCE.

Purification and Properties of Toluene
2-Monooxygenase

Toluene 2-monooxygenase  activity was monitored  in
vitro  via a sensitive radiometric assay using [14C]-tolu-
ene (2). Chromatography of cell-free  extracts revealed
that this was a three-component oxygenase system. All
three components have now been purified to homoge-
neity. In vitro  reconstitution  of the three proteins and
reduced  nicotinamide adenine  dinucleotide (NADH)
yielded an active enzyme system that oxidizes toluene
to ort/70-cresol  and this,  subsequently, to 3-methyl-
catechol. One component is a flavoprotein  containing
a 2Fe2S cluster that accepts  electrons  from  NADH
(Table  1). A  second component is  a  low molecular
weight protein that stimulates activity but has no obvious
redox-active  functional  group (Table 1).  The  largest
component has an a2p2Y2 subunit  structure (Table 1).
This  component is implicated as the hydroxylase com-
ponent as it alone will oxidize toluene in the presence of
dithionite + methyl viologen + O2 or hydrogen peroxide.
The  hydroxylase component contains four to six iron
atoms per holoenzyme. Spectroscopically, this compo-
nent resembles the soluble  methane monooxygenase
hydroxylase  component from Methylosinus trichospor-
iumOBZb (3).


Table 1.  Molecular Properties of Purified Components

                                Small
Property             Hydroxylase  Component  Reductase
Subunit structure
Subunit molecular
masses (kDa)
Molecular mass (kDa)
Gel filtration
Native PAGE
SDS-PAGE
Calculated (aa
quantitation)
Metal content
Iron content (mol/mol)
Inorganic S" content
(mol/mol)
FAD (mol/mol)
Pi
Absorption maxima


Specific activity
(units/mg)
Percent recovery
(apy)2
5.4, 37.7, 13.5


216
190
211
210


5.3
ND

ND
4.5
282 nm


1.7a

40
Monomer
10.4


19.3
—
10.5
10.4


ND
ND

ND
4.3
277 nm


79.4a

27
Monomer
40.0


45.8
—
41.8
40.0


2.3
2.9

1.2
5.8
270, 341 ,
423, 457
nm
512.0b

30
 	unit is defined as 1 nmol [  C]-toluene/min at 23°C.
b One unit is defined as 1 |imol cytochrome c reduced/min at 23°C.
ND = not detected.
— = not determined.
PAGE = polyacrylamide gel electrophoresis
SDS = sodium dodecylsulfate
FAD = flavin adenine dinucleotide
pi = isoelectric point
                                                 96

-------
In Vivo Studies With  P. Cepacia G4

P. cepacia G4 was shown to grow on  aromatic ring
compounds other than toluene and phenol. P. cepacia
also oxidized non-growth-supporting aromatic and  ali-
phatic substrates. Examples of the aromatic substrates
that were investigated in some detail include naphtha-
lene and indene. The oxidation of TCE by P. cepacia G4
has been studied in detail. The major oxidation product
is glyoxylic acid.  The effects of TCE on P.  cepacia G4
also were studied to determine how resistant the organ-
ism is to variable concentrations of TCE. Unlike Pseudo-
monas  putida F1, P.  cepacia G4  was  not detectably
toxified  by  low concentration of TCE or by metabolites
generated by oxidative mechanisms. High TCE concen-
trations, however, exerted a  solvent effect that  could
markedly depress cell  division rates  and even  cause
cell death.


References

1. Folsom, R.R., P.J. Chapman, and P.M. Pritchard. 1990. Phenol and
  trichloroethylene degradation by Pseudomonas cepacia G4: Kinet-
  ics and interactions between substrates. Appl. Environ. Microbiol.
  56:1,279-1,285.

2. Yeh, W.-K., D.T. Gibson, and T.N. Liu. 1977. Toluene dioxygenase:
  A  multicomponent  enzyme system.  Biochem. Biophys.  Res.
  Comm. 78:401-410.

3. Fox, B.C., W.A. Froland, J.E.  Dege, and J.D. Lipscomb. 1989.
  Methane monooxygenase from  Methylosinus trichosporium OB3b:
  Purification and properties of a three-component system with high
  specific activity from  a  type  II methanotroph.  J. Biol. Chem.
  264:10,023-10,033.
                                                     97

-------
      Biological and Nutritional Factors Affecting Reductive Dechlorination of
                               Chlorinated Organic Chemicals
                                             Dingyi Ye
                           National Research Council, Athens, Georgia

                                          W. Jack Jones
                     U.S. Environmental Protection Agency, Athens, Georgia
Introduction

Halogenated organic chemicals are of major public con-
cern because these  compounds are usually toxic and
persistent in the environment, and they tend to accumu-
late  in soils, sediments,  and biota.  Polychlorinated
biphenyls (PCBs) and organochlorine pesticides are en-
vironmental pollutants of great concern due to a history
of heavy use, toxicities, persistence in the environment,
and wide distribution in environmental media.

PCBs are a mixture of chlorinated biphenyls consisting
of 209 possible  congeners. These  compounds  were
widely used for almost 50 years, with several hundred
million pounds having been released into the environ-
ment. An organochlorine  pesticide of concern  is the
insecticide toxaphene, a complex mixture of chlorinated
camphenes.  Toxaphene consists of more than 177 de-
rivatives and was heavily used in  the United States
before 1982. Estimates indicate that about 233,688 met-
ric tons of toxaphene was manufactured in the United
States from  1964 to 1982. Toxaphene, like PCBs and
other organochlorines contaminants,  is a global pollut-
ant.  Many studies have shown toxaphene to be rela-
tively persistent and bioaccumulated  by biota (1).

The  objectives of the present  research were to study
factors affecting anaerobic transformation of PCBs and
organochlorine pesticides (e.g., toxaphene) and to de-
velop techniques to enhance their in situ bioremediation.
The  preliminary goals were 1) to characterize the an-
aerobic microbial  dechlorination of PCBs in Sheboygan
River, Wisconsin, sediment, 2) to examine the effects
of nutrients, Fe°, and electron carriers on dechlorination
of PCBs, and 3) to examine the anaerobic biotransfor-
mation of toxaphene using   indigenous  and  PCB-
dechlorinating microorganisms.
Toxaphene Biotransformation

Materials and Methods

A Hudson River (HR) pasteurized  enrichment culture
capable of reductive dechlorination  of PCBs was used
for initial toxaphene experiments. The enrichment cul-
ture was originally pasteurized at 85°C for 15 min and
subsequently transferred at monthly intervals (1 percent
v/v transfer, repasteurized at 90°C  for 10 min at each
transfer). The inoculum was  serially diluted, and the
highest dilution (10"6) retaining PCB-dechlorination ac-
tivity was inoculated to the medium without PCBs; this
culture  was used  as the  inoculum  for  PCB  and
toxaphene biotransformation studies. Two milliliters of
the  PCB-dechlorinating   inoculum  was anaerobically
transferred to 28-mL culture tubes containing 2 ml re-
vised  anaerobic mineral  medium (RAMM) and 1 g of
sterile, uncontaminated   (toxaphene-free)  pond  sedi-
ment. Toxaphene was subsequently added at a final
concentration of 500 u,g/g dry sediment. Control sedi-
ment samples were autoclaved three times (121°C for
1 hr each time) on consecutive days, with  addition of
toxaphene occurring on Day 4. All  cultures were incu-
bated at 25°C in the dark.

Results and Discussion

Anaerobic transformation  of toxaphene by the pasteur-
ized  HR enrichment was evident as indicated  by
changes  in the gas chromatography  (GC) isomer-
distribution  patterns. GC  chromatograms of the  auto-
claved control and  a sample inoculated with the HR
pasteurized inoculum after 7 months of incubation are
presented in Figure 1. Several peaks representative of
toxaphene isomers are numbered to facilitate compari-
son of the  control and experimental chromatograms.
Only a very minor change was observed for a late
                                                98

-------
Figure 1.  GC profile of toxaphene in A) autoclaved and B) live
         experimental microcosms (inoculated with enrich-
         ment from Hudson River) after 28 weeks of anaerobic
         incubation. Numbered peaks are for reference only.
eluting peak (#39) in the control and live experimental
samples; thus,  this peak was chosen as an internal
reference peak to which  other peaks were normalized
(based on peak height). In comparison with the sterile
control, appreciable changes in many peaks were noted
in  live samples after 7 months of incubation. The  per-
centage changes  in specific peaks for the live sample
compared with the sterile control are presented in Figure
2.  Increases in  some peaks were observed,  indicating
accumulation of some dechlorination  products.  Some
early eluting peaks also decreased,  however, suggest-
ing that in addition to reductive dechlorination, anaero-
bic degradation  may  have  occurred. The possible
anaerobic degradation of toxaphene will be further in-
vestigated by identifying polar transformation  products.

Changes in toxaphene isomer-distribution patterns were
also observed following GC analyses of autoclaved con-
trols (data not shown). As mentioned previously, sterile
controls were  prepared  by  autoclaving sediments at
121°C for 1 hr on  3 consecutive days. Thus, it is highly
unlikely that the observed transformations in the sterile
controls  were  biologically   mediated.  The  observed
changes are likely due to  either  abiotic (chemical)
o
0
CO
1
8
(D
Changes (tt
^
IQU
100 -
50 -
-50-
-100 -
c

''I
)

I


• 1 1
.
10
J Increasing Peaks
l.ll , ,
1 ' "ll'i'i I'jT |l
Decreasing Peaks
20 30 4
                                                                            Peak Number

                                                      Figure 2. Percentage change (relative to control sample) of se-
                                                              lect toxaphene peaks in live experimental microcosm
                                                              after 28 weeks of anaerobic incubation.
transformation or enhanced sorption of toxaphene to the
autoclaved soil matrix.  It has been previously reported
that specific toxaphene isomers were  transformed  in
sterile sediments and in a sand-Fe(ll)/Fe(lll) system (2).
Phototransformation  of toxaphene  has  also  been
reported  (2). On the  other hand, the high KOC (soil or-
ganic carbon  partition  coefficient) value reported for
toxaphene suggests that the chemical mixture should be
strongly sorbed to soil particulates (1); any differences
in KQC among the isomers may influence the toxaphene
isomer distribution patterns in long-term sediment incu-
bations. Thus,  the changes noted in the isomer distribu-
tion pattern (after correcting for abiotic transformations
observed  in  sterile  controls)  are  likely  isomers  of
toxaphene that were subject to transformation  by the
inoculated microorganisms and that were relatively re-
sistant to abiotic transformation. In these inoculated ex-
perimental cultures, CH4 production was not observed,
and a mete-directed  dechlorination  of amended PCB
congeners (Aroclor 1242) was confirmed  in separate
experiments. These results suggest  that the HR  pas-
teurized enrichment culture was  capable of  anaerobi-
cally transforming toxaphene.  The HR  pasteurized
enrichment is  easily maintained  and cultivated  and,
therefore, may be of potential use in the remediation  of
toxaphene-contaminated soils. Additional  studies are
under way to evaluate the effectiveness of this enrich-
ment culture for remediation of historically contaminated
soils.

PCB Biotransformation

Materials and Methods

PCB biotransformation  experiments were  performed
with PCB-contaminated (approximately 500 ppm) She-
boygan River (SR) sediment. For abiotic transformation
experiments, PCB-contaminated sediment was slurried
with anoxic site water inside an  anaerobic glove  box,
                                                   99

-------
homogenized, then amended with Fe°.  Of the  slurry
containing 1 g sediment (dry weight), 1.8 ml was trans-
ferred to replicate 28-mL serum tubes. Half of the tubes
were spiked with 300 u,g Aroclor 1242 as an available
PCB source.  Fe°, pyrite, and degassed, sterile distilled
water were then added. Control samples consisted of
autoclaved sediment slurries as described above. All
cultures were incubated at 25°C in the dark.

An experiment to assess the  effect of pasteurization on
microorganisms  eluted from SR  sediment was con-
ducted as  described by Ye  et  al. (3). Finally,  eluted
microorganisms from historically contaminated  (PCB)
SR sediment  were subjected  to pasteurization (85°C for
20 min) and used as inocula to assess their potential for
reductive  dechlorination  of  amended  Aroclor  1242,
1248, and 1254. Aroclors were added individually at a
final concentration of 500 u,g/g dry sediment.

Results and Discussion
Fe -Amended Experiments

Several studies document the Fe°-mediated reductive
dechlorination oftrichloroethylene (TCE) and other chlo-
rinated compounds (4). Our preliminary results of Fe°-
amended SR sediment slurries, however, indicated that
no dechlorination  of PCBs occurred after anaerobic in-
cubation for 2 weeks at 20°C in  either live or sterile
samples. Further, no evidence of reductive dechlorina-
tion of PCBs was observed in the SR sediment slurries
spiked with Aroclor 1242, indicating that bioavailability
of PCB congeners  was not  a limiting factor for the
Fe°-mediated dechlorination.

A prolonged  incubation time is usually necessary to
achieve biologically mediated reductive dechlorination
of PCBs. Thus, it was not surprising to find no evidence
of dechlorination in the live experimental samples, es-
pecially because the SR sediment had  been stored at
2°C to 4°C for approximately 1 yr before use. It is  likely
that  a significant  amount of time  is necessary for the
dechlorinating population to recover to a level to affect
significant dechlorination. Additional experimental  re-
sults  with SR sediment (without nutrient amendment)
indicated that approximately 4 weeks of incubation was
required before  detectable  PCB dechlorination was
observed.

Investigators  at  the  U.S. Environmental  Protection
Agency Athens  Research Laboratory  have recently
demonstrated Fe°-mediated reductive dechlorination of
other halogenated compounds. We have, however, no
evidence of PCB dechlorination in Fe°-amended  sam-
ples under similar experimental conditions. These re-
sults suggest that PCBs are more resistant to chemical
                                         A
                                         B
     -

      uliJL
Figure 3. GC profiles of Aroclor 1254 in A) pasteurized, B) auto-
        claved, and C) live experimental microcosms (inocu-
        lated with SR-eluted microorganisms) after 12 weeks
        of incubation.
(abiotic)  dechlorination than  other  chlorinated  com-
pounds examined to date.

Pasteurization of SR Sediment

CH4 production was not observed in experimental mi-
crocosms inoculated with the  pasteurized microorgan-
isms  from  SR  sediment; nonpasteurized  cultures,
however, were actively methanogenic. These observa-
tions are consistent with previous pasteurization experi-
ments  using  HR sediments  as  inocula  (3).  The
pasteurized cultures preferentially removed meta chlo-
rines, while the untreated cultures removed both meta
and para chlorines from selective PCB congeners. In the
present study, dechlorination of Aroclor 1254 was ob-
served by the untreated inocula after 6 weeks of incu-
bation; however, no significant dechlorination of Aroclor
1254 was evident in experiments inoculated with the
pasteurized inocula after 12 weeks of incubation (Figure
3). These results suggest that anaerobic, spore-forming
microorganisms  in the SR sediment exhibit  similar
dechlorinating pathways as the microorganisms  in the
                                                  100

-------
HR sediment. Stimulation Of in Situ  PCB dechlorination     2. Williams, R.R., and IF. Bidleman. 1978. Toxaphene degradation
may be possible through the addition of a suitable spore       in estuarine sediments. J. Agric. Food chem. 26:280-282.
germinant or growth Substrate.                               3. Ye, D., J.F. Quensen,  III, J.M. Tiedje, and S.A. Boyd.  1992. An-
                                                                  aerobic dechlorination  of polychlorolbiphenyls (Aroclor 1242) by
                                                                  pasteurized and ethanol-treated microorganisms from sediments.
RGfGTGnCGS                                                  Appl. Environ. Microbiol. 58:1,110-1,114.
                                                                4. Matheson, L.J., and P.G. Tratnyek. 1994. Reductive dechlorination
1.  Saleh, M.A.  1991. Toxaphene: Chemistry, biochemistry, toxicity       of chlorinated  methanes by  iron metal. Environ.  Sci. Technol.
   and environmental fate. Rev. Environ. Contam. Toxicol.  118:1-85.       28:2,045-2,053.
                                                            101

-------
     Predicting Heavy Metal Inhibition of the In Situ Reductive Dechlorination of
                     Organics at the Petro Processor's Superfund Site
                                           John H. Pardue
         Department of Civil and Environmental Engineering, Louisiana State University,
                                       Baton Rouge, Louisiana
Introduction
Transition metals and synthetic organic compounds are
common co-contaminants at waste sites that are candi-
dates for biological treatment. The inhibition of microbial
decomposition of natural organic matter by certain tran-
sition metals has been widely documented (1); however,
the inhibition of anaerobic degradation processes (e.g.,
reductive dechlorination) is poorly understood.
Inhibition Characteristics
Inhibition characteristics of a model heavy metal (cad-
mium,  Cd) on  a model  chlorinated  aromatic  (2,3,4-
trichloroaniline,  2,3,4-TCA)  was  determined  in the
laboratory.  Laboratory  microcosm  experiments were
conducted in three anaerobic flooded soils with varying
properties.  Dechlorination of 2,3,4-TCA to  monochlo-
roanilines occurred when total pore-water Cd concen-
trations were  below  a critical threshold level.  Inhibition
occurred across a continuum of Cd concentrations  in
several soils,  but a completely inhibited threshold con-
centration was readily identified (Figure 1). Dechlorina-
tion kinetics and metabolites differed with soluble metal
concentration. Speciation of soluble Cd was necessary
to predict whether inhibition would occur, particularly in
the  presence  of high concentrations of organic ligands
such as humic acids (Table 1). Estimation of metal pools
using selective  extractions and  measurement of acid-
volatile sulfide (AVS) provided additional information but
did  not adequately  predict whether  inhibition  of de-
chlorination would occur. These results demonstrated
the   importance of  quantification   and speciation  of
pore-water metals in predicting potential inhibition  of
anaerobic biodegradation reactions such as reductive
dechlorination.
                                                      a
                                                      o £?•
 .H  e
 s  s
 ? "3
 •S  §
                                                      =s fe.
1.4
1.2
1.0
0.8
0.6 -
0.4 -
0.2
00 -
o
o
0 0 ° 0
BLH •
0 Marsh o
Rice 0
•• o
*
o
	 , 	 CuDO 	 , — •-• 	 C>-T
                                                               1        10       100     1000     10000

                                                                   Soluble Cd concentration ((.ig/L)
Figure 1. Normalized 2,3,4-TCA dechlorination rates (k/kcontroi)
        versus soluble Cd concentrations  in  three flooded
        soils:  bottomland hardwood (BLH), rice paddy, and
        freshwater marsh.
Table 1.  MINTEQA2 Results From Pore Water of
        Representative Rice and Marsh Soil Suspensions
        (estimated pore-water humic acid concentrations
        were 1 mg/L in the rice soil and 55 mg/L in the
        marsh soil)
Soil
RS



MS

Equilibrated
Total Mass 2,3,4-TCA
Soluble Distribution, Dechlorination
Cd (mg/L) Cd Species % Inhibited
0.195 Cd+2
CdCI+
CdSO4 (aq)
Cd-Humate
0.350 Cd+2
Cd-Humate
43.7 +
5.9
1.6
48.5
1.0
98.7
                                                 102

-------
Site History

The Petro Processor's, Inc., site is a high-priority Super-
fund site near Baton Rouge, Louisiana. The site served
as a chemical waste pit from the early 1960s to the late
1970s. An estimated 60,000 tons of chlorinated organic
waste,  primarily  hexachlorobutadiene  and hexachlo-
robenzene (HCB),  was deposited  in several unlined,
diked pits. A spill event resulted  in contamination  of
stream sediments in an adjacent bottomland hardwood
wetland. Heavy metal contamination is contiguous with
chlorinated organic contamination  (primarily  HCB)  in
these sediments. These sediments are the site of a
bioremediation field trial directed at enhancing reductive
dechlorination of HCB (2).

Predicting Heavy Metal Inhibition

Sampling is being conducted to determine if metal inhi-
bition of reductive dechlorination can be predicted in the
field  at the Petro Processor's  site. Characteristics  of
inhibition of model  compounds  (described above) are
being used to develop a strategy for predicting inhibition.
Parallel laboratory studies are being used to confirm that
these same inhibition characteristics would be observed
for HCB in the sediments. Laboratory studies using
2,3,4-TCA  indicated that noninhibited soils could be
adequately predicted using the AVS/SEM (simultane-
ously extracted metal)  concept. This concept has been
used to predict the toxicity of metals to benthic organ-
isms (3). In studies with 2,3,4-TCA, soils in which molar
metal concentrations exceeded molar AVS were  not
always inhibited, requiring further metal speciation and
prediction of "free," uncomplexed metal concentrations
using MINTEQA2. Spatial and seasonal  information of
AVS and SEM and observations of lower chlorinated
benzene samples  are being collected in the field. Sec-
tions of the bayou where molar SEM exceeds molar AVS
are undergoing further metal speciation studies.

References

1. Duxbury, T. 1985. Ecological aspects of heavy metal responses in
  microorganisms. Adv. Microbiol. Ecol. 8:185-235.
2. Constant, W.D., J.H. Pardue,  R.D.  DeLaune, K. Blanchard, and
  G.A. Breitenbeck. 1995. Enhancement of  in situ  microbial degra-
  dation of chlorinated organic waste at the Petro  Processor's Su-
  perfund site. Env. Progress 14:51-60.
3. Di Toro, D.M.,  J.D. Mahony, D.J. Hansen,  K.J. Scott, M.B. Hicks,
  S.M. Mays, and  M.S. Redmond. 1990. Toxicity of cadmium in
  sediments: The role of  acid  volatile  sulfides.  Environ. Toxicol.
  Chem. 9:1,489-1,504.
                                                    103

-------
          Effect of Primary Substrate on the Reduction of 2,4-Dinitrotoluene
                               Jiayang Cheng and Makram T. Suidan
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio

                                         Albert D. Venosa
Risk Reduction Engineering Laboratory, U.S. Environmental Protection Agency, Cincinnati, Ohio
Introduction

2,4-Dinitrotoluene (DNT) is one of the priority pollutants
(1) commonly found in munitions wastes. It is recalci-
trant to  biological treatment in aerobic processes  (2),
such as the activated sludge system, but can be  de-
graded (3) in a sequential anaerobic/aerobic biosystem.
2,4-DNT is completely transformed to 2,4-diaminotolu-
ene (DAT) with ethanol as the primary substrate in an
anaerobic reactor. Subsequently, 2,4-DAT is readily min-
eralized (3) in an aerobic reactor. 2,4-DNT can not be
transformed in  the anaerobic reactor (4) without a  pri-
mary substrate. In this study, the anaerobic biotransfor-
mation of 2,4-DNT with ethanol, methanol, acetic acid,
or hydrogen as primary substrate was investigated. The
effect of the primary substrate on the reductive transfor-
mation of 2,4-DNT was also studied.

2,4-DNT-transforming anaerobic cultures were  accli-
mated with 2,4-DNT and ethanol, methanol, or acetic
acid as the feed organic substrates in three chemostats.
The concentrations of 2,4-DNT and  the primary sub-
strates in the feed to the three chemostats are  listed in
Table 1. The chemical oxygen demand (COD) loading
for all three chemostats was the same. Minerals and
nutrients were added to the chemostat feed to support
bacteria growth. Na2S • 9H2O (50 mg/L) was added to
maintain a reducing environment in the chemostats. The
pH and  the temperature in the chemostats were  main-
tained constant at 7.2 and 35°C, respectively. The  hy-
draulic retention time in the  chemostats was 40 days.

Table 1.  Concentrations of Substrates in Feed for the
        Chemostats
Chemostat
                      Ethanol    Methanol    Acetic
                       Fed       Fed     Acid Fed
2,4-DNT, mg/L
Primary substrate, mg/L
91.7
500
91.7
696
91.7
978
2,4-DNT was completely biotransformed to 2,4-DAT in
all three chemostats. All the primary substrates (ethanol,
methanol, and acetic acid) were converted to methane
and carbon dioxide.

After steady-state operation was achieved in the chemo-
stats, the mixed cultures from the chemostats were used
as the inocula for the batch tests to determine the kinet-
ics of anaerobic biotransformation of 2,4-DNT with dif-
ferent  primary substrates.  The  cultures were then
transferred  into the batch reactors  in  an oxygen-free
anaerobic chamber at 35°C. The pH and the tempera-
ture in the batch reactors were maintained the same as
those in the chemostats.  Different initial concentrations
of 2,4-DNT were used in the batch tests. To determine
the co-metabolic mechanism of the biotransformation of
2,4-DNT and the primary substrate, hydrogen was also
used as the primary substrate in the  batch tests.

Results and Discussion

All the batch tests were run in duplicate, and the devia-
tions of the results were less than 8 percent. The kinetics
of anaerobic biotransformation of 2,4-DNT with ethanol,
methanol, and acetic acid as the primary substrates are
illustrated  in  Figure 1  (a), (b), and (c), respectively.
2,4-DNT was completely biotransformed to 2,4-DAT via
4-amino-2-nitrotoluene  (4-A-2-NT) or 2-amino-4-nit.ro-
toluene (2-A-4-NT) under anaerobic conditions, regard-
less  of the  primary  substrate.  The  rate  of  the
biotransformation  of 2,4-DNT and the intermediates
(4-A-2-NT and 2-A-4-NT), however, was much higher in
the  presence of ethanol  than  that in the presence of
either methanol or acetic acid.  When ethanol was used
as the primary substrate,  hydrogen  was  produced
during  the  acetogenesis of ethanol.  The  hydrogen
then served as the electron donor for the reduction of
2,4-DNT to 2,4-DAT. The  bacteria also used ethanol for
their growth. When methanol or acetic acid was used as
                                                104

-------
          0.10
          0.10
          0.00
                                                                       2,4-DNT
                                                                       4-A-2-NT
                                                                       2-A-4-NT
                                                                       2,4-DAT
                                                                                  1400
                        200
400
600
800
1000
1200
1400
                                           4-A-2-NT
                                           2-A-4-NT
                                           2,4-DAT
                0       200      400      600      800
                                             Time, Hrs
                            1000
                            1200
                            1400
Figure 1. Anaerobic biotransformation of 2,4-DNT with (a) ethanol, (b) methanol, or (c) acetic acid as the primary substrate.
                                          105

-------
the primary substrate, the substrates were used in the
biosystem  to support the growth of the bacteria that
transformed 2,4-DNT to 2,4-DAT. Neither methanol nor
acetic acid was degraded until 2,4-DNT, 4-A-2-NT, and
2-A-4-NT were completely transformed to 2,4-DAT. The
hydrogen for the reductive  transformation of 2,4-DNT
and the intermediates was probably from bacterial en-
dogenous decay.  2,4-DNT was not biotransformed with-
out a  primary substrate (4) in both chemostat and batch
reactors. 2,4-DNT itself cannot support the growth of the
bacteria, and  the primary substrate is  necessary for
maintaining the biological  activities to transform 2,4-
DNT  to 2,4-DAT.  The rate of the biotransformation of
2,4-DNT was very low in the initial stage of the process,
indicating that 2,4-DNT inhibited its own  biotransforma-
tion. The presence of 2,4-DNT and its intermediates also
exhibited inhibition to the bioconversion of the primary
substrate (ethanol, methanol, or acetic acid). The higher
the initial concentration of 2,4-DNT, the longer was this
period of inhibition  to the  conversion  of the primary
substrate. Ethanol, methanol, and acetic acid were rap-
idly converted  by the bacteria after 2,4-DNT  and  its
biotransformation intermediates were completely trans-
formed to 2,4-DAT.

To prove that hydrogen was the electron donor for the
reductive biotransformation of 2,4-DNT, the same batch
test was conducted with hydrogen as the primary sub-
strate. The results are shown in Figure 2. In the control
reactors without  2,4-DNT, hydrogen with CO2 in  the
reactors was immediately converted to methane (Figure
2a). When 2,4-DNT was initially present in the reactors,
hydrogen was first consumed for the biotransformation
of 2,4-DNT to 2,4-DAT, and for supporting the growth of
the bacteria. Methane was  produced from the excess
hydrogen after 2,4-DNT, 4-A-2-NT, and 2-A-4-NT were
completely transformed to 2,4-DAT.  This phenomenon
indicates  that 2,4-DNT,  4-A-2-NT, and 2-A-4-NT also
inhibited the hydrogen-utilizing methanogenesis. The
higher the initial  concentration of 2,4-DNT,  the more
hydrogen was consumed for 2,4-DNT biotransformation
and the less hydrogen was  left for methane production
(Figure 2b-f).

References

1. Keither, L.H., and  W.A. Telliard. 1979. Priority pollutants I. A per-
  spective view. Environ. Sci. Technol. 13:416-423.
2. McCormick, N.G.,  J.H. Cornell, and A.M. Kaplan. 1978. Identifica-
  tion of biotransformation products from 2,4-dinitrotoluene. Appl.
  Environ. Microbiol. 35:945-948.
3. Berchtold, S.R., S.L. VanderLoop, M.T. Suidan, and S.W. Maloney.
  1995. Treatment of 2,4-dinitrotoluene using a two-stage system:
  Fluidized bed anaerobic GAC reactors  and  aerobic activated
  sludge reactors. Water Environ. Res. In press.
4. Cheng, J., Y. Kanjo, M.T. Suidan, and A.D. Venosa. 1995. Anaero-
  bic biotransformation of 2,4-dinitrotoluene with ethanol as primary
  substrate: mutual  effect of the substrates on their biotransforma-
  tion. Submitted for publication in Wat. Res.
                                                    106

-------
  o
  S
  o

 *O
  O

  P,
                                        2.0   0.02
20   40  60   80  100  120

        M
                                       2.0
                                     -  1,5
                                     -  1.0
                                     - 0.5
                                       0.0
  G
 •2  0.10
20   40  60   80  100  120


         (e)
     0.00
                                             0.01
                                0.00
0   20   40  60   80  100  120
                                0.20
                          0.0   0.00
                                                                                     O
                                                                                     CD
                                                                                     O
                                                                                     O
                                                                                     s-c
                                                                                     CX
                                                                                     CD
                                                                   0.0
0   20   40  60   80  100  120



              (f)
                                                                                     cd
                                    C/3
                                    d
                                    o
                                    o
OJO
o
                               0.0
         0   20   40  60   80  100  120           0    20  40   60  80   100 120

                                       Time, Hrs


       —•— 2,4-DNT   —— 4-A-2-NT    —•—  2-A-4-NT   —•— 2,4-DAT

                —o— Hydrogen Consumed    —°— Methane Produced



Figure 2. Anaerobic biotransformation of 2,4-DNT with hb as the primary substrate.
                                           107

-------
  Surfactants in Sediment Slurries: Partitioning Behavior and Effects on Apparent
                          Polychlorinated Biphenyl Solubilization
                    Jae-Woo Park, John F. Quensen, III, and Stephen A. Boyd
                        Michigan State University, East Lansing,  Michigan
It is generally believed that the biodegradation of poorly
water soluble compounds in soil or sediment systems is
limited by low bioavailability due to strong sorption of the
compounds to natural organic matter. The use of surfac-
tants to  increase the apparent water solubility of such
contaminants has often been suggested as a way  of
increasing their bioavailability to degrading microorgan-
isms. A possible limitation of this approach is that solu-
bility enhancement is much greater above the critical
micelle concentration (CMC) of the surfactant than be-
low it, and these supra-CMCs are often toxic or inhibitory
to bacteria. A few reports, however, indicate that sub-
CMC concentrations of surfactants may enhance the
anaerobic  dechlorination of aromatic compounds. The
goals of our present research efforts are to determine if
sub-CMCs of surfactants  can  enhance  the microbial
dechlorination of polychlorinated biphenyls (PCBs) and,
if so, by what mechanism(s). We have determined the
partitioning behavior of several surfactants in soil and
sediment slurries and their effects on PCB solubilization.
These experiments were undertaken to determine if an
increase in the apparent aqueous solubility of PCBs by
sub-CMCs of these surfactants is a plausible mechanism
for any observed enhancement of PCB dechlorination.

The sorption of four commercial nonionic surfactants
(Triton X-100, Triton X-405, Triton X-705, and Tween 80)
onto the Red Cedar River sediment used in  our PCB
dechlorination  assays  was  evaluated.  Sorption iso-
therms were plotted, and  Freundlich isotherms of the
form Cs=KCen were fitted to  the experimental data
where Cs is the sorbed  concentration of the surfactant
(mg/kg), Ce is the aqueous concentration of the surfac-
tant (mg/L), and K and  n are constants. K and n values
ranged from 1.193x 10'4to 1.009 x 10'3 and from  0.232
to 0.696, respectively. The Red Cedar River sediment
thus shows orders of magnitude less surfactant sorption
than has been reported for soils, as shown by the low K
value.
The distribution coefficients of three PCB congeners at
sub- and supra-CMC surfactant concentrations (up to
four times the CMC) were determined using [14C]labeled
PCBs.  The  aqueous-phase PCB  concentrations in-
creased at all surfactant  concentrations tested com-
pared  with  the  sediment-water   system  without
surfactants. Notably, this included an increase in the
aqueous-phase  concentrations of PCBs even  at the
lowest  surfactant concentration tested (0.05 times
CMC),  especially for the inherently less soluble hexa-
and tetra-CBs by Tween 80. In fact, Tween 80 increased
the solubility of 2,2',4,4',5,5'-CB by a factor of 3.3 at
25 percent of its CMC, and by a factor of 6.3 at 75
percent of its CMC.
The low sorption of the surfactants by Red Cedar River
sediments has important consequences for PCB solubi-
lization. Surfactant monomers sorbed to soils or sedi-
ments will increase the total organic matter content of
the solids and act as an additional sorptive phase. Con-
sequently, if surfactants strongly sorb to the sediments,
they may actually decrease the aqueous phase concen-
tration of nonionic compounds such as PCBs. When the
mass of sorbed  surfactant is small, however, as in the
case of the Red Cedar River sediment, most of the
surfactant mass exists in the water, and the PCB solu-
bilization  effect of aqueous phase surfactant micelles
and monomers dominates the sorptive capabilities of
sediment-associated surfactant and native organic mat-
ter. Therefore, the aqueous phase concentration of PCB
increases even when relatively small amounts of surfac-
tants are added to the system. While these solubility
enhancements are small  relative to those that occur
above  the CMCs of these surfactants, the increased
solubility may be enough  to significantly increase the
rate of PCB dechlorination, especially for the more chlo-
rinated and less water soluble congeners.
                                                108

-------
                   Partial Characterization of an Anaerobic, Aryl, and
                           Alkyl Dehalogenating Microorganism
                                          Xiaoming Zhang
                           National Research Council, Athens, Georgia

                                W. Jack Jones and John E. Rogers
                     U.S. Environmental Protection Agency, Athens, Georgia
Introduction

To better understand controls and pathways of anaero-
bic biotransformation  of organic  pollutants in contami-
nated environments, pure culture studies are beneficial.
To date, only a few strains of anaerobic dehalogenating
microorganisms have been isolated and characterized.
Among these, Desulfomonile tiedjei\s probably the most
widely studied (1).  In this study, we  report the partial
characterization of an anaerobic bacterium capable of
both  aryl and alkyl reductive dehalogenation.

Results and Discussion

An anaerobic bacterium, designated as strain XZ-1, was
isolated from  freshwater pond sediment near Athens,
Georgia. Isolate XZ-1  is a sporeforming, motile rod ca-
pable of reductive dehalogenation  of  chlorophenols.
Electron acceptors, including sulfite, thiosulfate, and ni-
trate  (but not sulfate), stimulated growth in the presence
of yeast extract and  pyruvate. None of the following
supported growth  or dehalogenation of chlorophenol
(CP): glucose, fructose, galactose,  rhamnose,  cello-
biose, xylan, ribose, citrate, fumarate,  acetate, peptone,
tryptone, casein hydrolysate, and casamino acids. The
addition  of 1  mM carbon  dioxide  reduced the  lag
time  before growth. No growth was  observed in  the
presence of  4 percent air  or  higher. Growth  was
completely inhibited  by  pentachlorophenol  (PCP)
(>32u,M),  2,3,4,5-tetraCP  (> 8 u,M),  3,4,5-triCP
(>16uM), 3,5-diCP(>120uM),  2,4-diCP (> 500 uM),
and 2-CP (>4,OOOu,M). The generation time of isolate
XZ-1 was 1.8 hr at pH 7.5 (optimal) and 30°C.

Isolate XZ-1  removed orffto-chlorines from  all ortho-
chlorine-containing phenols tested (e.g., 2-CP and pen-
tachlorophenol). Hydrogen, formate, ethanol, pyruvate,
and yeast extract served as electron donors for dehalo-
genation  of CPs. Only  pyruvate  and yeast extract,
however, stimulated growth either in  the absence or
presence of electron acceptors, including 3-chloro-4-hy-
droxyphenylacetate (an analog  of ortho-CP). The  aryl
dehalogenation  activity was  inducible, and  induction
was  inhibited by addition of chloramphenicol to  cell
suspensions.  Experiments  with D2O demonstrated
that  water  was the exclusive proton  source for  aryl
dehalogenation of chlorophenols. Proton nuclear mag-
netic resonance (NMR) studies indicated that hydrogen
was incorporated at the same position  where an ortho-
chlorine was removed. Product  solvent isotope effects
were 5.4 and 8.5 for dechlorination of 2,3-diCP  and
2-CP, respectively. An increase in the assay temperature
reduced the product solvent isotope effect in 2,3-diCP
dechlorinations.

Cell suspensions of isolate XZ-1 also were capable of
reduction  of 2,4,6-trinitrotoluene (TNT, 46 ppm) to
2,4,6-triaminotoluene via 2-amino-4,6-dinitrotoluene,
4-amino-2,6-dinitrotoluene, 2,6-diamino-4-nitrotoluene,
2,4-diamino-6-nitrotoluene, and several  unidentified
intermediates. The TNT transformation pattern was dif-
ferent in  aryl dehalogenation-induced cells  and  non-
induced cells.  The identified  intermediates of  TNT
reduction accumulated to lower levels in the induced
cells than in the noninduced cells. Addition of pyruvate
stimulated TNT transformation.  Heat-treated cell  sus-
pensions exhibited only traces  of TNT transformation
activity either with or without addition of pyruvate. Cell
suspensions of isolate XZ-1 also metabolized chloram-
phenicol in the presence of pyruvate. No intermediate(s)
of chloramphenicol transformation has been  identified
to date.

Both noninduced and aryl dehalogenation-induced cell
suspensions of isolate  XZ-1 dechlorinated  tetrachlo-
roethene to trichloroethene (TCE). A comparison of aryl
                                                 109

-------
and alkyl dehalogenation  rates in  noninduced and  in-   and to characterize further the physiology, nutrition, and
duced  cells suggests that at least two enzymes are   phylogeny of this anaerobe.
responsible for the two activities. Aryl dehalogenation-
induced cells also slowly  dechlorinated TCE to cis-1,   Reference
 - IC  oroe  ene.                                      1. Mohn, W.W, J.M. Tiedje. 1992. Microbiol. Rev. 56:482.
Additional studies are underway to identify the range of
transformation activities of dehalogenating isolate XZ-1
                                                   110

-------
      Microbial Degradation of Petroleum Hydrocarbons in Unsaturated Soils:
     The Mechanistic Importance of Water Potential and the Exopolymer Matrix
                                        Patricia A. Holden
    Department of Environmental Science, Policy, and Management, University of California,
                                        Berkeley, California

                                          James R. Hunt
          Department of Civil Engineering, University of California, Berkeley, California

                                        Mary K. Firestone
    Department of Environmental Science, Policy, and Management, University of California,
                                        Berkeley, California
Background

Total soil water potential, ₯, is the potential energy per
unit volume of unsaturated soil water and is commonly
reported in -MPa (1). Matric water potential, *₯m, is the
largest component of *P in most soils and arises from the
interaction of soil water with soil surfaces. Matric poten-
tial  determines water film thickness in soil, and thus
controls gas-phase mass transferthrough soil pores and
solution-phase mass transfer through water films. Bac-
teria in biofilms are in equilibrium with water in their
environment, and adaptation to a given soil water poten-
tial  or to a  changing water potential  condition during
wetting and drying will  affect intrinsic bacterial physiol-
ogy and biofilm characteristics (2). Because of its role in
both mass  transfer and bacterial reaction rates, soil
water potential is an important environmental factor con-
trolling  petroleum  biodegradation rates in unsaturated
soils.


Project Framework

Our biodegradation model for oil constituents at the
biofilm scale contains the following parameters that will
vary as a function of *P:
  qm = intrinsic molar removal rate per area of
       biofilm (moles/#-t)
   Ks = intrinsic half saturation constant (moles/L3)
   Lb = total biofilm thickness (L)
   De = effective diffusivity of contaminant through
       biofilm (L2/t)
   pm = number density of bacteria per mass of
       biofilm (#/m)
   pb = density of biofilm or mass of biofilm per
       biofilm volume (m/L3)
  Csat = aqueous solubility of petroleum hydrocarbon
       (moles/L3)

Experimental  protocols include  determining each pa-
rameter as a function of *P and determining the overall
removal rate in unsaturated soil as a function of *P. We
are also examining  how physicochemical properties of
the bacterial matrix are altered with *P to effect hydro-
carbon solubility and spreadability.

Biofilm Reactors and  Preliminary Results

Phenanthrene, hexadecane, and methyl-decalin are the
selected test substrates representing the three major
classes of petroleum constituents. Pristane is the con-
servative  tracer. Polyethylene glycol, a nonpermeating
solute with a molecular weight of 8,000 (PEG 8,000), is
used to set  matric  water potential in well-mixed  and
biofilm culture systems.

Custom-designed   biofilm  reactors  for  developing
biofilms under unsaturated  conditions have been con-
structed and are being tested using various growth sub-
strates.  Transmission electron  micrographs  taken
through biofilms grown under ^-controlled conditions
reveal architectural  changes,  specifically  cell packing
and morphological, with *P. Preliminary diffusion studies
suggest that diffusional mass  transfer through biofilms
is related to the ^-condition during growth.
                                                111

-------
References                                                  2- Harris, P.P. 1981. Effect of water potential on microbial growth and
                                                                     activity. In: Parr, J.F., W.R. Gardner,  and L.F. Elliott,  eds. Water
                                                                     potential relations in soil  microbiology, SSSA special  publication,
1.  Jury, W.A., W.R. Gardner, and WH. Gardner. 1991. Soil physics,       VoL 11> Number 6, Madison, Wl: Soil Science Society  of America.
   5th ed. New York: John Wiley & Sons.                                PP- 733-740-
                                                              112

-------
              Metabolic Indicators of Anaerobic In Situ Bioremediation of
                               Gasoline-Contaminated Aquifers
                     Harry R. Seller, Martin Reinhard, and Alfred M. Spormann
             Department of Civil Engineering, Stanford University, Stanford, California
Bioremediation is one of a limited number of options for
restoring aquifers contaminated with the hazardous aro-
matic hydrocarbons that  occur in unleaded gasoline,
such as benzene, toluene, ethylbenzene, and the xyle-
nes (BTEX). Considering the cost and technical difficulty
associated with introducing oxygen into some aquifers,
in situ bioremediation using indigenous, anaerobic bac-
teria merits serious consideration for some contami-
nated sites. A major impediment to the acceptance of in
situ bioremediation is the difficulty of demonstrating that
decreases in the concentrations of BTEX in  ground
watertruly represent bacterial metabolism of these com-
pounds rather than abiotic processes such as sorption,
dilution, or volatilization.

Work in our laboratory has included the characterization
of byproducts of alkylbenzene metabolism by pure and
mixed  anaerobic cultures (1, 2).  This research, which
has focused on sulfate-reducing cultures, has involved
the  extensive use of gas chromatography/mass spec-
trometry for metabolite characterization. We have re-
cently integrated such laboratory findings with field data
from a controlled-release  experiment conducted at the
Seal Beach Naval Weapons Station in California.

Based on the concordance of laboratory studies of an-
aerobic bacteria and field observations from the aquifer
in Seal Beach, we propose a group of compounds in-
cluding benzylsuccinic  acid, benzylfumaric acid (or  a
closely related isomer), and the  o-, m-,  and  p-methyl
homologs of these compounds as biogeochemical indi-
cators of in situ anaerobic alkylbenzene metabolism in
gasoline-contaminated  aquifers.  Under the controlled
conditions of the field study, a strong correspondence
was observed between  the disappearance of  alkylben-
zenes from ground water overtime and the appearance
of associated  metabolic byproducts. This correspon-
dence was both qualitative (i.e., only products specific
to the metabolism of toluene, o-xylene, and  m-xylene
were observed, and only these three hydrocarbons were
depleted)  and quantitative (i.e.,  metabolic byproduct
concentrations tended  to  increase as the associated
alkylbenzene concentrations decreased).
References

1. Seller, H.R., M. Reinhard, and D. GrbiCE-GaliCE. 1992. Metabolic
  byproducts of anaerobic toluene degradation by sulfate-reducing
  enrichment cultures. Appl. Environ. Microbiol. 58:3,192-3,195.

2. Seller, H.R. 1995. Anaerobic metabolism of toluene and other
  aromatic compounds by sulfate-reducing soil bacteria. Ph.D. dis-
  sertation.  Stanford University, Stanford, CA.
                                                  113

-------
            Contaminant Dissolution and Biodegradation in Soils Containing
                                   Nonaqueous-Phase Organics
    Larry E. Erickson, L.T. Fan, J. Patrick McDonald, George X. Yang, and Satish  K. Santharam
                             Kansas State University, Manhattan, Kansas
Several models have been developed to describe the
dissolution,  adsorption  to  soil,  and biodegradation of
nonaqueous-phase contaminants (i.e., hydrocarbons) in
the subsurface, and remediation times have been esti-
mated for various conditions (1-6). The significant rate-
limiting factors determining the required bioremediation
time appear to be the rates of transport of the electron
acceptor or oxygen and organic contaminants in pores
and  soil  aggregates in  the vicinity of the hydrocarbon
phase. The contaminants' solubilities  in  the aqueous
phase affect their dissolution and transport.  Contami-
nant dissolution and transport are more rapid  than oxy-
gen  transport for more  water-soluble compounds such
as toluene, and less rapid for less soluble compounds
such as  pyrene. As long as the nonaqueous phase is
present,  the  higher the solubility  of a compound, the
greater the extent of removal by pump-and-treat opera-
tions rather than by oxygen-limited biodegradation. The
sizes of aggregates and hydrocarbon blobs significantly
affect  remediation time, which  has  been found to  be
proportional to the square of the characteristic length of
the blob.
The  available experimental data for pyrene and anthra-
cene fit well with the results of simulation obtained with
one  of the models.  Besides  dissolution,  adsorption,
desorption, and biodegradation, this model takes into
account the hydrocarbon-phase size distribution; more-
over, it expresses the rate of biodegradation by Monod
kinetics.


References

1. Yang, X., I.E. Erickson, and L.T. Fan. 1993. Transport properties
  of toluene as a  non-aqueous phase liquid in ground water. In:
  Proceedings of the 8th  Conference on  Hazardous Waste Re-
  search. Manhattan, Kansas: Kansas State University, pp. 313-330.

2. McDonald, J.P., C.A. Baldwin, I.E. Erickson, and L.T. Fan. 1993.
  Modeling bioremediation of soil aggregates with residual NAPL
  saturation. In: Proceedings of the 8th Conference on Hazardous
  Waste Research. Manhattan, Kansas: Kansas State University, pp.
  346-365.

3. Gandhi, P., L.E. Erickson, and L.T. Fan. 1995. A simple method to
  study the effectiveness of bioremediation aided, pump-and-treat
  technology for aquifers contaminated by non-aqueous phase liq-
  uids, I. Single component systems. J. Haz. Mat. 39:49-68.

4. Gandhi, P., L.E. Erickson, and L.T. Fan. 1994. A simple method to
  study the effectiveness of bioremediation aided, pump-and-treat
  technology for aquifers contaminated by non-aqueous phase liq-
  uids, II. Multi-component systems. J. Haz. Mat. In press.

5. Yang, X., L.E. Erickson, and L.T. Fan. 1994. Astudy of dissolution
  rate-limited  bioremediation of soils contaminated by residual hy-
  drocarbons. J. Haz. Mat. In press.

6. Santharam, S.K., L.E. Erickson, and  L.T. Fan. 1994. Modeling the
  fate of polynuclear aromatic hydrocarbons in the rhizosphere. In:
  Proceedings of the 9th Annual Conference on Hazardous Waste
  Remediation. Manhattan,  Kansas: Kansas State University, pp.
  333-350.
                                                    114

-------
            Protein Expression in Mycobacteria That Metabolize Polycyclic
                                    Aromatic Hydrocarbons
                                       David E. Wennerstrom
                University of Arkansas for Medical Sciences, Little Rock, Arkansas

                                          Carl E. Cerniglia
                 National Center for Toxicological Research, Jefferson, Arkansas
Three species of mycobacteria have been isolated from
petroleum  contaminated  soil   (Mycobacterium  sp.
PYR-1) (1) or coal gassification  sites (Mycobacterium
sp.  PAH135 and M. gilvum)  (2,  3). These organisms
have potential application in the bioremediation of poly-
cyclic aromatic hydrocarbons (PAHs) because each  can
mineralize various PAHs, including  naphthalene, phen-
anthrene, pyrene, and fluoranthene. The present study
was initiated to  investigate the molecular basis for the
degradation of PAHs by these species of mycobacteria.
To determine part of the  physiological response of the
organisms to the presence of a  metabolizable  PAH in
the  environment, we have analyzed the expression of
proteins by each organism in response to pyrene using
two-dimensional sodium dodecylsulfate-polyacrylamide
gel  electrophoresis (SDS-PAGE). For each organism,
the  pattern of separated proteins was distinct, and pro-
teins increased  in expression following  addition of the
PAH. Major proteins increased in  induced cells of Myco-
bacterium sp. PYR-1 had approximate masses  of 105,
79,  53, 42, and  15 kDa.  In comparison, three proteins
were  induced in Mycobacterium sp. PAH135 (95,  70,
and 53 kDa) and in M. gilvum (72, 27, and 15 kDa). To
determine whether increased expression of these pro-
teins is associated with metabolism of pyrene in Myco-
bacterium sp. PYR-1, uninduced cells were incubated
with the PAH for varying  periods up to 8 hr, and  the
amounts of pyrene metabolism and protein expression
were quantified  by high-performance liquid chromatog-
raphy (HPLC) analysis and densitometry  of proteins
detected  in two-dimensional SDS-PAGE gels, respec-
tively. After  a delay  of  about  1 hr, uninduced cells
metabolized all of the pyrene within 8 hr. The 79 kDa
protein, undetectable in uninduced cells, was expressed
at 1.2 percent of proteins within 2 hr and was fully
expressed at about 2 percent of total protein within 4 hr.
Partial characterization of this protein by N-terminal se-
quencing  and hybridization  of  a  synthetic oligonu-
cleotide   probe   corresponding  to  the  amino  acid
sequence to Bamlll-digested Mycobacterium sp. PYR-1
deoxyribonucleic acid (DMA)  show that this protein is
similar to the kaIG gene product  (catalese-peroxidase)
expressed in many other mycobacteria. Kinetics of in-
creased expression of the 15 kDa protein followed those
for the 79 kDa protein. In contrast,  the 42 kDa  protein
was not increased until 6 hr and was not fully expressed
even at  8 hr after addition  of pyrene. A variant of the
organism was isolated that failed to metabolize  pyrene
and fluoranthene added to soft agar overlays or in liquid
cultures.  The variant retained the ability to metabolize
naphthalene and phenanthrene.  None of the proteins
studied was induced in this organism after exposure to
3 u,g/ml_  pyrene  for 24 hr. These results  indicate that
additional components are required for metabolism of
pyrene and fluoranthene compared with those for meta-
bolism of naphthalene and  phenanthrene in Mycobac-
terium sp. PYR-1. Our results suggest that the proteins
studied are associated with  metabolism of pyrene in
induced  cells of  this organism. These results provide
fundamental information about the  proteins expressed
by these mycobacteria during  PAH degradation. Clearly,
this information will be important for future application of
these mycobacterial strains as inoculants in the biore-
mediation of PAH-contaminated sites.
                                                 115

-------
References                                                   3-  Boldrin, B., A. Tiehm, and C. Fritzsche. 1993. Degradation of phen-
                                                                      anthrene, fluorene, fluoranthene, and pyrene by a Mycobacterium
1.  Heitkamp, M.A., W.  Franklin, and C.E. Cerniglia. 1988. Microbial        sp. Appl. Environ. Microbiol. 59:1,927-1,930.
   metabolism of polycyclic aromatic hydrocarbons:  Isolation  and
   characterization of pyrene-degrading bacterium. Appl. Environ. Mi-
   crobiol. 54:2,549-2,555.

2.  Grosser,  R., D. Warshawsky, and J.R. Vestal. 1991. Indigenous
   and enhanced mineralization of pyrene, benzo(a)pyrene, and car-
   bazole in soils. Appl. Environ. Microbiol. 57:3,462-3,469.
                                                              116

-------
                                    Section 6
                Hazardous Substance  Research  Centers
The Hazardous Substance Research Centers (HSRCs)  conduct research on bioremediation
under the direction of ORD's  National Center for Extramural Research and Quality Assurance.
Research is sponsored by the following centers: the Northeast Hazardous Substance Research
Center (Regions 1 and 2), the Great Lakes and Mid-Atlantic Hazardous Substance Research Center
(Regions 3  and 5), the South/Southwest Hazardous Substance Research Center (Regions 4 and
6), the Great Plains and Rocky Mountain Hazardous Research Center (Regions 7 and 8), and the
Western Region Hazardous Substance Research Center (Regions 9 and 10).

The symposium's poster session included presentations on the co-metabolic biodegradation kinet-
ics of trichloroethylene in unsaturated soils; the effect of water potential on biodegradation kinetics
and population dynamics; developments in anaerobic and aerobic bioventing; developments in the
treatability protocol for co-metabolic bioventing; the environmental safety of commercial  oil spill
bioremediation agents; the effectiveness of gas-phase bioremediation stimulating agents for un-
saturated zone in situ bioremediation; protein expression of mycobacteria that metabolize polycyclic
aromatic hydrocarbons; a field evaluation of pneumatic fracturing enhanced bioremediation; and
the solids suspension characteristics related to slurry biotreatment performance.
                                         117

-------
  Co-metabolic Biodegradation Kinetics of Trichloroethylene in Unsaturated Soils
                                Karen L. Skubal and Peter Adriaens
                            University of Michigan, Ann Arbor, Michigan
The ability of methanotrophicand heterotrophic bacteria
to aerobically transform chlorinated solvents is well es-
tablished.  Methane  monooxygenase (MMO) and aryl
monooxygenase enzymes, produced by these microor-
ganisms  respectively during growth on primary sub-
strates, catalyze the cooxidation and dehalogenation of
chlorinated ethenes including trichloroethylene  (TCE)
and  vinyl  chloride.  Bioventing may prove useful for
stimulating co-metabolism and achieving in situ  reme-
diation of vadose zone soils contaminated with chlorin-
ated alkenes. This possibility motivated an investigation
of co-metabolic dechlorination by indigenous microbial
populations in soils collected from Wurtsmith Air Force
Base (AFB) in Oscoda,  Michigan.
Contaminated aquifer and vadose regions at Wurtsmith
AFB contain perchloroethylene, TCE (up to 1 ,OOOu,g/L),
trans-dichloroethylene, vinyl  chloride,  dichlorobenzenes,
and benzene, toluene, ethylbenzene, and xylene (BTEX)
compounds.  High methane concentrations have also
been detected in soil gas at the site, indicating poten-
tially favorable conditions for methanotrophic bacteria.
Sandy soils from several depths are being characterized
and  studied  in  aerobic batch microcosm  systems at
room temperature to discern the relative importance of
methanotrophic and heterotrophic organisms, and to
optimize methods for their stimulation. Methanotrophs
are supplied  with  oxygen  and methane,  while  het-
erotrophs are supported on toluene as the  primary in-
ducing substrate. A range of environmentally relevant
concentrations is studied, and following an acclimation
period TCE is added  at  approximately one-tenth the
level of primary substrate. The effect of soil moisture on
biodegradation kinetics is examined by comparing mi-
crocosms containing soil maintained at the  local water
content of 4 percent to microcosms containing saturated
soil.  In addition, substrate degradation by soil-derived
cultures is monitored in liquid medium without soil.

Bacterial growth on methane and toluene has  been
stimulated,  and ongoing  work will  evaluate optimum
primary to co-metabolic substrate ratios and elucidate
the effect of moisture content on TCE co-metabolism in
soil systems. Through development of a simple method-
ology for screening soils and microbial populations in-
digenous to a particular site, this study may clarify the
potential of bioventing to enhance chlorinated solvent
transformation in unsaturated zones containing mixed
wastes.
                                                 119

-------
             The Effect of Water Potential on Biodegradation Kinetics and
                                      Population Dynamics
                                  Astrid Millers and Peter Adriaens
           Department of Civil and Environmental Engineering, University of Michigan,
                                        Ann Arbor, Michigan
Although bioventing is  currently being applied in the
field, much remains to be learned about the underlying
parameters controlling biological degradation kinetics in
these systems. These parameters  need to be system-
atically studied  to improve modeling  and design  of
bioventing applications.  In this investigation, the impact
of subsurface moisture content on biokinetic parameters
is studied,  and  the  applicability of biological  kinetics
obtained in saturated batch systems to the unsaturated
zone is evaluated. Specific emphasis is placed on study-
ing the effects of water potential on oxygen  availability,
microbial metabolism, and growth.
Mixed culture studies with indigenous microorganisms
derived from the unsaturated zone at the Wurtsmith field
site, Oscoda, Michigan, have been performed in batch
systems. No degradation of toluene was detected at a
field moisture of about 3 percent (by weight) even after
a month of incubation.  Moisture contents between 12
and 16 percent moisture exhibited the fastest degrada-
tion of toluene.  Differences in  biodegradation  kinetics
observed as a function of moisture content and  inde-
pendent of population shifts are  being verified using
pure cultures of a toluene-degrading  microorganism,
isolated from the same unsaturated soil  samples.
Water potential, the thermodynamic variable expressing
water activity and therefore water availability for the
microorganisms, is used as the experimental variable
rather than the gravimetric moisture content. Varying
water contents of the soil as a result of drying due to
airflow in bioventing operations  influence the different
components of the water potential in the soil matrix. The
osmotic and  matric water potential components are
studied separately in their effect on bacterial growth,
energy production, and degradation  kinetics. Bacteria
isolated from an unsaturated zone below Wurtsmith Air
Force Base are grown in  liquid  culture on  toluate at
different concentrations of membrane diffusable solutes
(NaCI) and nondiffusable solutes (polyethyleneglycol,
PEG).  PEG is  used to simulate the effect of  matric
potential independent of the effect of mass transfer limi-
tations resulting from moisture changes in porous me-
dia. Salt additions to the liquid medium resulted in higher
growth rates of toluate degraders up to 0.2 M NaCI and
increased  CO2 evolution.  The amount of adenosine
triphosphate produced appeared to be  independent of
the salt addition.  Studies will  be extended  to assess
growth in homogenous solids of defined pore structure,
and mixed  population studies will be performed to as-
sess the separate effect of population shifts. The results
from these  studies will serve as  a model for water po-
tential induced microbial stress in unsaturated soil hori-
zons during bioventing.
                                                 120

-------
                       Anaerobic-Aerobic Bioventing Development
                                         Gregory D. Sayles
                         National Risk Management Research Laboratory,
                      U.S. Environmental  Protection Agency, Cincinnati, Ohio

                                Munish Gupta and Makram T. Suidan
  Department of Civil and Environmental Engineering, University of Cincinnati, Cincinnati, Ohio
Motivation

Surface spills and  leaking pipelines and underground
storage tanks can result in vadose zone soils contami-
nated with hazardous chemicals. The vadose zone must
be cleaned to minimize contamination of ground water
and emissions of volatile  organic chemicals to the at-
mosphere.

Aerobic bioventing is now  a common approach to treat-
ing aerobically biodegradable contaminants (e.g., fuels)
in the vadose zone. Many highly chlorinated aromatics
and aliphatics can be destroyed microbiologically, most
rapidly  by sequential anaerobic-aerobic treatment. Usu-
ally, the biochemical pathway providing the highest rate
for the  initial steps of microbial destruction of the highly
chlorinated organic  is  anaerobic reductive dechlorina-
tion.  Once partially dechlorinated, the resulting  com-
pounds typically degrade faster under aerobic, oxidizing
conditions. For example, perchloroethylene, wood treat-
ing wastes containing  pentachlorophenol (PCP), poly-
chlorinated biphenyls  (PCBs), and  many  chlorinated
pesticides can be at least partially dechlorinated by micro-
bial consortia under anaerobic conditions. The resulting
dechlorinated product often can  be destroyed microbi-
ologically under aerobic conditions.

For treatment of large volumes of soil at depth, the most
cost-effective approach is  likely to be an in situ process
that takes advantage of the above well-known anaerobic
and aerobic biological activities. Our approach to this
challenge is to develop "anaerobic/aerobic bioventing."

Approach

Anaerobic-aerobic bioventing requires development of
a new  process:  anaerobic bioventing.  Conditions must
be established  in the contaminated vadose  zone to
induce  anaerobic microbial biodegradation. Anaerobic
conditions may be established by injecting nitrogen into
the vadose zone to displace all oxygen. Avolatile cosub-
strate in the nitrogen stream may be needed to induce
the soil microorganisms to consume all the oxygen and
to establish a low oxidation-reduction  potential (ORP).
The low ORP should induce dechlorination of the con-
taminant. The cosubstrates must  be volatile and con-
sumable  by anaerobic  microorganisms (e.g., ethanol,
acetone,  hydrogen).

Once the contaminant is fully or mostly dechlorinated,
aerobic bioventing is initiated by  injecting air into the
vadose zone. The aerobic microorganisms should then
complete the mineralization of the contaminant.

To move the technology to the field, the following ques-
tions  are  currently being  addressed with pilot-scale
tests:

• Can venting with nitrogen  at a  low rate establish
  adequate anaerobic conditions  for dechlorination in
  unsaturated soil? What is the cost to supply nitrogen
  at the required flow rate?

• Can a  cosubstrate  introduced  with  the  nitrogen
  stream  promote the  dechlorination of target com-
  pounds such as PCP in unsaturated soil?

• What are the most effective volatile cosubstrates, bio-
  logically and  economically,  to promote anaerobic
  dechlorination?  (Several  will be evaluated.)

• Will switching to an aerobic environment (by replac-
  ing nitrogen and the primary substrate with air in the
  injection  stream) promote  mineralization  of  the
  dechlorination byproducts?

• What classes of compounds are amenable to anaero-
  bic/aerobic bioventing? (Several will be evaluated.)
                                                 121

-------
• Will the addition of hydrogen gas (H2) to the nitrogen
  stream aid  in the  development of reducing environ-
  ments and  promote dechlorination?

Initial  experiments were conducted beginning  in June
1995. The tests were conducted in  pilot-scale  soil col-
umns. Two 4-ft long, 4-in. diameter glass columns were
built to simulate an in situ horizontal column of unsatu-
rated  soil. Gas sampling ports  were placed  along the
length of the column. The column was filled with sand
inoculated with secondary effluent to add biomass and
moisture. Essential inorganic nutrients were also added.
Nitrogen  gas passed through an oxygen scrubber was
used to flush the columns of oxygen. At field scale, the
nitrogen  could be supplied  by tank  or, perhaps more
economically, by separating it from atmospheric air us-
ing  onsite  molecular  sieves.  The  first  cosubstrates
evaluated were ethanol and acetone.
                                                 122

-------
              Development of Co-metabolic Bioventing:  Laboratory Tests
                                         Gregory D. Sayles
                         National Risk Management Research Laboratory,
                      U.S. Environmental Protection Agency, Cincinnati, Ohio

                                Alan D. Zaffiro and Jennifer S. Platt
                                  IT Corporation, Cincinnati, Ohio
Background

The objective of this project is to accumulate the infor-
mation necessary to write a protocol for determining
site-specific treatability of co-metabolic bioventing. "Co-
metabolic bioventing" is a bioventing-like process that
delivers air and a volatile co-metabolite to a vadose
zone  contaminated  with  chlorinated  solvent  (e.g.,
trichloroethylene) to induce in situ biodegradation of the
contaminant.

The process attempts to induce the following biologically
catalyzed (unbalanced) reactions:

           Cosubtrate + O2 => CO2 + H2O

           TCE + 02 + H20 => C02 + HCI

where the cosubstrate is chosen because it stimulates
the productions of enzymes in the microbial culture that
oxidize trichloroethylene  (TCE). Known volatile cosub-
strates include methane, propane,  butane,  toluene,  jet
fuel, gasoline, and isoprene. The co-metabolic biovent-
ing system  delivers  the cosubstrate  and oxygen from
gas injection wells to induce the in situ  biodegradation
of the contaminant TCE.

The  Remediation  Technology Development  Forum
Workgroup  on In-Situ Bioremediation  of  Chlorinated
Solvents, a  research team with members from industry
and government, plans to conduct field demonstrations
of three  in situ  bioremediation  technologies for  the
cleanup of chlorinated solvents at two sites. Laboratory
studies using site-contaminated soils are supporting the
field research. The three technologies are 1) co-metabolic
bioventing of  TCE  and dichloroethylene contaminated
vadose zone  soils, 2) accelerated  anaerobic biotreat-
ment  of chlorinated ethylene contaminated ground
water, and 3) intrinsic bioremediation of chlorinated eth-
ylene contaminated aquifers.  Dover Air Force Base
(AFB), in Delaware was selected as the first site.

One task under the co-metabolic bioventing project was
established to  develop a protocol to test site-specific
treatability of co-metabolic bioventing. The  U.S. Envi-
ronmental Protection Agency's  National Risk Manage-
ment  Research   Laboratory has assumed  primary
responsibility for this task. Information from laboratory
testing using soil from Dover AFB and two to four other
sites will be used to generate the  recommended proto-
col. The  efficacy of the  protocol  will be evaluated  by
comparing treatability test results with results of at least
two field demonstrations.

Approach

To simulate co-metabolic bioventing in a closed reactor,
a biodegradation test using soil from Dover AFB was
conducted in the following manner:

• Reactors were closed bottles containing 100 g of soil.

• Soil was unsaturated and not mixed.

• Cosubstrate  (toluene or propane) was added  to the
  bottle.

• No  moisture,  nutrients,  or  microorganisms were
  added.

• Killed controls were  established.

• Two sets of bottles were established:
  - Reactors that were sacrificed in triplicate at various
    times during the test.
  - Reactors that were monitored by automated respi-
    rometry.
                                                 123

-------
• Contaminant and cosubstrate loss with  time  were
  monitored with the sacrificial reactors.

• Oxygen use and carbon dioxide production with time
  were monitored by the respirometer.

The biodegradation test was conducted during June and
July  1995.  Earlier  tests conducted  with  clean  soils
spiked with  TCE and cosubstrates indicated that care
must be taken in the reactor design to minimize abiotic
losses of TCE.  Bottles that  expose the soil  and its
atmosphere only to glass and  Teflon showed significant
(greater than  40 percent) loss of TCE in killed controls
within a week. This  loss was  attributed  to sorption into
the Teflon gaskets. We have  redesigned the bottles to
maximize glass surfaces and virtually eliminate Teflon
surface area.  Preliminary tests showed that these reac-
tors minimized abiotic TCE loss very well.
                                                  124

-------
          Evaluating the Environmental Safety of Using Commercial Oil Spill
                                    Bioremediation Agents
                                      Jeffrey L. Kavanaugh
      Center for Environmental Diagnostics and Bioremediation, University of West Florida,
                                        Pensacola, Florida

                              C. Richard Cripe and Carol B. Daniels
       Gulf Ecology Division, U.S. Environmental Protection Agency, Gulf Breeze, Florida

                                         Rochelle Araujo
     Ecosystems Research Division, U.S. Environmental Protection Agency, Athens, Georgia

                                           Joe E. Lepo
      Center for Environmental Diagnostics and Bioremediation, University of West Florida,
                                        Pensacola, Florida
The use of commercial bioremediation agents (CBAs)
for reducing the ecological impact of oil spills  raises
several risk assessment questions.  The presence of
petroleum  hydrocarbons may contribute some toxicity;
CBAs, with their associated chemical constituents (e.g.,
nutrients, dispersants, enzymes),  microbes,  and inert
ingredients, may add to this toxicity either directly or
indirectly through decreases in dissolved oxygen or in-
creases in particulates. In addition, interaction of CBAs
with oil may have other environmental  effects,  either
through increasing the amount of petroleum  hydrocar-
bons available to aquatic organisms (i.e., through  bio-
surfactant activity) or by generation of toxic metabolites.
A related issue is whether use  of a CBA could reduce
the toxicity of the oil (an  efficacy issue).
A tiered approach, with increasing complexity, cost,  and
effort, has been proposed to address the environmental
safety of CBA usage. Originally developed for assessing
effluents, 7-day chronic estimator tests  using  a  fish
(Menidia beryllina) and a crustacean (Mysidopsis bahia)
were adapted to evaluate  CBAs; the tests utilize end-
points of survival, growth, and, in the case of the mysids,
a measurement of egg production. Tier II evaluates the
toxicity of the  CBA, alone  and in the presence of a
water-soluble fraction of oil, to provide baseline informa-
tion on CBA toxicity and potential synergism with petro-
leum  hydrocarbons. Tier III examines effluents from
flow-through test systems that model a variety of aquatic
habitats  (open  water, beach, marsh) to assess toxicity
under more realistic conditions, where a CBA and oil are
allowed to interact.

Data are presented on the toxicity of a variety of CBAs
classified by vendors as microbial,  nutrient,  enzyme,
dispersant, and "other." In the flow-through test systems,
the CBAs exhibited relatively low toxicity, either by them-
selves or in the presence of an artificially weathered oil.
During a particular period, an apparent interaction be-
tween one CBA and oil appeared  to increase toxicity in
the marsh system. Toxicity reduction in the sand com-
ponent of the beach test system could not be developed
into an efficacy endpoint because very small quantities
of oil produced measurable  effects  on  a  benthic
amphipod.
                                                125

-------
        UNIFAC Phase Equilibrium Modeling To Assess the Bioavailability of
               Multicomponent Nonaqueous-Phase Liquids Containing
                            Polycyclic Aromatic Hydrocarbons
                                       Catherine A. Peters
        Department of Civil Engineering and Operations Research, Princeton University,
                                      Princeton, New Jersey
This work is part of a project to evaluate bioremediation
of contaminants that  are  nonaqueous-phase  liquid
(NAPL) mixtures of polycyclic aromatic hydrocarbons
(PAHs). This poster presents the first phase of this work,
aimed at gaining a thorough understanding of multicom-
ponent NAPL/water-phase equilibria for very complex
mixtures. This provides basic information about maxi-
mum bioavailable concentrations of PAHs.
The thermodynamics of multicomponent NAPL/water-
phase equilibria can be described with knowledge of the
mixture composition, NAPL-phase activity coefficients,
and pure solute aqueous solubilities. This analysis in-
volves  application of the UNIFAC  model  to  predict
NAPL-phase activity coefficients  for constituent com-
pounds in four different tar materials for which detailed
composition data are available. This group contribution
method has proven to  be useful for complex mixtures
such as coal tars because a mixture is represented by
a relatively small number of functional groups, making
thermodynamic  analysis  using excess  Gibbs  energy
models tractable. Forthis work, the molecular structures
of the uncharacterized portions of the tars are approxi-
mated through nonparametric regressions of functional
group characteristics with molecular weight. The UNI-
FAC model was found to predict nearly ideal behavior
for most tar constituents. The activity coefficients range
from 0.14  (quinoline) to 1.27 (ethylbenzene), but the
vast majority of the constituents are predicted to have
activity coefficients between 0.9 and 1.1.
These results provide a firm theoretical basis for making
an assumption of solution ideality for many tar constitu-
ents (i.e., Raoult's  law). The robustness of this conclu-
sion is  indicated through  comparable results  across
different tar materials, and through a sensitivity analysis
to the estimated characteristics of the uncharacterized
fractions. These results, in conjunction with laboratory
measurements of PAH biodegradation rates (for individ-
ual  compounds and for multiple substrate NAPL sys-
tems), will  eventually be integrated into a mathematical
model describing the rate of biotransformation of PAH-
containing  NAPL contaminants and the dynamics of the
composition of the  NAPL residual.
                                                126

-------
         Field Evaluation of Pneumatic Fracturing Enhanced Bioremediation
                                      Sankar N. Venkatraman
           Department of Chemical and Biochemical Engineering, Rutgers University,
                                      Piscataway, New Jersey

                             Thomas M. Boland and John R. Schuring
    Department of Civil and Environmental Engineering, New Jersey Institute of Technology,
                                       Newark, New Jersey

                                         David S. Kosson
           Department of Chemical and Biochemical Engineering, Rutgers University,
                                      Piscataway, New Jersey
In situ bioremediation  is often  limited by the rate  of
transport of nutrients and electron acceptors (e.g., oxy-
gen, nitrate) to the microorganisms mediating the proc-
ess, particularly in soil formations with moderate to low
permeability. To overcome these rate limitations, an in-
vestigation was conducted to integrate the process  of
pneumatic fracturing with bioremediation. Pneumatic
fracturing is  an innovative technology that uses high-
pressure air to create artificial fractures in contaminated
geologic formations, resulting in enhanced subsurface
air flow and transport  rates. Following  fracturing, the
pneumatic fracturing system  can also be used to inject
electron acceptors and other biological  amendments
directly into a formation to stimulate biodegradation. The
specific bioremediation process evaluated in this project
used amendment injections and low-rate in situ vapor
extraction to provide oxygen and other supplements,
which resulted in  the formation of aerobic, denitrifying,
and methanogenic biodegradation zones, spatially dis-
tributed with increasing distance from the fracture inter-
faces. A "countercurrent" bioremediation process was
thus established with respect to the diffusion of contami-
nants towards the fracture interface.

Afield pilot demonstration of the integrated technologies
was carried  out  at a gasoline  refinery site  over  a
20-month period. Initial site characterization indicated
the presence of BTX at concentrations of up to 1,500
mg/kg soil,  as well  as other hydrocarbons. The soil at
the site was overconsolidated clayey silt with very low
permeability. The site was pneumatically fractured fol-
lowed by periodic injections of subsurface amendments
over a period of 50 weeks. Results demonstrated that
fracturing increased subsurface permeability by an av-
erage of 36 times.  Information gained from periodic
vapor sampling indicated that following subsurface in-
jections, the production of carbon dioxide was enhanced
due  to increased biological activity. Following a  lag
phase, the  methanogens became active,  and  an in-
crease in methane production was observed. There was
no carbon dioxide or methane detected  in the prede-
monstration vapor samples. The mass of carbon con-
verted to carbon dioxide and methane was used as an
independent measure for the depletion of total carbon.
Based on this balance, the Generated to Cbi0degraded ratio
was computed to be 3.3:1, indicating that other carbon
sources in gasoline also served as substrate and par-
ticipated in the biodegradation process. After 1 year of
process operation  and monitoring,  soil  samples ob-
tained from the site indicated a 79-percent reduction in
soil-phase BTX concentrations, and over 85 percent of
the BTX reduction was attributed to biodegradation.
                                                127

-------
 Solids Suspension Characteristics Related to Slurry Biotreatment Performances
                      J.-W. Jim Tzeng, Paul T. McCauley, and John A. Glaser
    National Risk Management Research Laboratory, U.S. Environmental Protection Agency,
                                          Cincinnati, Ohio
Introduction

Slurry biotreatment has been  demonstrated to be an
effective  process  for bioremediation of contaminated
soils, sediments, and sludges (1). Solid-phase biotreat-
ment such as composting or formal land treatment units
cannot compete with slurry treatment for the extent of
treatment in short timeframes. Slurry biotreatment has
been commonly conducted in reactor systems such as
agitated tanks or lined lagoons. Slurry systems provide
conditions of improved  contact between the  pollutant
and the  microorganisms  responsible for the desired
biotransformation.  The extent of particle suspension by
agitation  is a crucial factor controlling treatment effi-
ciency. Power input from an impeller system dictates not
only the homogeneity of the slurry medium but also the
degree of particle  suspension. Equally importantly, the
power input requirement is an important economic factor
in  assessing the  feasibility of bioslurry treatment for
particular solids to be treated (2). Very little attention is
given, however, to this important component of the treat-
ment system. This work presents our current technical
efforts  on identifying the  conditions for optimal slurry
agitation  for bioremediation of contaminated  solids.

Results

Four flow regimes in terms of particle suspension char-
acteristics have been identified; in  increasing order of
impeller  power  input,   they   are:  nonsuspension,
semisuspension, off-bottom suspension, and complete-
suspension regimes. Experimental results indicate that
unique relationships exist between the flow regimes and
power input. In addition, kinetic energy rather than im-
peller rotational  speed dictates particle suspension dy-
namics in a slurry medium. A flow regime map (Figure
1)  is constructed using power  input as the primary pa-
rameter. At extremely low power inputs, particles remain
stationary and settle on  the tank bottom. As the power
input exceeds a certain value, the upper layer of the
settled particles starts to become mobile. With a further
      N on-Suspension

           Semi-Suspension
                           Favorable
                           Operating
                           Range
0)
c
1
0
_J
in
TJ
0
CO
                  Power Input

Figure 1. Flow regime map of particle suspension in slurry agi-
        tation systems.

increase in power, the layer of settled particles reduces
in thickness, and eventually all  particles are mobilized
with a portion of particles moving along the tank bottom.
Such  a state corresponds to the minimum off-bottom
suspension as conventionally reported in the literature.
The suspended particles tend to fall back to the tank
bottom, however, due to insufficient momentum trans-
ferred from the liquid medium to particles. Dynamics of
particles in this regime can be described by an up-and-
down  motion, and the particle distribution is nonuniform
along the  height of the slurry tank. An increase in the
impeller power input reduces the degree of nonunifor-
mity, and the particle distribution becomes rather uni-
form  as   the  power  input  exceeds the minimum
complete-circulation value.


Location of impellers (e.g.,  bottom  clearance) greatly
affects the particle suspension.  A substantial reduction
in power input required for both on-bottom and off-bot-
tom particle  suspension is obtained as an impeller is
placed near the tank bottom. The conventional design
of agitated tanks,  with bottom clearance equal to the
                                                 128

-------
impeller diameter or tank radius,  is inadequate for par-
ticle suspension applications.

Conclusion

Feasibility  of  slurry bioreactors  for bioremediation of
contaminated  solids depends on  the energy consump-
tion required to achieve adequate fluid mixing and par-
ticulate suspension in the slurry medium. Four particle
suspension regimes are identified with respect to  the
power input to the  slurry medium in an agitated tank.
Solids properties affecting the suspension behavior in-
clude 1)  size  and shape distribution, 2) density differ-
ence,  and  3) solids  loading.  Operation  under  the
complete  suspension  regime for  achieving  maximum
uniformity of particle suspension may not be necessary
because the treatment efficiency may only be marginally
improved. This is because mass transfer resistance be-
tween particles and the bulk liquid  phase is not the only
rate-limiting step in the soil-slurry treatment process.

References

1. U.S.  EPA.  1990.  Engineering  bulletin:  Slurry biodegradation.
  EPA/540/2-90/016. Cincinnati, OH.
2. Muskett, M.J., and A.W.  Nienow. 1988.  Capital vs. running costs:
  The economics of mixer  selection. In: The Institution of Chemical
  Engineers, ed. Fluid Mixing III.
                                                    129

-------