EPA/600/R-93/124
July 1993
In-Situ Bioremediation of Ground Water
and Geological Material:
A Review of Technologies
by
Robert D. Morris, Robert E. Hinchee, Richard Brown,
Perry L. McCarty, Lewis Semprini, John T. Wilson,
Don H. Kampbell, Martin Reinhard, Edward J. Bouwer
Robert C. Borden, Timothy M. Vogel,
J. Michele Thomas and C. H. Ward
68-C8-0058
Project Officer
John E. Matthews
Chief, Applications and Assistance Branch
Robert S. Kerr Environmental Research Laboratory
Ada, Oklahoma 74820
ROBERT S. KERR ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
ADA, OKLAHOMA 74820
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BIBLIOGRAPHIC INFORMATION
PB93-215564
Report Nos; none
Title: In-situ Bioremediation of Ground Water and Geological Material: A Review of
Technologies.
Date; Jul 93
Authors; R. D. Morns, R. E. Hichee, R. Brown, P. L. McCarty, and L. Semprini.
Performing Organization; Dynamac Corp., Ada, OK.
Performing Organization Report Nos; EPA/600/R-93/124
Sponsoring Organization: *Robert S. Kerr Environmental Research Lab., Ada, OK.
Type of Report and Period Covered; Research rept.
Supplementary Notes; See also PB93-146850. Sponsored by Robert S. Kerr
Environmental Research Lab., Ada, OK.
NTIS Field/Group Codes; 68C, 68D, 57K
Price; PC A12/MF A03
Availability; Available from the National Technical Information Service,
Springfield, VA. 22161
Number of Pages; 252p
Keywords; *Ground water, *Water pollution control, ^Remediation, *Land pollution
control, Hazardous materials, Waste disposal, Hydrocarbons, Microorganisms,
Biological treatment, Biodegradation, Activated sludge process, Biological
industrial waste treatment, Oxidation, *Bioremediation, Contaminated soils,
Chlorinated solvents.
Abstract; The report provides the reader with a detailed background of the
technologies available for the bioremediation of contaminated soil and ground
water. The document has been prepared for scientists, consultants, regulatory
personnel, and others who are associated in some way with the restoration of soil
and ground water at hazardous waste sites. It provides the most recent scientific
understanding of the processes involved with soil and ground-water remediation, as
well as a definition of the state-of-the-art of these technologies with respect to
circumstances of their applicability and their limitations. In addition to
discussions and examples of developed technologies, the report also provides
insights to emerging technologies which are at the research level of formation,
ranging from theoretical concepts, through bench scale inquiries, to limited
field-scale investigations. The report centers around a number of bioremediation
technologies applicable to the various subsurface compartments into which
contaminants are distributed. The processes which drive these remediation
Continued on next page
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BIBLIOGRAPHIC INFORMATION
Continued... PB93-215564
technologies are discussed in depth along with the attributes which direct their
applicability and limitations according to the phases into which the contaminants
have partitioned. These discussions include in-situ remediation systems, air
sparging and bioventing, use of electron acceptors alternate to oxygen, natural
bioremediation, and the introduction of organisms into the subsurface. The
contaminants of major focus in the report are petroleum hydrocarbons and
chlorinated solvents.
11
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DISCLAIMER
The information in this document has been funded wholly or in part by the United States
Environmental Protection Agency under Contract No. 68-C8-0058 to Dynamac Corporation.
This report has been subjected to the Agency's peer and administrative review, and has been
approved for publication as an EPA document. Mention of trade names or commercial products
does not constitute endorsement or recommendation for use.
n -»-
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FOREWORD
EPA is charged by Congress to protect the Nation's land, air and water systems. Under
a mandate of national environmental laws focused on air and water quality, solid waste
management and the control of toxic substances, pesticides, noise and radiation, the Agency
strives to formulate and implement actions which lead to a compatible balance between human
activities and the ability of natural systems to support and nurture life.
The Robert S. Kerr Environmental Research Laboratory is the Agency's center of
expertise for investigation of the soil and subsurface environment. Personnel at the laboratory
are responsible for management of research programs to: (a) determine the fate, transport and
transformation rates of pollutants in the soil, the unsaturated and the saturated zones of the
subsurface environment; (b) define the processes to be used in characterizing the soil and
subsurface environment as a receptor of pollutants; (c) develop techniques for predicting the
effect of pollutants on ground water, soil, and indigenous organisms; and (d) define and
demonstrate the applicability and limitations of using natural processes, indigenous to the soil
and subsurface environment, for the protection of this resource.
In-situ bioremediation of subsurface environments involves the use of microorganisms
to convert contaminants to less harmful products and sometimes offers significant potential
advantages over other remediation technologies. This report provides the most recent scientific
understanding of the processes involved with soil and ground-water bioremediation and
discusses the applications and limitations of the various in-situ bioremediation technologies.
Clinton W. Hall
Director
Robert S. Kerr Environmental
Research Laboratory
111
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IV
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CONTENTS
Foreword iii
Figures xi
Tables riii
Executive Summary 1
Section 1. Introduction 1-1
Section 2. In-situ Bioremediation of Soils and Ground Water Contaminated with
Petroleum Hydrocarbons
2.1. Introduction 2-1
2.2. Fundamental Principles 2-1
2.3. Historical Perspective 2-3
2.4. Repositories of Expertise 2-5
2.5. General Designs 2-5
2.6. Laboratory Testing 2-6
2.7. Contamination Limits 2-6
2.8. Site Characterization 2-8
2.9. Favorable Site Conditions 2-9
2.9.1 Solubility 2-9
2.9.2 Volatility 2-9
2.9.3 Viscosity 2-9
2.9.4. Toxicity 2-10
2.9.5. Permeability of Soils and Subsurface Materials 2-10
2.9.6. Soil Type 2-10
2.9.7. Depth to Water 2-10
2.9.8. Mineral Content 2-10
2.9.9. Oxidation/Reduction Potential 2-10
2.9.10. pH 2-11
2.10. Infrastructure and Institutional Issues 2-11
2.11. Performance 2-12
2.12. Problems 2-12
2.13. Site Properties vs Cost 2-14
2.13.1. Mass of Contaminant 2-14
2.13.2. Volume of Contaminated Aquifer 2-14
2.13.3. Aquifer Permeability/Soil Characteristics 2-14
2.13.4 Final Remediation Levels 2-14
2.13.5. Depth to Water 2-15
2.13.6. Monitoring Requirements 2-15
2.13.7. Contaminant Properties 2-15
2.13.8. Location of Site 2-15
2.14. Previous Experience with Costs 2-16
2.15. Regulatory Acceptance 2-17
2.16. Knowledge Gaps 2-18
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Section 3. Bioventing of Petroleum Hydrocarbons
3.1. Fundamental Principles 3-1
3.1.1. Review of the Technology 3-1
3.1.2. Maturity of the Technology 3-3
3.1.3. Repositories of Expertise 3-3
3.2. Contamination that is Subject to Treatment 3-5
3.3. Special Requirements for Site Characterization 3-6
3.3.1. Soil Gas Survey 3-6
3.3.2. Soil Gas Permeability and Radius of Influence 3-6
3.3.3. In-Situ Respiration 3-7
3.4. Impact of Site Characteristics on Applicability 3-9
3.5. Process Performance 3-10
3.5.1. Case Study: Hill AFB Site 3-12
3.5.2. Case Study: Tyndall AFB Site 3-14
3.5.3. Performance of Other Sites 3-16
3.6. Problems Encountered with the Technology 3-17
3.7. Costs 3-17
3.8. Regulatory Acceptance 3-17
3.9. Knowledge Gaps and Research Opportunities 3-17
Section 4. Treatment of Petroleum Hydrocarbons in Ground Water by Air Sparging
4.1. Introduction 4-1
4.2. Development of Air Sparging 4-2
4.3. Principles of the Technology 4-3
4.4. Benefits of Air Sparging 4-5
4.5. Dangers of Air Sparging 4-8
4.6. Barriers to Flow 4-8
4.7. Control of Flow 4-10
4.8. Summary of Limitations 4-11
4.9. System Application and Design 4-13
4.9.1. Nature and Extent of Site Contaminants 4-13
4.9.2. Hydrogeologic Conditions 4-14
4.9.3. Potential Ground-Water and Vapor Receptors 4-14
4.10. Field Pilot Testing 4-14
4.11. Design Data Requirements 4-15
4.12. System Elements 4-17
4.13. System Examples 4-19
4.14. Cost Factors 4-22
4.15. Conclusion 4-23
Section 5. Ground-Water Treatment for Chlorinated Solvents
5.1. Introduction 5-1
5.2. BiotransformationofCAHs 5-3
5.2.1. Primary Substrates and Cometabolism 5-3
5.2.2. CAH Usage as Primary Substrates 5-6
5.2.3. Anaerobic Cometabolic Transformations of CAHs 5-6
5.2.4. Aerobic Microbial Transformation of
Chlorinated Aliphatic Hydrocarbons 5-7
VI
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5.3. Processes Affecting Chemical Movement and Fate 5-9
5.3.1. Effect of Sorption 5-9
5.4. Field Pilot Studies of CAH Transformation 5-11
5.4.1. Results with Methanotrophs 5-12
5.4.2. Results with Phenol Utilizers 5-14
5.4.3. Comparison Between the Methane and Phenol Studies 5-16
5.4.4. Anaerobic Transformation of Carbon Tetrachloride 5-16
5.5. Procedures for Introducing Chemicals into Ground Water 5-17
5.6. The Effect of Site Conditions on Remediation Potential 5-21
5.6.1. Microorganism Presence 5-23
5.7. Summary 5-24
Section 6. Bioventing of Chlorinated Solvents for Ground-Water
Cleanup through Bioremediation
6.1. Fundamental Principles 6-1
6.2. Maturity of the Technology 6-2
6.3. Primary Repositories of Expertise 6-2
6.4. Contamination Subject to Treatment 6-3
6.5. Special Requirements for Site Characterization 6-4
6.6. Site Characteristics that are Particularly Favorable 6-5
6.7. Site Characteristics that are Particularly Unfavorable 6-5
6.8. Performance under Optimal Conditions 6-6
6.8.1. Importance of the Rate Law 6-6
6.8.2. Importance of Partitioning 6-6
6.9. Problems Encountered with the Technology 6-9
6.10. Relevant Experience with System Design 6-9
Section 7. In-Situ Bioremediation Technologies for Petroleum Derived
Hydrocarbons Based on Alternate Electron Acceptors (other than
molecular oxygen)
7.1. Fundamental Principles 7-1
7.1.1. Comparison of Oxygen and Alternate Electron Acceptor
Based In-Situ Bioremediation Technologies 7-1
7.1.2. Hydrocarbon Transformation Based on Alternate
Electron Acceptors 7-3
7.1.2.1. Laboratory Studies 7-3
7.1.2.2. Large Scale Bioremediation Studies Using Nitrate 7-6
7.1.3? Maturity of the Technology 7-6
7.1.4. Primary Repository of Expertise 7-7
7.2. Contamination That is Subject to Treatment 7-8
7.2.1. Chemical Nature 7-8
7.2.2. Range of Concentration 7-8
7.3. Requirements for Site Characterization and Implementation
of the Technology 7-8
7.4. Favorable Site Characteristics 7-9
7.5. Unfavorable Site Characteristics 7-9
7.5.1. Chemical and Physical Nature of the Contamination 7-9
7.5.2. Site Hydrogeology and Source Characteristics 7-9
7.5.3. Infrastructure and Institutional Issues 7-9
vn
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7.6. Optimal Site Conditions 7-9
7.7. Problems Encountered with the Technology 7-10
7.8. Properties of the Site and the Contaminants Determining
the Costs of Remediation 7-10
7.9. Previous Experience with Cost of Implementing the Technology 7-11
7.10. Factors Determining Regulatory Acceptance of the
Technology 7-11
7.11. Primary Knowledge Gaps and Research Opportunities 7-11
7.11.1. Degradation Under Ideal Conditions 7-12
7.11.2. Degradation Under Ground-Water and Soil Conditions 7-12
7.11.3. Degradation at Sites 7-12
7.11.4. Degradation at the Hydrocarbon/Water Interface
and Within the Nonaqueous Phase 7-13
7.11.5. Methods for Monitoring Performance of
Bioremediation Process 7-13
7.11.6. Design of Optimal Nutrient and Electron Acceptor Systems 7-13
Section 8. Bioremediation of Chlorinated Solvents using Alternate
Electron Acceptors
8.1. Introduction 8-1
8.2. Metabolism and Alternate Electron Acceptors 8-2
8.3. Biotransformation of Chlorinated Solvents in the
Presence of Alternate Electron Acceptors 8-3
8.3.1. Carbon Tetrachloride Biotransformation 8-7
8.3.2. Tetrachloroethene and Trichloroethene Biotransformation 8-7
8.4. Approaches for Treatment 8-8
8.5. Field Experience 8-10
8.6. Sequential Anaerobic/Aerobic Transformations of
Chlorinated Solvents 8-10
8.7. Performance 8-11
8.7.1. Physical/Chemical Properties 8-11
8.7.2. Concentration Range 8-12
8.7.3. Favorable Redox Conditions 8-13
8.8. Biotransformation Stoichiometry 8-15
8.9. Biotransformation Rates 8-18
8.10. Limitations 8-19
8.11. Research Needs 8-21
8.12. Concluding Remarks 8-21
Section 9. Natural Bioremediation of Hydrocarbon-Contaminated Ground Water
9.1. General Concept of Natural Bioremediation 9-1
9.2. Hydrocarbon Distribution, Transport and Biodegradation
in the Subsurface 9-2
9.2.1. Petroleum Hydrocarbon Biodegradation 9-3
9.2.2. Subsurface Microorganisms 9-3
9.2.3. Use of Different Electron Acceptors for Biodegradation 9-4
9.2.3.1. Aerobic Biodegradation 9-4
9.2.3.2. Biodegradation via Nitrate Reduction 9-5
9.2.3.3. Biodegradation using Ferric Iron 9-5
vin
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9.2.3.4. Biodegradation via Sulfate Reduction and
Methanogenesis 9-6
9.2.4. Effect of Environmental Conditions on Biodegradation 9-6
9.3. Natural Bioremediation of a Hydrocarbon Plume 9-7
9.4. Case Studies of Natural Bioremediation 9-9
9.5. Site Characterization for Natural Bioremediation 9-10
9.5.1. Is the Contaminant Biodegradable? 9-11
9.5.2. Is Biodegradation Occurring in the Aquifer? 9-11
9.5.3. Are Environmental Conditions Appropriate for Biodegradation?.... 9-12
9.5.4. If the Waste Doesn't Completely Biodegrade, Where Will It Go? ....9-12
9.6. Monitoring Natural Bioremediation Systems 9-13
9.7. Performance of Natural Bioremediation Systems -..9-14
9.8. Predicting the Extent of Natural Bioremediation 9-14
9.9. Issues That May Affect the Costs of This Technology 9-16
9.10. Knowledge Gaps and Research Opportunities 9-16
Section 10. Natural Bioremediation of Chlorinated Solvents
10.1. Summary 10-1
10.2. Fundamental Principles 10-1
10.3. Chemical Reactions 10-3
10.4. Microbiological Reactions 10-4
10.4.1. Aerobic 10-4
10.4.2. Anaerobic 10-6
10.5. Predictions of Product Distribution 10-8
10.6. Rationale for Technology 10-11
10.7. Practical Implications 10-15
10.8. Special Requirements for Site Characterization 10-16
10.9. Favorable Site Characteristics 10-16
10.10. Unfavorable Site Characteristics 10-18
10.11. Cost Evaluations 10-20
10.12. Knowledge Gaps 10-20
10.13. Conclusion 10-20
Section 11. Introduced Organisms for Subsurface Bioremediation
11.1. Fundamental Principles of the Technology 11-1
11.1.1. Review of the Development of the Technology 11-1
11.1.2. Matrix Properties that Affect Transport 11-2
11.1.3. Properties of Organisms that Affect Transport 11-4
11.1.4. Operational Factors that Affect Transport 11-5
11.1.5. Environmental Factors that Affect Survivability of
Added Organisms 11-7
11.1.6. Field Demonstrations of Microbial Transport 11-7
11.1.7. Inoculation to Enhance Biodegradation of Hydrocarbons 11-9
11.1.8. Inoculation to Enhance Biodegradation of
Chlorinated Compounds 11-10
11.2. Maturity of the Technology 11-11
11.3. Primary Repositories of Expertise 11-12
IX
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11.4. Other Factors Concerning Application •. 11-12
11.5. State of the Art of Transport of Microorganisms with Specialized
Metabolic Capabilities and Research Opportunities 11-13
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FIGURES
Figure 1. Distribution of contaminants in the subsurface 2
Figure 2. Contaminant locations treated by in situ bioremediation 2
Figure 3. Contaminant locations treated by bioventing 5
Figure 4. Contaminant locations treated by air sparging 6
Figure 5. Contaminant locations treated by alternate electron acceptors 8
Figure 6. Contaminant locations treated by aerobic natural bioremediation 9
Figure 7. Profile of a typical hydrocarbon plume undergoing natural bioremediation 10
Figure 1.1. Distribution of petroleum hydrocarbons in the subsurface 1-2
Figure 1.2. Migration of DNAPL through the vadose zone to an impermeable
boundary in relatively homogenous subsurface materials 1-3
Figure 1.3. Perched and deep DNAPL reservoirs from migration through
heterogeneous subsurface materials 1-3
Figure 2.1. Bioremediation in the saturated zone 2-2
Figure 3.1. Impact of physicochemical properties on potential for bioventing 3-5
Figure 3.2. Gas injection/soil gas sampling monitoring point used
by Hinchee and Ong (1992) in their in-situ respiration studies 3-8
Figure 3.3. Average oxygen utilization rates measured at four test sites 3-8
Figure 3.4. Conceptual layout of bioventing process with air injection only 3-10
Figure 3.5. Conceptual layout of bioventing process with air withdrawn
from clean soil 3-11
Figure 3.6. Conceptual layout of bioventing process with soil gas reinjection 3-11
Figure 3.7. Conceptual layout of bioventing process with air injection into con-
taminated soil, coupled with dewatering and nutrient application 3-12
Figure 3.8. Cumulative hydrocarbon removal from the Hill AFB
Building 914 soil venting site 3-13
Figure 3.9. Results of soil analysis at Hill AFB before and after venting 3-13
Figure 3.10. Results of soil analysis from Plot V2 at Tyndall AFB before
and after venting 3-15
Figure 3.11. Cumulative percent hydrocarbon removal at Tyndall AFB for
Sites VI and V2 3-15
Figure 4.1. Diagram of air sparging system 4-1
Figure 4.2. Differences between old and new air sparging technologies 4-3
Figure 4.3. The effects of air flow in saturated environment as a
function of air flow rate 4-4
Figure 4.4. Air sparging partitioning and removal mechanisms as a
function of volatility 4-7
Figure 4.5. Air sparging removal mechanisms as a function of product volatility 4-7
Figure 4.6. Inhibited vertical air flow due to impervious barrier 4-9
Figure 4.7. Channeled air flow through highly permeable zone 4-9
Figure4.8. Effect of injection pressure on air flow 4-10
Figure 4.9. Agreement between sparge parameters in estimating
the radius of influence 4-16
Figure 4.10. Nested sparge well 4-18
Figure 4.11. Monitoring point for sparging systems 4-18
Figure 4.12. System layout Site A 4-20
Figure 4.13. Layout of Site B air sparging/vent system 4-22
Figure 5.1. Anaerobic transformations of CAHs 5-6
XI
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Figure 5.2. Effect of advection, dispersion, and sorption on contaminant
movement from an experiment at Borden Air Force Base 5-10
Figure 5.3. Methane and oxygen utilization by methanotrophs at the
Moffett test facility 5-13
Figure 5.4. CAH transformation by methanotrophs at the Moffett test facility 5-13
Figure 5.5. CAH transformation and dissolved oxygen changes resulting
from phenol addition at the Moffett test facility 5-15
Figure 5.6. CT transformation under anaerobic conditions at the Moffett test site 5-17
Figure 5.7. Pump-extract-reinject method for mixing of chemicals with ground water 5-18
Figure 5.8. Simulation modeling of pump-extract-reinject method for
methanotrophiccometabolismofVC 5-19
Figure 5.9. Simulation modeling of pump-extract-reinject method for
methanotrophiccometabolismoftrans-DCE 5-20
Figure 5.10. Subsurface recirculation system for chemical introduction
and mixing with ground-water contaminants 5-20
Figure 5.11. CAH contamination in a relatively homogeneous
subsurface environment 5-21
Figure 5.12. CAH contamination in a relatively nonhomogeneous
subsurface environment 5-22
Figure 6.1. Two hypothetical implementations of in-situ bioventing
of chlorinated solvents 6-2
Figure 6.2. Soil bed constructed by S.C. Johnson & Son, Inc., in Racine, Wisconsin,
to treat an airstream containing 2,000 to 3,500 ppm of propellent gas 6-10
Figure 6.3. Effect of flow rate on removal efficiency of propellant
gas hydrocarbons in a soil bed reactor 6-11
Figure 8.1. Important electron donors and acceptors in
biotransformation processes 8-3
Figure 8.2. Chemical requirements and products of anaerobic bioremediation
for one cubic meter soil contaminated with PCE using
microbial system reported by de Bruin et al. (1992) 8-16
Figure 8.3. Chemical requirements and products of anaerobic bioremediation
for one cubic meter soil contaminated with PCE using
microbial system reported by DiStefano et al. (1991) 8-17
Figure 9.1. Profile of a typical hydrocarbon plume undergoing natural bioremediation 9-8
Figure 9.2. Plan view of a typical hydrocarbon plume undergoing
natural bioremediation 9-8
Figure 10.1. PCE anaerobic transformations 10-6
Figure 10.2. Abiotic and biotic transformations of 1,1,1-trichloroethane 10-7
Figure 10.3. Reductive dechlorination of trichloroethylene (TCE) under
hypothesized anaerobic field or laboratory conditions 10-9
Figure 10.4. Chemical degradation of 1,1,1-trichloroethane (TCA) 10-9
Figure 10.5. Chemical and microbial degradation of TCA (lower microbial activity) 10-10
Figure 10.6. Chemical and microbial degradation of TCA (higher microbial activity) 10-10
Figure 10.7. Chemical and microbial degradation of both TCE and TCA 10-11
Figure 10.8. Schematic illustrating the reductive dechlorination of polychlorinated
compounds in an anaerobic biofilm and subsequent mineralization of
the products of anaerobic treatment in an aerobic biofilm 10-12
Figure 10.9. Relationships between degree of chlorination and anaerobic reductive
dechlorination, aerobic degradation and sorption onto subsurface
material 10-19
xii
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TABLES
Table 3.1. Darcy velocity in relation to soil type 3-7
Table 3.2. Comparison of biodegradation rates obtained by the
in-situ respiration test with other studies 3-16
Table 4.1. Oxygen availability, Ib/day 4-5
Table 4.2. Henry's constant for selected hydrocarbons 4-6
Table 4.3. Water table mounding and collapse 4-11
Table 4.4. Limits to the use of air sparging 4-13
Table 4.5. Site and pilot test data needed for design 4-16
Table 4.6. Air sparging system elements 4-17
Table 4.7. Approximate cost factors 4-23
Table 5.1. Common halogenated aliphatic hydrocarbons 5-2
Table 5.2. Potential for CAH biotransformation as a primary substrate or through
cometabolism 5-4
Table 5.3. Transformations of CAHs (after Vogeletal., 1987) 5-5
Table 6.1. The common chlorinated organic compounds occurring as
contaminants of ground water 6-3
Table 6.2. Effect of the concentration of trichloroethylene on the rate of
biodegradation of aviation gasoline vapors in soil microcosms 6-4
Table 6.3. Partitioning of chlorinated organic compounds between air,
water, and solids 6-7
Table 6.4. Kinetics of depletion of natural hydrocarbons in unsaturated soil
and subsurface material 6-7
Table 6.5. Mineralization of vapors of chlorinated solvents in soil acclimated
to degrade vapors of natural hydrocarbons 6-8
Table 6.6. Removal of vapors of trichloroethylene and vinyl chloride in
subsurface material under optimal conditions 6-9
Table 7.1. Field studies where denitrification has been evaluated 7-7
Table 8.1. Anaerobic transformation of selected chlorinated solvents in
microcosms and enrichment cultures under different
redox conditions 8-5
Table 8.2. Reductive dehalogenation reactions catalyzed by pure
cultures of bacteria 8-6
Table 8.3. Favorable and unfavorable chemical and hydrogeological site
conditions for implementation of in-situ bioremediation 8-9
Table 8.4. Physical-chemical properties of chlorinated solvents common
to ground-water contamination 8-12
Table 8.5. Standard reduction potentials at 25°C and pH 7 for some redox
couples that are important electron acceptors in microbial
respiration and for some half-reactions involving chlorinated solvents 8-14
Table 8.6. Stoichiometric relationships for possible bioremediation
reactions involving complete dechlorination of PCE to ethene 8-17
Table 8.7. PCE dechlorination rates by different anaerobic bacteria 8-19
Table 10.1. Production, proposed maximum contaminant levels, and
toxicity ratings of common halogenated aliphatic compounds 10-2
Table 10.2. Relative rates of degradation by methanogenic cultures 10-7
Xlll
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Table 10.3. Products of anaerobic dechlorination 10-13
Table 10.4. Products of aerobic degradation 10-13
Table 10.5. Proposed nutrients for bioremediation 10-14
Table 10.6. Some information needed for prediction of organic contaminant
movement and transformation in ground water 10-17
Table 11.1. Possible applications of introduced microorganisms 11-13
xiv
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EXECUTIVE SUMMARY
INTRODUCTION
It is the intent of this report to provide the reader with a detailed background of the
technologies available for the bioremediation of contaminated soil and ground water. The document
has been prepared for scientists, consultants, regulatory personnel, and others who are associated
in some way with the restoration of soil and ground water at hazardous waste sites.
The reader is served by this presentation in that it provides the most recent scientific
understanding of the processes involved with soil and ground-water remediation, as well as a
definition of the state-of-the-art of these technologies with respect to circumstances of their
applicability and their limitations. In addition to discussions and examples of developed tech-
nologies, the report also provides insights to emerging technologies which are at the research level
of formation, ranging from theoretical concepts, through bench scale inquiries, to limited field-scale
investigations.
In order for the information in this document to be of maximum benefit, it is important that
the reader understand how contaminants are distributed among the various subsurface
compartments. This distribution, or phase partitioning of contaminants, is dependent upon a
number of factors including the characterization of the contaminants themselves and that of the
subsurface environment. This distribution is exemplified in Figure 1 where contaminants are
shown to be associated with the vapor phase in the unsaturated zone, a residual phase, or dissolved
in ground water.
The report centers around a number ofbioremediation technologies applicable to the various
subsurface compartments into which contaminants are distributed. The processes which drive
these remediation technologies are discussed in depth along with the attributes which direct their
applicability and limitations according to the phases into which the contaminants have partitioned.
These discussions include in-situ remediation systems, air sparging and bioventing, use of electron
acceptors alternate to oxygen, natural bioremediation, and the introduction of organisms into the
subsurface. The contaminants of major focus in this report are petroleum hydrocarbons and
chlorinated solvents.
IN-SITU BIOREMEDIATION OF SOIL AND GROUND WATER
Bioremediation of excavated soil, unsaturated soil, or ground water (Figure 2) involves the
use of microorganisms to convert contaminants to less harmful species in order to remediate
contaminated sites. In order for these biodegradative processes to occur, microorganisms require
the presence of certain minerals, referred to as nutrients, and an electron acceptor. Several other
conditions, i.e. temperature, pH, etc., impact the effectiveness of these processes. The use of
biooxidation for environmental purposes has existed for many years and has led to considerable
information regarding the biodegradability of specific classes of compounds, nutrient and electron
acceptor requirements, and degradation mechanisms. Activated sludge and other suspended
growth systems have been used for decades to treat industrial and municipal wastes. Land
treatment processes for municipal wastewater and petroleum refinery and municipal wastewater
sludges have also been practiced for several decades and have generated a great deal of information
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Water
Table
Residual
Saturation
Capillary
^Fringe
Dissolved
Contaminants
Figure 1. Distribution of contaminants in the subsurface.
Water
Table
Residual
Saturation
ved
Contaminants
Figure 2. Contaminant locations treated by in-situ bioremediation.
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on nutrient requirements, degradation rates, and other critical parameters affecting biological
oxidation.
In the 1970s, tests were conducted to evaluate biological degradation of petroleum
hydrocarbons in aquifers. Results from these tests demonstrated thatin-situ bioremediation could
reduce levels of hydrocarbons in aquifers, and provided considerable information concerning the
processes which take place and the requirements necessary to drive these processes.
Although a variety of minerals are required by the microorganisms, it is usually necessary
to add only nitrogen and phosphorus. The most common electron acceptor used in bioremediation
is oxygen. Stoichiometrically, approximately three pounds of oxygen are required to convert one
pound of hydrocarbon to carbon dioxide. Nutrient requirements are less easily predicted. If all
hydrocarbons are converted to cell material, however, it can be assumed that nutrient requirements
of carbon to nitrogen to phosphorus ratios are in the order of 100:10:1. In some cases where the levels
of contaminants are low, sufficient nitrogen and phosphorus are naturally present, and only oxygen
is required for the biological processes to proceed.
In-situ bioremediation systems for aquifers typically consist of extraction points such as
wells or trenches, and injection wells or infiltration galleries. In most cases, the extracted ground
water is treated prior to the addition of oxygen and nutrients, followed by subsequent reinjection.
Critical to the design of an in-situ bioremediation system is the ground-water flow rate and
flow path. The ground-water flow must be sufficient to deliver the required nutrients and oxygen
according to the demand of the organisms, and the amended ground water should sweep the entire
area requiring treatment. This is a critical point in that it is often the hydraulic conductivity of the
ground-water system itself or the variability of the aquifer materials which limits the effectiveness
of in-situ technologies or prevents its utility entirely. A suggested target for in-situ remediation
technologies is a hydraulic conductivity of at least 10" cm/sec (100 ft/yr). The results of a number
of referenced studies suggest that in-situ bioremediation of the subsurface is usually limited to
formations with hydraulic conductivities of 10"* cm/sec (100 ft/yr) or greater to overcome the
difficulty of pumping fluids through contaminated formations.
In-situ bioremediation systems are often integrated with other remediation technologies
either sequentially or simultaneously. For example, if free phase hydrocarbons are present, a
recovery system should be used to reduce the mass of free phase product prior to the implementation
of bioremediation. In-situ vapor stripping can be used to both physically remove volatile
hydrocarbons and to provide oxygen for bioremediation. These systems can also reduce levels of
residual phase hydrocarbons as well as constituents adsorbed to both unsaturated soils and soils
which become unsaturated during periods when the water table is lowered.
As a class, petroleum hydrocarbons are biodegradable. The lighter soluble members are
generally biodegraded more rapidly and to lower residual levels than are the heavier, less soluble
members. Thus monoaromatic compounds such as benzene, toluene, ethylbenzene, and the xylenes
are more rapidly degraded than the two-ring compounds such as naphthalene, which are in turn
more easily degraded than the three-, four-, and five-ring compounds.
Polyaromatic hydrocarbons are present in heavier petroleum hydrocarbon blends and
particularly in coal tars, wood treating chemicals, and refinery waste sludges. These compounds
have only limited solubility in water, adsorb strongly to soils, and degrade at rates much slower than
monoaromatic hydrocarbons.
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Nonchlorinated solvents used in a variety of industries are generally biodegradable. For
example, alcohols, ketones, ethers, carboxylic acids, and esters are readily biodegradable but may
be toxic to the indigenous microflora at high concentrations due to their high water solubility.
Lightly chlorinated compounds such aschlorobenzene, dichlorobenzene, chlorinated phenols,
and the lightly chlorinated PCBs are typically biodegradable under aerobic conditions. The more
highly chlorinated analogs are more recalcitrant to aerobic degradation but more susceptible to
degradation under anaerobic conditions.
Chlorinated solvents and their natural transformation products represent the most
prevalent organic ground-water contaminants in the country. These solvents, consisting primarily
of chlorinated aliphatic hydrocarbons, have been widely used for degreasing aircraft engines,
automotive parts, electronic components, and clothing.
In-situ biodegradation of most of these solvents depends upon cometabolism and can be
carried out under aerobic or anaerobic conditions. Cometabolism requires the addition of an
appropriate primary substrate to the aquifer and perhaps an electron acceptor, such as oxygen or
nitrate, for its oxidation.
In the early 1980s there were few companies that had experience in the bioremediation of
soil and ground water. Since that time many companies have utilized bioremediation technologies,
although claims of experience are frequently overstated. There now exists a number of organizations
and specialists that are knowledgeable in the field of in-situ bioremediation. Several environmental
companies have staffs that are experienced in the application of this technology. Many large
corporations, especially the oil and chemical companies, have also developed in-house expertise.
Some of the U.S. Environmental Protection Agency laboratories as well as Department of Defense
and Department of Energy groups have conducted laboratory research and field demonstration
studies concerning bioremediation.
BIOVENTING
Bioventing is the process of supplying air or oxygen to soil to stimulate the aerobic
biodegradation of contaminants. This technology is applicable to contaminants in the vadose zone
and contaminated regions of an aquifer just below the water table (Figure 3). This in-situ process
may be applied to the vadose zone as well as an extended unsaturated zone caused by dewatering.
Bioventing is a modification of the technology referred to as soil vacuum extraction, vacuum
extraction, soil gas extraction, and in-situ volatilization.
Laboratory research and field demonstrations involving bioventing began in the early
1980s, with particular emphasis to the remediation of soil contaminated with hydrocarbons. Early
on, researchers concluded that venting would not only remove gasoline by physical means, but
would also enhance microbial activity and promote the biodegradation of gasoline. Much of the
success of this technology is because the use of air as a carrier of oxygen is 1,000 times more efficient
than water. It is estimated that various forms of bioventing have been applied to more than 1,000
sites worldwide, however, little effort has been given to the optimization of these systems.
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Water
Table
Residual
Saturation
Contaminants
Figure 3. Contaminant locations treated by bioventing.
Bioventing is potentially applicable to any contaminant that is more readily biodegradable
aerobically than anaerobically. Although most applications have been to petroleum hydrocarbons,
applications to PAH, acetone, toluene, and naphthalene mixtures have been reported. In most
applications, the key is biodegradability versus volatility. If the rate of volatilization significantly
exceeds the rate of biodegradation, removal essentially becomes a volatilization process.
In general, low-vapor pressure compounds (less than 1 mm Hg) cannot be successfully
removed by volatilization, and can only be biodegraded in a bioventing application. Higher vapor
pressure compounds (above 760 mm Hg) are gases at ambient temperatures and therefore volatilize
too rapidly to be biodegraded in a bioventing system. Within this intermediate range (1 - 760 mm
Hg) lie many of the petroleum hydrocarbon compounds of regulatory interest, such as benzene,
toluene, and the xylenes, that can be treated by bioventing.
In addition to the normal site characterization required for the implementation of this or any
other remediation technology, additional investigations are necessary. Soil gas surveys are
required to determine the amount of contaminants, oxygen, and carbon dioxide in the vapor phase;
the latter are needed to evaluate in-situ respiration under site conditions. An estimate of the soil
gas permeability along with the radius of influence of venting wells is also necessary to design full-
scale systems, including well spacing requirements, and to size blower equipment.
Although bioventing has been performed and monitored at several field sites, many of the
effects of environmental variables on bioventing treatment rates are still not well understood. In-
situ respirometry at additional sites with drastically different geologic conditions has further
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defined environmental limitations and site-specific factors that are pertinent to successful bioventing.
However, the relationship between respirometric data and actual bioventing treatment rates has
not been clearly determined. Concomitant field respirometry and closely monitored field bioventing
studies are needed to determine the type of contaminants that can successfully be treated by in-situ
bioventing and to better define the environmental limitations to this technology.
AIR SPARGING
Air sparging is the injection of air under pressure below the water table to create a transient
air-filled porosity by displacing water in the soil matrix. Air sparging is a remediation technology
applicable to contaminated aquifer solids and vadose zone materials (Figure 4). This is a relatively
new treatment technology which enhances biodegradation by increasing oxygen transfer to the
ground water while promoting the physical removal of organics by direct volatilization. Air
sparging has been used extensively in Germany since 1985 but was not introduced to the United
States until recently.
When air sparging is applied, the result is a complex partitioning of contaminants between
the adsorbed, dissolved, and vapor states. Also, a complex series of removal mechanisms are
introduced, including the removal of volatiles from the unsaturated zone, biodegradation, and the
partitioning and removal of volatiles from the fluid phase. The mechanisms responsible for removal
are dependent upon the volatility of the contaminants. With a highly volatile contaminant, for
example, the primary partitioning is into the vapor phase, and the primary removal mechanism is
through volatilization. By contrast, contaminants of low volatility partition into the adsorbed or
dissolved phase, and the primary removal mechanism is through biodegradation.
Water
Table
Residual
Saturation
Cajillary
ringe
Ground Water
Dissolved
Contaminants
itur
Figure 4. Contaminant locations treated by air sparging.
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One of the problems in applying air sparging is controlling the process. In either bioventing
or ground-water extraction, the systems are under control because contaminants are drawn to the
point of collection. By contrast, air sparging systems cause water and contaminants to move away
from the point of injection which can accelerate and aggravate the spread of contamination.
Changes in lithology can profoundly affect both the direction and velocity of air flow. A second
problem in air sparging is accelerated vapor travel. Since air sparging increases the vapor pressure
in the vadose zone, any exhausted vapors could be drawn into receptors such as basements. As a
result, in areas with potential vapor receptors, air sparging should be done with vent systems which
allow an effective means of capturing sparged gases.
As with any technology, there are limitations to the utility and applicability of air sparging.
The first is associated with the type of contaminants to be removed. For air sparging to work
effectively, the contaminant must be relatively volatile and relatively insoluble. If the contaminant
is soluble and nonvolatile, it must be biodegradable. The second limitation to the use of air sparging
is the geological character of the site. The most important geological characteristic is the
homogeneity of the site. If significant stratification is present, there is a danger that sparged air
could be held below an impervious layer and spread laterally, thereby resulting in the spread of
contamination.
Another constraint of concern is depth related. There is both a minimum and maximum
depth for a sparge system. A minimum depth of 4 feet, for example, may be required for a sufficient
thickness to confine the air and force it to "cone-out" from the injection point. A maximum depth
of 30 feet might be required from the standpoint of control. Depths greater than 30 feet make it
difficult to predict where the sparged air will travel.
ALTERNATE ELECTRON ACCEPTORS
Bioremediation using electron acceptors other than oxygen is potentially advantageous for
overcoming the difficulty in supplying oxygen for aerobic processes. Nitrate, sulfate, and carbon
dioxide are attractive alternatives to oxygen because they are more soluble in water, inexpensive,
and nontoxic to microorganisms. The demonstration of this technology in the field is limited,
therefore, its use as an alternate electron acceptor for bioremediation must be viewed as a
developing treatment technology. Figure 5 illustrates the location of contaminants that may be
remediated by introduction of alternate electron acceptors.
Some compounds are only transformed under aerobic conditions, while others require
strongly reducing conditions, and still others are transformed in both aerobic and anaerobic
environments. In the absence of molecular oxygen, microbial reduction reactions involving organic
contaminants increase in significance as environmental conditions become more reducing. In this
environment, some contaminants are reduced by a biological process known as reductive
dehalogenation. In reductive dehalogenation reactions, the halogenated compound becomes the
electron acceptor. In this process, a halogen is removed and is replaced with a hydrogen atom.
Bioremediation with alternate electron acceptors involves the stimulation of microbial
growth by the perfusion of electron donors, electron acceptors, and nutrients through the formation.
Addition of alternate electron acceptors other than nitrate for bioremediation has not been
documented at field scale but has been widely studied at laboratory scale. Nitrate as an electron
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Water
Table
Residual
Saturation
Capillary
.Fringe
Dissolved
Contaminants
Figure 5. Contaminant locations treated by alternate electron acceptors.
acceptor has been used for bioremediation of benzene, toluene, ethylbenzene, and xylenes in ground
water and on aquifer solids. As for other in-situ remediation technologies, formations with
hydraulic conductivities of 10"4 cm/sec (100 fiYsec) or greater are most amenable to bioremediation.
The combination of an anaerobic process followed by an aerobic process has promise for the
bioremediation of highly chlorinated organic contaminants. Generally, anaerobic microorganisms
reduce the number of chlorines on a chlorinated compound via reductive dechlorination, and
susceptibility to reduction increases with the number of chlorine substitutes. Conversely, aerobic
microorganisms are more capable of transforming compounds with fewer chlorinated substitutes.
With the removal of chlorines, oxidation becomes more favorable than does reductive dechlorination.
Therefore, the combination of anaerobic and aerobic processes has a potential utility as a control
technology for chlorinated solvent contamination.
NATURAL BIOREMEDIATION
The basic concept behind natural bioremediation is to allow naturally occurring
microorganisms to degrade contaminants that have been released into the subsurface. It is not a
"no action" alternative, as in most cases it is used to supplement other remediation techniques. In
some cases, only the removal of the primary source may be necessary. In others, conventional
ground-water remediation techniques such as pump and treat may be used to reduce contaminant
concentrations within the aquifer.
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Natural bioremediation is capable of treating contaminants aerobically in the vadose zone
and at the margins of plumes (Figure 6), where oxygen is not limiting. Some sites have shown that
anaerobic bioremediation processes also occur naturally and can significantly reduce contaminant
concentration on aquifer solids and in ground water. Benzene, toluene, ethylbenzene, and xylene
can be removed anaerobically in methanogenic or sulfate-reducingenvironments; highly chlorinated
solvents can undergo reductive dechlorination in anaerobic environments.
While there are no "typical" sites, it may be helpful to consider a hypothetical site where a
small release of gasoline has occurred from an underground storage tank (Figure 7). Rainfall
infiltrating through the hydrocarbon-contaminated soil will leach some of the more soluble
components including benzene, toluene, and xylenes. As the contaminated water migrates
downward through the unsaturated zone, a portion of the dissolved hydrocarbons may biodegrade.
The extent of the biodegradation will be controlled by the size of the spill, the rate of downward
movement, and the appropriateness of requisite environmental conditions. Dissolved hydrocarbons
that are not completely degraded in the unsaturated zone will enter the saturated zone and be
transported downgradient within the water table where they will be degraded by native
microorganisms to an extent limited by available oxygen and other subsurface conditions. The
contaminants that are not degraded will move downgradient under anaerobic conditions. As the
plume migrates, dispersion will mix the anaerobic water with oxygenated water at the plume
fringes. This is the region where most natural aerobic degradation occurs.
One of the major factors controlling the use of natural bioremediation is the acceptance of
this approach by regulators, environmental groups, and the public. The implementation of these
systems differs from conventional techniques in that a portion of the aquifer is allowed to remain
Water
Table
Residual
Saturation
Ground Water
Dissolved
Contaminants
Figure 6. Contaminant locations treated by aerobic natural bioremediation.
9
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Aerobic - Uncontammated Ground Water
Figure 7. Profile of a typical hydrocarbon plume undergoing natural bioremediation.
contaminated. This results in the necessity of obtaining variances from regulations, and some type
of risk evaluation is usually required. Even when public health is not at risk, adjoining land owners
may have strong concerns about a contaminant plume migrating under and potentially impacting
their property. Therefore, control of plume migration at these sites, usually utilizing some type of
hydraulic system, is often necessary. Although natural bioremediation imposes few costs other than
monitoring and the time for natural processes to proceed, the public may perceive that this is a "no
action" alternative. These various factors may generate opposition to selecting natural bioremediation
rather than conventional technologies.
There is almost no operating history to judge the effectiveness of natural bioremediation. In
addition, there are currently no reliable methods for predicting its effectiveness without first
conducting extensive field testing. This is often the primary reason why natural bioremediation is
not seriously considered when evaluating remedial alternatives. At many low priority sites,
regulators may have assumed that natural bioremediation would control the migration of
dissolved contaminants. Often, these sites have not been adequately characterized nor have they
been monitored to determine the effectiveness of this remediation technology. At present, there are
no well-documented, full-scale investigations of natural bioremediation, but there is a considerable
amount of ongoing research concerning the processes which drive this potentially effective
remediation alternative.
INTRODUCED ORGANISMS
Historically, the movement of microorganisms in the subsurface was first discussed in the
mid- 1920s in relation to the enhanced recovery of oil by the production of biological surfactants and
10
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gases. Later, the transport of bacteria through soil was studied to measure the effectiveness of soil-
based sewage treatment facilities such as pit latrines and septic tanks in terms of the removal of
pathogens. In recent years, research has been directed toward the introduction of microorganisms
to soil and ground water to introduce specialized metabolic capabilities, to degrade contaminants
which resist the degradative processes of indigenous microflora, or when the subsurface has been
sterilized by contaminants. In these attempts to introduce microorganisms to the subsurface, it is
often difficult to differentiate their activities from indigenous populations. The use of introduced
microorganisms has proven most successful in surface bioreactors when treating extracted ground
water in closed-loop recirculation systems.
For added organisms to be effective in contaminant degradation, they must be transported
to the zone of contamination, attach to the subsurface matrix, survive, grow, and retain their
degradative capabilities. There are a number of phenomena which affect the transport of microbes
in the subsurface including grain size, cracks and fissures, removal by sorption in sediments high
in day and organic matter, and the hydraulic conductivity. Many other factors affect the movement
of microorganisms in the subsurface including their size and shape, concentration, flow rate, and
survivability.
The use of microorganisms with specialized capabilities to enhance bioremediation in the
subsurface is an undemonstrated technique. However, research has been conducted to determine
the potential for microbial transport through subsurface materials, public health effects, and
microbiaJ enhanced oil recovery.
SUMMARY
This report has been prepared by leading soil and ground-water remediation scientists in
order to present the latest technical, institutional, and cost considerations applicable to subsurface
remediation systems. It is aimed at scientists, consultants, regulatory officials, and others who are,
in various ways, working to achieve efficient and cost-effective remediation of contaminants in the
subsurface environment.
The document contains detailed information about the processes, applications, and
limitations of using remediation technologies to restore contaminated soil and ground water. Field
tested as well as new and innovative technologies are discussed. In addition, site characterizations
requirements for each remediation technology are discussed along with the costs associated with
their implementation. A number of case histories are presented, and knowledge gaps are pointed
out in order to suggest areas for which additional research investigations are needed.
11
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SECTION 1
1.1. INTRODUCTION
The purpose of this report is to provide the reader with a detailed background of
the fundamentals involved in the bioremediation of contaminated surface soils,
subsurface materials, and ground water. A number of bioremediation technologies are
discussed along with the biological processes driving those technologies. The application
and performance of these technologies are also presented. These discussions include
in-situ remediation systems, air sparging and bioventing, use of electron acceptors
alternate to oxygen, natural bioremediation, and the introduction of organisms into the
subsurface. The contaminants of major focus in this report are petroleum hydrocarbons
and chlorinated solvents.
Location of the contamination in the subsurface is critical to the implementation
and success of in-situ bioremediation. Also important to success is the chemical nature
and physical properties of the contaminant(s) and their interactions with geological
materials.
In the unsaturated zone, contamination may exist in four phases (Huling and
Weaver, 1991): (1) air phase - vapor in the pore spaces; (2) adsorbed phase - sorbed to
subsurface solids; (3) aqueous phase - dissolved in water; and (4) liquid phase -
nonaqueous phase liquids (NAPLs). Contamination in saturated material can exist as
residual saturation sorbed to the aquifer solids, dissolved in the water, or as a NAPL.
Contaminant transport occurs in the vapor, aqueous, and NAPL phases. The
interactions between the physical properties of the contaminant influencing transport
(density, vapor pressure, viscosity, and hydrophobicity) and those of the subsurface
environment (geology, aquifer mineralogy and organic matter, and hydrology) determine
the nature and extent of transport.
NAPL existing as a continuous immiscible phase has the potential to be mobile,
resulting in widespread contamination. Residual saturation is the portion of the bulk
liquid retained by capillary forces in the pores of the subsurface material; the NAPL is no
longer a continuous phase but exists as isolated, residual globules. Residual phase
saturation will act as a continuous source of contamination in either saturated or
unsaturated materials due to dissolution into infiltrating water or ground water, or
volatilization into pore spaces.
Liquids less dense than water, such as petroleum hydrocarbons, are termed light
nonaqueous phase liquids (LNAPLs). LNAPLs will migrate vertically until residual
saturation depletes the liquid or until the capillary fringe is reached (Figure 1.1). Some
spreading of the bulk liquid will occur until the head from the infiltrating liquid is
sufficient to penetrate to the water table. The hydrocarbons will spread laterally and float
on the surface of the water table, forming a mound that becomes compressed into a
spreading lens due to upward pressure of the water (Hinchee and Reisinger, 1987).
Fluctuations of the water table due to seasonal variations, pumping, or recharge can
result in movement of bulk liquid further into the subsurface with significant residual
contamination present beneath the water table. The more soluble constituents will
dissolve from the bulk liquid into the water and will be transported with the migrating
ground water.
1-1
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Water,
Table
Residual
Saturation
Capillary
Fringe
Dissolved
ontaminants
Figure 1.1. Distribution of petroleum hydrocarbons in the subsurface.
Vertical migration of dense nonaqueous phase liquids (DNAPLs) will continue
through soils and unsaturated materials under the forces of gravity and capillary
attraction until the capillary fringe or a zone of lower permeability is reached. The bulk
liquid spreads until sufficient head is reached for penetration into the capillary fringe to
the water table. Because the density of chlorinated solvents is greater than that of water,
DNAPLs will continue to sink within the aquifer until an impermeable layer is reached
(Figure 1.2). The chlorinated solvents will then collect in pools or pond in depressions on
top of the impermeable layer. DNAPL contamination in heterogeneous subsurface
environments (Figure 1.3) is difficult to both identify and remediate.
Bioremediation of ground waters, aquifer solids, and unsaturated subsurface
materials is widely practiced for contaminants derived from petroleum products.
Currently, the most important techniques for bioremediating petroleum-derived
contaminants are based on enhancement of indigenous microorganisms by delivery of an
appropriate electron acceptor plus nutrients to the subsurface. These techniques are
in-situ bioremediation, bioventing, and air sparging; natural bioremediation of
petroleum hydrocarbons is also discussed. This paper presents sections devoted to each
of the above-mentioned techniques authored by experts actively engaged in
bioremediation and research. The sections are: Section 2, In-situ Bioremediation of Soils
and Ground Water Contaminated With Petroleum Hydrocarbons; Section 3, Bioventing of
Petroleum Hydrocarbons; Section 4, Treatment of Petroleum Hydrocarbons in Ground
Water By Air Sparging; Section 7, In-situ Bioremediation Technologies for Petroleum-
Derived Hydrocarbons Based on Alternate Electron Acceptors (other than molecular
oxygen); and Section 9, Natural Bioremediation of Hydrocarbon-Contaminated Ground
Water.
1-2
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Figure 1J2. Migration of DNAPL through the vadose zone to an impermeable boundary in
relatively homogenous subsurface materials (Ruling and Weaver, 1991).
CLAY
After. Waterloo Centre for Groundwater Research, 1989
Figure 1.3. Perched and deep DNAPL reservoirs from migration through heterogeneous
subsurface materials (Hiding and Weaver, 1991).
1-3
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Chlorinated solvents are more difficult to bioremediate than petroleum
hydrocarbons, and bioremediation efforts are still in the research and development stage.
Biological processes for most chlorinated compounds, whether aerobic or anaerobic,
require the presence of a primary substrate for cometabolism. Both enhanced and
natural bioremediation of chlorinated compounds are widely investigated in laboratory,
pilot-scale, and field-scale studies. Results of these efforts are presented in the following
sections: Section 5, Ground-water Treatment for Chlorinated Solvents; Section 6,
Bioventing of Chlorinated Solvents for Ground-water Cleanup through Bioremediation;
Section 8, Bioremediation of Chlorinated Solvents Using Alternate Electron Acceptors;
and Section 10, Natural Bioremediation of Chlorinated Solvents.
The introduction of microorganisms to the subsurface for bioremediation
purposes is discussed in Section 11, Introduced Organisms for Subsurface
Bioremediation. Although not considered a successful technique at this time due to
concerns about survivability of introduced microorganisms, this method may someday be
useful at sites sterilized by contamination.
Discussed in each section are basic biological and nonbiological processes
affecting the fate of the compounds of interest, documented field experience,
performance, repositories of expertise, primary knowledge gaps and research
opportunities, favorable and unfavorable site conditions, regulatory acceptance, special
requirements for site characterization, and problems encountered with the technology.
Although the focus of this paper is bioremediation, remediation of most sites will require
use of other technologies not discussed, such as pump and treat, soil washing, etc. The
place of bioremediation in the cleanup of hazardous waste sites is still evolving, and
evaluation of its effectiveness is under investigation by regulators, researchers, and
remediation firms.
Hinchee, R.E., and H.J. Reisinger. 1987. A practical application of multiphase
transport theory to ground water contamination problems. Ground Water
Monitoring Review. 7(l):84-92.
Ruling, S.G., and J.W. Weaver. 1991. Dense Nonaqueous Phase Liquids. Ground Water
Issue Paper. EPA/540/4-91-002. Robert S. Kerr Environmental Research
Laboratory. Ada, Oklahoma.
1-4
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SECTION 2
IN-SITU BIOREMEDIATION OF SOILS AND GROUND WATER
CONTAMINATED WITH PETROLEUM HYDROCARBONS
Robert D. Morris
Eckenfelder, Inc.
227 French Landing Drive
Nashville, Tennessee 37228
Telephone: (615)255-2288
Fax: (615)25^^332
2.1. INTRODUCTION
This chapter discusses the use of in-situ bioremediation processes to treat ground
water and aquifer solids contaminated with petroleum hydrocarbons under aerobic
conditions using indigenous microorganisms. Natural bioremediation, the use of nitrate
as an electron acceptor, introduced organisms, air sparging to provide oxygen, and
treatment of the unsaturated zone are all addressed in other sections. Still other sections
address the same topics for chlorinated solvents. Discussions of issues covered in the
other sections are included in this document only to the extent that is necessary to
adequately address some topics.
2£. FUNDAMENTAL PRINCIPLES
Bioremediation, whether of excavated soils, aquifer solids or unsaturated
subsurface materials, is the use of microorganisms to convert harmful chemical
compounds to less harmful chemical compounds in order to effect remediation of a site
or a portion of a site. The microorganisms are generally bacteria but can be fungi.
Indigenous bacteria that can degrade a variety of organic compounds are present in
nearly all subsurface materials. The use of introduced microorganisms, as discussed in
Section 11, has not been shown to be of significant benefit. Microorganisms require
certain minerals, usually referred to as nutrients, and an electron acceptor. While a
variety of minerals such as iron, magnesium, and sulfur are required by the
microorganisms, it is usually only necessary to add nitrogen and phosphorus sources.
The other minerals are needed in trace amounts, and adequate amounts are normally
found in most ground waters and subsurface materials. The most common electron
acceptor used in commercial bioremediation processes is oxygen. Other electron
acceptors such as nitrate can be used for some contaminants such as most aromatic
hydrocarbons, although restrictions may apply to the levels of nitrate that may be
introduced to ground water. This topic is covered in Sections 7 and 8.
Bioremediation systems supply nitrogen, phosphorus, and/or oxygen to bacteria
that are present in the contaminated aquifer solids and ground water.
Stoichiometrically, it would take approximately three pounds of oxygen to convert one
pound of hydrocarbon to carbon dioxide and water. Experience with wastewater
treatment indicates that the expressed oxygen requirements are usually near half of the
stoichiometric amount. Conversely, some of the oxygen introduced for biooxidation of the
contaminants may be consumed by other reactions or is lost through inefficient
2-1
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distribution. As a result, first approximations of oxygen requirements are typically
based on the three-to-one ratio.
Nutrient requirements are less easily predicted. If all of the hydrocarbon mass
were converted to cell material, the nutrient requirements based on the mass of
hydrocarbon to be consumed would be approximated by a ratio of carbon to nitrogen to
phosphorous of 100:10:1. The nutrient requirement will be less than this ratio to the
extent that direct conversion of hydrocarbons to carbon dioxide and water occurs.
Nutrients already exist in the subsurface materials and ground water, nitrogen is fixed
by the indigenous bacteria, and nutrients are recycled from dead bacteria. However,
adsorption of nutrients by geologic materials can substantially increase the amount of
nutrients that have to be introduced in order to distribute nutrients across the
contaminated zone. Adsorption may be modest in clean sands but may consume most of
the nutrients in silts and clays, especially if the solids have a high natural organic
content.
In some instances, nitrogen, phosphorus, or oxygen may be present in sufficient
quantities to support degradation of the constituents of interest. In those cases, only one
or two of the three elements would need to be added. This is most likely to be the case
where low levels of contamination are present. In such cases there may be sufficient
nitrogen and phosphorus sources present and only oxygen needs to be provided.
In-situ bioremediation systems for aquifers typically consist of a combination of
injection wells (or galleries or trenches) and one or more recovery wells as shown in
Figure 2.1. In most instances, the recovered ground water will be treated prior to
amendment with nutrients and/or an oxygen source and reinjection. Ground-water
treatment has frequently consisted of an air stripper tower or activated carbon but may
incorporate an oil/water separator, a biological treatment unit, an advanced oxidation
unit, or combinations of treatment units. Treatment of the ground water is likely to be
necessary based on regulatory considerations and is beneficial from a process economics
perspective when the recovered ground water contains more than a few ppm of
biodegradable substances. When the recovered ground water contains low levels of
readily degradable constituents, the biodegradable constituents will be degraded within a
short distance of the injection point and will not add significantly to the oxygen and
nutrient requirements.
Oxygen
Source
Nutrients
Ground-Waier
Treatment
Figure 2.1. Bioremediation in the saturated zone.
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Critical to the design of an in-situ bioremediation system are the ground-water
flow rate and flow path. Ground-water flow must be sufficient to deliver the required
amounts of nutrients and oxygen in a reasonable time frame. The amended ground
water should sweep the entire area requiring treatment, and the recovery wells should
capture the injected ground water to prevent migration outside the designated treatment
zone. In order to ensure that adequate control can be maintained over the ground water,
usually only a portion of the recovered ground water is reinjected. The other portion is
then discharged by an acceptable method.
In-situ bioremediation systems can be integrated with other remediation
technologies either sequentially or simultaneously (Morris et al., 1990). If free phase
hydrocarbons are present, a free phase recovery system such as a dual phase pump or
skimmer should be used to reduce the mass of the free phase hydrocarbons prior to
implementation of bioremediation. In-situ vapor stripping (ISVS) systems (U.S. EPA,
1991a) can serve to both physically remove volatile hydrocarbons and to provide oxygen for
biodegradation. ISVS systems can remove residual free phase hydrocarbons as well as
constituents adsorbed to both unsaturated materials and aquifer solids exposed during
periods of lower water table levels. Depending on the air flow and nutrient availability,
hydrocarbons in subsurface solids (including the capillary fringe just above the water
table) will undergo biooxidation. The combined mechanisms can serve to reduce
significantly the mass of hydrocarbons. As a result, the time and cost of providing
nutrients and oxygen through injection of amended ground water may be substantially
reduced.
£3. HISTORICAL PERSPECTIVE
The use of biooxidation for environmental purposes has been practiced for many
years. Biological processes have been used to treat wastewater for nearly sixty years
(Eckenfelder, 1967). Activated sludge and suspended growth systems have become
commonplace for waste treatment in many industries and for municipal waste. This
use of biological degradation of organic compounds led to the generation of a wide body of
information regarding biodegradability of specific compounds and classes of chemicals,
nutrient and electron acceptor requirements, and oxidation mechanisms. Land
treatment processes for wastewater, refinery, and municipal wastes have also been
practiced for several decades and have generated additional information on nutrient
requirements, degradation rates, and other critical parameters affecting biological
oxidation (Overcash and Pal, 1979).
In the 1970s, several studies sponsored by the American Petroleum Institute were
conducted using the method developed by Richard L. Raymond, Sr., then at Sun Tech., to
biologically degrade hydrocarbons in aquifers (Bauman, 1991). This method involved the
recovery of ground water with treatment using an air stripper tower and subsequent
reinjection following amendment with nitrogen and phosphorus sources (ammonium
chloride and sodium orthophosphate salts). Oxygen was generally provided by sparging
air at the bottom of the injection well (Raymond et al., 1976). Many of these early tests
were conducted prior to the enactment of state mandated cleanup levels. As a result,
these tests demonstrated that in-situ bioremediation could reduce the levels of petroleum
hydrocarbons in an aquifer, but did not generate documentation of the ability to reach the
ground-water quality standards that are necessary in today's regulatory environment.
It was soon recognized that the rate at which oxygen could be introduced by
sparging air in a ground-water injection well would limit the effectiveness of the
technology. Hydrogen peroxide was identified as a potential method of introducing
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oxygen (Brown et al., 1984). The solubility in water limits the amount of dissolved oxygen
that can be delivered from air to 8 to 10 ppm, unless injection occurs substantially below
the water table. Use of pure oxygen in place of air can increase the rate of introduction of
oxygen fivefold. Hydrogen peroxide, which decomposes to oxygen and water, is
completely soluble in water. Practical considerations, including toxicity towards
bacteria, limit hydrogen peroxide concentrations to 100 to 1,000 ppm. Hydrogen peroxide
could thus theoretically provide oxygen at 5 to 50 times faster than could sparging air in
the injection wells and should result in shorter remediation times. However, the
efficiency of delivering oxygen by this method has been quite variable even when favorable
results were obtained from laboratory screening tests (Lawes, 1991; Huling et al., 1990;
Hinchee and Downey, 1988; and Flathman et al., 1991). Further, as microbial
populations decrease as a function of decreasing food source (the contaminants),
tolerance toward hydrogen peroxide may also decrease. As a result, hydrogen peroxide
may not be the most appropriate oxygen source for many sites.
More recently, many practitioners have used ground-water sparging techniques
to introduce oxygen, as discussed in Section 4. In this method, wells or drive points are
screened over a narrow interval several feet below the water table. Air is forced into the
aquifer under pressure resulting in saturation of the ground water in the vicinity of the
injection point. This procedure can also strip volatiles from the ground water and can
cause increased rates of migration. Generally, it should be used in conjunction with
ground-water capture and/or in-situ vapor stripping systems.
Other approaches to providing an electron acceptor include the use of surfactants
to create microbubbles (Michelsen et al., 1990), on site generation of oxygen (Prosen et al.,
1991), or use of alternate electron acceptors such as nitrate, as discussed in Sections 7
and 8.
During the time when technology to deliver oxygen was evolving, nutrient sources
were being developed, and an understanding of hydrogeological considerations was
evolving. Initially, the salts blend developed by Richard L. Raymond, Sr. was used
(approximately equal amounts of ammonium chloride and sodium orthophosphate)
(Raymond et al., 1978). Some practitioners have changed to potassium salts to reduce the
potential for swelling of clays and to tripolyphosphate which will solubilize rather than
precipitate iron, calcium, and magnesium (Brown and Morris, 1988).
The need for detailed understanding and control of the site hydrogeology has long
been recognized. Many early designs were developed with limited aquifer hydrology test
data, and well locations were determined using logic or limited calculations to predict
the areas of influence of injection and recovery wells. It has become more common to
conduct aquifer tests and use computer models (analytical models suffice for most
smaller sites) to locate injection and recovery wells (Falatico and Norris, 1990). This
approach can be used to predict remediation times based on oxygen and nutrient
demands estimated from contaminant concentrations and ground-water recirculation
rates. Models can also be used to evaluate the feasibility of bioremediation at a particular
site (Rifai and Bedient, 1987). For readily degradable substances, modeling efforts are
more beneficial than laboratory treatability studies and may be less costly. More
sophisticated models that also address contaminant and nutrient transport and oxygen
uptake are also available (Borden, 1991).
In the mid-1980s, there were few companies with experience in bioremediation of
aquifers or soils. Since that time many companies have utilized bioremediation
technologies, although claims of experience are frequently overstated. In the last few
years this technology has gained the support of the U.S. EPA, as evidenced by the many
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supportive public statements made by Administrator Reilly, the support of many of the
U.S. EPA laboratories, and the creation of U.S. EPA sponsored research, committees,
and seminars that have promoted the use of bioremediation.
2.4. REPOSITORIES OF EXPERTISE
There now exists a large number of organizations and people who are
knowledgeable about in-situ bioremediation. However, many important practical
findings have not been adequately shared. Several environmental companies have staff
who are experienced and/or knowledgeable in the application of this technology. Some
companies have tended to specialize in aboveground systems and may have limited
experience in remediation of aquifers. Many large corporations, especially the major oil
and chemical companies, have also developed in-house expertise. Several U.S. EPA
laboratories (e.g., Robert S. Kerr Environmental Research Laboratory, Ada, Oklahoma),
DOD groups (e.g., U.S. Air Force) and universities (e.g., Rice University) have conducted
laboratory and, in some cases, field studies on in-situ bioremediation. Typically, some
environmental companies, some large site owners, and some EPA laboratories have
more experience with the actual application of the technology, while some other groups
have more in-depth understanding of the science involved.
25. GENERAL DESIGNS
The most common design is a system that uses a combination of injection and
recovery wells, as shown in Figure 2.1. Recovered ground water is treated, typically
using an air stripper tower, amended with nutrients, and reinjected. Oxygen is supplied
using air sparging in the injection well or by introduction of hydrogen peroxide.
Amended ground water can also be introduced through injection galleries or
trenches. This approach is most likely to be used in shallow aquifers.
The above systems can be modified to introduce oxygen by using air spargers
located directly within the aquifer, as discussed in Section 5, either in combination with
in-situ vapor stripping or using the unsaturated zone as a biofilter.
For shallow aquifers with sandy material, nutrients can be introduced from the
surface, allowing percolation of rain water or added water to carry the nutrients into the
aquifer. If oxygen is introduced by air sparging, ground-water recovery systems are only
required to prevent migration of contaminated ground water.
The design of the ground-water recirculation system is best done using a ground-
water flow model (Falatico and Norris, 1990). For most sites, a two-dimensional
analytical flow model will be sufficient. The model will allow several design concepts to
be evaluated and the most favorable selected. These models can be more effective than
laboratory treatability studies to determine feasibility and can be used to make midcourse
modifications to operating conditions.
Operating plans should include maintenance of wells and equipment, monitoring
schedules, reporting schedules, and milestones for evaluation of system performance so
that modifications in operating procedures can be made and, if necessary, additional
wells installed.
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Each of the systems should incorporate an appropriate monitoring system.
Monitoring wells are required to determine the distribution of nutrients and oxygen, and
to monitor pH and other ground-water chemistry parameters, ground-water elevations,
and changes in contaminant concentrations.
2.6. LABORATORY TESTING
Laboratory tests can be used as screening tests to determine site feasibility, as
treatability tests to determine the rate and extent of biodegradation that might be attained
during remediation, and as engineering tests to provide design criteria (U.S. EPA,
1991b).
Screening tests include pH and plate counts to determine if existing conditions are
favorable to microbial growth. Respirometer tests, which measure oxygen uptake but do
not normally measure disappearance of the contaminant(s), provide confirmation that
the microbial population is metabolically active. These tests can be run under a number
of nutrient conditions to provide an indication of nutrient effects.
Treatability studies are generally conducted with soil/ground-water slurries.
Several conditions are usually tested including unmodified microcosms, nutrient
amended microcosms, and biologically inhibited conditions. These tests can measure
the rate of change of the constituents of concern as well as changes in pH and microbial
populations. The tests provide data on the rate and extent of conversion of contaminants.
During bioremediation of hydrocarbons in aquifers, the rate of degradation is usually
controlled by the rate of supply of nutrient and oxygen. Under these conditions,
laboratory rate data do not extrapolate directly to the field. However, the laboratory data
on the rate and extent of removals of hazardous constituents are important for the
heavier hydrocarbons, such as heavy crude oil, bunker oil, or coal gas tars. Removal of
compounds from these materials is often limited by the reaction kinetics of the
microorganisms rather than the rate of supply of some essential nutrient. The extent of
biodegradation of oily phase hydrocarbons to microbial biomass or metabolic end
products is very site specific.
2.7. CONTAMINATION LIMITS
The range of contaminant concentrations that are amenable to bioremediation
depends on a number of factors. The distribution of contamination may allow remedy
through in-situ bioremediation alone. However, if contamination is distributed both
above and below the water table, it may be more practical to use other remediation
technologies or to combine in-situ bioremediation with other technologies such as free
phase recovery, ground-water sparging, and in-situ vapor stripping.
As a class, petroleum hydrocarbons are biodegradable (Gibson, 1984). The lighter,
more soluble members are generally biodegraded more rapidly and to lower residual
levels than are the heavier, less soluble members. Thus monoaromatic compounds such
as benzene, toluene, ethylbenzene, and the xylenes are more rapidly degraded than the
two-nng compounds such naphthalene, which are in turn more easily degraded than
the three-, four- and five-ring compounds. The same is true for aliphatic compounds
where the smaller compounds are more readily degraded than the larger compounds.
Branched hydrocarbons degrade more slowly than the corresponding straight-chain
hydrocarbons.
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Typically, site remediation is concerned with commercial blends of petroleum
hydrocarbons. As for individual compounds, the lighter blends are more readily
degraded than the heavier blends. For example, gasoline can be biodegraded to low levels
under many conditions. Heavier products such as number 6 fuel oil or coal tar, however,
contain many higher molecular weight compounds such as five-ring aromatic
compounds. These mixtures degrade much more slowly than gasoline and, as a result,
significantly lower rates and extents of biodegradation should be anticipated.
Polyaromatic hydrocarbons are present in heavier petroleum hydrocarbon blends
and particularly in coal tars, wood treating chemicals, and refinery wastes. These
compounds have only limited solubility in water, adsorb strongly to subsurface
materials, and degrade at rates much slower than monoaromatic hydrocarbons and
most aliphatic and alicyclic compounds found in refined petroleum hydrocarbon
products. Because of their low solubility and strong adsorption to solids, their availability
for degradation is often the limiting factor in treatment (Brubaker, 1991). They are more
likely to be biodegraded in mixtures with more soluble and thus more readily degradable
hydrocarbons because the more readily degradable species will support a larger
microbial population (McKenna and Heath, 1976).
Because petroleum hydrocarbons are frequently found in the presence of other
organic constituents, it is necessary to consider the degradability of other classes of
compounds. Nonchlorinated solvents used in a variety of industries are generally
biodegradable. Alcohols, ketones, esters, carboxylic acids and esters, particularly the
lower molecular weight analogs, are readily biodegradable, but may be toxic at high
concentrations due to their high water solubilities. Lightly chlorinated compounds such
as chlorobenzene (U.S. EPA, 1986), dichlorobenzene, chlorinated phenols and the lightly
chlorinated PCBs are typically biodegradable under aerobic conditions. The more highly
chlorinated analogs are more recalcitrant to aerobic degradation but are more
susceptible to degradation under anaerobic conditions.
Several of the common chlorinated solvents (chlorinated ethanes and ethenes) can
be degraded under aerobic conditions, as discussed in Section 5. This requires the
addition of a cometabolite unless certain chemical species such as toluene or phenol are
present with the chlorinated species. It is reasonable to expect that some aerobic
biodegradation of chlorinated solvents will occur in the presence of petroleum
hydrocarbon blends, particularly those containing appreciable amounts of toluene. This
is, however, a very site-specific phenomenon and one for which there is not enough
documentation to make reliable predictions. Further, many chlorinated solvents can
inhibit biodegradation of petroleum hydrocarbons if the solvent species is present at high
enough concentrations. Data on biodegradability and other properties of environmental
interest are available from several handbooks (Montgomery, 1991; Montgomery and
Wilkom, 1990; Howard, 1969 and 1990; and Verschueren, 1983).
Petroleum hydrocarbons can generally be mineralized; i.e., converted to carbon
dioxide and water. The extent of conversion that is likely to occur is greatest for the
lighter molecular weight constituents. For gasoline, the extent of conversion is largely
limited by the efficiency and completeness of the distribution of nutrients and an electron
acceptor. For the heavier petroleum hydrocarbons, especially polynuclear aromatic
hydrocarbons (PAHs), the limiting factor may be the rate of solubilization, the release
from interstitial pore spaces, or the rate of degradation of the higher molecular weight
constituents.
Concentrations of contaminants that are toxic or large quantities of oily phase
material that do not permit penetration of nutrients and/or an electron acceptor are not
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appropriate for in-situ bioremediation. Toxicity seldom occurs where petroleum
hydrocarbons are the only contaminants. Toxicity can occur with some chlorinated
solvents and with very soluble compounds such as alcohols. The levels at which a
specific compound is toxic will be to some extent site specific as the microbial
communities have substantial capacity to adapt. Approximations of concentrations at
which specific compounds are toxic can be obtained from several handbooks
(Montgomery, 1991; Montgomery and Wilkom, 1990; Howard, 1989 and 1990; and
Verschueren, 1983).
Bioremediation is not generally applicable to metals but may incidentally mobilize
or immobilize various metals. Generally, the presence of metals has little direct effect on
the bioremediation process (Robert Norris, personal experience). While some metals
such as mercury can be toxic to bacteria, the microbial population frequently adapts to
the concentrations present. The effect of metals or organics on the microbial population
of a specific site can be tested through plating experiments or treatability tests.
When viscous materials such as the heavier fuel oil blends are present at
concentrations that prevent the flow of water and diffusion of nutrients and electron
acceptors, bioremediation will be impractical. The concentration at which this occurs
will vary with soil type but generally will be above 20,000 mg/kg.
Very high concentrations of contaminants will create very high oxygen and/or
nutrient demands. Meeting these demands might require excessively longer times and
higher costs than other technologies (Piontek and Simpkin, 1992). High levels of
contamination are a bigger problem with low or marginally permeable aquifers than
with highly permeable aquifers. In some cases, provision of oxygen through ground-
water sparging may provide a suitable approach to higher levels of contamination.
2& SITE CHARACTERIZATION
Two important aspects of site characterization frequently receive less attention
than they should. While implementation of this or any on-site or in-situ technology
requires delineation of the extent of contamination, including the presence and extent of
oily phase material, the concentrations of contaminants on aquifer solids is often
overlooked. The solubility of petroleum hydrocarbons is low and thus the preponderance
of the hydrocarbon mass is associated with the solids and not in the dissolved phase. For
sites contaminated with petroleum hydrocarbons, quantities associated with aquifer
solids are far more important than ground-water concentrations. Even when numerous
samples of both cores and ground waters have been analyzed by currently available
standard analytical methods, the total mass of hydrocarbons may not be accurately
determined. The total mass calculated from component specific analyses for volatiles,
semivolatiles, polynuclear aromatics, base neutrals, and acid extractables do not account
for the total mass. Nonspecific analysis such as Total Petroleum Hydrocarbon (TPH)
analyses can measure components that are not of interest; e.g., asphalt particles, do not
measure the most volatile compounds, and can yield highly variable results as shown in
studies where split samples have been sent to different laboratories (Anonymous, 1992).
Even without analytical considerations, obtaining representative data is
sometimes difficult, particularly for sites with heterogeneous conditions and/or multiple
sources. Even with extensive sampling, it is quite likely that the total mass of
contaminants at a specific site will not be known within 50 percent; however, if analyses
of the aquifer solids are not conducted, the uncertainty can be an order of magnitude or
more.
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The second important aspect of site characterization that is frequently slighted is
site hydrogeology. Because the rate of remediation of petroleum hydrocarbons in
saturated materials is almost always controlled by the rate of distribution of the nutrients
and oxygen source, and thus the rate of ground-water recirculation, aquifer
hydrogeological properties are critical. For easily biodegraded materials such as
gasoline, it is more important to model the ground-water flow than it is to conduct
laboratory treatability studies. Site characterizations should include aquifer tests such
as 24-hour pump tests. Relatively simple analytical flow models can provide a good
approximation of ground-water recovery and injection capabilities and thus the
feasibility of providing nutrients and oxygen in an acceptable time frame.
It is necessary to identify nearby ground-water receptors in order to design a
capture system for the injected water that protects adjacent ground-water supplies and to
be able to evaluate the potential impact of residual nutrients that may discharge to
surface water.
The concentrations of other potential contaminants that are not biodegradable,
such as heavy metals, should be determined because bioremediation processes are not
likely to appreciably change the concentration of these species. If they are present above
regulated levels, it may be necessary to combine bioremediation with another technology
or select another remediation strategy.
Microbial populations offer an indication of whether site conditions will support
microbial activity that will degrade petroleum hydrocarbons. Tests can be made for
heterotrophic (total) microbial populations or for bacteria that can utilize the
contaminant of interest. This can be a useful tool to screen for conditions where bacteria
have been negatively impacted by the site conditions. Although failure of soils or aquifer
solids to contain a viable microbial community capable of degrading a range of petroleum
hydrocarbons is rare («1% of sites), early identification of such a problem is important.
2.9. FAVORABLE SITE CONDITIONS
2.9.1. Solubility
The more soluble hydrocarbons are readily biodegraded and can be partially
captured by recovery wells for surface treatment. Petroleum hydrocarbons are not
sufficiently soluble to be treated by pump and treat alone.
2.9.2. Volatility
Volatility does not affect biodegradation; however, the volatility of the contaminant
does determine if it can be treated by ground-water sparging combined with
biodegradation in the unsaturated zone, or with in-situ vapor stripping, or in-situ vapor
stripping combined with dewatering.
2.9.3. Viscosity
Highly viscous hydrocarbons are not as easily biodegraded because it is difficult to
establish contact among contaminant, bacteria, nutrients, and an electron acceptor.
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2.9.4. Tootitity
Contaminants may be toxic or inhibitory to the microbial community. Frequently,
the bacteria have adapted to the presence of these compounds. This can usually be
readily determined by performing plate counts of subsurface materials and ground-
water samples or conducting treatability tests to determine the effect of potential
toxicants on the rate and extent of biodegradation.
2.9.5. Permeability of Soils and Subsurface Materials
The greater the permeability, the easier it is to distribute nutrients and an electron
acceptor to the contaminated solids and ground water. Of course, these conditions also
tend to lead to greater extent of contamination. The importance of permeability increases
with the mass of contaminant to be addressed and the urgency of completing the
remediation process.
2.9.6. Soil Type
In addition to permeability, soil type also impacts the degree of adsorption of
contaminants and nutrients by the soils. Sand and gravel are the most favorable soil
types for nutrient transport; clays are the least favorable. Karst formations allow for
rapid recovery and introduction of amended ground water; however, prediction and
control of flow paths may be difficult or severely limited. The soil organic matter content
(e.g., humates) impacts the movement of petroleum hydrocarbons through the aquifer.
2.9.7. Depth to Water
Depth to ground water should be considered not so much as a favorable or
unfavorable characteristic but as a factor to be taken into consideration in designing a
system. The greater the depth to water, the greater head that can be provided at injection
points, and thus the greater the potential injection rates that can be obtained. Shallower
water tables limit the head that can be attained and are more favorable to the use of
injection galleries. Air sparging, when used as an oxygen source, also has the potential
to transfer volatiles to the unsaturated zone and thus the surface air. Efficient capturing
of these gases requires an adequate unsaturated interval if an in-situ vapor stripping
system is used or if the unsaturated materials are used as a biofilter. Significant depths
to water can add to the cost of installation, but will also add to the cost of other
alternatives as well.
2.9.8. Mineral Content
Calcium, magnesium, and iron can cause precipitation of nutrients and caking
in water lines. This can be minimized by using tripolyphosphates, which sequester these
minerals. However, tripolyphosphate will form precipitates with these minerals unless
present in amounts equal to or greater than a 1:1 molar ratio.
2.9.9. Oxidation/Reduction Potential
Iron can also be a problem because natural biooxidation of petroleum
hydrocarbons can consume nearly all of the available oxygen in the ground water. As a
result of these reduced conditions, ferric iron can serve as an electron acceptor for
anaerobic degradation of some hydrocarbons. In this process, ferric iron is reduced to
ferrous iron, which is more soluble. When oxygenated ground water is introduced into
the formation, the less soluble oxidized form of iron (Pe*3) will form and precipitate. This
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can reduce the permeability of the formation. (However, if the aquifer is maintained 'in
the oxidized state, further dissolution of iron should not occur. In The Netherlands,
ground water in the vicinity of production wells is routinely oxygenated to reduce the
amount of iron in the production well water.)
2.9.10. pH
Bioremediation is favored by near neutral pH values (6 to 8). However, in aquifers
where natural pH values are outside this range, biodegradation may proceed without
hindrance. Biodegradation appears to proceed quite well, for instance, in the New Jersey
Pine Barrens where pH values of 4.5 to 5 are common (Brown et al., 1991). Where the pH
has been shifted away from neutral by manmade changes, biodegradability is likely to be
impaired (Robert Norris, personal experience).
2.10. INFRASTRUCTURE AND INSTITUTIONAL ISSUES
One advantage of in-situ treatment systems is the ability to install and conduct
remediation with minimal disruption to the site. Implementation does, however, require
the installation of wells, transfer lines, aboveground systems for amending the injection
water with nutrients and an electron acceptor source and, if necessary, a treatment
system for removal or reduction of the contaminant in the recovered ground water prior
to reinjection and/or discharge to an alternate receptor. The size of the treatment area
will depend on the ground-water flow and the treatment design. In most cases the size
and appearance of the aboveground treatment infrastructure are acceptable.
The acceptance of in-situ bioremediation as a remediation technology by the
public and various regulatory agencies has been generally favorable over the last seven or
eight years and has improved significantly over the last two or three years with the
support of the U.S. EPA, many state agencies, as well as favorable publicity in trade
journals and the popular press. As an in-situ technology that is viewed as a natural
process that results in destruction rather than relocation of the contamination, in-situ
bioremediation meets many of the objectives of State and Federal agencies. Questions of
efficacy (biodegradability) and production of toxic intermediates have infrequently been
an issue with the treatment of petroleum hydrocarbons.
Issues tend to be mostly site-specific. One frequent issue is the discharge of
treated or untreated ground water. In many instances, several permits are required.
Work has been delayed because a particular permit was delayed or was altogether
unobtainable. In order to maintain hydraulic control over the aquifer, it is generally
necessary to reinject only a portion of the recovered ground water. Some states regulate
reinjection wells and galleries as Class V wells. The remaining water can be discharged
to a municipal sewer. Where sewer or water treatment systems are near or exceed
capacity, sewage discharge permits may not be obtainable.
If the remaining ground water is discharged to surface water, a National
Pollutant Discharge Elimination System (NPDES) permit is usually required. States in
arid regions usually require a special permit for extraction of ground water. While site
remediations conducted under Superfund allow work to proceed without formally
obtaining state and local permits, the standards and requirements of the permits must be
met.
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2.11. PERFORMANCE
The ability to meet relevant regulatory end points depends both on the end points
and the limits of the technology. End points can be State mandated levels, risk based
levels, Federal mandated levels, or Toxic Characteristic Leaching Potential (TCLP) based
levels. The targeted end points can vary significantly, and the specific levels set for a
given site often determine whether end points will be met by the specific technology
employed. Particularly troublesome are State regulations that set levels at or below the
detection limit or at background. Because it is a nonspecific analysis, using background
TPH levels as the remediation goal can create difficulties in interpreting data and lead to
misleading conclusions regarding the performance of the system.
Under ideal conditions, in-situ bioremediation can reduce petroleum hydrocarbon
levels to nondetectable levels (10 mg/kg). This is more easily obtained with the lighter
blends in permeable and homogeneous formations where placement of injection and
recovery wells (galleries, etc.) is unencumbered. Generally, for lighter petroleum
blends, the hardest regulatory end point to meet is the benzene limit. Although benzene
is highly biodegradable, MCLs for benzene are at least an order of magnitude lower than
for other specific light hydrocarbon constituents. As a result, if the benzene end point
can be reached, the level for the other components will most probably be met as well.
For heavier petroleum hydrocarbons, BTEX compounds (benzene, toluene,
ethylbenzene, and xylenes) may not be present in significant concentrations to be of
concern. Typically, TPH will be the target analysis to be met. The heavier the petroleum
mixture, the more probable there will be residuals of very slowly degraded components.
These components tend to have low water solubilities, which can limit their rate of
degradation. If TPH is the only criterion, the measurements will not determine which
petroleum hydrocarbon components have gone untreated. Compounds that are not of
environmental concern may contribute to reported TPH values and thus complicate
interpretation.
Polyaromatic hydrocarbons can be difficult to treat to the regulated levels. The
MCLs for many of these compounds are low because they are suspected carcinogens.
The rate of release of PAHs from subsurface solids may be too slow to support an active
microbial population and degradation rates may be impractically slow. Fortunately, the
degradability of these compounds is better in mixtures containing lower molecular
weight compounds found in many commercial petroleum products. Available data on
the limits of PAH degradation under in-situ bioremediation conditions are limited and
contradictory, and thus predictions of treatment limits are likely to be unreliable.
2.12. PROBLEMS
Inadequate characterization of a site can result in a bioremediation system being
underdesigned. If the total mass of contamination is underestimated, a specific design
will take longer to achieve the remediation goals than predicted from the available data.
If the site hydrogeology is not adequately characterized, the production rate of recovery
wells or, more likely, the rates at which injection wells can receive water may be
overestimated. If this latter situation occurs, it will take longer to provide the required
nutrients and electron acceptor. Since provision of nutrients and/or oxygen is frequently
the rate controlling step, the remediation may take proportionately longer.
The capacity of recovery wells, and particularly injection wells, tends to decrease
with time. The deterioration of injection wells can result from movement of fines,
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precipitation of minerals, or from excessive microbial growth in or in the immediate
vicinity of the screened interval of the injection well. Proper selection of a gravel pack
and installation and development of the wells will reduce the propensity for problems.
Mineral clogging of the well screen and formation can occur because the
chemistry of the injection water is different from that of the ground water. Typically,
ground water in an aquifer contaminated with petroleum hydrocarbons will have a low
oxidation potential because natural biodegradation will have utilized most of the
dissolved oxygen. Frequently this results in elevated dissolved minerals, especially iron.
Recovery and treatment of ground water typically introduces oxygen into the water even
if an oxygen source is not added. Reinjection of this water can result in precipitation of
iron and other metals when the injection water mixes with the ground water.
In some instances it may be preferable to discharge all of the recovered ground
water and reinject clean water from another source. Other sources of water that have
been used are city water and uncontaminated ground water from another part of the
same or adjacent aquifer.
Calcium, magnesium, and iron will form precipitates with orthophosphates
when orthophosphate salts are used as nutrients. The formation of precipitates should
be evaluated with laboratory tests during the design phase. The use of adequate levels of
tripolyphosphate salts can alleviate precipitation problems.
Reduced aquifer permeability can also result from swelling of clays if sodium salts
of phosphates are used as the nutrient source. In such materials potassium salts should
be used.
Biological growth can reduce permeability and/or restrict flow through well
screens. This can be addressed by periodically adding higher levels of hydrogen peroxide
and surging the wells.
Use of dilute hydrochloric acid to clean the wells may also work, particularly
when mineral deposits are the primary problem (Driscoll, 1986). For treatment of
excessive microbial growth, however, hydrogen peroxide has the advantage that the dead
microbial mass is in the form of particles as opposed to the slimy material that can form
following acidification. Removal of the biological mass will be facilitated by a more
flocculent mass as opposed to a slimy mass.
It has been suggested that the addition of nutrients in high concentration batches
instead of continuous addition at low concentration might reduce the tendency for
microbial growth in the well bore and the immediate vicinity of the injection point.
After the system has been in operation for an extended period, it may become
apparent that the distribution of nutrients and oxygen is not as anticipated. Frequently
this can be corrected by adjusting the relative rates of ground-water recovery or
reinjection in the various wells. These adjustments are more efficiently made using a
ground-water flow model. Determination of ground-water elevations in monitoring wells
will determine within a few days whether or not the adjustment in flows is having the
desired effect. Statistically significant changes in contaminant, nutrient, or dissolved
oxygen levels are likely to take several weeks to a few months to be observed. In some
instances, adjusting the flows between wells may not produce the desired effect. It may
then be necessary to add an additional well(s).
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Identifying problems with nutrient distribution and thus impact on contaminant
levels in a timely fashion requires that monitoring wells be properly located. Wells need
to be located so that flows in different directions can be determined and at distances that
produce changes in water chemistry within a reasonable time frame. The distance
between injection wells and the nearest monitoring wells should be based on predicted
flow times rather than distance. The time of travel for a conservative tracer between the
injection well and the nearest monitoring well should be on the order of one week.
If either nutrients or contaminants appear in monitoring wells that are outside
the treatment zone, it may be necessary to change the relative flows in the injection and
recovery wells. In particular, it may be necessary to reduce the fraction of the recovered
ground water that is being reinjected.
Some state regulations (e.g., New Jersey) require that the final nutrient
constituent levels in the ground water be at or below background levels at the completion
of the project. In order to avoid problems meeting this requirement, it is necessary to use
the minimum amount of nutrients that are needed to complete biodegradation across the
site. Since nutrient requirements are a function of many factors, it is difficult to
determine the total amount required a priori. It is necessary to monitor nutrient
distribution across the site during remediation and adjust nutrient addition rates to
balance requirements of the bacteria, nonbiological removal in the aquifer, and the
nutrient concentrations required to close the site.
2.13. STIE PROPERTIES vs COSTS
In-situ bioremediation costs are dependent on a number of factors including site
conditions, remedial goals, the design of the system, and the operating and monitoring
schedule.
2.13.1. Mass of Contaminant
The greater the mass of contamination present, the greater the nutrient and
electron acceptor requirements. This increases not only the chemical costs, but
increases either the time to achieve remediation or requires greater capital expenditure
for wells, pumps, and aboveground treatment.
2.13.2. Volume of Contaminated Aquifer
The greater the volume of aquifer solids and ground water subject to treatment,
the greater the number of injection and recovery points that will be required, or the time
to achieve remediation will be longer.
2.13.3. Aquifer Permeability/Soil Characteristics
For a given size plume and contaminant mass, it will generally be more
expensive to remediate a low permeability aquifer than to remediate a more permeable
aquifer because either longer remediation times or more injection and recovery points
will be required.
2.13.4. Final Remediation Levels
The more stringent the remediation goals, the more costly will be the remediation
in most cases. For gasoline and other light petroleum hydrocarbon spills in relatively
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permeable and homogeneous aquifers, the time to proceed from a less stringent
remediation goal to a more stringent remediation goal might not be very long. However,
for the heavier hydrocarbon blends, particularly those containing PAHs, the slow
dissolution and thus biodegradation of the least soluble components may limit the rate of
bioremediation. Long periods of time may be required to meet stringent ground-water
quality standards. A small change in the remediation goal could thus make a large
change in the time to attain the remediation goal and may also increase the size of the
aquifer zone that requires active treatment.
2.13.5. Depth to Water
The depth to water will affect the design of the system as well as the cost. For
systems using injection wells, the greater the depth to water, the more head pressure
and thus the greater the flow of injected water that can be introduced. This can reduce
the number of wells needed or shorten the remediation time. However, the cost of
installing wells increases with depth. Very shallow aquifers may be treated using
injection galleries, or through percolation of nutrients with air sparging and thus be
relatively inexpensive to construct and operate.
2.13.6. Monitoring Requirements
Monitoring costs can be substantial. The number of wells to be monitored, the
frequency of monitoring, the number and type of parameters all contribute to the costs.
Monitoring should be designed to provide a basis for evaluation of progress, to identify
conditions that require process modifications, and to ensure that the system is under
hydraulic control. Monitoring data that does not serve as a basis for making decisions
should be avoided. Unnecessary monitoring adds to the costs of data acquisition, data
interpretation, and report writing and reading. Unnecessary data can also impede
interpretation of the critical issues.
2.13.7. Contaminant Properties
The extent to which a compound will be recovered with captured ground water is
dependent on its solubility or, more precisely, its octanol/water partition coefficient as
well as the organic content of the solids. The larger the proportion of the contaminant
mass that is recovered, the less time and expense required to provide sufficient nutrients
and electron acceptor. On the other hand, treatment costs for the recovered ground
water may increase with the concentration of the contaminant in the recovered water.
2.13.8. Location of Site
The location of a site can also impact the costs. Remote sites will have higher
costs of providing labor due to travel and housing costs. This is typically much more
important for small sites, particularly for a system whose operations are highly
automated and technicians are not required on a daily basis.
The use of air sparging techniques offers the potential to reduce the costs of in-situ
bioremediation. The depth to water, type of subsurface material, and saturated interval
of the aquifer will all affect the costs. Shallow aquifers beneath sandy materials permit
nutrients to be added from the surface. Large, saturated intervals permit large radius of
influences of sparging wells and thus smaller numbers of sparge wells. Stratification of
subsurface materials also affects the radius of influence. For greater detail on air
sparging see Section 4.
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2.14. PREVIOUS EXPERIENCE WITH COSTS
Costs for bioremediation are not easily generalized. As previously discussed,
many factors affect the cost of remediation. The number of completed and documented in-
situ bioremediation projects with readily available cost data is small compared to the
variables affecting the costs. Frequently, available cost data, especially from larger sites,
includes but does not define costs for many other activities associated with the site
remediation.
For the same site conditions and contaminant distribution, the cost of
bioremediation can vary significantly depending on the specific design. For instance,
incorporating more recovery and injection wells will increase the capital costs but may
reduce the operating and maintenance costs by reducing the total time of remediation.
The choice of an oxygen source (or an alternate electron acceptor) may have a
large impact on costs. Using hydrogen peroxide instead of oxygen will increase monthly
operating costs, but may reduce overall operating costs by shortening the period of
operation and thus the time over which operating, monitoring, and reporting costs are
incurred. Provision of oxygen through air sparging has the potential to substantially
shorten the time of remediation and costs, especially for heavier hydrocarbons. For
lighter petroleum hydrocarbons, the reduced time and cost of supplying oxygen may be
offset by the additional costs for a system to capture and treat air from the unsaturated
zone. Air sparging using the unsaturated zone as a vapor phase biotreatment system
could prove to be the lowest cost system for volatile hydrocarbons.
In addition to differences in designs, the contractor can impact costs through the
selection of equipment as well as efficiency of construction, permitting, and operation.
Generally, the use of good quality components and automated equipment will minimize
overall costs.
Limited anecdotal (R.A. Brown, 1992) and personal information indicate that in-
situ bioremediation of light petroleum products at leaking underground storage tank
sites has cost from one to 1.5 million dollars for 0.5 to one-acre sites and required from
less than one to up to five years. Costs per acre would be expected to decrease
significantly with scale-up. Two systems that included an in-situ venting system which
served to address a large portion of the contamination at the capillary zone, cost 0.7 to one
million dollars and took approximately three years. Extrapolation from an air sparging
system that treated chlorinated solvents through physical removal suggests costs of 0.3 to
0.5 million dollars and a time frame of one to 1.5 years.
A recent U.S. EPA document (U.S. EPA, 199Ic) provides some cost information for
ongoing current projects. This cost data may not be inclusive of all costs and may not
sort out costs where multiple technologies are being used.
New York: Gasoline contamination over approximately one acre with a depth to
water of ten feet. System consists of an infiltration trench for nutrients and
hydrogen peroxide and three 80 gpm recovery wells. Initiated in January 1989.
Costs are reported to be $250K
Iowa: Ground water contaminated with PAHs and BTEX. In-situ system.
Construction costs reported as $149K with anticipated additional costs of $1.5M.
Kansas: Approximately TOOK cubic feet of aquifer contaminated with BTEX.
Combined in-situ soil flushing with bioremediation using nitrate. Bioventing is
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also being considered. Reported expenditures were $275K with anticipated
additional costs of $650K
California: Approximately 3,000 cubic yards of aquifer contaminated with diesel
and gasoline. System consisted of a closed loop system with hydrogen peroxide
as the oxygen source. In-situ vapor stripping and soil flushing were also
incorporated. Started November 1988 and completed March 1991. Costs reported
as $1.6M.
Michigan: An approximately 1/4-acre site was contaminated with gasoline.
System consisted of infiltration gallery and injection wells. Air and hydrogen
peroxide were used as oxygen sources. Source area treatment lasted
approximately 1.5 years. Costs were approximately $600K, of which sampling
and analytical costs were a significant portion. Some residual in downgradient
edge of plume will be addressed with air sparging.
Texas: Approximately 20 acres contaminated with BTEX, some chlorinated
solvents, and other organics. In-situ bioremediation is being used to augment
pump and treat. Both pure oxygen and nitrate are being used as electron
acceptors. Recovered ground water is being treated in an aboveground
bioreactor. Ground water from a clean portion of the aquifer is being used for
injection. Capital costs were approximately $5M, including the water treatment
plant. Upgrading the initial pump-and-treat system to include in-situ
bioremediation cost $200K in capital and $150Kin pilot testing and engineering.
2.15. REGULATORY ACCEPTANCE
Acceptance of the technology on a specific site is, as for any technology, impacted
by those criteria normally used to evaluate technologies:
Is it appropriate for the contamination of concern?
Is it a permanent remedy?
Is it implementable under the specific site conditions?
Can the technology reach the site specific remediation goals?
Is the technology innovative?
Has the technology been demonstrated?
Can the technology be implemented without violating the intent of local or
state regulations?
- Will its application be protective of public health during its construction and
as a result of its performance?
Specifically favorable to in-situ bioremediation are the benefits of being able to
avoid bringing contaminated subsurface materials to the surface, minimization of
interruption of ongoing commercial operations, low profile of operations, destruction
rather than transport of the contaminants and, frequently, costs.
Areas of concern specific to in-situ bioremediation are: (1) State regulations
prohibiting the injection of water not meeting drinking water standards; (2) residual
levels of nutrient components; (3) potential for formation of nitrate from ammonium; and
(4) concern for maintaining hydraulic control over the contaminated ground water and
the injected nutrients.
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2.16. KNOWLEDGE GAPS
Knowledge gaps include both those items that are not understood well by anyone
and the myriad of small pieces of information that are known by various individuals and
consultants that have not been disseminated in any organized fashion. Information
obtained from one site tends to be generalized for all sites. In some instances conclusions
drawn from performance of one design are used to evaluate systems of such different
design characteristics that the conclusions are not at all valid. Many knowledge gaps
could be considerably narrowed if an efficient exchange of information could be achieved
and maintained.
Specific areas where increased understanding would be very beneficial to the
implementation of in-situ bioremediation are (Bartha, 1991):
- Identification of the real cause and effects of the difficulties that have been
observed on some sites.
- Better correlation between laboratory test results and field performance.
Greater understanding of nutrient transport. How do different nutrient
sources move through various types of subsurface material and what can be
done to facilitate nutrient transport?
Selection, control, and enhancement of oxygen distribution under a variety of
site conditions.
Greater understanding of conditions under which aquifer permeability will
be reduced and how to prevent aquifer blockage.
- Better models of nutrient, oxygen, and contaminant transport and
biodegradation rates.
Better understanding of natural attenuation as an alternative to active
remediation or subsequent to active remediation.
Better cost data on completed projects.
Methods of addressing low permeability aquifers.
Methods of solubilizing the higher molecular weight hydrocarbons.
Increased understanding of the effects and limits set by site conditions.
Improved methods of engineering and site management.
- Improved methods for estimating contaminant mass, including analytical
procedures.
Better understanding of degradation pathways even though degradation
pathways are much better understood for petroleum hydrocarbons than for
most other classes of compounds.
Better assessment protocols.
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Consultant. January/February, pp. 1.7- 1.10.
Bartha, R. 1991. Utilizing bioremediation technologies: Difficulties and approaches.
Bioremediation Workshop. Interdisciplinary Bioremediation Working Group.
Rutgers University, New Brunswick, New York.
Bauman, B. 1991. Biodegradation research of the American Petroleum Institute.
Presented at: In Situ Bioreclamation: Application and Investigation for
Hydrocarbons and Contaminated Site Remediation. San Diego, California.
March 19-21, 1991.
Borden, R.C. 1991. Simulation of enhanced in situ biorestoration of petroleum
hydrocarbons. In: In Situ Bioreclamation: Application and Investigation for
Hydrocarbons and Contaminated Site Remediation. Eds., R.E. Hinchee and R.F.
Olfenbuttel. Butterworth-Heinemann. pp. 529- 534.
Brown, R.A. 1992. Personal communication. Groundwater Technology, Inc. Trenton,
New Jersey.
Brown, R.A., Dey, J.C. and McFarland, W.E. 1991. Integrated site remediation
combining groundwater treatment, soil vapor extraction, and bioremediation. In:
In Situ Bioreclamation: Application and Investigation for Hydrocarbons and
Contaminated Site Remediation. Eds., R.E. Hinchee and R.F. Olfenbuttel.
Butterworth-Heinemann. pp. 444-449.
Brown, R.A., and R.D. Norris. 1988. U.S. Patent 4,727,031. Nutrients for Stimulating
Aerobic Bacteria.
Brown, R.A., Norris, R.D., and Raymond, R.L. 1984. Oxygen transport in contaminated
aquifers. In: Proceedings of the Petroleum Hydrocarbon and Organic Chemicals
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Association. Houston, Texas.
Brubaker, G.R. 1991. In situ bioremediation of PAH-contaminated aquifers. In:
Proceedings of the Petroleum Hydrocarbons and Organic Chemicals in Ground
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Driscoll, F.G. 1986. Groundwater and Wells. Johnson Division. St. Paul, Minnesota.
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Eckenfelder, W. Wesley, Jr. 1967. Industrial Water Pollution Control. McGraw-Hill
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Falatico, R.J., and Norris, R.D. 1990. The necessity of hydrogeological analysis for
successful in situ bioremediation. In: Proceedings of the Haztech International
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Flathman, P.E., K.A. Khan, D.M. Barnes, J.H. Caron, S.J. Whitehead, and J.S. Evans.
1991. Laboratory evaluation of hydrogen peroxide for enhanced biological
treatment of petroleum hydrocarbon contaminated soil. In: In Situ
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Bioreclamation: Application and Investigation for Hydrocarbons and
Contaminated Site Remediation. Eds., R.E. Hinchee and R.P. Olfenbuttel.
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Gibson, D.T. 1984. Microbial Degradation of Organic Compounds. Microbiology Series.
Marcel Dekker, Inc. New York, New York.
Hinchee, R.E., and D.C. Downey. 1988. The role of hydrogen peroxide in enhanced
bioreclamation. In: Proceedings of the Petroleum Hydrocarbons and Organic
Chemicals in Ground Water: Prevention, Detection and Restoration. Vol 2.
National Water Well Association, pp. 715-721.
Howard, P.H. 1989. Handbook of Environmental Fate and Exposure Data for Organic
Chemicals: Volume I. Large Production and Priority Pollutants. Lewis
Publishers. Chelsa, Michigan.
Howard, P.H. 1990. Handbook of Environmental Fate And Exposure Data For Organic
Chemicals: Volume II. Solvents. Lewis Publishers. Chelsa, Michigan.
Huling, S.G., B.E. Bledsoe, and M.V. White. 1990. Enhanced Bioremediation Utilizing
Hydrogen Peroxide as a Supplemental Source of Oxygen: A Laboratory and Field
Study. Final Report. NTIS PB90-183435/XAB. EPA/600/2-90/006.
Lawes, B.C. 1991. Soil-induced decomposition of hydrogen peroxide. In: In Situ
Bioreclamation: Application and Investigation for Hydrocarbons and
Contaminated Site Remediation. Eds., R.E. Hinchee and R.F. Olfenbuttel.
Butterworth-Heinemann. pp. 143-156.
McKenna, E.J., and R.D. Heath. 1976. Biodegradation of Polynuclear Aromatic
Hydrocarbon Pollutants by Soil and Water Microorganisms. University of Illinois
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Michelsen, D.L., M. Lofti, and D.L. Violette. 1990. Application of air microbubbles for
treatment of contaminated groundwater. In: Proceedings of the HMCRI - 7th
National RCRA/Superfund Conference and Exhibition. St. Louis, Missouri.
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Publishers. New York, New York.
Montgomery, J.H., and L.M. Wilkom. 1990. Groundwater Chemical Desk Reference.
Vol. I. Lewis Publishers. New York, New York.
Norris, R.D., S.S. Sutherson, and T.J. Callmeyer. 1990. Integrating different
technologies to accelerate remediation of multiphase contamination. Presented at
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Massachusetts. October 17-19,1990.
Overcash, M.R., and D. Pal. 1979. Design of Land Treatment Systems for Industrial
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Piontek, K.R., and T.S. Simpkin. 1992. Factors challenging the practicability of in situ
bioremediation at a wood preserving site. In: Proceedings of the 85th Annual
Meeting and Exhibition of the Air and Waste Management Association. Kansas
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Prosen, B.J., W.M. Korreck, and J.M. Armstrong. 1991. Design and preliminary
performance results of a full scale bioremediation system utilizing an on-site
oxygen generation system. In: In Situ Bioreclamation: Applications and
Investigations for Hydrocarbons and Contaminated Site Remediation. Eds., R.E.
Hinchee and R.F. Olfenbuttel. Butterworth-Heinemann. pp. 523-528.
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73:390-404.
Raymond, R.L., V.W. Jamison, J.O. Hudson, R.E. Mitchell, and V.E. Farmer. 1978.
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hydrocarbon biodegradation and in situ restoration. In: In Proceedings of the
NWWA/API Conference on Petroleum Hydrocarbons and Organic Chemicals in
Ground Water: Prevention Detection, and Restoration. National Water Well
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Chlorinated Aromatic Compounds. EPA 600/2-86/090.
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Studies Under CERCLA: Aerobic Biodegradation Remedy Screening. Interim
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2nd Edition. Van Nostrand Reinhold. New York, New York.
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SECTIONS
BIOVENTING OF PETROLEUM HYDROCARBONS
Robert E. Hinchee
Battelle Memorial Institute
505 King Avenue
Columbus, Ohio 43201-2693
Telephone: (614)424-4698
Fax: (614)424-3667
3.1. FUNDAMENTAL PRINCIPLES
Bioventing is the process of supplying air or oxygen to the unsaturated zone to
stimulate aerobic biodegradation of a contaminant. Bioventing is applicable to any
contaminant that is biodegradable aerobically. Air can be injected through boreholes
screened in the unsaturated zone, or air can be extracted from boreholes, pulling air
from the surface into a contaminated area.
For the purposes of this manuscript, the term "bioventing" will be reserved to
processes occurring above the water table. The term "air sparging" as discussed in
Section 4 is a separate technology designed to treat contamination below the water table.
The two technologies are often used in combination. Obviously, once sparged air rises
above the water table, it can also biovent the unsaturated zone. This section will focus on
in-situ applications to the vadose zone and the use of dewatering to extend the vadose
zone. Bioventing may also be used to introduce methane or other hydrocarbons to
stimulate cometabolic degradation of chlorinated compounds, as addressed in Section 5.
3.1.1. Review of the Technology
The first documented evidence of bioventing was reported by the Texas Research
Institute, Inc., in a study for the American Petroleum Institute (Texas Research
Institute, 1980; 1984). A large-scale model experiment was conducted to test the
effectiveness of a surfactant treatment to enhance the recovery of spilled gasoline. Only
30 1 of the 250 1 originally spilled could be accounted for, and thus questions were raised
about the fate of the gasoline. Subsequently, a column study was conducted to determine
a diffusion coefficient for soil venting. This column study evolved into a biodegradation
study that concluded that as much as 38% of the fuel hydrocarbon was biologically
mineralized. Researchers concluded that venting would not only remove gasoline by
physical means, but also could enhance microbial activity and promote biodegradation of
the gasoline (Texas Research Institute, 1980; 1984).
The first actual field-scale bioventing experiments were conducted by van Eyk for
Shell Oil. In 1982 at van Eyk's direction, Delft Geotechnics in The Netherlands initiated a
series of experiments to investigate the effectiveness of bioventing for treating
hydrocarbon-contaminated soils. These studies are reported in a series of papers
(Anonymous, 1986; Staatsuitgeverij, 1986; van Eyk and Vreeken, 1988; van Eyk and
Vreeken, 1989a; van Eyk and Vreeken, 1989b).
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Wilson and Ward (1986) suggested that using air as a carrier for oxygen could be
1,000 times more efficient than using water, especially in deep, hard-to-flood unsaturated
zones. They made the connection between soil venting and biodegradation by observing
that "soil venting uses the same principle to remove volatile components of the
hydrocarbon." In a general overview of the soil venting process, Bennedsen et al. (1987)
concluded that soil venting provides large quantities of oxygen to the unsaturated zone,
possibly stimulating aerobic degradation. They suggested that water and nutrients
would also be required for significant degradation and encouraged additional investiga-
tions.
Biodegradation enhanced by soil venting has been observed at several field sites.
Investigators claim that at a soil venting site for remediation of gasoline-contaminated
soil, significant biodegradation occurred (measured by a temperature rise) when air was
supplied. Investigators pumped pulses of air through a pile of excavated soil and
observed a consistent rise in temperature, which they attributed to biodegradation. They
claimed that the pile was cleaned up during the summer primarily by biodegradation
(Conner, 1988). However, they did not control for natural volatilization from the
aboveground pile, and not enough data were published to critically verify their
biodegradation claim.
Ely and Hefiher (1988) of the Chevron Research Company patented a bioventing
process. They did not provide their experimental design and data, but they did present
their findings graphically. At a site contaminated by gasoline and diesel oil, they
observed a slightly higher removal through biodegradation than through evaporation. At
a gasoline-contaminated site, results indicated that about two-thirds of the hydrocarbon
removed was due to volatilization and one-third due to biodegradation. At a site
containing only fuel oils, approximately 75 I/well/day were biodegraded, while removals
by volatilization were low due to low vapor pressures of the fuel oil. Ely and Heffner
claimed that the process is more advantageous than strict soil venting because removal
is not dependent only on vapor pressure. In the examples stated in the patent, CQz was
maintained between 6.8% and 11% and O2 between 2.3% and 11% in vented air. The
patent suggested that the addition of water and nutrients may not be acceptable because
of flushing of the contaminants to the water table, but nutrient addition is included as
part of the patent. The patent recommends flow rates between 50 and 420 m3/min per
well and states that air flows higher than those required for volatilization may be
optimum for biodegradation. The Chevron patent is the only patent directly related to the
bioventing process. However, other soil venting patents may be relevant.
At Traverse City, Michigan, researchers from the Robert S. Kerr Environmental
Research Laboratory (U.S. Environmental Protection Agency) observed a decrease in the
toluene concentration in unsaturated zone soil gas, which they measured as an indicator
of fuel contamination in the unsaturated zone. They assumed that advection had not
occurred, and attributed the toluene loss to biodegradation. The investigators concluded
that because toluene concentrations decreased near the oxygenated ground surface, soil
venting is an attractive remediation alternative for biodegrading light volatile
hydrocarbon spills (Ostendorf and Kampbell, 1989).
This work was followed by a field-scale bioventing pilot study (Kampbell et al.,
1992a, Kampbell et al., 1992b; The Traverse Group Inc., 1992). Two experimental
configurations were evaluated. In one plot, air was injected near the water table in one
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row of wells, extracted in another row near the water table, and reinjected at an
intermediate depth in a third row of wells. In the second plot, air was injected only at the
water table. The plots were thirty feet wide and fifty feet long. The water table was 15 feet
below land surface. Air was supplied at 5 cubic feet per minute, resulting in an average
residence time of air in the unsaturated zone of 24 hours.
After one year of operation, hydrocarbon vapor concentrations in the unsaturated
zone were reduced from 2,000 mg/1 to 14 mg/1. The concentration of hydrocarbon vapors
in the air that escaped to the atmosphere was less than 0.5 ug/1 throughout the entire
demonstration. In the plot with direct injection of air, the mass of gasoline above the
water table was reduced from 530 g/m2 plan surface area to 3.3 g/m2. Below the water
table, the mass of gasoline was reduced from 2450 g/m2 to 1920 g/m2. In the plot with
reinjection of air, the mass above the water after one year of bioventing was 0.3 g/m2, and
the quantity below the water table was 907 g/m2.
Although most of the gasoline persisted below the water table, benzene was
depleted. After one year of bioventing, ground water in contact with the gasoline
contained less than 5 ug/1 of benzene.
To date, the best documented full-scale bioventing study was initiated in 1988 at
Hill Air Force Base (AFB) in Utah (Dupont et al., 1991; Hinchee and Arthur, 1991;
Hinchee et al., 1991). This work was followed by a thoroughly documented field pilot-
scale study atTyndall AFB in Florida (Miller, 1990; Miller etal., 1991).
3.1.2. Maturity cf die Technology
Bioventing has been applied for remediation of sites since the early- to mid-1980s
in the form of soil venting. This process is also known as soil vacuum extraction,
vacuum extraction, soil gas extraction, and in-situ volatilization. At most if not all sites
where soils are ventilated, oxygen is supplied and biodegradation is stimulated, and in
many cases biodegradation is a significant contributor to remediation. In early
applications of soil venting, this was not recognized or documented. More recently, soil
venting vendors have begun to monitor and document biodegradation, and some are now
designing and operating soil venting systems to optimize biodegradation, either in
addition to volatilization or to minimize volatilization.
Various forms of undocumented bioventing have been applied to more than 1000
sites worldwide. However, biodegradation has been confirmed by monitoring at no more
than 90% of these sites, and optimized at far fewer.
3.1.3. Repositories of Expertise
A number of groups are involved in bioventing research and application. The
separation into groups specializing in research vs. application is based upon the author's
impressions. This list is not intended to be all inclusive, but only to be representative of
organizations known to the author to be significantly involved in bioventing work. The
author knows of many other organizations doing bioventing work, and no doubt other
organizations unknown to the author are doing bioventing work.
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Bioventing Research
Bioventing Application
Battelle Memorial Institute
Columbus, Ohio
Contact: Robert E. Hinchee
Phone: (614)424-4698
Fax: (614)424-3667
University of Karlsruhe
Karlsruhe, Germany
Contact: Nils-Christian Lund
Phone (49)0721-594016
Fax: (49)0721-551729
U.S. EPA, Robert S Ken-
Environmental Research Laboratory
Ada, Oklahoma
Contact: John T.Wilson
Phone: (405)436-8532
Fax: (405)436-8529
U.S. EPA, Risk Reduction
Engineering Laboratory
Cincinnati, Ohio
Contact: Richard C. Brenner
Phone: (513)569-7657
Fax: (513)569-7787
U.S. Air Force Civil
Engineering Support Agency
Contact: Catherine M. Vogel
Phone: (904)283-6036
Fax: (904)283-6004
U.S. Navy Civil
Engineering Laboratory
Port Hueneme, CA
Contact: Ronald Hoeppel
Phone. (805)982-1655
Fax: (805)982-1409 or -1418
Utah Water Research Laboratory,
Utah State University
Logan, Utah
Contact: Ryan Dupont
Phone: (801)750-3227
Fax: (801)750-3663
Engineering-Science, Inc.
Denver, Colorado
Contact: Doug Downey
Phone: (303)831-8100
Fax: (303)831-8208
Groundwater Technology, Inc.
Trenton, New Jersey
Contact: Richard A. Brown
Phone: (609)587-0300
Fax: (609)587-7908
Integrated Science and Technology
Atlanta, Georgia
Contact: H. James Reisinger, II
Phone: (404)425-3080
Fax: (404)425-0295
The Traverse Group
Ann Arbor, Michigan
Contact: John Armstrong
Phone: (313)747-9300
Fax: (313)483-7532
U.S. Air Force Center for
Environmental Excellence
Brooks AFB, Texas
Contact: Ross N. Miller
Phone: (512)536-4331
Fax: (512)536-4330
Woodward-Clyde Consultants
San Diego, CA
Contact: Desma S. Hogg
Phone: (619)294-9400
Fax: (619)293-7920
TAUW Infra Consult bv
Devemter, The Netherlands
Contact: Leon G.C.M. Urlings
Phone: (31)5700-99528
Fax: (31)5700-99666
Delft Geotechnics
Delft, The Netherlands
Contact: J. van Eyk
Phone: (31)015-693707
Fax: (31)015-610821
-------
3.2. CONTAMINATION THAT IS SUBJECT TO TREATMENT
Bioventing is potentially applicable to any contaminant that is more readily biode-
gradable aerobically than anaerobically, such as most petroleum hydrocarbons (Atlas,
1981). To date, most applications have been to petroleum hydrocarbons (Hoeppel et al.,
1991); however, application to PAHs (Lund et al., 1991; Hinchee and Ong, 1992) and to an
acetone, toluene, and naphthalene mixture (Hinchee and Ong, 1992) have been reported.
In most applications, the key is biodegradability vs. volatility. If the rate of
volatilization significantly exceeds the rate of biodegradation, removal becomes more of a
volatilization process. Figure 3.1 illustrates the general relationship between a com-
pound's physicochemical properties and its potential for bioventing.
io2
H - Henry's Law Coefficient (atm • m3mole-')
Gasoline
Vapor Pressure Too
High to Biovent
Vapor Pressure
Amenable to
Volatilization lnme"iy
di
•diipeihylphenoh
e
4-melhylphenol
acenaphlhene
fluorcne
Vapor Pressure
Too Low to Volatilize
10
10'7 IO'6 10'5
loroelhane
neihyleiher
10° 10'2 lO'1 I 10 100 1000
Solubility (uM)
Figure 3.1. Impact of physicochemical properties on potential for bioventing.
-------
In general, compounds with a low vapor pressure (below ~ 1 mm Hg) cannot be
successfully removed by volatilization, but can be biodegraded in a bioventing application.
Higher vapor pressure compounds, above - 760 mm Hg, are gases at ambient tem-
peratures. These compounds volatilize too rapidly to be biodegraded in a bioventing sys-
tem. Compounds with vapor pressures between 1 and 760 mm Hg may be amenable to
either volatilization or biodegradation in a bioventing system. Within this intermediate
range lie many of the petroleum hydrocarbon compounds of greatest regulatory interest
such as benzene, toluene, and the xylenes. As can be seen in Figure 3.1, various
petroleum fuels are more or less amenable to bioventing. Some components of gasoline
are too volatile to easily biodegrade. Most of the diesel constituents are sufficiently
nonvolatile to preclude volatilization, whereas the constituents of JP-4 jet fuel are
intermediate in volatility.
3.3. SPECIAL REQUIREMENTS FOR SITE CHARACTERIZATION
Normal site characterization data are required for implementation of this or any
other remedial technology and will not be addressed here. In general, three site
characterization tests not typically performed are required for application of bioventing.
These tests include: (1) a soil gas survey incorporating measurements of oxygen and
carbon dioxide, (2) a pneumatic conductivity test, and (3) an in-situ respiration test. In
addition, soil samples should be collected for nutrient analysis, and microbial
characterization may be desirable.
3.3.1. Soil Gas Survey
Soil gas surveys are now commonly practiced as part of many site
characterizations. The methods for sample collection are well documented in the
literature and will not be discussed here (Kerfoot, 1987; Marrin and Kerfoot, 1988). Any
method that assures collection of a soil gas sample from discrete depths should be
sufficient. Soil gas samples should be analyzed for the contaminant hydrocarbon as well
as for oxygen and carbon dioxide. For bioventing to be successful in stimulating
biodegradation, the contaminated area must be oxygen deficient. If it is not, the addition
of more oxygen will have no effect.
3.3.2. Soil Gas Permeability and Radius of Influence
An estimate of the soil's permeability to air flow (k) and the radius of influence (Rj)
of venting wells are both important elements of full-scale bioventing design. On-site
testing provides the most accurate estimate of the soil's permeability to air. On-site
testing can also be used to determine the radius of influence that can be achieved for a
given well configuration, flow rate and air pressure. These data are used to design full-
scale systems. Specifically, they are needed to space venting wells, to size blower
equipment, and to ensure that the entire site receives a supply of oxygen-rich air to
sustain in-situ biodegradation.
The permeability of soils to the flow of gas (k) varies according to grain size, soil
uniformity, porosity, and moisture content. The value of k is a physical property of the
soil; k does not change with different extraction/injection rates or different pressure
levels. Soil gas permeability is generally expressed in the units cm2 or darcy (1 darcy = 1
x 10-8 cm2). Like hydraulic conductivity, soil gas permeability may vary by more than an
order of magnitude on the same site due to soil heterogeneity. The range of typical k
values to be expected with different soil types is given in Table 3.1.
-------
The radius of influence (Rj) is defined as the maximum distance from the air
extraction or injection well where measurable vacuum or pressure (soil gas movement)
occurs. R] is a function of soil properties, but is also dependent on the configuration of
the venting well and extraction or injection flow rates, and is altered by soil stratification.
On sites with shallow contamination, the radius of influence can also be increased by im-
permeable surface barriers such as asphalt or concrete. These paved surfaces may or
may not act as vapor barriers. Without a tight seal to the natural soil surface, the
pavement will not significantly impact soil gas flow.
TABLE 3.1. DARCY VELOCITY IN RELATION TO SOIL TYPE"
Sail Type kinDany
Coarse Sand 100-1000
Medium Sand 1- 100
Fine Sand 0.1 - 1.0
Silts/Clays <0.1
"Source: Johnson et al. (1990)
Several field methods have been developed for determining soil gas permeability
(Sellers and Fan, 1991). The most favored field test method is probably the modified field
drawdown method developed by Paul Johnson of Shell Development Company (Johnson,
1991). This method involves the injection or extraction of air at a constant rate from a
single venting well while measuring the pressure/vacuum changes over time at several
monitoring points in the soil away from the venting well.
a&3. In-Situ Respiration
An on-site, in-situ respiration test was developed by Hinchee and associates at
Battelle (Hinchee and Ong, 1992). The test has been used at numerous sites throughout
the United States. To conduct the test, narrowly screened soil gas monitoring points are
placed into the unsaturated zone. The soils are vented with air containing an inert tracer
gas for a given period of time. The apparatus for the respiration test is illustrated in
Figure 3.2.
In a typical experiment, two monitoring point locations ~ the test location and a
background control location -- are used. A cluster of three to four probes is usually
placed in the contaminated soil of the test location. A l-to-3% concentration of inert gas is
added to the air, which is injected for about 24 hours. The air provides oxygen to the soil,
while inert gas measurements provide data on the diffusion of Oz from the ground
surface and the surrounding soil and assure that the soil gas sampling system does not
leak. A background control location is placed in an uncontaminated site with air
injection to monitor natural background respiration.
3-7
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2 5 or More Feel
3-Wjy Vdlvmg
0 5 in 2 Feel
Ga-. Sampling Pon
Gnxind Surface
Inen Gas
Small Dumeier
Probe
Screen
Gas injection/soil gas sampling monitoring point used by Hinchee and Ong ( 1992)
in their in-situ respiration studies.
Figure 3.2.
Measurements of COz and 0% concentrations in the soil gas are taken before any
air and inert gas injection. After air and inert gas injection are turned off, CO 2 and 02
and inert gas concentrations are monitored over time. The monitoring points in
contaminated soil at sites amenable to bioventing show a significant decline in 0(2 over a
40- to 80-hour monitoring period. Figure 3.3 illustrates the average results from four
such sites, along with the corresponding O2 utilization rates in terms of percent of QZ
consumed per hour. In general, little or no O2 utilization was measured in the
uncontaminated background monitoring point.
25
20-
10 -
5 -
Paiuxeni River NAS. MD
k-OI.Vfc/hr
Tyndall AFB. FL
k-04.1%/hr
Eielvm AFB. AK
k - 0.22%/hr
0
20
40 60
Time (hours)
Figure &3. Average oxygen utilization rates measured at four test sites.
80
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The biodegradation rates measured by the in-situ respiration test appear to be
representative of those for a full-scale bioventing system. Miller (1990) conducted a 9-
month bioventing pilot project at Tyndall AFB at the same time Hinchee and Ong (1992)
were conducting their in-situ respiration test. The OQ utilization rates (Miller, 1990),
measured from nearby active treatment areas, were virtually identical to those measured
in the in-situ respiration test.
CO 2 production proved to be a less useful measure of biodegradation than 02
disappearance. The biodegradation rate in milligrams of hexane-equivalent/kilograms
of soil per day based on CO 2 appearance is usually less than can be accounted for by the
02 disappearance. The Tyndall AFB site was an exception. That site had low-alkalinity
soils and low-pH quartz sands, and C02 production actually resulted in a slightly higher
estimate of biodegradation (Miller, 1990). In the case of the higher pH and higher
alkalinity soils at Fallon NAS and Eielson AFB, little or no gaseous CO% production was
measured (Hinchee and Ong, 1992). This could be due to the formation of carbonates
from the gaseous evolution of CO2 produced by biodegradation at these sites. A similar
problem was encountered by van Eyk and Vreeken (1988) in their attempt to use C02
evolution to quantify biodegradation associated with soil venting. As a rule, O2 utilization
is a more reliable measure of bioventing-induced biodegradation than is CO2
consumption.
3.4. IMPACT OF SITE CHARACTERISTICS ON APPLICABILITY
Assuming contaminants are present that are amenable to bioventing, gas
permeability is probably the most important site characteristic. Soils must be sufficiently
permeable to allow movement of enough gas to provide adequate oxygen for
biodegradation. Gas permeability is a function of both soil structure and grain size, as
well as of soil moisture content. Even in a coarse sand, if soil moisture content is high,
adequate gas flow may not be possible. The site must be sufficiently permeable to allow a
minimum of approximately one soil gas exchange per week. Typically, permeability in
excess of 1 darcy is adequate. When the permeability falls below -0.1 darcy, gas flow is
primarily through either secondary porosity (such as fractures) or any more permeable
strata that may be present (such as thin sand lenses).
The feasibility of bioventing in these low-permeability soils is a function of the
distribution of flow paths and diffusion of air to and from the flow paths within the
contaminated area. In a soil with reasonably good diffusion, a maximum separation of 2
to 4 feet between flow path and contaminant may still result in treatment. This is
obviously a very site-specific characteristic. Bioventing has been successful in some low-
permeability soils, a silty clay site at Fallon NAS in Nevada (Hinchee, unpublished data),
and a silty site on Eilson AFB in Alaska (Leeson et al., 1992). At a clay site on Tinker
AFB in Oklahoma, there has been less success (Hinchee and Ong, 1992).
In addition to gas permeability, hydraulic conductivity may be important if it is
necessary to either add nutrients or dewater a site.
Another important site characteristic is contaminant distribution. Bioventing is
primarily a vadose zone treatment process. The vadose zone may be extended through
dewatering, but contamination below the water table cannot be treated.
3-9
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3.5. PROCESS PERFORMANCE
In contrast to soil vacuum extraction, which maximizes the flow of air to speed
removal of the contaminant, bioventing to enhance biodegradation provides flow of air
through hydrocarbon-contaminated soils at rates and configurations that will ensure
adequate oxygenation for aerobic biodegradation, and will minimize or eliminate the
production of a hydrocarbon-contaminated off-gas. The addition of nutrients and
moisture may be desirable to increase biodegradation rates; however, field research to
date indicates that these additions may not always be needed (Dupont et al., 1991; Miller
et al., 1991). If necessary, nutrient and moisture addition could take any of a variety of
configurations.
The supply of oxygen to contamination in the capillary fringe or below the water
table may not be adequate. Dewatering may at times be necessary, depending on the
distribution of contaminants relative to the normal water table.
An important feature of good bioventing design is the use of narrowly screened
soil gas monitoring points to sample gas in short vertical sections of the soil. These
points are utilized to monitor local oxygen concentrations, as oxygen levels in the vent
well are not representative of local conditions.
Typically, a soil venting system is installed to draw air from a vent well in the area
of greatest contamination. This configuration allows straightforward monitoring of the
off-gases. However, its disadvantage is that hydrocarbon off-gas concentrations are
maximized and could require permitting and treatment.
Figure 3.4 shows a bioventing system that involves air injection only. Although
this is the lowest cost configuration, careful consideration must be given to the fate of
injected air. The objective is for most, if not all, of the hydrocarbons to be degraded, and
for CO 2 to be emitted at some distance from the injection point. If a building or subsur-
face structure were to exist within the radius of influence of the well, hydrocarbon vapors
could be forced into that structure. Thus, protection of subsurface structures may be
required. A bioventing system with this configuration was installed at Hill AFB in 1991
and is currently under study by U.S. EPA REEL and the U.S. Air Force.
Cutoff Well ui Prevent
Figure 3.4. Conceptual layout of bioventing process with air injection only.
3-10
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Figure 3.5 is an illustration of a configuration in which air is injected (the
injection may also be by a passive well) into the contaminated zone and withdrawn from
clean soils. This configuration allows the more volatile hydrocarbons to degrade prior to
being withdrawn, thereby eliminating contaminated off-gases. This configuration
typically does not require an air emission permit, although site-specific exceptions may
apply. Work with a configuration similar to this at a gasoline site near Atlanta, Georgia,
currently is under way.
Air Injection
(Optional)
Figure JUi. Conceptual layout of bioventing process with air withdrawn from clean soiL
Figure 3.6 illustrates a configuration that may alleviate the threat to subsurface
structures while achieving the same basic effect as air injection alone. In this
configuration, soil gas is extracted near the structure of concern and reinjected at a safe
distance. If necessary, makeup air can be added before injection. Implementation of a
configuration similar to this is occurring on a site at Eglin AFB in Florida.
Monitoring
Figure 5U5. Conceptual layout of bioventing process with soil gas reinjection.
3-11
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Figure 3.7 illustrates a conventional soil venting configuration at sites where
hydrocarbon emissions to the atmosphere are not a problem. This may be the preferred
configuration. Dewatering, nutrient, and moisture additions are also illustrated.
Dewatering will allow more effective treatment of deeper soils. The optimal
configuration for any given site will, of course, depend on site-specific conditions and
remedial objectives.
Nutrient Application
rA tf'"~~"-
kV. it, «..
V\ "'
Ml HI
Figure 3.7. Conceptual layout of bioventing process with air injection into contaminated soil,
coupled with dewatering and nutrient application.
3.5.1. Case Study: Hill AFB Site
A spill of approximately 100,000 1 of JP-4 jet fuel occurred when an automatic
overflow device failed at Hill AFB in Ogden, Utah. Contamination was limited to the
upper 20 m of a delta outwash of the Weber River. This surficial formation extends from
the surface to a depth of approximately 20 m and is composed of mixed sand and gravel
with occasional clay stringers. Depth to regional ground water is approximately 200 m;
however, water may occasionally be found in discontinuous perched zones. Soil moisture
averaged less than 6% in the contaminated soils.
The collected soil samples had JP-4 fuel concentrations up to 20,000 mg/kg, with
an average concentration of approximately 400 mg/kg (Oak Ridge National Laboratory,
1989). Contaminants were unevenly distributed to depths of 20 m. Vent wells were
drilled to approximately 20 m below the ground surface and were screened from 3 to 18 m
below the surface. A background vent was installed in an uncontaminated location in
the same geological formation approximately 200 m north of the site.
Venting was initiated in December 1988 by air extraction at a rate of -40 m^fhr.
The off-gas was treated by catalytic incineration, and it was initially necessary to dilute
the highly concentrated gas to remain below explosive limits and within the incinerator's
hydrocarbon operating limits. The venting rate was gradually increased to-2,500 m3/hr
as hydrocarbon concentration levels dropped. During the period between December 1988
and November 1990, more than 1.0 x K)6 m3 of soil gas was extracted from the site. In
3-12
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November 1989, ventilation rates were reduced to between -500 and 1000 m3/hr to provide
aeration for bioremediation while reducing off-gas generation. This change allowed
removal of the catalytic incinerator, saving ~$6,000 per month.
During extraction, oxygen and hydrocarbon concentrations in the off-gas were
measured. To quantify the extent of biodegradation at the site, the oxygen was converted
to an equivalent basis. This was based on the stoichiometric oxygen requirement for
hexane mineralization. JP-4 hydrocarbon concentrations were determined based on
direct readings of a total hydrocarbon analyzer calibrated to hexane. Based on these
calculations, the mass of the JP-4 fuel as carbon removed was -50,000 kg volatilized and
40,000 kg biodegraded. Figures 3.8 and 3.9 illustrate these results.
1988
1990
Date
Figure 3.8. Cumulative hydrocarbon removal from the Hill AFB Building 914 soil venting site
(Dupont, et aL, 1991).
Hill AFB Building 914 Soil Samples
Depth
(meters)
100 1000
Hydrocarbon Concentration (mg/kg)
Q Before E3 Intermediate • After
Figure 3.9. Results of soil analysis at Hill AFB before and after venting. (Each bar represents
the average of 14 or more samples) (Dupont, et aL, 1991).
3-13
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Hinchee and Arthur (1991) conducted bench-scale studies using soils from this
site and found that, in the laboratory, both moisture and nutrients became limiting after
aerobic conditions were achieved. This led to the addition of first moisture and then
nutrients in the field. The results of these field additions are shown in Figure 3.8.
Moisture addition clearly stimulated biodegradation; nutrient addition did not.
The failure to observe an effect of nutrient addition could be explained by a number
of factors, including:
• The nutrients failed to move in the soils; this is a problem partic-
ularly for ammonia and phosphorus (Aggarwal et al., 1991).
Remediation of the site was entering its final phase, and there was
not enough time for the microbes to respond to the nutrient addition.
• Nutrients may not have been limiting.
3.5.2. Case Study: TyndallAFB Site
As a follow-up to the Hill AFB research, a more controlled study was designed at
Tyndall AFB (Miller et al., 1991). The experimental area in this study was located at a
site where past JP-4 fuel storage had resulted in contaminated soils. The nature and
volume of fuel spilled or leaked were unknown. The site soils are a fine- to medium-
grained quartz sand. The depth to ground water was 0.5 to 1.5 m.
Four test cells were constructed to allow control of gas flow, water flow, and
nutrient addition. Test cells VI and V2 were installed in the hydrocarbon-contaminated
zone; the other two were installed in uncontaminated soils. Initial site characterization
indicated the mean soil hydrocarbon levels were 5,100 and 7,700 mg of hexane-equiva-
lent/kg in treatment plots VI and V2, respectively. The contaminated area was
dewatered, and hydraulic control was maintained to keep the depth to water at ~2 m.
This exposed more of the contaminated soil to aeration. During normal operation,
airflow rates were maintained at approximately one air-filled void volume per day.
Biodegradation and volatilization rates were much higher at the Tyndall AFB site
than those observed at Hill AFB; these higher rates were likely due to higher average
levels of contamination, warmer temperatures, and the presence of moisture. After 200
days of aeration, an average hydrocarbon reduction of -2,900 mg/kg was observed. This
represents a reduction in total hydrocarbons of approximately 40%.
The study was terminated because the process monitoring objectives had been
met; biodegradation was still vigorous. Although the total petroleum hydrocarbons had
been reduced by only 40%, the low-molecular-weight aromatics - benzene, toluene, ethyl-
benzene, and xylenes (BTEX) - were reduced by more than 90% (Figure 3.10). It appears
that the bioventing process more rapidly removes BTEX compounds than other JP-4 fuel
constituents.
Another important observation of this study was the effect of temperature on the
biodegradation rate. Miller (1990) found that the van Hoff-Arrhenius equation provided
an excellent model of temperature effects. In the Tyndall AFB study, soil temperature
varied by only ~7°C, yet biodegradation rates were approximately twice as high at 25°C
than at 18°C.
3-14
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Figure 3.10. Results of soil analysis from Plot V2 at Tyndall AFB before and after venting,
(Each bar represents the average of 21 or more soil samples. Miller et aL, 1991).
In the Tyndall AFB study, the effects of moisture and nutrients were observed in a
field test. Two side-by-side plots received identical treatment, except that one (V2)
received both moisture and nutrients from the initiation of the study while the other plot
(VI) received neither for 8 weeks, then moisture only for 14 weeks, followed by both
moisture and nutrients for 7 weeks. As illustrated in Figure 3.11, no significant effect of
moisture or nutrients was observed. The lack of moisture effect contrasts with the Hill
AFB findings, but is most likely the result of contrasting climatic and hydrogeologic
conditions. Hill AFB is located on a high-elevation desert with a very deep water table.
Tyndall AFB is located in a moist subtropical environment, and at the site studied, the
water table was maintained at a depth of approximately 2 m.
-D— % Removal by BiodegraJjimn. VI
__ % Removal by Bindegndumn - V2
30 60 90 120 ISO 180 210
Venting Time (Days)
Figure 3.11. Cumulative percent hydrocarbon removal at lyndall AFB for Sites VI and V2
(Miller etal., 1991).
3-15
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The nutrient findings support Held observations at Hill AFB that the addition of
nutrients does not stimulate biodegradation. Based on acetylene reduction studies,
Miller (1990) speculates that adequate nitrogen was present due to nitrogen fixation.
Both the Hill and Tyndall AFB sites were contaminated for several years before the
bioventing studies, and both sites were anaerobic. It is possible that nitrogen fixation,
which is maximized under these conditions, provided the required nutrients. In any
case, these findings show that nutrient addition is not always required.
In the Tyndall AFB study, a careful evaluation of the relationship between air flow
rates and biodegradation and volatilization was made. It was found that extracting air at
the optimal rate for biodegradation resulted in 90% removal by biodegradation and 10%
removal by volatilization. It was also found that the volatilized contaminants were
completely biodegraded after the air was passed through clean soil.
3.5.3. Performance of Other Sites
In addition to the Hill AFB and Tyndall AFB sites, numerous other site studies
are reported in the literature. A summery of several other studies is presented in Table
3.2.
TABLE 3.2. COMPARISON OF BIODEGRADATION RATES OBTAINED BY THE iN-Srru RESPIRATION
TEST WITH OTHER STUDIES
Sue
Venous (8
locations)
Hill AFB. Utah
Tyndall AFB,
Florida
Netherlands
Netherlands
Undefined
Undefined
Undefined
New Zealand
Traverse City
Scale of
Application
In situ respiration
tests
Full-scale, 2 years
Field pilot, lyear
and in situ respira-
tion tests
Undefined
Field pilot, 1 year
Full-scale
Full-scale
Full-scale
Pilot-scale/
Full-scale
Pilot-Scale
Respiration
Contaminants
Various
JP-4 Jet Fuel
JIM Jet Fuel
Undefined
Diesel
Gasoline and
Diesel
Diesel
Fuel Oil
Diesel
Spent Oil
Aviation Gasoline
Estimated
Rates Cb Biodegradation
Otlhour) Rates
0.02 - 0.99 04-19 mg/kg/day
up to 0 52 up to 10 mg/kg/day
01-10 2-20 mg/kg/day
0 1 - 0.26 2- 5 mg/kg/day b
042 8 mg/kg/day
SO kg/well/dayc
100 kg/well/dayc
60 kg/well/dayc
0.2- 20 mg/kg/day
4 mg/kg/day b
References
Hmchee and Ong, 1992
Hmchee et al . 1991
Miller, 1990
Urhngs el al. 1990
van Eyk and Vreeken,
1969b
Ely and Heffher, 1988
Ely and Heffner. 1988
Ely and Heffner, 1988
Hogg etal, 1992
Wilson, 1992
Rates reported by Hmchee et al (1991) were first order with respect to oxygen, for comparison purposes, these have been
converted to zero order with respect to hydrocarbons at an assumed oxygen concentration of 10 percent
Rates reported as oxygen consumption rates, these have been converted to hydrocarbon degradation rates assuming a
31 oxygen-to-hydrocarbon ratio
Units are in kilograms of hydrocarbon degraded per 30 standard cubic feet per minute (scfm) extraction vent well per
day
3-16
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3.6. PROBLEMS ENCOUNTERED WITH THE TECHNOLOGY
The primary problems encountered with bioventing are:
Accurate estimate of emissions - One of the key variables in bioventing cost
and design is the need or lack of need for off-gas treatment. To permit an
emission with or without off-gas treatment, regulators typically require an
estimate of emission rate.
Time required for remediation - Bioventing that is primarily dependent on
biodegradation is a slow process requiring two or more years for
remediation. At many sites this can be a problem.
Determination of effectiveness on nonpetroleum hydrocarbons -- Many sites
are contaminated with a mixture of chemical wastes, and little is known of
the effectiveness of bioventing on nonpetroleum hydrocarbons.
Regulatory acceptance - With this technology, as with many emerging
technologies, obtaining regulatory acceptance can be difficult.
3.7. COSTS
As with many emerging technologies, not much published experience exists to
precisely determine cost; however, some general guidelines are available. The basic cost
of soil venting equipment has been estimated to be as low as $10 per yd3. Off-gas
treatment costs can run $30 to 50/yd3(Long, 1992). The key to the cost of bioventing is the
monitoring costs, and these are very site specific. Due to the relatively long time period
required for biodegradation, a bioventing site optimized for biodegradation as opposed to
volatilization will incur higher monitoring costs. On some sites this may be offset by not
having gas treatment costs, thus reducing total remediation costs.
3.8. REGULATORY ACCEPTANCE
The primary obstacle to regulatory acceptance is demonstrating potential
effectiveness. Many regulators are unfamiliar with the technology and skeptical of
accepting a remedial approach that may require 2 to 3 years to show results. Generally
after the responsible regulators) agree to allow the use of bioventing, permitting is not
difficult.
3.9. KNOWLEDGE GAPS AND RESEARCH OPPORTUNITIES
Bioventing has been performed and monitored at several field sites contaminated
with middle distillate fuels, mainly JP-4 jet fuel. Yet the effects of environmental
variables on bioventing treatment rates are not well understood. In-situ respirometry at
additional sites with drastically different geologic conditions has further defined
environmental limitations and site-specific factors that are pertinent to successful
bioventing. However, the relationship between respirometric data and actual bioventing
treatment rates have not been clearly determined. Additional field respirometry and
closely monitored field pilot bioventing studies at the same sites are needed to determine
what types of contaminants can be successfully treated in situ by bioventing and what the
3-17
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environmental limitations are. Studies are also needed on a wide variety of
contaminants. Studies to date clearly show that many preconceptions regarding the
factors that control bioventing rates may be incorrect. For example, active respiration at
a subarctic site at Eielson AFB near Fairbanks, Alaska, suggests that good hydrocarbon
degradation can occur in situ at locations that are continually subjected to a cold
environment. Failure to accelerate biodegradation rates by adding nitrogen fertilizer to
biovented soils that contain low nitrogen levels indicates that nutrient addition at some
sites may not be required. Also, fine-grained moist clayey soils have been readily aerated
and showed aerobic respiration, indicating that bioventing may be feasible in soils having
low permeabilities. Other low permeability sites have not proven amenable to bioventing,
and better procedures to evaluate sites are needed.
Vapor phase biodegradation occurs and can take place in situ. The question of
how soil sorption and partitioning of volatile organic compounds into soil air affects
biodegradation rates was addressed earlier by McCarty (1987). This question needs
further attention as the movement of the vapor phase in soils is complex and dependent
on changing soil environmental conditions.
Bioventing rates need to be determined under varying vapor extraction rates since
an important purpose for bioventing is to biodegrade the vapor within the soil profile.
Minimal soil aeration levels that provide for high degradation rates must be determined
under different soil conditions. Interaction of the vapor phase with soil particles and
microorganisms in the uncontaminated soil profile needs further research in both the
laboratory and in the field.
3-18
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Aggarwal, P.K., J.L. Means, and R.E. Hinchee. 1991. Formulation of nutrient solutions
for in situ bioremediation. In: In Situ Bioreclamation. Applications and
Investigations for Hydrocarbon and Contaminated Site Remediation. Eds., R.E.
Hinchee and R.F. Olfenbuttel. Butterworth-Heinemann. Stoneham, Massa-
chusetts, pp. 51-66.
Anonymous. 1986. In situ reclamation of petroleum contaminated sub-soil by
subsurface venting and enhanced biodegradation. Research Disclosure. No.
26233, 92-93.
Atlas, R.M. 1981. Microbial degradation of petroleum hydrocarbons: An environmental
perspective. Microbiol. Rev. 45: 180-209.
Bennedsen, M.B., J.P. Scott, and J.D. Hartley. 1987. Use of vapor extraction systems for
in situ removal of volatile organic compounds from soil. In: Proceedings of
National Conference on Hazardous Wastes and Hazardous Materials.
Washington, DC. pp. 92-95.
Conner, J.S. 1988. Case study of soil venting. Poll. Eng. 7: 74-78.
Dupont, R.R., W. Doucette, and R.E. Hinchee. 1991. Assessment of in situ bioremedi-
ation potential and the application of bioventing at a fuel-contaminated site. In:
In Situ and On-Site Bioreclamation. Eds., R.E. Hinchee and R.F. Olfenbuttel.
Butterworth-Heinemann. Stoneham, Massachusetts, pp. 262-282.
Ely, D.L., and D.A. Heffner. 1988. Process for in-situ biodegradation of hydrocarbon
contaminated soil. U.S. Patent Number 4,765,902.
Hinchee, R.E., and M. Arthur. 1991. Bench-scale studies of the soil aeration process for
bioremediation of petroleum hydrocarbons. J. Appl. Biochem. Biotech. 28/29:901-
906.
Hinchee, R.E., D.C. Downey, R.R. Dupont, P. Aggarwal, and R.N. Miller. 1991. Enhanc-
ing biodegradation of petroleum hydrocarbon through soil venting. J. Hazardous
Materials. 27:315-325.
Hinchee, R.E., and S.K. Ong. 1992. A rapid in situ respiration test for measuring
aerobic biodegradation rates of hydrocarbons in soil. Submitted to the Journal of
the Air & Waste Management Association. 21 pp.
Hoeppel, R. E., R.E. Hinchee, and M.R. Arthur. 1991. Bioventing soils contaminated
with petroleum hydrocarbons. J. Industrial Microbiology. 8:141-146.
Hogg, D.S., R.J. Burden, and P.J. Riddell. 1992. In situ vadose zone bioremediation of
soil contaminated with non-volatile hydrocarbon. Presented at HMCRI
Conference. February 4. San Francisco, California.
Johnson, P.C., M.W. Kemblowski, and J.D. Colthart. 1990. Quantitative analysis for the
cleanup of hydrocarbon-contaminated soils by in-situ soil venting. Ground Water.
28(3):413-429. May-June.
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Johnson, P.O. 1991. HyperVentilate Users Manual. Shell Development. Houston,
Texas. 3 pp.
Kampbell, D.H., J.T. Wilson, and C.J. Griffin. 1992a. Bioventing of a gasoline spill at
Traverse City, Michigan. In: Bioremediation Of Hazardous Wastes. EPA/600/R-
92/126. Office of Research and Development.
Kampbell, D.H., J.T. Wilson, C.J. Griffin, and D.W. Ostendorf. 1992b. Bioventing
reclamation pilot project-aviation gasoline spill. In: Abstracts, Subsurface
Restoration Conference. National Center for Ground Water Research. June 21-
24,1992. Dallas, Texas, p. 297.
Kerfoot, H.B. 1987. Soil-gas measurement for detection of groundwater contamination by
volatile organic compounds. Environ. Sci. Technol. 21(1): 1022-1024.
Leeson, A., R.E. Hinchee, J. Kittle, G. Sayles, C.M. Vogel, and R.N. Miller. 1992.
Optimizing bioventing in shallow vadose zones and cold climates. Submitted for
Publication in the Proceedings of the In-Situ and On-Site Bioremediation
Symposium. Niagra on the Lake, Canada.
Long, G. 1992. Bioventing and vapor extraction: Innovative technologies for contam-
inated site remediation. J. Air and Waste Mgt. Assoc. 43(3):345-348.
Lund, N.-Ch., J. Swinianski, G. Gadehus, and D. Maier. 1991. Laboratory and field
tests for a biological in situ remediation of a coke oven plant. In: In Situ
Bioreclamation. Applications and Investigations for Hydrocarbon and
Contaminated Site Remediation. Eds., R.E. Hinchee and R.F. Olfenbuttel.
Butterworth-Heinemann. Stoneham, Massachusetts, pp. 396-412.
Marrin, D.L. and W.B. Kerfoot. 1988. Soil gas surveying techniques. Environ. Sci.
Technol. 22(7):740-745.
McCarty, P.L. 1987. Bioengineering issues related to in situ remediation of
contaminated soils and groundwater. In: Proceedings, Conference on Reducing
Risk from Environmental Chemicals Through Biotechnology. Seattle,
Washington. July.
Miller, R.N. 1990. A field scale investigation of enhanced petroleum hydrocarbon
biodegradation in the vadose zone combining soil venting as an oxygen source
with moisture and nutrient additions. Ph.D. Dissertation. Utah State University.
Logan, Utah.
Miller, R.N., R.E. Hinchee, and C. Vogel. 1991. A field-scale investigation of petroleum
hydrocarbon biodegradation in the vadose zone enhanced by soil venting at Tyndall
AFB, Florida. In: In Situ Bioreclamation. Applications and Investigations for
Hydrocarbon and Contaminated Site Remediation. Eds., R.E. Hinchee and R. F.
Olfenbuttel. Butterworth Publishers. Stoneham, Massachusetts, pp. 283-302.
Oak Ridge National Laboratory. 1989. Soil characteristics: Data summary, Hill Air
Force Base Building 914 fuel spill soil venting project. An unpublished report to
the U.S. Air Force.
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Ostendorf, D.W., and D.H. Kampbell. 1989. Vertical profiles and near surface traps for
field measurement of volatile pollution in the subsurface environment. In:
Proceedings of NWWA Conference on New Techniques for Quantifying the
Physical and Chemical Properties of Heterogeneous Aquifers. National Water
Well Association. Dublin, Ohio.
Sellers, K., and C.Y. Fan. 1991. Soil vapor extraction: Air permeability testing and
estimation methods. In: Proceedings of the 17th REEL Hazardous Waste
Research Symposium. EPA/600/9-91/002, April.
Staatsuitgeverij. 1986. Proceedings of a Workshop, 20-21 March, 1986. Bodembescher-
mingsreeeks No. 9. Biotechnologische Bodemsanering. pp. 31-33. Rapportnr.
851105002. ISBN 90-12-054133. Ordernr. 250-154-59. Staatsuitgeverij Den Haag:
The Netherlands.
Texas Research Institute. 1980. Laboratory Scale Gasoline Spill and Venting Experi-
ment. American Petroleum Institute. Interim Report No. 7743-5:JST.
Texas Research Institute. 1984. Forced Venting to Remove Gasoline Vapor from a Large-
Scale Model Aquifer. American Petroleum Institute. Final Report No. 82101-
F:TAV.
The Traverse Group, Inc. 1992. Biouenting Reclamation Pilot Program. U.S. Coast
Guard Air Station. Traverse City, Michigan. Final Report. Prepared for the U.S.
Coast Guard and the U.S. EPA (Robert S. Kerr Environmental Research
Laboratory).
van Eyk, J., and C. Vreeken. 1988. Venting-mediated removal of petrol from subsurface
soil strata as a result of stimulated evaporation and enhanced biodegradation.
Med. Fac. Landbouww. Riiksuniv. Gent. 53(4b): 1873-1884.
van Eyk, J., and C. Vreeken. 1989a. Model of petroleum mineralization response to soil
aeration to aid in site-specific, in situ biological remediation. In: Groundwater
Contamination: Use of Models in Decision-Making, Proceedings of an
International Conference on Groundwater Contamination. Ed., G. Jousma.
Kluwer Boston/London, pp. 365-371.
van Eyk, J., and C. Vreeken. 1989b. 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, Enuirotech
Vienna. Vol. 2, Session 3. ISBN 389432-009-5. Westarp Wiss., Essen, pp. 475-485.
Wilson, J.T. 1992. Technologies for contaminant destruction: Enhanced biological
electron acceptor I^CV In: Abstracts, Subsurface Restoration Conference
National Center for Ground Water Research. June 21-24, 1992. Dallas, Texas.
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Wilson, J.T., and C.H. Ward. 1986. Opportunities for bioremediation of aquifers contami-
nated with petroleum hydrocarbons. J. Ind. Microbiol. 27:109-116.
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SECTION 4
TREATMENT OF PETROLEUM HYDROCARBONS IN GROUND WATER
BY Am SPARGING
Richard Brown
Groundwater Technology, Inc.
310 Horizon Center Drive
Trenton, New Jersey 08691
Telephone: (609)587-0300
Fax: (609)587-7908
4.1. INTRODUCTION
Petroleum hydrocarbon contamination of ground water is a significant and often
complex problem. The complexity results from the fact that ground-water contamination
is a product of soil contamination. Especially with older spills, significant amounts of
hydrocarbons can be trapped below the water table. Traditional pump and treat is
ineffective because of the low solubility of trapped oily phase hydrocarbons. Typically,
less than five percent of a hydrocarbon spill ever enters the dissolved phase. The bulk of
contamination is sorbed to soil and/or aquifer solids. Venting is ineffective because of the
water saturation. Bioremediation, while effective in treating this trapped hydrocarbon,
is often quite expensive when relying on chemical oxygen carriers such as hydrogen
peroxide. There has been a need, therefore, to develop a technology which could more
cost effectively address petroleum hydrocarbon contamination of ground water. A
technology that offers the most promise is air sparging.
Air sparging, simply viewed, is the injection of air under pressure below the
water table. This creates a transient air filled porosity by displacing water in the soil
matrix (Figure 4.1). The minimum pressure that is required to displace water in an air
sparging system is that which is needed to overcome the resistance of the soil matrix to
Air
Monitoring
Prohe
Vapor
Extraction /Sparger
Wcll\ ' "'-"
Monitoring
Probe
Venl Radius - f(Vacuum)
Sparge Radius - f(Depth)(Pressure)
Contaminated
Soil
Transient Air
Filled Porosirv
Figure 4.1. Diagram of air sparging system.
4-1
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air flow. This resistance to flow is a function of the height of the water column that needs
to be displaced and of the flow restriction (air/water permeability) of the soil matrix.
When this "break-out" pressure is achieved, air enters the soil matrix, travels
horizontally and vertically through the soil matrix displacing water, and eventually exits
into the vadose zone.
Air sparging is a relatively new treatment technology for addressing
contamination below the water table. By displacing water in the soil matrix and creating
a transient air filled porosity, air sparging provides two benefits. First, air sparging
enhances biodegradation by increasing oxygen transfer to the ground water. Second, it
can enhance the physical removal of organics by direct volatile (vapor phase) extraction.
4.2. DEVELOPMENT OF AIR SPARGING
Soil vapor extraction (SVE, or venting), the inducement of air flow by the
application of a vacuum, has long been recognized as one of the more effective means of
treating volatile organic compounds (VOCs), particularly petroleum fuels in the vadose
zone (Hoag and Marley, 1984). SVE addresses VOCs through two primary mechanisms.
First, SVE stimulates biodegradation by supplying oxygen for aerobic metabolic
processes. Second, it physically removes the contaminants by removing vapors
associated with adsorbed contaminants. Both mechanisms are dependent on the
effective movement of air through the subsurface.
Because SVE technology depends on the flow of air through a soil matrix, it has
been obviously limited to the treatment of unsaturated soils. SVE cannot directly treat
VOC-contaminated soils below the water table or contaminated ground water because
there is no air filled porosity below the water table, and, therefore, no air to move. SVE
has, however, been used to indirectly stimulate the biodegradation of dissolved
contaminants by increasing oxygen content in the vadose zone, and therefore diffusion
from the vadose zone to ground water (Clayton et al., 1989). To directly treat saturated
zone contaminants most effectively with SVE, however, generally requires that the site be
effectively dewatered so that a vacuum can be applied and air flow induced.
The difficulty of dewatering and the costs and problems of treating the extracted
ground water has made coupled dewatering-SVE systems less than an optimal solution.
As a result, there has been a search for technology that could effectively extend the utility
of SVE to saturated systems.
Air sparging effectively removes contamination below the water table. Air
sparging is the injection of air directly into a saturated formation. Air sparging can
successfully treat VOCs and petroleum hydrocarbons in ground-water aquifers through
direct volatile removal and biodegradation. Air sparging has been extensively used in
Germany since 1985 (Killer and Gudemann, 1988) and was successfully introduced in
the United States in 1990 (Brown et al., 1991; Marley et al., 1990; Middleton and Hiller,
1990).
Air sparging, as practiced today, should not be confused with older systems,
which were also called air sparging and were used in early bioremediation projects
(Raymond et al., 1975). The difference between the two technologies is where the air is
injected. With older technologies the air was injected into the water column in the well.
The air, in this case, travels through the water column and does not directly contact the
formation matrix. With modern air sparging, the air injection pressure is greater than
4-2
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the hydraulic head, thus the well contains no water and air is directly injected into the
formation. Figure 4.2 shows the differences between the two "air sparging" technologies.
Water Table —•
Water Column
Formation
Air Bubbles
Diffuser
Water Table
Injected Air
Column
Formation
Air Bubbles
Old
Air Sparging
(In Well Sparging)
New
Air Sparging
(Air Injection)
Figure 4.2. Differences between old and new air sparging technologies.
Air sparging is an emerging technology for the treatment of ground water
contaminated with volatile organic compounds. It is being used to increasingly greater
extents to treat petroleum hydrocarbon contaminated ground-water aquifers, overcoming
the limitation of SVE for treating saturated zone contaminants and improving the
efficacy of bioremediation. The benefits and limitations of this technology are still being
defined both in field application and research.
4.3. PRINCIPLES OF THE TECHNOLOGY
When air is injected into a contaminated soil/water matrix (i.e., an aquifer), there
are a number of phenomena that result from the air movement. Some are beneficial -
they remove contamination; and some are actually or potentially detrimental -they
increase or spread contamination. The following is a list and description of these
phenomena:
Enhanced
Oxygenation:
Enhanced
Dissolution:
Air traveling through the aquifer dissolves in the soil water and
replenishes oxygen that may have been depleted by chemical or
biological processes. Normal oxygen replenishment is slow, as it
relies on diffusion from the surface of the water table. Sparged air,
which is distributed throughout the aquifer, has a short diffusion
path length. Enhanced oxygenation is a beneficial phenomenon as it
can stimulate biodegradation.
Air traveling through the aquifer causes turbulence in the soil
pores. This mixes the water and adsorbed VOCs and enhances their
partitioning into the water phase. Normal water/soil contact is static
and dissolution is diffusion-limited. Enhanced dissolution is
beneficial if the ground water is collected, but detrimental if the
contaminated plume is not captured or treated (by in-situ stripping).
Dissolution can also help promote biodegradation.
-------
Volatilization:
Ground-water
Stripping:
Physical
Displacement:
Adsorbed phase contaminants will evaporate into the air stream and
be carried into the vadose zone. The extent of volatilization is
governed by the vapor pressure of the VOC. Volatilization is
prevented in normal saturated environments because there is no air
phase. Volatilization is a beneficial process as it can remove a
significant mass of contaminants.
The aerated aquifer can act as a crude air stripper if sufficient air
flow is passed through the soil matrix. VOCs with a sufficiently
high Henry's Law constant will volatilize from the water into the air
stream and be removed. This is a generally beneficial process.
At very high air flow rates, water can be rapidly and physically
displaced. This is observed often in air-rotary drilling. The
displaced water, if contaminated, will spread contamination in any
direction and is thus not easily captured by existing ground-water
systems. Displacement is a generally detrimental phenomenon and
should be avoided.
These phenomena are all the result of air passing through the aquifer matrix.
Which process is active is generally a function of the amount of air passing through the
soil matrix. These phenomena do not all occur at the same air flow rates. As shown in
Figure 4.3, oxygenation and dissolution occur at essentially all air flow rates.
Volatilization and stripping require moderate rates of air flow. Physical displacement
generally only occurs at high pressures or flows. To maximize the benefits of air
sparging and minimize the detriments requires an optimization of air flow. Too low an
air flow will not effectively remove VOCs and may increase ground-water
concentrations; too high a flow can rapidly and physically spread the contamination.
Optimizing the air flow will maximize mass removal while minimizing the potential
spread of contamination.
Enhanced Oxygenation
Enhanced Partitioning
Volatilization
I Ground-Water Stopping
.Optimum Operating Range,
Physical Displacement
High
Air Flow Rate, SCFM
Key
r Generally Beneficial Effect
ii. Potentially Detrimental Effect
— —•• Generally Detrimental Effect
Figure 4.3. The effects of air flow in saturated environment as a function of air flow rate.
44
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4.4. BENEFITS OF AIR SPARGING
Air sparging is a potentially effective means of treating petroleum hydrocarbons.
That is because air sparging promotes two significant removal mechanisms -
biodegradation and volatilization.
A primary benefit in treating petroleum hydrocarbons with air sparging is that it
is an effective means of supplying oxygen to the saturated zone. This benefit leads to a
key application of air sparging, i.e., enhancing aerobic bioremediation. Air sparging
results in efficient aeration as a result of several factors. First, there is penetration of
air into the contaminated saturated zone. Under normal conditions air only contacts the
surface of the aquifer. With air sparging the contact is distributed over the entire sparged
interval. Second, because air sparging creates air-filled porosity in the soil matrix, the
diffusive path length of air (oxygen) into the water is considerably shortened compared to
normal ground-water conditions. Under normal conditions the distance between the air
and water phases can be on the order of meters; with air sparging, the distance will be,
at most, only several times greater than a soil pore, i.e., a few millimeters. Third, the
"turbulence" caused by air sparging enhances the dissolution and distribution of oxygen
into the water phase. Since biodegradation is critically dependent on oxygen supply, the
efficient aeration engendered by air sparging will enhance bioremediation.
Given this aeration efficiency, an advantage of air sparging is the amount of
oxygen that it can provide for biodegradation compared with the use of hydrogen
peroxide. Even assuming limited utilization of oxygen in the sparged air, air sparging
can supply significant amounts of oxygen for bioremediation. Table 4.1 compares the
amount of oxygen supplied to bioremediation from air sparging or from the use of
hydrogen peroxide.
TABLE 4.1. OXYGEN AVAILABILITY, Db/day
AIR SPARGING HYDROGEN PEROXIDE (1000 PPM)
Flow Utilization Flow Utilization
SCFM 100% 50% 10% GPM 100% 50% 10%
10
25
50
236
590
1182
118
295
590
24
59
118
10
25
50
56
140
280
28
70
140
6
14
28
As can be seen, a total sparge flow rate of 25 CFM at only a 10% utilization
provides as much oxygen as injecting 10 gpm of 1000 ppm HzOz assuming 100%
utilization. One of the problems with hydrogen peroxide, in addition to the relatively low
oxygen content at proper use rates, is that peroxide can be quite unstable in some soils.
In such cases its utilization is much less than 100% due to premature decomposition. In
these situations air sparging would have an even greater advantage as an oxygenation
source.
4-5
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In addition to effective oxygenation, air sparging can also remove contaminants
through volatilization, either directly, by "evaporating" the adsorbed phase, or,
indirectly, by stripping contaminated ground water.
In the first volatilization process, direct extraction, air bubbles that form during
air sparging traverse horizontally and vertically through the soil column, creating
transient air-filled regimes in the saturated soil matrix. Volatile hydrocarbons that are
exposed to this sparged air environment "evaporate" into the gas phase and are carried
by the air stream into the vadose zone where they can be captured by a vent system.
Whether a compound is extractable by an air sparge system is, as with soil vapor
extraction (SVE), determined by its vapor pressure. The practical vapor pressure limit
for an air sparging system, as it is for SVE, is -1 mm Hg.
In the second volatilization process, ground-water stripping, the key to successful
treatment is attaining good contact between the injected air and contaminated ground
water. Given good air-to-water contact, the effectiveness of air sparging in ground-water
treatment is then determined by the Henry's Law constant of the dissolved VOC
contaminants being treated by the air sparger system. A Henry's Law constant, KH (atm-
mS-mole-1), greater than 10-5 indicates a volatile constituent that can be removed by air
sparging. Table 4.2 lists the Henry's Law constant for several volatile hydrocarbon
constituents.
TABLE 4JL HENRY'S CONSTANT FOR SELECTED HYDROCARBONS
CONSTITUENT HENRY'S CONSTANT,
Cyclohexane 1.9 x 102
Benzene 5.6 x 10-3
Ethylbenzene 8.7x10^
Toluene 6.3 x 10^
Xylene 5.7 x HHJ
Naphthalene 4.1x104
Phenanthrene 2.5 x
The olefins and BTEX compounds are easily stripped from water by air sparging,
as indicated by their Henry's Law constants being much greater than 10-5 (atm-m3-
mole-1); heavier compounds such as PAHs are more difficult to remove.
When air sparging is applied, the result is a complex partitioning of the
petroleum hydrocarbon between the adsorbed, dissolved and vapor state, as well as a
complex series of removal mechanisms that may be engendered - removal as a vapor,
biodegradation, and removal as a solute in ground water. Which mechanisms are the
primary removal mechanisms and which are secondary depends on the volatility of the
contaminant. As shown in Figure 4.4, with a highly volatile product, the primary
partitioning is into the vapor state, and the primary removal mechanism is through
volatilization. By contrast, with a low volatility product, partitioning is primarily to the
4-6
-------
adsorbed or dissolved state, and the primary removal mechanism is through
biodegradation.
High Volatility Hydrocarbon Mixtures
Vapor Removal
Low Volatility Hydrocarbon Mixtures
Vapor Removal
*
A,
''Dissolved "i . Migranim /
Phase ' ~" Dllullnn
Biodegradauon
Migration/
' •*" Diluiiim
Key
Primary Mechanism
-----»• Secondary Mechanism
Figure 4.4. Air sparging partitioning and removal mechanisms as a function of volatility.
Because air sparging can both stimulate biodegradation as well as remove
hydrocarbon vapors, it can treat a wide range of petroleum hydrocarbon products. As
shown in Figure 4.5, the treatment mechanisms, however, vary with the type of product.
Heavy products such as No. 6 fuel oil are treated primarily through biodegradation.
Light products such as gasoline are treated more through simple volatilization than
through biodegradation.
100
80
i
! 60
o
S 40
20
\\\\\\\\\\
• Biodegradation'//
No 6 Fuel Waste Diesel Jel Fuel Mineral Gasoline
Oil Oil Spirits
Volatility
Figure 4J&. Air sparging removal mechanisms as a function of product volatility.
4-7
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4.5. DANGERS OF AIR SPARGING
A fundamental issue in any remedial process is having control over the process.
With processes that are based on extraction such as SVE or ground-water recovery, the
process begins with the system under control because contaminants are being drawn to a
point of collection. By contrast, injection systems such as air sparging start with no
control because flow is away from the injection point. Control must be gained and
maintained. Therefore, with air sparging, anything that affects the control of the flow of
air can limit the application of air sparging. There are two fundamental concerns with
the efficacy and utilization of air sparging. These concerns may be categorized as
structural and operational.
First, with respect to structural concerns, air sparging is based on the controlled
injection of air into a saturated soil matrix. The injected air traverses horizontally and
vertically through the soil. Anything that impedes the flow of air will impact the utility of
air sparging. Flow impedance may be caused by Hthological barriers that block the
vertical flow of the air. It may also be caused by channelization where the horizontal air
flow is "captured" by high permeability channels. The issue in considering structural
limitations to air sparging is understanding barriers to flow.
Second, with respect to operational concerns, it is important to keep in mind that
the injection of air can displace both vapors and water. Unless control is established and
maintained, this displacement can accelerate and aggravate the spread of
contamination. The issue in considering operational issues is understanding the control
of flow.
4.6. BARRIERS TO FLOW
The effectiveness of air sparging is dependent on the unrestricted flow of air
horizontally and vertically through the soil matrix. Anything that restricts or channels
air flow limits air sparging.
Geological barriers will obviously impact air flow. With a sparge system, air flow
must be both horizontal and vertical. The vertical travel is important for the ultimate
removal of the volatilized contaminant. If the geology restricts vertical air flow, then
sparging can push the dissolved contamination downgradient as shown in Figure 4.6.
Any less pervious zone (such as a clay barrier) above the zone of air injection may restrict
vertical air flow and severely reduce the effectiveness of air sparging. The barriers do not
have to be nonpervious but may simply be a gradation to material with a lower
permeability, which can restrict vertical air flow. The presence or absence of such
barriers should be determined during installation of the system and through a pilot test
study.
A second potential impact of geology is the presence of soil layers having higher
permeability than the sparging zone, which may intercept and channel air flow as
shown in Figure 4.7. Such channels are likely when the soil matrix is layered or highly
heterogeneous. The greater the degree of heterogeneity, the higher the risk of channeled
flow. Channeled air flow may cause the uncontrolled spread of contamination. To
minimize the risk of channelization, a complete Hthological profile of the sparging area
should be developed before the system is installed. The importance of channelized flow
should be evaluated during a pilot test.
-------
Contaminated Soil
Dissolved Panicles
Air/Contaminant Migration
Figure 4.6. Inhibited vertical air flow due to impervious barrier.
Figure 4.7. Channeled air flow through highly permeable zone.
4-9
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4.7. CONTROL OF FLOW
There are two potential concerns with the use of air sparging. Injection of air can
displace both vapors and water, and this displacement can accelerate and aggravate the
spread of contamination. Therefore, air flow must be controlled during a sparging
operation.
There are two operational conditions that may potentially cause a spread of
dissolved contaminants. The first is the injection pressure and flow. The second is
water table mounding.
A potential cause of dissolved contaminant migration, is over pressurizing the
sparge system. As discussed above, too high an air flow rate can physically displace
water. An illustration of this is afforded by air rotary drilling. An air rotary rig uses
high pressure/volume air (-100-200 psi, 300-600 cfm) to lift and remove cuttings. When
drilling below the water table, air rotary rigs have been known to "pop" covers off of
adjacent monitoring wells and cause water geysers.
Ostensibly, the minimum injection pressure is that which is required to overcome
the water column (i.e., 1 psi for every 2.3 feet of hydraulic head). As pressure is
increased above this minimum, air is "injected" laterally into the aquifer. As seen in
Figure 4.8, there is an initial linear relationship between the sparge pressure and
direction of air travel. At low sparge pressure (injection pressure equal to hydraulic
head) the air travels 1-2 feet horizontally for every foot of vertical travel. As the sparge
pressure increases, the degree of horizontal travel also increases. Enhanced horizontal
travel allows a single well to treat a greater area of aquifer. However, increasing the
pressure does not always provide a benefit. Increased pressure may cause air flow to
become turbulent, and the added pump energy is wasted. The danger under turbulent
conditions is that a dissolved plume of contaminants could be pushed away from the
sparge well. Figure 4.8 shows a point of inflection, where the increase in injection
pressure does not give a corresponding increase in air flow radius. This transition to
turbulent flow is also observed in venting systems where high vacuum can result in
frictional heating of the vent gases.
I 90n
s| 8°~
3*
|2 6°-
OIM jo-
40-
II
r* •¥
30-
20-
10-
Field Measurements
Turbulem Flow
Controlled Flow I (Potential for
Water
Displacement)
00 20 40 60 80 100 120 140 160 180 200
Ratio of Horizontal Radius vs Sparge Depth
Figure 4.8. Effect of iiqection pressure on air flow.
4-10
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A second potential cause of increased dissolved contaminant migration is water
table mounding. Air sparging does raise the level of the water table. Normally, ground
water would flow away from a mound. However, the mounding produced by sparging is
caused by the displacement of water with air. Flow away from the mound may not be
induced because the net density of the water column is decreased, thus counteracting the
mounding. This lowered density is dramatically seen by taking water table
measurements after the sparge system was shut off as shown in Table 4.3.
TABLE 43. WATER TABLE MOUNDING AND COLLAPSE
Depth to Water (ft) ©
Well # - Distance from Static Sparging 5 Min 10 Min
Sparge Point Water Level Water Level After After
MW-7
SE-1919
S-2629
NE-13
5
6.42
6.71
13
6.46
6.20
6.55
6.52
4.09
6.93
6.96
6.11
10.03
6.54
6.77
7.44
6.96
6.75
The water table collapses back to its original elevation after air injection is
stopped. This collapse shows the displacement of water by air during sparging.
Mounding and collapse is greater for monitoring points close to the sparge point.
Because of this density compensation, mounding may not spread any contamination.
Additionally, if the air flow rate is high enough during sparging, any dissolved
constituents can be stripped before they migrate away from the treatment area. Sparging
has been successfully applied with no evidence of ground-water contaminant migration
(Brown et al., 1991; Middleton and Miller, 1990).
The second "danger" of sparging is accelerated vapor travel. This is of concern
when the product is volatile and where there are receptors. Since air sparging increases
pressure in the vadose zone, any exhausted vapors can be drawn into building
basements. Basements are generally low pressure areas, and this can lead to
preferential vapor migration and accumulation in basements. As a result, in areas with
potential vapor receptors, air sparging should be done with a concurrent vent system. A
vent system provides an effective means of capturing sparged gases.
4.8. SUMMARY OF LIMITATIONS
As with any technology, there are limitations to the utility and applicability of air
sparging. Understanding those limitations is important to the proper development and
4-11
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use of air sparging. As discussed above, there are several types of limitations to air
sparging technology. The first is the type of contaminant. For air sparging to be
effective as a removal mechanism, the contaminant must be volatile and insoluble. If the
contaminant is soluble or nonvolatile, the contaminant must be biodegradable. In the
case of volatile or insoluble contaminants, air sparging functions as an extractive
process as well as a biodegradative process. In the case of biodegradable contaminants,
air sparging is a destructive process.
The second limitation to air sparging technology is the geological character of the
site. The most important geological characteristic is structural homogeneity or
heterogeneity. If there is stratification present in the saturated zone, there is a possibility
that sparged air could be held below an impervious layer and thus spread laterally,
causing the contamination to spread. To guard against such occurrences, any bore hole
to be used as a sparge point should be logged by continuous coring over its entire depth
before installation of the well. If stratification is present and sparging is to be used, all
lithological units above the sparge interval should be of equal or greater permeability
compared to the unit in which the sparge point is screened. Optimally the permeability
of the geologic material above the screened interval of the sparge well should increase
with increasing elevation until the water table is reached.
The second most important geological characteristic is permeability. In many
geological environments, the permeability in the vertical direction is less than
permeability in the horizontal direction. The permeability should be sufficient to allow
the sparge air to move through the aquifer matrix, both horizontally and vertically. If
flow is impeded in either direction because of low permeability, then sparging may be
precluded. If the ratio of horizontal to vertical permeability is low(<2:l), sparging can be
effective even though the general permeability is also low (>1O5 cm/sec). If the ratio of
horizontal to vertical permeability is high (>3:1), then the general permeability must be
higher (>1(H cm/sec) for sparging to be effective.
Finally, there are physical constraints on the operation of a sparge system. The
constraints are primarily depth related. There is both a minimum and maximum depth
for a sparge system. The minimum depth, 4 feet, is the saturated thickness required to
confine the air and force it to "cone-out" from the injection point. If there is insufficient
saturated thickness, then the air could short-circuit around the sparge point. The
maximum sparge depth, 30 feet, is important from the standpoint of
control/predictability. At depths greater than 30 feet it is difficult to predict where the
sparge air will travel, making it difficult to design a control system for containing the
sparged air with the area being treated and/or to capture the sparge air once it exits the
saturated zone. Any small layers with low permeability lying above the sparge interval
could have a drastic effect on the air movement. The ability to detect such layers becomes
increasingly more difficult with greater sparge depths. Thus the risk of improper
design/control also increases. A second depth related constraint is the depth to water.
There needs to be sufficient unsaturated soils to allow for the installation of a soil vapor
extraction system so that the VOCs mobilized by sparging can be captured. The
minimum depth that is required for installation of a soil vapor extraction system is four
(4) feet.
To assure the effectiveness of a sparge system, proper consideration must be given
to the potential limitations to the technology. As discussed above, these limitations are
based on the properties of the contaminant being treated, the geological characteristics of
the site, and on the physical limitations to the technology. Table 4.4 summarizes the
limitations to sparging.
4-12
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TABLE 4.4. LIMITS TO THE USE OF Am SPARGING
FACTOR
PARAMETER
LIMIT I DESIRED RANGE
Contaminant
Geology
Physical
Volatility
Solubility
Biodegradability
Heterogeneity
Hydraulic
conductivity
Sparge Depth
Depth to Water
>5mm Hg
<20,000 mg/1
BOD5 >.01 mg/1,
BOD5:ThOD >.001
No impervious layers
above sparge point
If layering present
hydraulic conductivity increases
above sparge point
>1O6 cm/s if horizontal:
vertical is <2:1
>1CH cm/s if horizontal:
vertical is >3:1
>4 feet, <30 feet
>4feet
4.9. SYSTEM APPLICATION AND DESIGN
The best assurance for effective system performance is proper system design. Air
sparging systems can only be properly designed through collection of appropriate site
data and field pilot testing. Field pilot testing necessitates the installation of test
sparge/vent points. The installation of the pilot test system provides an opportunity to
identify any subsurface barriers or irregularities that may restrict air flow. An air
sparging system can be correctly designed only if sufficient data concerning site
conditions have been determined. The data requirements consist of:
1. the nature and extent of site contaminants,
2. specifics of the site hydrogeology, and
3. thorough knowledge of potential ground water and vapor receptors.
4.9.1. Nature and Extent of Site Contaminants
The volatility of the petroleum hydrocarbon being treated should be determined.
The higher the volatility, the more vapor transport will be a factor. Vapor transport can
be beneficial in that it accelerates treatment by physically removing volatile petroleum
4-13
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hydrocarbons. However, vapor transport can also be a problem in that it must be
controlled and treated.
The mass distribution of the site contamination must be known in order to
effectively utilize air sparging. The vertical extent of adsorbed phase contaminants at or
below the water table must be determined in order to effectively determine the depth of
sparging wells. The lateral extent of adsorbed phase contamination below the water table
must be known to ensure complete remedial system coverage. In addition, the down-
gradient dissolved ground-water concentrations should be delineated in order to allow
monitoring of the plume during sparging operation and placement of recovery or sparge
wells for ground-water treatment.
4.9.2. Hydrogeologic Conditions
Several hydrogeologic parameters are of great concern in ensuring correct design
and operation of an air sparge system. The soil texture must allow for air transmission
in order for volatilization or biodegradation to occur. In general a hydraulic conductivity
of >1O5 cm/sec is necessary for effective air sparging. Poorly compacted fill materials are
also a poor choice for an air sparging system as they may exhibit settling if subjected to
high air pressures.
Of even more importance is the homogeneity of the site soils. Permeability
contrasts due to natural stratigraphic changes or differential filling by human activity
will alter the air flow. Lower permeability lenses will create a barrier to the upward
moving air and can cause lateral spread of the contaminants. High permeability
channels may "capture" the air stream and also cause contaminants to spread.
4.9.3. Potential Ground-water and Vapor Receptors
Since injected air flow can displace both vapors and liquids, the proximity of vapor
or ground-water receptors should be determined before a sparging system is installed
and operated. If such receptors are present, system safeguards should be used.
The first safeguard is the use of soil vent systems. Soil vent systems are
mandatory where volatile hydrocarbons are being treated and there are potential vapor
receptors, or where vapor phase controls are required. The vent system should be
designed to have a greater flow than the sparge system and should have a greater radius
of influence. Barrier soil vent systems can also be placed between the sparge system and
potential vapor receptors.
The second safeguard is to utilize ground-water control to prevent the migration of
dissolved contaminants. Active pumping systems or effective barriers should be
installed. Air flow does not follow hydraulic gradients. Therefore, ground-water control
should be installed where receptors exist and not just downgradient. Ground-water
controls may be necessary in areas where the geology is heterogeneous or of low
hydraulic conductivity (<1(H cm/sec). Ground-water controls may consist of water
collection systems or of an outer sparge system used as a barrier system.
4.10. FIELD PILOT TESTING
Air sparging requires a balanced air flow. Too low a flow can result in a loss of
remedial effectiveness; too high a flow can result in a loss of control. Because of the
potential for loss of control, an air sparge system should never be installed without a pilot
4-14
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test. Installation of a sparge system requires proper design of the separate components -
the vent system (if volatile hydrocarbons are present) and the sparge system, as well as
balancing of the two components. The basic design data to be determined by a pilot test
are:
1) The radius of influence of the air sparging system conducted at different
injection flows/pressures.
2) The radius of influence of the vacuum extraction system.
3) The pressure and vacuum requirements for effective treatment and effective
capture of volatilized materials.
The field tests consist of up to three sequential tests. The first test is a sparge
radius of influence test. The second is a vacuum radius of influence test. The third is a
combined sparge/vent test. The second and third are required at sites where vapor levels
are a concern.
A number of different parameters can be measured during the tests to determine
radius of influence. These include:
- Vacuum or pressure vs. distances. This is an indication of radius of influence.
- VOC concentrations in soil vapor or ground water. This is an indication of
what is being removed and of areas being impacted. Concentrations should be
determined before, during (with and without the system running) and after
each test.
- CO 2 and 0% levels in soil vapor. This is an indication of biological activity.
These measurements need to be taken before, during and after each pilot test
under static as well as pumping conditions.
- Dissolved oxygen (DO) levels in water. This is a good indicator of effectiveness.
In areas contaminated with petroleum hydrocarbons, static DO levels are
generally < 2 mg/1. Sparging should raise the DO level substantially. Good
initial DO measurements are required to determine changes.
- Water levels before and during test. Air flow during sparging will cause some
mounding. Levels should be recorded before the test to determine background.
Using multiple parameters allows for cross correlation during design. With this
cross correlation, it is possible to determine effective air flow through the area of
contamination and ensure capture of the volatilized materials. There is generally good
agreement among parameters as shown in Figure 4.9.
4.11. DESIGN DATA REQUIREMENTS
At the conclusion of the site characterization and pilot test, a complete set of
design data should have been collected. Table 4.5 lists the different data required and
their significance for design.
4-15
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Flow Flow Flow
I 1 3
Background (T-0)
Distance From Sparge Point, Ft.
Figure 49. Agreement between sparge parameters in estimating the radius of influence.
TABLE 4.5. SITE AND PILOT TEST DATA NEEDED
DATA
IMPACT ON DESIGN
Lithological Barriers
Vertical Extent of Contamination
Horizontal Extent of Contamination
Volatility of Contaminant
Sparge Radius of Influence
Optimal Flow Rates
Vent Radius of Influence
Vacuum/Pressure Balance
Vapor Levels
Feasibility/Sparging Depth
Sparging Depth
Number of Sparge Wells
Vapor Control (Venting)
Well Spacing/Flow Requirement
Compressor Size
Well Spacing
Blower Size/Well Placement
Vapor Treatment
4-16
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4.12.
A sparge system consists of a number of different elements. Some of them are
essential, while others are optional. Their use is dictated by the type of product or by the
site conditions. Table 4.6 includes a list of the system elements and their importance for
sparging.
TABLE 46. Am SPARGING SYSTEM ELEMENTS
COMPONENT
IMPORTANCE
DESCRIPTION
Sparging Well
Essential
Air Compressor
Essential
Monitoring System Essential
Heat Exchanger Optional
Muffler Optional
SVE System Optional
Vapor Treatment
Optional
Ground-water Control Optional
Sparging wells consist of a small section of pervious pipe (slotted pipe
or diffuser) placed at the bottom of a borehole, set below the zone of
contamination (Figure 4.10) A sparge interval should be set every 10-
15 ft below the water table The borehole is grouted above the sparge
interval Above 15 psi, well material should be steel (Compressed Gas
Association, 1989)
The compressor should be capable of delivering 10-20 cfm per well at 1-
3 times breakout pressure Breakout pressure = (Depth/2.3) Flow rate
is determined by ground-water flow and soil volume being treated A n
air to water ratio of 10-20.1 is desirable
A mom ton ng system is necessary to achieve and maintain control
The basic system (Figure 411) should consist of a shallow monitoring
well to measure water table elevation. DO, VOCs, and pressure as well
as vadose zone vapor probes t> monitor VOCs and pressure/vacuum
A heat exchanger, i e , thermal vanes, should be used with PVC (Air
Inlet) systems, depending on injection pressure
In urban areas, a muffler may be required to meet noise abatement
requirements
Where volatile hydrocarbons are being treated and vapor receptors or
vapor control requirements exist, an SVE system is necessary to
capture the VOCs mobilized by the sparge system The SVE system
should be designed such that a net negative pressure is maintained in
the treatment area. The total flow of the SVE system should be at least
twice the total flow of the sparge system SVE system elements include
wells er trenches and a vacuum pump or blower
If air quality is a concern, vapor treatment may be required for the SVE
system. Vapor treatment options include thermal treatment (catalytic
or thermal oxidation) as well as biotreatment Petroleum hydrocarbon
vapors are generally biodegradable and may be effectively treated by
diffusion through a soil bed a- compost bed.
If ground-water receptors or control of existing migration are an issue,
or if there is a risk of ground-water migration due to high
heterogeneity, then a ground-water control system may be required
This may consist of barrier or interceptor wells, or of a barrier sparge
system With petroleum hydrocarbons, dissolved contaminants will
be stripped by the air flow and will be biodegraded by enhanced
oxygenation Therefore long term ground-water controls are not an
essential part of the system The primary issue in deciding if ground-
water controls are required is the need for immediate containment of
the contamination
4-17
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Cement —»•
Grout —••
Sand •
Grout •
Bentonite -
Sand-
-asing
Figure 4.10. Nested sparge welL
.Sparge Screen (1-2')
.02 Slot
Sand
Vapor Probe
(.04 Slot)
Vapor Probe
(.04 Slot)
Figure 4.11. Monitoring point for sparging systems.
4-18
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4.13. SYSTEM EXAMPLES
The following two examples show the installation of an air sparging system for
shallow ground-water aquifers:
Site A-
The subsurface environment at the site generally consists of fill material overlying
a continuous sheet of naturally occurring Quaternary sediments (medium sands).
Within the southern geographical portion of the property, the Quaternary sediments rest
unconformably on top of the sediments of the Potomac Formation, which in turn overlie
the basement complex consisting of a volcanic intrusive rock, probably granodiorite. The
geology observed during drilling activity indicates that the saturated Quaternary
sediments are relatively homogeneous across the property. A natural barrier (clays of
the Potomac Formation) exists, which locally minimizes the potential for vertical
downward migration of dissolved phase total petroleum hydrocarbon and chlorinated
VOCs present in the shallow water-bearing zone into deeper water-bearing units.
The flow of shallow ground water at the property appears to trend northwest (NW)
to southeast (SE), under an average hydraulic gradient of 0.021 ft/ft across the property.
The gradient does not vary appreciably across the property, ranging from 0.015 ft/ft to
0.026 ft/ft.
The bulk of the contamination appears to be located within two soil horizons; one
shallow (-3-9 ft), and one above, at, and just below the water table (15-18+ ft). Thus, it
may be concluded that there is soil contamination in both the unsaturated (vadose) and
saturated zones (water table aquifer). Second, the soil contamination is primarily
isolated to the former tank field area and extends hydrogeologically laterally and
downgradient in the direction of shallow ground-water flow. Based on analysis of the
soil, it is estimated that approximately 300 to 500 pounds of contamination exist in the
upper horizon and an additional 200 to 300 pounds exist in the lower horizon.
Using data obtained during pilot testing, a pattern of vent and sparge points was
developed to provide overlapping influence (negative net pressure) and favorable site
coverage for the treatment system. Additional probe nests were strategically placed to
monitor system performance. A complete list of treatment and monitoring points
installed at the site is specified below, and pictured in Figure 4.12.
7 Combination vapor extraction/air sparge points (AS/VP1-AS/VP7); to be
installed.
1 Vapor extraction only point (VP1).
- 7 Sparge only points (AS1-AS7).
8 Vapor monitoring probe nests (PR1-PR8).
The 7 vent/sparge points form a rough ellipse surrounding the former tank field
area and extending to the property perimeter. The 7 innermost sparge only points were
specified to complete coverage and to provide concentrated treatment within the former
tank field area, where contaminant levels are highest.
The vent system was operated at -40 inches of water vacuum and -60 CFM per
point for a total flow of -500 CFM. The sparge system was operated at 10 psi and a flow of
16 CFM per point for a total flow of -225 CFM. Sparge system flow rate was designed to be
-------
f~] PR4 - Monitor Probe
A AS I-Sparge Pi.
O ASV3 - Sparge/Vent PC.
O VPI - Vent Pt
MW-7
Main St.
i r
Figure 4.12. System layout Site A.
SiteB:
Based upon observations made during drilling, the site geology material consists
of approximately eight to fifteen feet of a brown to black sandy, silty clay with minor
occurrences of ash, gravels and other construction debris. The fill material rests
unconformably atop the Marcellus Shale, which is continuous to at least 35 feet below
surface grade. The Marcellus Shale exhibits a deteriorated surface at the contact
between the unconsolidated and consolidated material, indicating that the surface was
previously exposed to weathering processes. On the northern portion of the facility, the
shale exhibits minor fracturing, but the shale is competent on the southern portion of the
facility.
4-20
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Ground water occurs within the unconsolidated sediments at a depth ranging
from approximately three to six feet below surface grade under water table conditions.
Shallow monitoring well elevational data indicate the major component of ground-water
flow is approximately north to south. A minor component of ground-water flow appears
to occur preferentially northeast to southwest. The hydraulic gradient at the water table
is estimated to average 0.015 ft/ft across the property.
Air sparging tests were conducted at three pressure levels; 10, 12 and 14 psi. The
test used a 100 psi diesel-powered air compressor. The sparge test points (SP1 and SP2)
were installed to a depth of 11 and 11.5 feet, respectively, below the water table and are
screened for the bottom two feet. Sparge tests were performed by injecting air into the
sparge test point at 10, 12 and 14 psi; the flow rates that correspond to these pressure
levels ranged from 40 to 53 CFM. The resulting pressure and VOC levels were measured
in the seven vapor probes, which were previously utilized for the vapor extraction tests.
Sparging influence was considered present at an induced pressure level of 0.01 inches of
water column.
The sparge results appeared to be radial and did not indicate any directional
orientation/correlation. Increasing the applied pressure did not appear to have any effect
on the radius of influence at a given test point. The change in VOC concentrations
during the sparge tests was significant. In most probes the VOC levels, as indicated by
OVA readings, increased significantly with sparging. During the sparge tests a 31 to 40
foot radius of influence (at 0.01 inches of water column induced pressure) was developed.
This test demonstrated the feasibility of sparging for VOC treatment at and below the
water table.
Based on data obtained during the hydrogeologic investigation, soil gas survey and
pilot tests, a pattern of vapor extraction trenches and sparge points was developed to
provide overlapping influence (negative net pressure) and favorable site coverage for the
treatment system. A complete list of treatment and monitoring points for the system is
specified below.
36 sparge points (to be installed);
• 2,200 feet of vapor extraction trenches (to be installed);
12 stainless steel vapor monitoring drive point probes (PR1-PR12 to be
installed);
• 11 shallow monitoring wells (MWl-3, 48, 5-7 existing; MW8-11 to be installed);
• Ideep monitoring weU (MW-4D existing); and,
5 potential sump locations.
In order to provide assurance that adequate vacuum would be induced across the
site, the pattern of vent locations necessary for full coverage was determined by
assuming a maximum trench spacing of 45 feet. This is one-half the calculated radius of
influence (ROD of the vapor extraction trench network. This treatment system layout is
designed to maintain a net negative pressure and thus capture VOC contaminated soil
gas both on and off the property. Figure 4.13 indicates the proposed location of the vapor
extraction trenches and the sparging wells.
Vapor extraction will be accomplished using a 50 Hp blower, having a capacity of
2,500 CFM at 60 inches of water column vacuum. Influent vacuum/flow rate will be
controlled with an ambient air intake valve. A liquid knockout tank, particulate filter
and muffler will be placed on the influent line to eliminate or reduce water generated
during system operation, solids, and noise, respectively. An effluent muffler was
4-21
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specified to further reduce noise levels to minimize the impact to nearby residents. A
12,000 pound granular activated carbon (GAG) unit was specified on the vapor extraction
effluent to remove contaminants from the extracted air prior to discharge.
In order to ensure favorable site coverage, a 30 foot radius of influence was
assumed in designing the pattern of points to be utilized for the final system. A total of 96
sparge locations were specified to provide this coverage (Figure 4.13). The proposed
location of the sparge points (to the east and south of the facility building) was based on
ground water and soil analytical results obtained during the hydrogeologic investigation.
Operation of 36 sparge points at 40 CFM each results in a potential total flow rate of 1,440
GFM. However, as with vapor extraction, significant competition between points will
result in reduced air flow rates. Based on this fact and experience from a similar site, a
flow rate of 860 CFM was used for the design. The vapor extraction system should be able
to easily capture the sparge air since the sparge system design flow rate is roughly one-
third of the vapor extraction system design flow rate of 2,500 CFM. The sparge air will be
provided by a 75 Hp rotary lobe-type blower capable of delivering 860 CFM at 10 psi.
u
i
(/>
&
MW-ll
Residence Service Station
Figure 4.13. Layout of site B air sparging/Vent system.
4.14. COST FACTORS
The cost for a sparge system is dependent on the size of the site, the degree of
contamination, the application depth, the geology (permeability, heterogeneity) of the
site, and the permitting requirements. Table 4.7 lists the approximate costs for a one acre
site having shallow ground water (DTW <20 ft), and a moderate permeability (fine to
medium sand).
4-22
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TABLE 4.7. APPROXIMATE COST FACTORS
FACTOR
COST, $ OPTIONAL
COST INFLUENCES
Sparge Well (PVC)
Sparge Well (Steel)
Total Wells (10)
Vent Wells (5 additional)
Piping & Trenching (PVC)
Piping & Trenching (Steel)
Compressor (300 cfm, 20 psi)
Vacuum Blower (1000 cfm, 100")
Vapor Treatment (Carbon)
Vapor Treatment (Thermal)
Permitting (Air)
Permitting (Water)
Pilot Test
Design
Construction
O&M (per Year)
Reporting (per Year)
Total without Options
Total with Vapor Control
Total Maximum
3,000
5,000
30,000
12,000
45,000
60,000
15,000
20,000
180,000
120,000
15,000
25,000
20,000
25,000
50,000
75,000
30,000
260,000
432,000
570,000
N
Y
N
Y
N
Y
N
Y
Y
Y
Y
Y
N
N
N
N
Y
Depth, Diameter
Depth, Diameter
HOP, Area
ROIa, Sparge flow, Area, DTWb
Area, Depth
Area, Depth
Flow, Pressure
Flow, Vacuum
Flow, Concentration
Flow, Concentration
Regulations
Regulations
Depth, Geology
Flow, VOCs, Area, Regulations
Flow, VOCs, Area, Regulations
Regulations, Flow, VOC
Concentrations
Regulations
a Radius of influence
Depth to water
4.15. CONCLUSION
While soil vapor extraction has long been recognized as an effective means of
removing volatile organics from subsurface soils, it has been limited to treatment of
unsaturated soils. Where contamination exists below the water table, soil vapor
extraction is limited and can only be used with an extensive and often costly dewatering
operation.
Air sparging is a means of extending the utility of vapor extraction technology to
the saturated regime. With air sparging, air is injected under pressure below the water
table, creating a transient air-filled porosity. This enhances biodegradation as well as
volatilization of petroleum hydrocarbon contaminants from the soil and ground water.
The net result is a rapid and significant decrease in contaminant levels.
Air sparging has two inherent dangers. First, the VOC-laden air stream can
rapidly migrate through the vadose zone to low pressure zones such as basements,
causing a vapor hazard. To prevent this occurrence, a sparge system should be operated
in conjunction with a vent system. A second danger is that the injected air can mobilize
ground-water contaminants rather than stripping them, causing accelerated
4-23
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downgradient migration. This can occur if there are vertical barriers to air migration
causing the air to be trapped producing lateral spread. It can also occur if too much
pressure is used, physically displacing the water column.
Because of these dangers, proper design is essential. This necessitates a field pilot
test and careful site delineation. With proper design, the use of sparging can
substantially and rapidly remediate ground-water contamination. The key to the effective
use of air sparging is proper design. This entails first understanding the distribution of
contaminants across the site. Second, the flow dynamics of air must be determined in
the vadose zone and in the saturated zone. This can only be done through a properly
designed pilot study. Once both the contaminant distribution and the flow dynamics are
known, the number, location and type of treatment wells can be specified. This, in turn,
leads to the equipment specifications. Through a careful and phased design process, air
sparging can be an effective remedial system.
To further expand the utility of sparging, there are a number of questions that
need to be addressed. These questions are:
What are the limitations to air sparging technology?
How does air sparging impact the site hydrogeology and contaminant
transport?
What are the most effective means of determining the radius of influence,
pressure requirements, and effectiveness of a sparge system to minimize
detrimental effects?
With effective design and careful monitoring, air sparging can be an important
remedial tool. If it is applied in a simplistic fashion, air sparging can be ineffectual at
best or counter-productive at worse.
4-24
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REFERENCES
Brown, R.A., C. Herman, and E. Henry. 1991. The use of aeration in environmental
clean-ups. In: Proceedings, Haztech International Pittsburgh Waste Conference.
Pittsburgh, Pennsylvania. May 1991.
Clayton, W.S., R.A. Brown, and K.L. Brody. 1989. The reduction of groundwater
contamination by vapor extraction of volatile organics from the vadose zone. In:
New England Environmental Expo. Boston, Massachusetts. May 1989.
Compressed Gas Association. 1989. Handbook of Compressed Gases. Van Nostrand
Reinhold. New York, New York. 657 p.
Hiller, D., and H. Gudemann. 1988. In situ remediation of VOC contaminated soil and
groundwater by vapor extraction and groundwater aeration. In: Proceedings,
Haztech '88 International. Cleveland, Ohio. September 1988.
Hoag, G.E., and M.C. Marley. 1984. Induced soil venting for the recovery/restoration of
gasoline hydrocarbons from the vadose zone. In: Proceedings of the Petroleum
Hydrocarbons and Organic Chemicals in Ground Water Conference. National
Water Well Association, American Petroleum Institute. Houston, Texas.
November 1984.
Marley, M.C., M.T. Walsh, and P.E. Nangeroni. 1990. Case study on the application of
air sparging as a complementary technology to vapor extraction at a gasoline spill
site in Rhode Island. In: Proceedings, HMC Great Lakes 90. Hazardous
Materials Control Research Institute. Silver Spring, Maryland.
Middleton, A.C., and D.H. Hiller. 1990. In situ aeration of ground water - a technology
overview. In: Proceedings, Conference on Prevention and Treatment of Soil and
Groundwater Contamination in the Petroleum Refining and Distribution
Industry. Montreal, Quebec, Canada. October 1990.
Raymond, R.L., V.W. Jamison, and J.O. Hudson. 1975. Biodegradation of high-octane
gasoline in groundwater. Development in Industrial Microbiology. Volume 16.
American Institute of Biological Sciences. Washington, DC.
4-25
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SECTION 5
GROUND-WATER TREATMENT FOR CHLORINATED SOLVENTS
Perry L. McCarty
Lewis Semprini
Western Region Hazardous Substance Research Center
Stanford University, Stanford, California 94305-4020
Telephone: (415)72*4131
Fax: (415)725-8662
5.1. INTRODUCTION
Chlorinated solvents and their natural transformation products represent the
most prevalent organic ground-water contaminants in the country. These solvents,
consisting primarily of chlorinated aliphatic hydrocarbons (CAHs), have been used
widely for degreasing of aircraft engines, automobile parts, electronic components, and
clothing. Once dirty, chlorinated solvents often have been disposed into refuse sites,
waste pits and lagoons, and storage tanks. Because of their relative solubility in water
and their somewhat poor sorption to soils, they tend to migrate downward through soils,
contaminating water with which they come into contact. Being denser than water, their
downward movement is not impeded when they reach the water table, and so they can
penetrate deeply beneath the water table. CAHs have water solubilities in the range of 1
g/1, or several orders of magnitude higher than the drinking water standards for those
that are regulated.
The major chlorinated solvents used in the past are carbon tetrachloride (CT),
tetrachloroethene (PCE), trichloroethene (TCE), and 1,1,1-trichloroethane (TCA). These
compounds can be transformed by chemical and biological processes in soils to form a
variety of other CAHs, including chloroform (CF), methylene chloride (MC), cis- and
trans- 1,2-dichloroethene (cis-DCE, trans-DCE), 1,1-dichloroethene (1,1-DCE), vinyl
chloride (VC), 1,1-dichloroethane (DCA), and chloroethane (CA). These chemicals, their
solubilities in water, and drinking water maximum contaminant limits (MCL), if
applicable, are listed in Table 5.1. This is the group of chemicals generally to be
addressed as a result of chlorinated solvent contamination of ground water.
Just over one decade ago, most of the compounds listed in Table 5.1 were
considered to be nonbiodegradable. Transformation products of the chlorinated solvents
then started to be found in ground waters, and this led to expanded efforts to determine
the chemical and biological processes responsible. It was found that most of the CAHs
can in fact be transformed by biological processes, but generally, the microorganisms
responsible cannot obtain energy for growth from the transformations. The
transformations are brought about by cometabolism, or through interactions of the CAHs
with enzymes or cofactors produced by the microorganisms for other purposes. There
are now widespread efforts to take advantage of cometabolism for the transformation of
CAHs in ground water, but this is a much more complicated process than the usual
biological treatment processes that have been used for years, in which organic compound
destruction is accomplished by organisms that use the compounds as primary substrates
for energy and growth. In cometabolism, other chemicals must be present to serve as
primary substrates to satisfy the energy needs of the microorganisms, and indeed must
be tailored so that they can stimulate the production of the biological agents that affect
cometabolism of the CAHs.
5-1
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TABLE 5.1. COMMON HALOGENATED ALIPHATIC HYDROCARBONS
Compound
Carbon Tetrachloride
Chloroform
Methylene Chloride
1, 1, 1-Trichloroethane
1 , 1 -Dichloroethane
1,2-Dichloroethane
Chloroethane
Tetrachloroethene
Trichloroethene
cis-l,2-Dichloroethene
trans-1,2-
Dichloroethene
1 , 1 -Dichloroethene
Vinyl chloride
Formula
CC14
CHC13
CH2C12
CHsCCls
CHsCHC^
CH2C1CH2C1
CH3CH2C1
CC12=CC12
CHC1=CC12
CHC1=CHC1
CHC1=CHC1
CH2=CC12
CH2=CHC1
Acronym
CT
CF
MC
TCA
1,1-DCA
1,2-DCA
CA
PCE
TCE
cis-DCE
trans -
DCE
1,1-DCE
VC
Density
1.595
1.485
1.325
1.325
1.175
1.253
1.625
1.462
1214
1.214
Water
Solubility
(mgll)
800
8,200
13,000
950
5,500
8,700
150
1,000
400
400
U.S. Drinking
Water MCL
(1*11)
5
100
200
5
5
5
70
100
7
2
Much has already been learned about cometabolism of CAHs. However, full-scale
field applications of this process are greatly limited, and there are virtually no
sufficiently well-documented full-scale applications at present that can be used to guide
design and application or that can be used to evaluate costs. Thus, any application of
bioremediation for chlorinated solvent destruction in the field must be considered as a
research activity and should be evaluated as such. As with any new and untested
process, failure to reach desired goals should be anticipated, and surprises can be
5-2
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expected. Nevertheless, the understanding of the process is now at a stage where full-
scale experimentation is desirable, and indeed is a necessity if biodegradation of
chlorinated solvents is to become a reality rather than just a laboratory curiosity.
The purpose of this chapter is to provide background information on the state of
knowledge of CAH biodegradation, to discuss field as well as laboratory testing of the
process, to summarize the potential application of biological destruction for various
CAHs, and to discuss the effect of site conditions on the probability for success of in-situ
field applications.
5.2. BIOTRANSFORMATIONOFCAHS
5.2.1. Primary Substrates «»*»d Cometabolism
Organic compounds can be biotransformed by microorganisms through two
basically different processes: (1) use as a primary substrate, and (2) cometabolism. In
the first process biodegradation occurs when the organism consumes the organic
compound as a primary substrate to satisfy its energy and organic carbon needs. This is
the usual process for organic decomposition in nature and the process generally
captured in the vast majority of biological treatment processes designed for municipal
and industrial wastewaters. Knowledge about organism growth and kinetics of primary
substrate utilization is quite extensive.
Cometabolism, on the other hand, is the fortuitous transformation of an organic
compound by enzymes or cofactors produced by organisms for other purposes. Here, the
organisms obtain no obvious or direct benefit from the transformation. Indeed, it may be
harmful to them. Cometabolism is also a natural process, but it has not been used
extensively for treatment of organic wastes, and knowledge of the process and its
practical application are by comparison quite limited. Cometabolism, however, is the
process by which most of the CAHs can be biotransformed. Because of the potential
usefulness of biotransformation of CAHs, there is much ongoing research to learn how
cometabolism might be applied. For cometabolism to occur, an active population of
microorganisms having the cometabolizing enzymes or cofactors must be present. This
means that the appropriate primary substrates for growth and maintenance of these
organisms must also be present. This is an aspect that adds greater complexity and cost
to cometabolic biotransformations.
Biotransformation through primary substrate utilization or through
cometabolism may occur under either aerobic or anaerobic conditions. Table 5.2 contains
a summary of the CAHs that have been shown to be degraded by the two different
processes under aerobic or anaerobic conditions. Relative information on transformation
rates under the different processes is also indicated, and transformation rates are
discussed subsequently in greater detail.
Transformations of CAHs in the natural environment also can occur both
chemically (abiotic) and biologically (biotic). The major abiotic and biotic transformation
processes occurring in natural systems are summarized in Table 5.3 (Vogel et al., 1987).
The abiotic processes most frequently occurring under either aerobic or anaerobic
conditions are hydrolysis and dehydrohalogenation. Abiotic transformations generally
result in only a partial transformation of a compound and may lead to the formation of a
new compound that is either more readily or less readily biodegraded by
microorganisms. Biotic transformation products are different under aerobic than
anaerobic conditions. When used as a primary substrate, organic chemicals are
5-3
-------
11-
generally completely mineralized under both aerobic and anaerobic conditions. However,
with cometabolism, as with abiotic transformations, CAHs are generally transformed
only partially by the biological process. The eventual fate depends upon other abiotic or
TABLE 5.2. POTENTIAL FOR CAH BIOTRANSFORMATION AS A PRIMARY SUBSTRATE OR
THROUGH COMETABOLISM
Primary Substrate Cometabolism
Compound Aerobic Anaerobic Aerobic Anaerobic
Potential Potential Potential" Potential"
CC14
CHClg
CH2C12 Yes
CH3CC,3
CH3CHC12
CH2C1CH2C1 Yes
CH3CH2C1 Yes
CC12=CC12
CHC1=CC12
CHC1=CHC1
CH2=CC12
CH2=CHC1 Yes
0
X
Yes XXX
X
X
X
XX
0
XX
XXX
X
xxxx
xxxx
XX
xxxx
XX
X
b
XXX
XXX
XX
XX
X
CAH
Product
CHC13
CH2C12
CH,CHC,2
CHaCHzCl
CH 3Cri2Cl
CHC1=CC12
CHC1=CHC1
CH2=CHC1
CH2=CHC1
8 0- very small if any potential; X - some potential; XX - fair potential; XXX • good
potential; XXXX - excellent potential.
>> Readily hydrolyzed abiotically, with half-life on order of one month.
5-4
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biotic reactions that might occur. Aerobic biotic transformations generally are
oxidations and are classified as hydroxylation, or the substitution of a hydroxyl group on
the molecule, or epoxidation, in the case of unsaturated CAHs. The anaerobic biotic
processes generally are reductions that involve either hydrogenolysis, the substitution of
a hydrogen atom for chlorine on the molecule, or dihaloelimination, where two adjacent
chlorine atoms are removed, leaving a double bond between the respective carbon atoms.
TABLE 5& TRANSFORMATIONS OF CAHs (AFTER VOGEL FT AL., 1987)
Reactions
\. Substitution
a solvolysis. hydrolysis
RX + H20 - ROM +• HX
b. other nucleophilic reactions
RX + N- — RN + X-
II. Oxidation
a a - hydroxylation
I I
-C-X + H2O — -C-X
I I
H OH
b. epoxidation
y 0 V
C - C +H2O —^C-VC/+2H++2e-
/ \ ' N
III. Reduction
a. hydrogenolysis
nv , • tA . 0—- DU j^ V*
KA T n T *c ^* Kn T A
b dihaloelimination
II \ /
-C— C- + 2e- — C = C +2X-
II / ^
X X
c coupling
2 RX + 2e- -* R - R + 2X'
IV. Dehydrohalogenation
I I \ /
•C—C C = C +HX
I I / \
X H
Examples
CH3CH2CH2Br + H2O -* CH3CH2CH2OH + HBr
CH3CH2Br + HS' -> CH3CH2SH + BR'
CH3CHCI2 + H2O — CH3 CCI2 OH + 2H+ + 2e-
CHCICCI2 + H2O -» CHCIOCCI2 + 2H+ i- 2e"
CCU + H+ + 2e- — CHCI3 + Cr
CCI3CCI3 + 2e- -* CCI2 CCI2 + 2CI'
2 CCI4 + 2e- -* CCI3 CCI3 + 2CI'
CCI3CH3 - CCI2 CH2 + HCI
5-5
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5.2.2. CAH Usage as Primary Substrates
Few of the CAHs have been shown to serve as primary substrates for energy and
growth by some microorganisms. Among the Cj compounds, dichloromethane (DM) can
be used as a primary substrate under both aerobic and anaerobic conditions. DM can be
completely mineralized while serving as a primary substrate under anaerobic conditions
by municipal digesting sludge microorganisms (Rittmann and McCarty, 1980a,b;
Klecka, 1982; Freedman and Gossett, 1991). Pure cultures of the genera Pseudomonas
and Hyphomicrobium have been isolated that can grow aerobically on DM as the sole
carbon and energy source (Brunner and Leisinger, 1978; Brunner et al., 1980; Stucki et
al., 1981; La Pat-Polasko et al., 1984; Kohler-Staub and Leisinger, 1985).
The two-carbon saturated CAH, 1,2-dichloroethane (1,2-DCA), can also be used as
a primary energy source under aerobic conditions (Stucki et al., 1983, Janssen et al.,
1985). One unsaturated two-carbon CAH, VC, also has been shown to be available as a
primary substrate for energy and growth under aerobic conditions (Hartmans et al.,
1985; Hartmans and de Bont, 1992). These few exceptions noted to date indicate that only
the less halogenated one- and two-carbon CAHs might be used as primary substrates for
energy and growth, and that the organisms that are capable of doing this are not
necessarily widespread in the environment. The biological transformation of most of the
CAHs depends upon cometabolism.
5.2.3. Anaerobic Cometabolic Transformation of CAHs
In 1981, the potential for anaerobic biological cometabolism of brominated and
chlorinated (halogenated) aliphatic hydrocarbons was demonstrated (Bouwer et al.,
1981). Subsequently, CAHs, in general, have been found to transform under a variety of
environmental conditions in the absence of oxygen. Figure 5.1 illustrates the various
anaerobic biotic and abiotic pathways that chlorinated aliphatic compounds may undergo
at contaminated sites. For example, the chlorinated solvent TCA may be transformed
CCI3CCI3
Figure 5.1. Anaerobic Transformations of CAHs (after Vogel et al., 1987).
5-6
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abiotically to form 1,1-DCE and acetic acid. The rates are relatively slow, with a half-life
for TCA on the order of one year (Vogel et al., 1987; Cline and Delfmo, 1989; Jeffers et al.,
1989). Also, under anaerobic conditions, TCA may be biologically transformed into 1,1-
DCA, which can be further reduced to CA. CA is relatively stable biologically, but
abiotically can be transformed into ethanol and chloride, thus rendering it relatively
nontoxic. Thus, when TCA is discharged to soil, a variety of abiotic and biotic
transformation products may be found there in later years. As another example, TCE
can be reduced anaerobically to either cis- or trans -DCE, both of which can be further
transformed into VC. Recent research has indicated that VC can even undergo
reduction into ethylene (Freedman and Gossett, 1989; DiStefano et al., 1991), which is
essentially harmless.
The pathways outlined in Figure 5.1 suggest that it is possible to render harmless
essentially any chlorinated aliphatic compound under anaerobic conditions. While this
is true, there are several problems that hinder this potential approach to bioremediation.
First, the biotic transformations generally involve cometabolism such that other organic
compounds must be present to serve as primary substrates for organism growth.
Second, the rates of anaerobic transformation are much greater for the highly
chlorinated compounds than for the less-chlorinated compounds, so that the less-
chlorinated ones persist longer in the environment. Third, some of the anaerobic
transformation products are more hazardous than the parent compounds. Examples
here are TCE transformation to VC and TCA transformation to 1,1-DCE. Fourth,
reaction rates tend to be greater under highly reducing conditions associated with
methane formation than under the less reducing conditions associated with
denitrification (Bouwer and Wright, 1988). The latter is the main anaerobic process
occurring when excess nitrates are present. Reductive transformation rates are
somewhat intermediate between the two under conditions favoring sulfate reduction
(sulfate, but no nitrate present). Fifth, when proper environmental conditions are
present, microorganisms that can bring about the transformations through
cometabolism must also be present. Thus, with anaerobic conditions, one cannot count
upon sufficiently high rates and complete transformation to harmless products to occur
in ground water unless all the right conditions are present. On the other hand,
anaerobic transformation processes do frequently occur, converting chlorinated aliphatic
compounds into less chlorinated species that are more readily transformed by aerobic
microorganisms. It is for this reason, as well as to help understand the environmental
fate of compounds, that knowledge of anaerobic pathways is important.
5.2.4. Aerobic Microbial Transformation of Chlorinated Aliphatic Hydrocarbons
Although some CAHs, particularly those with few chlorines on the molecule,
were shown to be biodegradable by microorganisms some time earlier, knowledge that a
broader range of CAHs can be oxidized aerobically through cometabolism is rather
recent. Wilson and Wilson (1985) showed for the first time that TCE may be susceptible to
aerobic degradation through use of soil microbial communities fed natural gas. The
processes involved are illustrated by the following equations for TCE cometabolism by
methanotrophic bacteria, organisms that oxidize methane for energy and growth:
Methane Oxidation:
CH4 M^Ofc. CH3OH - fr- H2CO - ^— »» HCOOH — ^ » CO2
^
^
V
NADH, O2 Synthesis NADH NADH (1)
5-7
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TCE Epoxidation:
O
MMO / \
CCI2 - CHCI "^ > CI2 C CHCI »» »• CO2,C|-,H2O
NADH, O2 (2)
Methanotrophs use an oxygenase (methane monooxygenase or MMO) to catalyze
the oxidation of methane to methanol. This requires energy or reducing power in the
form of NADH. MMO also oxidizes TCE fortuitously to form TCE epoxide (Little et al.,
1988; Fox et al., 1990), an unstable compound that chemically undergoes decomposition to
yield a variety of products, including carbon monoxide, formic acid, glyoxylic acid, and a
range of chlorinated acids (Miller and Guengerich, 1982). In mixed cultures, as occurs
in nature, cooperation between the TCE oxidizers and other bacteria occurs, and TCE is
further mineralized to carbon dioxide, water, and chloride (Fogel et al., 1986; Henson et
al., 1989; Roberts et al., 1989; Henry and Grbic-Galic, 1991a).
Since the report of Wilson and Wilson (1985) with TCE cometabolism, much
scientific research addressing this phenomenon has been performed. The groups of
aerobic bacteria currently recognized as being capable of transforming TCE and other
CAHs through cometabolism comprise not only the methane oxidizers (Fogel et al., 1986;
Little et al., 1988; Mayer et al., 1988; Oldenhuis et al., 1989; Tsien et al., 1989; Henry and
Grbic-Galic, 1990; Alvarez-Cohen and McCarty, 1991a,b; Henry and Grbic-Galic, 1991a,b;
Lanzarone and McCarty, 1990; Oldenhuis et al., 1991), but also propane oxidizers
(Wackett et al., 1989), ethylene oxidizers (Henry, 1991), toluene, phenol, or cresol
oxidizers (Nelson et al., 1986, 1987,1988; Wackett and Gibson, 1988; Folsom et al., 1990;
Harker and Kim, 1990), ammonia oxidizers (Arciero et al., 1989; Vannelli et al., 1990),
isoprene oxidizers (Ewers et al., 1991), and vinyl chloride oxidizers (Hartmans and de
Bont, 1992). These microorganisms all have catabolic oxygenases that catalyze the initial
step in oxidation of their respective primary or growth substrates and have potential for
initiating the oxidation of CAHs. There is currently insufficient information on the
relative advantages and disadvantages of the different oxygenase systems to recommend
definitively one over the other, but each may have its place. Most research to date has
been conducted with the methane oxidizers and the group of bacteria containing toluene
oxygenase, which can be induced with primary substrates such as toluene, phenol, and
cresol.
The oxygenases for the above organisms are often nonspecific and fortuitously
initiate oxidation of a variety of compounds including most of the CAHs. The exceptions
are highly chlorinated CAHs such as CT and PCE. In general, oxygenases act on
unsaturated CAHs such as TCE, by adding oxygen across the double bond to form an
epoxide. With saturated CAHs such as CF or TCA, a hydroxyl group is generally
substituted for one of the hydrogen atoms in the CAH molecule. Frequently, the
resulting products from CAH oxidation are chemically unstable and decompose as
described above for TCE, yielding products that are further metabolized by other
microorganisms present in nature.
-------
5.3. PROCESSES AFFECTING CHEMICAL MOVEMENT AND FATE
In order to apply in-situ bioremediation of CAHs, an understanding of factors
affecting the movement and fate of contaminants in ground water is needed. Once
percolated from the land surface to ground water, organic contaminants such as the
chlorinated solvents and petroleum hydrocarbons are subject to a variety of influences
that lead to a complex pattern of behavior. The major processes influencing the
transport, distribution, and fate of these chemicals in ground water include the following
(McCarty et al., 1982 ):
1. Advection: the miscible transport in aqueous solution under the influence of
the hydraulic potential gradient;
2. Dispersion: the mixing and spreading of concentration fronts that arise largely
from differential rates of movement along the myriad individual flow paths
through the porous medium;
3. Sorption: the partitioning of a compound between the moving solution and the
stationary solid phase;
4. Transformation: the result of chemical reactions or microbial activity that may
convert an organic compound into stable products or into another intermediate
product;
5. Immiscible transport: the migration of slightly soluble chemicals as a separate
liquid phase, often driven downward by density difference in the case of
chlorinated solvents;
6. Diffusional transport: the slow migration of solute molecules into the matrix
rock or dead-end pores under the influence of a concentration driving force.
The influence of these factors on contaminant behavior has been summarized in
several reviews (NAS, 1984; Mackay et al., 1985; Goltz and Roberts, 1986, 1987). The
following brief discussion focuses on the principles underlying the sorption and
transformation processes, with CAHs being used for illustration.
5.3.1. Effect of Sorption
CAHs do not tend to sorb to soils and aquifer materials as readily as do many
hazardous chemicals such as pesticides, PAHs, and PCBs. Nevertheless, sorption in
aquifer systems is sufficient to retard the rate at which they move in ground water in
relation to the movement of ground water itself. This relative movement can be
expressed mathematically by the retardation equation (Freeze and Cherry, 1979):
v/vc = 1 + ptKd/n (3)
where, v = average linear velocity of ground water
vc = average linear velocity of the contaminant
pb = bulk mass density of solids in aquifer
n = porosity
K
-------
The term (1 + pbKj/n) is commonly known as the retardation factor. For aquifer
materials, Pb is approximately 1.8 g/cm3, and n generally varies between 0.2 and 0.4
(Freeze and Cherry, 1979). With these units, K
-------
The aquifer material at Borden had a low organic content (0.02%). Measured
values of IQ for CT and PCE were 15xlO« m3/g and 45x10* m3/g, respectively (Curtis et
al., 1986a). These values corresponded to retardation factors of 1.9 and 3.6. Although
retardation factors inferred from short-term field observations (10-30 days) were
consistent with the laboratory-measured Kj values, the study also showed that the
retardation factors for these two compounds increased with time and distance from the
point of injection to high values of 2.5 for CT and 5.9 for PCE. This appears to have been
related partially to a slow rate of diffusion of the contaminants into the aquifer solids, the
characteristic time scale which can be measured in terms of weeks to months, rather
than hours as commonly assumed. This suggests that short-term laboratory evaluations
are not adequate for determining retardation factors in the field. Another factor probably
causing the increased retardation with time is aquifer heterogeneities.
Some have indicated that a good correlation exists between Kj and the aquifer
organic content and the contaminant's octanol/water partition coefficient, K<,w
(KarickhofT et al., 1979; Schwarzenbach and Westall, 1981). However, this correlation
appears relatively poor for aquifers with low organic content (foc<0.1%) (McCarty et al.,
1981). Generally, aquifers are fairly poor in organic matter content so that the
retardation noted appears to be more a function of sorption to inorganic rather than
organic materials (Curtis et al., 1986b).
Retardation is an important process in ground waters for at least two reasons.
First, since chemicals have different sorptive properties, their relative rates of movement
through aquifers will differ widely (Roberts et al., 1982). Thus, if an aquifer is
contaminated with several compounds at one location, each contaminant will move at a
different speed, and they will arrive at a downgradient well at different times. The other
aspect of importance is that knowledge of the retardation factor provides a basis for
estimating the relative amount of the contaminant present in the aqueous phase as
compared with that sorbed to the aquifer solids. For example, for a retardation factor of
5, one-fifth of the contaminant is present in the aqueous phase and four-fifths is sorbed
onto the aquifer solids. Restoration of a contaminated aquifer requires that the
contaminant be removed from the solid phase as well as from the liquid phase. In
addition, sorption also tends to reduce the contaminant transformation rate by making
the contaminant inaccessible to microorganisms. The effects of these different factors on
in-situ bioremediation are discussed later in the section on nutrient introduction and
mixing.
5.4. FIELD PILOT STUDIES OF CAR TRANSFORMATION
There is no well-documented full-scale experience with in-situ bioremediation of
CAHs upon which to base full-scale application. However, limited small field-scale pilot
studies have been conducted in order to determine the effectiveness of certain approaches
to remediation. Three efforts have been conducted using the Moffett Naval Air Base pilot
facility in Mountain View, California, to evaluate the capacity of native microorganisms
(i.e., bacteria indigenous to the ground-water zone) to aerobically and anaerobically
cometabolically degrade CAHs when proper conditions were provided to enhance
bacterial growth.
The Moffett Field studies were conducted in a shallow confined aquifer under
conditions typical of ground-water contamination by CAHs. Two aerobic systems with
oxygen as the electron accepter have been tested: (1) methanotrophs that use methane as
a primary substrate and cometabolically transform CAHs with methane monooxygenase
5-11
-------
(MMO); and (2) phenol-utilizers in which toluene oxygenase (TO) serves as the
cometabolizing enzyme. Target contaminants in these studies were TCE, cis-DCE, trans-
DCE, and VC, in the concentration range of 40 to 120 ug/1. The third study conducted at
Mofiett Field was an anaerobic, or anoxic, study in which acetate was the primary
substrate applied to obtain transformation of CT under denitrification conditions.
The experimental approach taken was similar to that proposed for bioremediation
in the field (see Section 5.5). Extracted ground water from the treatment zone was
amended with the growth substrate and oxygen, and reinjected to stimulate indigenous
growth. CAHs in the extracted ground water were reinjected into the biostimulated
zone. Bioremediation conducted in this manner promoted the degradation of inplace
contaminants as well as contaminants that were extracted and reinjected, thus obviating
aboveground treatment.
The experiments were performed as a series of stimulus-response tests. The
stimulus was the injection of the compounds of interest and the response was their
concentration history at monitoring locations. The tests included bromide tracer tests to
study advection and dispersion; transport tests with the CAHs to study the retardation
process due to sorption and to evaluate whether transformation occurred in the absence
of active biostimulation; and biostimulation and biotransformation tests to evaluate CAH
transformation following the introduction of the primary substrates and electron
acceptors.
5.4.1. Results with Metfaanotrophs
Detailed discussions of the experimental methodology and results of the
methanotrophic studies are presented by Roberts et al. (1990) and Semprini et al. (1990,
(1991a). Indigenous methanotrophs were stimulated in three successive field seasons
through the addition of ground water saturated with methane (16 to 20 mg/1) and oxygen
(33 to 38 mg/1), which were introduced into the test zone in alternating pulses without any
other supplementary nutrients (N and P).
Figure 5.3 shows the concentration history of methane and oxygen at the 82
observation well during the initial biostimulation experiment along with model
simulations of Semprini and McCarty (1991). During the period of 200 to 430 hr, methane
and oxygen concentrations rapidly decreased, indicating the growth of methane-
utilizers. In order to control the clogging of the injection well and borehole interface, the
alternate pulse injection of methane and oxygen containing ground water was initiated
at 430 hr, with a pulse cycle time of 4 and 8 hr, respectively. The model simulations,
represented by the solid line, matched the field observations using a reasonable set of
biological and transport input parameters. Simulation modeling supported the
conclusions that methanotrophic bacteria were stimulated in the test zone, that
biofouling of the near well-bore region was limited by the pulsing methodology, and that
these processes can be simulated when appropriate rate and transport equations are
used.
Figure 5.4 shows the response at theS2 well of the target contaminant compounds
in the third season and model simulations (Semprini and McCarty, 1992).
Transformation of the organic target compounds ensued immediately following the
introduction of methane at time zero, increasing with time as the bacterial population
grew. Rapid transformation of VC and trans-DCE were observed, followed by cis-DCE
and TCE (not shown). TCA was a ground-water contaminant at the field site, and its
possible transformation was followed as well. No transformation of TCA was found; its
concentration was about 100 ug/1. Competitive inhibition of VC and trans-DCE
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transformation by methane was indicated in response to the dynamic pulsing of methane
and oxygen that was initiated at 20 hr. In order to effectively simulate the field
observations, both competitive inhibition kinetics and rate limited sorption were required
in the biotransformation model, thus reinforcing the conclusion that these processes
were occurring and need to be considered with in-situ remediation. The simulations
indicated that the overall rate of decrease in VC concentration may have been limited by
the rate of its desorption from the aquifer solids. The simulations also indicated that
physical processes, such as desorption, can limit times of cleanup by an enhanced
microbial process.
200
600
Time (hours)
Figure 5.3.
Methane and oxygen utilization by methanotrophs at the Moffett test facility
(after Semprini and McCarty, 1991).
200
Time (hour)
Figure 5.4.
CAH transformation by methanotrophs at the Moffett test facility (after
Semprini and McCarty, 1992).
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The comparison of the cometabolic rate parameters obtained from the modeling
exercise indicated that VC and trans -DCE were transformed at rates similar to that of
methane, the primary substrate added for growth, while cis-DCE and TCE were
transformed at rates one to two orders of magnitude slower that methane. The
simulations indicate that trans-DCE concentration decreased more slowly than VC since
it was more strongly sorbed (higher Kj), and thus a greater contaminant mass had to be
degraded. The order of magnitude difference in rates for cis-DCE and trans-DCE shows
that a small change in chemical structure can have a large effect on the cometabolic
transformation rate.
The specific conclusions from this study were:
1) The stimulation of indigenous methanotrophs could be accomplished through
methane and DO addition.
2) The rates and extents of transformation of CAHs were compound specific.
3) The percentage transformations achieved in a 2 m biostimulated zone were:
TCE, 20%; cis-DCE, 50%; frans-DCE, 90%; and VC, 95%.
4) The cometabolic transformation was strongly tied to methane utilization; upon
stopping methane addition, transformation rapidly ceased.
5) The cometabolic transformation was competitively inhibited by methane, an
effect that reduces the transformation rate.
6) Only a temporary enhancement of the cometabolic transformation could be
achieved by substituting formate (a noncompetitive substrate) for methane.
7) The rate of transformation was limited by the rate of desorption from the
aquifer solids, especially for the more rapidly degraded VC and trans-DCE.
8) The results agreed with those obtained in soil microcosm studies, performed
under conditions that mimicked the field tests.
5.4.2. Results with Phenol Utilizers
In the methanotrophic studies, limited degradation of TCE and cis-DCE was
achieved The objective of this study was to evaluate TO for in-situ biodegradation of
TCE, cis-DCE and trans-DCE at the Moffett test site. This was accomplished through the
introduction of phenol as a primary growth substrate and oxygen as an electron acceptor.
The evaluation was performed at the same site as the methanotrophic study, using the
same experimental methodology and at a similar contaminant concentration range,
permitting a direct comparison of the TO system with the MMO system. Active
biostimulation was initiated through the pulsing of phenol at time-averaged
concentrations ranging from 6 to 12 mg/1.
The concentration responses of DO, TCE, and cis-DCE at the SSE2 well, 2 m from
the injection well, are shown in Figure 5.5 (Hopkins et al., 1992). The biostimulation with
phenol is indicated by the DO decreases, which were small during the periods of low
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phenol addition but increased after higher phenol concentrations were added. Decreases
in cis-DCE and TCE concentrations were associated with decreases in OO, indicating
cometabolic transformations resulted from biostimulation. Significant degradation of cis-
DCE, and TCE were observed, with cis-DCE being more rapidly degraded than TCE. The
cis-DCE concentration decreased by approximately GO to 70% and TCE by 20 to 30% during
the period of low phenol addition. Doubling the phenol injection concentration resulted
in a greater transformation of both TCE and cis-DCE, with 85 to 90%, and over 90%,
transformed, respectively. Here, trans-DCE (not shown) was the least transformed of the
three compounds studied. Upon decreasing the amount of phenol added, the TCE
concentration increased, indicating the extent of transformation was related to the
amount of phenol added. As with the methanotrophic study, no significant
transformation of the TCA ground-water contaminant was observed.
0.0
200
400
600
1000
1200
Time (hours)
Figure 5.5. CAH transformation and dissolved oxygen changes resulting from phenol
addition at the Moffett test facility (after Hopkins et al., 1992).
The specific conclusions from this study were:
1) The stimulation of indigenous phenol-utilizers was accomplished through
phenol and oxygen addition.
2) The enhanced population effectively degraded TCE and cis-DCE but was less
effective in degrading trans-DCE.
3) Transformations achieved in a 2 m biostimulated zone were: TCE, 85%; and
cis-DCE over 90%.
4) The cometabolic transformation was competitively inhibited by phenol.
5) The cometabolic transformation was strongly tied to the amount of phenol
utilized.
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6) The results agreed with microcosm studies that were performed under
similar conditions to the field study.
5.4.3. Comparison Between the Methane and Phenol Studies
The pilot-scale studies both demonstrated in-situ biodegradation of CAHs could be
achieved. The degradation was shown to be compound specific in both cases. The phenol-
utilizers more effectively degraded TCE and cis-DCE, while the methane-utilizers more
effectively degraded trans-DCE, on a percentage basis. Both studies showed that
cometabolic transformation was closely associated with primary substrate utilization,
and that the primary substrate also competitively inhibited the transformation, slowing
the transformation rates. The concentrations of the CAHs were relatively low (<100 jig/1),
thus the results should not be extrapolated to higher concentrations. There is a need for
studies over a range of contaminant concentrations. Future studies need to explore CAH
concentration effects on degradation efficiency and on other primary substrates that may
stimulate more effective oxygenase systems.
5.4.4. Anaerobic Transformation of Carbon Tetrachloride
A Moffett study was also performed to evaluate CT transformation under
anaerobic conditions (Semprini et al., 1991b). CT is not transformed through aerobic
cometabolism, but anaerobic transformation has been reported and extensively studied.
The goals of this field evaluation were: to determine whether reductive transformation of
CT could be accomplished under the mildly reducing conditions of denitrification, what
factors affect the rates and extents of transformation, and what transformation
intermediates might be formed. In addition, other contaminants were present at the site
as ground-water contaminants, including TCA and two chlorofluorocarbons, CFC-11,
and CFC-113. The disappearance of these compounds was followed as well.
Acetate was first introduced at a time-averaged concentration of 25-46 mg/1, in the
presence of the nitrate (25 mg/1) and sulfate (700 mg/1), which were present in the ground
water and served as potential electron acceptors. Nitrate utilization commenced
immediately after the introduction of acetate and was complete within 100 hours, while
acetate utilization also commenced immediately but was expressed somewhat more
slowly and less completely because of the stoichiometric excess applied. The onset of CT
transformation was observed after approximately 350 hours (Figure 5.6), after which the
CT concentrations at the monitoring wells gradually declined, more rapidly at the more
distant (S2) well than at the nearer (81) well. Chloroform (CF) appeared as an
intermediate product of the CT transformation at all of the sampling points in an amount
corresponding to approximately one-half to two-thirds of the CT that disappeared. The
CF response and simulation modeling indicated that CF also was transformed, but more
slowly than CT.
The pattern of CT concentrations suggested that the CT transformation proceeded
more rapidly further downgradient from the injection well at a location just beyond
where nitrate became depleted. To test the hypothesis that the absence of nitrate would
enhance the CT transformation, nitrate was removed from the recycled water prior to
injection, beginning at 1260 hours. The CT concentration then declined abruptly at the Si
monitoring well. During this period without nitrate feed, the fractional yield of the CF by
product declined to about one-third of the CT transformed. Substantial acetate utilization
persisted in the absence of nitrate feed, suggesting that sulfate (present at 700 mg/1 in the
native ground water) may have served as an electron acceptor; however, no sulfide was
detected.
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0}
=1 02-
I
w
§01
U
00
-*- Nitrate (x 10-3)
^-CT
-o- Chloroform
MO 750
Time (Hours)
1000
1250
Figure 5.6. CT transformation under anaerobic conditions at the Moffett test site (after
Semprini et al., 1991b).
The background contaminants, including 1,1,1-TCA, CFC-11, and CFC-113, were
also partially transformed under the influence of anoxic biostimulation.
The specific conclusions from this study were:
1) The stimulation of indigenous acetate-utilizing denitrifiers was accomplished
through acetate addition to ground water containing nitrate but no oxygen.
2) Enhanced in-situ reductive transformations can be promoted through the
addition of an appropriate primary substrate.
3) Although steady-state conditions were not reached, average transformations
achieved over the 2.2 m test distance were: CT, 95%; CFC-11, 68%; CFC-113,
20%, and TCA, 15%.
4) CF was formed as an intermediate product of CT transformation, consistent
with laboratory studies.
5.5. PROCEDURES FOR INTRODUCING CHEMICALS INTO GROUND WATER
Perhaps the most significant challenge for in-situ bioremediation of CAHs is the
introduction into the subsurface environment of chemicals needed by microorganisms
for growth and mixing them with the contaminants to be degraded. Unless a suitable
primary substrate for stimulating cometabolic biotransformation of CAHs is already
present in the aquifer, one will need to be added. This may not be so difficult in the case
of cometabolism under anaerobic conditions, but it may be under aerobic conditions
where oxygen must be introduced as well. If methanotrophic cometabolism is desired,
then both methane and oxygen must be added. These gases are of limited solubility in
water. Therefore, added concentrations, together with other gaseous components, such
as molecular nitrogen, collectively must be below the saturation partial pressure in the
aquifer, which may not be much higher than 1 atm in shallow ground waters. A
common procedure is to mix the chemicals of interest with water and introduce them
into the ground water. The introduced water will push away the native ground water
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containing the contaminants, so that the required mixing between the contaminants and
the introduced chemicals will not occur. However, if the contaminants sorb somewhat to
the soils, then a benefit will be obtained as the sorbed contaminants will desorb into the
introduced water, thus bringing the contaminants and introduced chemicals together.
Another engineering consideration is the problem of excessive microbial growth
near the point where the chemicals are introduced into the ground water.
Microorganisms will tend to grow near the point of chemical introduction where
concentrations are the highest. In order to avoid such clogging, a strategy is needed that
will make growth difficult near the point of injection. The periodic introduction of
inhibitory chemicals such as chlorine or ozone may be required. A strategy such as
pulsing of the primary substrate and oxygen, as performed in the Moffett Field
experiments described previously, so that they do not occur together near the injection
point is a possibility. Full-scale experience with such strategies is limited and so
documented guidelines for application are not available.
-'°t
As with other forms of bioremediation presented in this section, delivery of
essential nutrients is a prerequisite for effective in-situ bioremediation. One potential
method for accomplishing the mixing of nutrients with ground-water contaminants is to
use a pump-and-treat extraction and injection system with bioremediation added, as
depicted in Figure 5.7 (McCarty et al., 1991). Here, all of the biological treatment is
carried out in the aquifer itself. Ground water is extracted through a series of wells
spanning the contaminant plume in a direction perpendicular to that of ground-water
flow. At the surface, methane and oxygen are added to the extracted ground water,
either together or in alternating pulses, depending upon the extent to which a stimulated
methanotrophic population has been developed.
The alternating pulses are used to distribute the microbial growth throughout the
test zone. The ground water containing the appropriate primary substrates and electron
acceptors (i.e., methane and oxygen) and extracted contaminants are reinjected into the
treatment zone through a series of wells distributed parallel to the extraction wells. In
the subsurface biotreatment zone, both the in-place contaminants and the reinjected
contaminants are biologically degraded. Another alternative is to use a combination of
aboveground treatment and ih-situ treatment. In this case, the contaminants would be
removed at the surface, and only the required amendments would be added to the
extracted ground water prior to reinjection.
Methane and Oxygen
Addition in
Chlorinated Ethenes Alternating Pulses
Reinjected into the
Biostimulated Zone
Figure 5.7. Pump-extract-reinject method for mixing of chemicals with ground water (after
McCarty et al., 1991).
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The above in-situ biotreatment system was directly compared with a pump-and-
treat system through model simulations. In the latter case, an unspecified form of
surface treatment was assumed to remove contaminants quantitatively before the water
is reinjected. Otherwise, the two systems were assumed to operate identically, allowing
direct comparison through model simulations (Semprini and McCarty, 1991, 1992). The
results of model simulations are shown in Figure 5.8 for VC, a weakly sorbed
contaminant that is rapidly degraded by the methanotrophic process. Here the in-situ
process was shown to be as effective as pump and treat. The model simulations of Figure
5.8 also indicate the advisability of recycling the contaminants through the treatment
zone rather than removing them through treatment at the surface. The line with
triangles (Biostim+pump) shows a simulation in which the contaminants in the
extracted water were removed using surface treatment before reinjection. Methane and
oxygen were added to the surface-treated reinjected water. Some, but not a significant,
enhancement of removal was achieved by adding surface treatment. Moreover, the in-
situ bioremediation process also degraded the reinjected contaminants to nontoxic end
products, which is an advantage over some forms of surface treatment.
Figure 5.8.
20
80 100 120
Time (days)
140
160 180 200
Simulation modeling of pump-extract-reinject method for methanotrophic
cometabolism of VC (after McCarty et al., 1991).
Model simulations are quite useful in evaluating the different approaches and
potential effectiveness of proposed treatment schemes. Simulations for trans-DCE
(Figure 5.9), a more strongly sorbed compound than VC but one that is as rapidly
degraded by methanotrophs, shows the in-situ process becoming even more attractive
than pump and treat. However, for compounds that were less effectively degraded by the
methanotrophic process, such as cis-DCE and TCE, in-situ treatment was found through
simulation modeling to be less effective in reducing the cleanup times or the amount of
water extracted compared to normal pump-and-treat procedures.
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"Sb
ID
u
Q
Figure 5.9. Simulation modeling of pump-extract-reinject method for methanotrophic
cometabolism of trans-DCE.
Another possible system for delivering the needed chemicals is subsurface ground-
water recirculation (Figure 5.10). This eliminates the need to pump contaminated
ground water to the surface. This mixing method, which is under development, uses a
subsurface recirculation unit with an upper and lower screen, and a pump to induce
flow through the unit and through the porous formation. For methanotrophic treatment,
methane and oxygen would be introduced directly into the recirculating ground water.
This method would eliminate pumping the contaminated ground water to the surface,
surface treatment, and subsequent reinjection. One possibility is that several
recirculation units could span perpendicularly across a plume, making a biologically
reactive barrier through which ground water would flow and be treated.
Oxygen —| ,— Methane
Recirculaiion Unit-
Seal
Vadnse Zone
Ground Water
Figure 5.10. Subsurface recirculation system for chemical introduction and mixing with
ground-water contaminants.
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The mixing system illustrated in Figure 5.7 might be used not only with other
forms of aerobic treatment, but also with anaerobic treatment as well. One possibility is
to use in-situ treatment to augment a pump-and-treat remediation that is already in
place.
5.6. THE EFFECT OF SITE CONDITIONS ON REMEDIATION POTENTIAL
Figure 5.11 illustrates contamination of soil and ground water by leakage from a
storage tank, a common way by which contamination with liquids occurs (McCarty,
1990). As the liquid is pulled downward by gravity, residuals left behind contaminate the
surface soil, the unsaturated (vadose) zone, and finally the aquifer containing the ground
water itself. After the leakage is found and stopped, and the most highly contaminated
soil around the tank is excavated, one must then deal with a lower-concentration
residual in the soil, the vadose zone, and the ground water. If the contaminating liquid
is a mixture of many different compounds, then each may move and be transformed at
different rates. Biological and chemical transformations may not lead to mineralization,
but may result in producing other organic chemicals that may be either less or more
harmful than the original. Organics may become strongly sorbed onto subsurface
minerals or may penetrate into cracks so that they are not accessible by microorganisms
or their enzymes.
Figure 5.11. CAH contamination in a relatively homogeneous subsurface environment (after
McCarty, 1990).
The relatively homogeneous subsurface environment indicated in Figure 5.11 is
ideal, but seldom encountered. In such a case, ground-water flow direction and rate
might be determined from relatively few observations of piezometric heads and data from
pumping tests. Subsurface environments often are much more complex than this, some
perhaps as illustrated in Figure 5.12. Layering of permeable (sands and gravels) and
less-permeable (silts, clays, rock) strata is common and may contain discontinuities that
could result from faults or large-scale stratigraphic features. Conductivity of water and
contaminants through rocks and other such barriers may result from joints and
fractures that are difficult to locate and to describe. The mixture of gravel, sand, silt,
clay, and organic matter of which the subsurface environment consists can vary widely
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from location to location, as can the grain-size distribution and mineral composition
within each broad class of subsurface strata.
In addition, abandoned wells can often provide passageways between separated
aquifers. Recalcitrance of contaminants in such systems may result from high
concentrations that are toxic, from presence of the organics in fissures, strong sorption
to particle surfaces, or diffusion into small pore spaces in minerals, rendering the
contaminants inaccessible to microorganisms and their enzymes. Sorbed compounds
often desorb slowly, and this often becomes the limiting factor affecting rates of
biodegradation as well as removal by pump-and-treat methods (McCarty, 19901.
Recalcitrance may also result from insufficient levels of required nutrients, such as
nitrogen and phosphorus for bacterial growth (Knox et al., 1985; Thomas and Ward,
1989). For aerobic treatment, optimal concentrations of ammonia or nitrate nitrogen are
in the range of 2 to 8 pounds per 100 pounds of organic material, while inorganic
phosphorus requirements are about one-fifth of this (McCarty, 1988). When these
nutrients are below optimum levels, rates of biodegradation slow considerably and may
be more dependent upon rates of nitrogen and phosphorus regeneration than on other
factors. With anaerobic degradation, nutrient needs are generally less, but organism
growth rates are slower. The absence of suitable electron acceptors is another factor that
can affect biodegradability. For aromatic hydrocarbons, degradation rates are generally
enhanced through aerobic decomposition. Thus, introduction of oxygen can be useful.
Generally, the quantity of oxygen required is similar to the mass of contaminants
present. In complex subsurface systems, getting the oxygen to the areas of need can
prove difficult.
Figure 5.12. CAH contamination in a relatively nonhomogeneous subsurface environment
(after McCarty, 1990).
Frequently, when environmental conditions are not appropriate for
biodegradation to occur, potential solutions often involve addition of chemicals (Knox et
al., 1985; Roberts et al., 1989; Thomas and Ward, 1989; McCarty et al., 1991). This is
perhaps not difficult with surface contamination but may be nearly impossible with some
subsurface contamination, depending upon the hydrogeology. With the latter, conditions
that make pump and treat difficult render efforts at bioremediation difficult as well. If it
is difficult to pump contaminants out of the ground, then it is also difficult to pump
chemicals or microorganisms into the ground to reach the contaminants. In such
cases, biological approaches may not offer significant time advantages over pump and
treat. The main advantage of bioremediation is likely to be an environmental one, i.e.,
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the contaminants are destroyed with a minimum of disruption of the surface
environment. In some cases, costs may be significantly reduced as well.
In some cases, proper environmental conditions may be obtained by moving
contaminants to effect dilution or mixing with natural chemicals in the subsurface
system. Dilution by mixing of contaminated and uncontaminated ground water can
reduce contaminant toxicity. Also with dilution, alternate electron acceptors such as
oxygen or nitrates, or essential nutrients such as nitrates, phosphates and iron, that are
present in uncontaminated water, may be brought together with the contaminants for
better biodegradation. Again, better methodologies for predicting the outcome of this
strategy are needed.
Where environmental conditions are suitable and where the proper microbial
populations are present, complete mineralization of organic contaminants can occur,
even within the most complex hydrogeological environments. Even where
environmental conditions are not ideal, degradation of many organic chemicals may
take place at reduced rates, with half-lives on the order of one or two years. In such
cases, the correct strategy may be to leave the contaminants alone and allow the problem
to be rectified by natural processes. Environmentally, this may be the best position to
take. The difficulty here is in obtaining evidence that would convince us, the regulatory
authorities, and the public that such natural processes are indeed occurring. Also
difficult is making good estimates of the time-frame for natural purification to occur.
Currently, we do not know what evidence to collect to prove the occurrence of natural
degradative processes, nor how to collect it This is a most important area of need.
5.6.1. Microorganism Presence
With contaminants that are known to be readily biodegradable, the absence of a
suitable microbial population may also be a factor. Methodologies for determining
microorganism presence are under development. Some include the simple exposure of
aseptically obtained soil to the contaminants of concern under ideal chemical conditions
for biodegradation. If the microorganisms are naturally present, then degradation of the
contaminant will occur. Other approaches are to attempt to identify the presence of
species known to biodegrade the compounds of interest, or to use molecular probes that
can identify the presence of specific microorganisms, nucleic acid sequences, or enzymes
that are key to compound degradation. These more sophisticated techniques are not yet
fully developed, but may offer promise for the future.
If appropriate organisms are not present, then they may be introduced into the
surface or subsurface environment (Omenn et al., 1988). Such organisms may be
natural, but not ubiquitous in nature. Their growth and introduction into a new system
may thus be acceptable. An important question is whether such specialized organisms
can survive in the new environment, and if so, can they be transported to the place of
need? If the hydrogeology is complex, then this may be most difficult. In other research,
attempts are being made to engineer microorganisms that are capable of degrading
organic compounds that are inherently recalcitrant. The potential use of such
organisms raises societal concerns as well as the physical and biological barriers to
successful organism introduction into the environment. Nevertheless, such approaches
deserve to be explored as they will add to our overall knowledge of the biodegradation
process.
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5.7. SUMMARY
In-situ biodegradation of most CAHs depends upon cometabolism and can be
carried out under aerobic or anaerobic conditions. Cometabolism requires that an
appropriate primary substrate be added to the aquifer, and perhaps an electron acceptor
such as oxygen or nitrate for its oxidation. Only a few CAHs can serve as primary
substrates for biodegradation. In order to apply in-situ bioremediation, conditions must
be appropriate. The aquifer should be relatively homogeneous so that chemicals can be
mixed with ground water. Sufficient primary substrate must be added to satisfy the
needs of the respective bacteria. Generally, the more highly chlorinated CAHs should be
converted by anaerobic in-situ treatment to less chlorinated forms that can be degraded
through aerobic cometabolism. It is necessary to determine whether the appropriate
microorganisms are present as indigenous organisms in the aquifer. Generally, this
requires laboratory studies on aseptically obtained aquifer material. Sufficient
characterization of the aquifer is desirable so that the injection and distribution of
chemicals can be modeled and decomposition of CAHs can be reasonably well-predicted.
The formation of halogenated intermediate products, which may be of public
health concern, poses an obstacle to the deployment of the anaerobic approach for aquifer
bioremediation with the more highly chlorinated species. Recent laboratory and field
investigations, however, show PCE and TCE can be completely dehalogenated to ethene,
which is encouraging. Research work therefore needs to focus on determining how
effectively enhanced reductive dehalogenation to nontoxic end products can be
accomplished in the complex subsurface environment and the potential benefits as well
as disadvantages of forming less halogenated intermediates.
Until full-scale experience is available, the best approach might be to attempt
remediating sites that are relatively simple hydrogeologically and contain more readily
degradable contaminants. An ideal case would be to degrade VC through aerobic
cometabolism with methanotrophs. VC is difficult to remove with the normal pump-and-
treat system because it is a known human carcinogen with relatively low MCL, and it
does not sorb well to activated carbon or other sorbers. Thus, surface treatment is
difficult and expensive. However, VC can be used as a primary substrate, if the
appropriate organisms are present, or at least can be destroyed through cometabolism by
methanotrophic bacteria. Here, the ratio of methane addition and VC degradation is
quite low, about two kg of methane are required per kg of VC destroyed. Experience is
also needed with systems for introducing chemicals into the subsurface environment
and for mixing them with the contaminants of concern. Once experience with the easier
cases is available, then application in more complex situations can be attempted.
Without full-scale application, little can be said about the cost of such treatment. Thus,
there is much yet to be learned.
ACKNOWLEDGEMENT
This report contains information that resulted from studies conducted through
the Western Region Hazardous Substance Research Center through EPA Grant No.
R815738.
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Kohler-Staub, D., and T. Leisinger. 1985. Dichloromethane dehalogenase of
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Lanzarone, N.A., and P.L. McCarty. 1990. Column studies on methanotrophic
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LaPat-Polasko, L.T., P.L. McCarty, and A.J.B. Zehnder. 1984. Secondary substrate
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in groundwater: A critical review. Environ. Sci. Technol. 19(5):384-392.
Mackay, D.M., D.L. Freyberg, P.V. Roberts, and J.A. Cherry. 1986. A natural-gradient
experiment on solute transport in a sand aquifer. I. Approach and overview of
plume movement. Water Resour. Res. 22(13):2017-2029.
Mayer, K.P., D. Grbic-Galic, L. Semprini, and P.L. McCarty. 1988. Degradation of
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5-27
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Sons, Inc. NewYork, New York. pp. 191-209.
Miller, R.E., and P.P. Guengerich. 1982. Oxidation of trichloroethylene by liver
microsomal cytochrome P-450: Evidence for chlorine migration in a transition
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metabolism of trichloroethylene by a bacterial isolate. Appl. Environ. Microbiol.
52(2):383-384.
Nelson, M.J.K., S.O. Montgomery, W.R. Mahaffey, and P.H. Pritchard. 1987.
Biodegradation of trichloroethylene and involvement of an aromatic
biodegradative pathway. Appl. Environ. Microbiol. 53(5):949-954.
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Environ. Microbiol. 54(2):604-606.
Oldenhuis, R., R.L.J.M. Vink, D.B. Janssen, and B. Witholt. 1989. Degradation of
chlorinated aliphatic hydrocarbons by Methylosinus trichosporium OB3b
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Oldenhuis, R., J.Y. Oedzes, J.J. van der Waarde, and D.B. Janssen. 1991. Kinetics of
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5-29
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Appl. Environ. Microbiol. 49(l):242-243.
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SECTION 6
BIOVENTING OF CHLORINATED SOLVENTS FOR GROUND-WATER CLEANUP
THROUGH BIOREMEDIATION
John T. Wilson
Don H. Kampbell
U.S. Environmental Protection Agency
Robert S. Kerr Environmental Research Laboratory
P.O. Box 1198
Ada, Oklahoma 74820
Telephone: (405)436-8532,8564
Fax: (405M36-8529
6.1. FUNDAMENTAL PRINCIPLES
Chlorinated solvents such as tetrachloroethylene, trichloroethylene, carbon
tetrachloride, chloroform, 1,2-dichloroethane, and dichloromethane (methylene chloride)
can exist in contaminated subsurface material as (1) the neat oil, (2) a component of a
mixed oily waste, (3) a solution in soil water, or (4) a vapor in soil air. Spills of such
materials to subsurface material are frequently treated by soil vacuum extraction to
remove volatile oils and vapors in soil air, or by air stripping of contaminated ground
water. Both physical treatment processes produce a waste stream of contaminant vapors
in air. At present, these wastes are discharged to the atmosphere, or treated with
activated carbon, catalytic combustion or incineration.
Bioventing refers to biological treatment of oils in the vadose zone, supported by
oxygen delivered to the contaminated subsurface material through the advective flow of
air. In areas contaminated with oily phase material, oxygen delivered in air can support
direct biological degradation of the oily material. Any material volatilized into the air
may be swept away before treatment can occur (see section 3.2 for a discussion). If there
is adequate residence time of the vapors in material without oily phase contaminants, the
contaminant vapors can be degraded as the air moves away from the source areas.
In theory, a similar process can be used to support biological destruction of certain
chlorinated solvents. Many chlorinated solvents are not subject to direct biodegradation,
but can be cooxidized during microbial growth on another hydrocarbon. Frequently, the
extent of destruction of the chlorinated solvent is limited by the low solubility of the
growth substrate, and of oxygen, in water. Air is an ideal medium to deliver oxygen and
hydrocarbons for microbial growth to contaminated subsurface environments.
The chlorinated solvents, particularly trichloroethylene, are toxic to organisms
that cooxidize them. Toxic effects of trichloroethylene become important above 6 mg/1
water or 2 mg/1 air (Table 6.2; Broholm et al., 1990; Broholm et al., 1991). Air or water in
contact with oily phase trichloroethylene frequently exceeds the toxic limit. Further,
biodegradation supported in one pass through the unsaturated zone may not reduce the
concentration of contaminants to acceptable limits. In the near term, successful
implementation of bioventing for chlorinated solvents will most likely include (1) physical
transfer of the contaminant to air through soil vacuum extraction or air sparging of
ground water, (2) dilution of contaminants, if necessary, by addition of make up air,
6-1
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followed by reinjection into subsurface material that is not contaminated with oily phase
solvent. This idealized implementation is illustrated in Figure 6.1.
Surface ^
Em.ss.on Substrate
Sampler *
Primary
Substrate &
Pumps ^ Make Up Air
injection Well
\ri & Primary
Substrate
Soil Gas Monitoring
Figure 6.1. Two hypothetical implementations of in-situ bioventing of chlorinated solvents.
6\2. MATURITY OF THE TECHNOLOGY
As of this writing (1993) the technology is emerging. Bench-scale systems have
been evaluated in a number of laboratories, but no pilot- or field-scale system is yet in
place. The technology is essentially a blend of bioventing as an engineered activity, and
biotechnology for cooxidation of chlorinated solvents. Progress in bioventing is rapid, and
considerable effort is being expended on the biology, physiology, and biochemistry of
chlorinated solvent cooxidation in ground water. Bioventing of chlorinated solvents
awaits the linkage of bioventing and the microbiology of cooxidation of the chlorinated
solvents.
There is a good prospect that the States and U.S. Environmental Protection
Agency will see permit applications for bioventing of chlorinated solvents starting in 1993
or 1994.
6.3. PRIMARY REPOSITORIES OF EXPERTISE
The best University research to date has been done at the University of Texas at
Austin in the laboratories of Gerald Speitel and Ray Loehr. Don Kampbell at the Ken-
Laboratory at Ada, Oklahoma, did the seminal work in this area, and continues an active
laboratory-scale program. Harvey Read and Thomas Stocksdale with S.C. Johnson &
Son, Inc. (Racine, Wisconsin) and Hinrich L. Bohn of the University of Arizona (Tuscon,
Arizona) have the most experience with field-scale biopiles designed to oxidize
6-2
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hydrocarbon vapors. These biopiles are very similar to systems that would be designed
for cooxidation of chlorinated solvents.
Hinrich L. Bohn
Dept. of Soil & Water Science
University of Arizona
Tuscon, AZ 85721
Phone- (602)621-1646
Fax: (602)621-1647
DonH Kampbell
U.S.EPA
Robert S. Kerr Laboratory
Ada, OK 74820
Phone: (405)436-8564
Fax- (405)436-8529
Raymond C Loehr
Dept of Civil Engineering
The University of Texas at Austin
Austin, TX 78712-1076
Phone (512)471-5602
Fax: (512)471-0592
Harvey W. Read
S.C Johnson & Son, Inc.
1525 Howe Street
Racine, WI 53403
Phone (414)631-2000
Fax- (414)631-2167
Gerald E Speitel
Dept. of Civil Engineering
The University of Texas at Austin
Austin, TX 78712-1076
Phone- (512)471-5602
Fax (512)471-0592
Thomas T. Stocksdale
S.C Johnson &Son,Inc
1525 Howe Street
Racine, WI 53403
Phone (414)631-2000
Fax- (414)631-2167
6.4. CONTAMINATION SUBJECT TO TREATMENT
Table 6.1 lists the nine most important chlorinated alkanes in ground water in a
survey of 358 hazardous wastes sites (Plumb and Pitchford, 1985). Vinyl chloride in
ground water is almost invariably produced from the reductive dechlorination of
tetrachloroethylene or trichloroethylene. The other compounds are used as solvents and
find their way into ground water through improper disposal.
TABLE 6.1. THE COMMON CHLORINATED ORGANIC COMPOUNDS OCCURRING AS
CONTAMINANTS OF GROUND WATER
Compound
Trichloroethylene
Tetrachloroethylene
Chloroform
Methylene Chloride
1,1, 1-Tnchloroe thane
1,2-Dichloroethane
Vinyl Chloride
Chlorobenzene
Carbon Tetrachloride
Detection
frequency
(% of 358 sites)
51
36
28
19
19
14
8.7
not ranked*
not ranked
Average
concentration
in all samples
(mgll)
20
35
041
22
0.24
0.90
not ranked
not ranked
not ranked
Average
concentration
in samples
where detected
(mgll)
38
97
not ranked
11.2
not ranked
6.3
not ranked
not ranked
not ranked
a "Not ranked* means not ranked in the top twenty organic contaminants at the 358 sites surveyed
-------
Trichloroethylene, chloroform, 1,1,1,-trichloroethane, 1,2-dichloroethylene, and
dichloromethane (methylene chloride) can be biologically cooxidized during growth on a
variety of substrates (methane, propane, toluene) that exist as vapors and can be
delivered to the subsurface environment through the flow of air (Wilson and White, 1986;
Henson et al., 1988, Wackett and Gibson, 1988). Chlorobenzene, dichloromethane, 1,2-
dichloroethane, and vinyl chloride can also serve as primary substrates for microbial
growth (Janssen et al., 1985; Davis and Carpenter, 1990; Rittmann and McCarty, 1980);
they do not necessarily require a cosubstrate for biodegradation, although removal in the
presence of a cosubstrate may be more rapid. There is no known pathway for aerobic
biodegradation of tetrachloroethylene or carbon tetrachloride.
With the exception of trichloroethylene and vinyl chloride, the average
concentrations listed in Table 6.1 can be easily tolerated by heterotrophic bacteria.
Cooxidation of trichloroethylene and vinyl chloride can occur through an epoxide that is
chemically reactive, and is considerably more toxic than the parent compound.
Trichloroethylene concentrations below 3 mg/1 air do not inhibit the rate of
oxidation of the primary substrate (Table 6.2). Less is known about toxic effects of vinyl
chloride; concentrations of 1.0 mg/1 air are generally tolerated in laboratory microcosms.
TABLE 6.2. EFFECT OF THE CONCENTRATION OF TRICHLOROETHYLENE ON THE RATE OF
BIODEGRADATION OF AVIATION GASOLINE" VAPORS IN SOIL MICROCOSMS
Concentration of Trichloroethylene Degradation of Gasoline Vapors
(mglkg soil) (mg/lair, if all (mg/kg day)
trichloroethylene
volatilized)
45,600
4,500
13
4.1
L2
0.0
1,000
3
0.9
0.3
0.0
0.06
0.11
24
26
31
33
8 The hydrocarbons in gasoline support the cooxidation of trichloroethylene.
&5. SPECIAL REQUIREMENTS FOR SITE CHARACTERIZATION
Sites should be carefully mapped to separate areas with oily phase liquids from
areas that only have contamination in air or water. Areas containing oily phase liquids
have a great capacity to contaminate soil air or ground water. In regions containing oily
phase liquids, the mass of contaminant removed through biological treatment is trivial
compared to the mass of contaminant that will partition into air or water and be carried
away. Air or water should not be injected into geological material containing oils unless
there is an intent to recover the effluent through pump and treat or soil vacuum
extraction.
64
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Many contaminated subsurface materials contain chlorinated solvents sorbed to
organic materials. These materials can act as source areas for contamination of ground
water or soil air and are good candidates for in-situ bioventing. Core material should be
extracted to determine the total mass of sorbed chlorinated solvents that are subject to
remediation. Desorption isotherms can be useful to determine the extent of remediation
required to meet cleanup goals for ground water or soil air.
6.6. SITE CHARACTERISTICS THAT ARE PARTICULARLY FAVORABLE
Transmissive material is better than less transmissive material, because
injection and extraction wells can be spaced on wider intervals, and the power
requirement for blowers is less. A deep water table provides more volume of unsaturated
material, and thus more residence time on a surface area basis, for treatment of vapors.
Direct metabolism or cooxidation of chlorinated solvents ultimately produces
hydrochloric acid. Unless the geological matrix contains carbonates or other natural
buffers, the pH will drop to levels that inhibit oxidation of the primary substrate. If
ground-water pH is controlled by a carbonate/bicarbonate buffering system sustained by a
source of carbonate in the aquifer matrix, biological activity can proceed for much longer
periods of time.
In-situ treatment need not impede the beneficial use of infrastructure at a site.
Sites with infrastructure of great economic value (such as refineries or factories) or sites
of great importance to health and safety (such as highways, hospitals, certain military
installations) are good candidates for in-situ remediation. In-situ remediation is also
particularly appropriate if contamination has migrated to property owned by a second
party.
6.7. SITE CHARACTERISTICS THAT ARE PARTICULARLY UNFAVORABLE
Material that contains secondary porosity (such as cracks, channels or fissures)
allows short circuiting of gases. These secondary passages will form if the material has
enough clay or organic matter to form water-stable aggregates.
The primary substrates are hydrocarbons supplied at concentrations in air that
are near or within the explosive range. The potential for migration of explosive vapors in
basements, sewers, utility conduits, and other underground excavations should be
assessed.
Material that has wide variations in texture is difficult to treat with any
technology that circulates fluids. Above the water table, fine textured materials tend to
retain organic liquids by capillary attraction. Because fine textured materials have more
clay and organic matter, they also have a greater sorptive capacity. Remedial fluids tend
to flow around rather than through the most contaminated material.
Sites with unfavorable characteristics that are being remediated prior to sale or
transfer to a second party may not be remediated in an acceptable time frame.
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a8. PERFORMANCE UNDER OPTIMAL CONDITIONS
6.8.1. Importance of the Rate Law
The active microbial populations in unsaturated soil behave as if they are
contained in a thin film of water in chemical equilibrium with soil air and ground water.
At high concentrations, the enzymatic machinery of the organisms degrading the
compound are saturated. The rate of degradation is proportional to the amount of active
biomass, but is independent of the concentration of organic compound. Disappearance of
the organic contaminant is described with zero-order kinetics, where the decrease in
contaminant mass is linear with time. The rate of biodegradation is most conveniently
normalized to the dry weight of soil in contact with contaminated soil air or ground
water.
At low concentrations, the supply of organic compound is limiting, and the rate of
degradation is proportional to the concentration of the organic compound in contact with
the active microorganisms as well as the amount of active biomass. Disappearance of
the organic compound is described by first-order kinetics expressed as a half-life, or by a
first-order rate constant.
6.&2. Importance of Partitioning
Interpretation of first-order rate laws, and particularly extrapolation of
experimental data to other systems, is complicated by partitioning of the organic
compound between soil air, water and solids. Laboratory experiments done with
columns usually have a natural ratio of air, water, and solids; but batch laboratory
experiments usually do not. If the batch laboratory systems have a greater proportion of
air, the kinetics of depletion in the laboratory system will be slowed with respect to the
depletion that would be seen at field scale. The factor by which kinetics are slowed can be
estimated by dividing the ratio of air to solids in the experimental system by the ratio of
air to solids in the natural system. In Table 6.3, simple chemodynamic theory was used
to predict the partitioning of the chlorinated solvents between air, water, and solids in a
representative subsurface material. For most compounds, the majority of contaminant
mass is in the soil water and therefore available to microorganisms. However,
biodegradation of vinyl chloride will be strongly influenced by partitioning, particularly
in cold subsurface material.
After acclimation, the biological removal of vapors of natural hydrocarbons is very
rapid. Typical vadose zone subsurface material, as depicted in Table 6.3, has at most 100
ml of air per kg of soil. The air could contain at most 25 mg of oxygen gas, which would
support metabolism of approximately 7 or 8 mg/kg of hydrocarbon. Table 6.4 presents the
rate of hydrocarbon oxidation in fertile, well acclimated subsurface material. Oxygen
would be exhausted in a few hours in these soils at typical air content.
Table 6.5 presents laboratory data on the kinetics of mineralization of
trichloroethylene, chloroform, and 1,2 dichloroethane vapors. To provide a common
basis for comparison, all rates were calculated as zero-order rates normalized to the
mass of soil. In general, the rate of removal of chlorinated solvents was one to ten
percent of the removal of natural hydrocarbons.
-------
TABLE &3. PARTITIONING OFCHLORINATED ORGANIC COMPOUNDS BETWEEN Ant, WATER,
AND SOLIDS IN A HYPOTHETICAL SUBSURFACE MATERIAL WITH AN AIR-FILLED
POROSITY OF 0.2, A WATER-FILLED POROSITY OF 0.2, AND AN ORGANIC CARBON
CONTENT OF 100 ME/KG
Compound
Trichloroethylene
Tetrachloroethylene
Chloroform
Methylene Chloride
1,1, 1-Tnchloroe thane
1,2-Dichloroethane
Vinyl Chloride 25°C
Vinyl Chlonde 10°C
Chlorobenzene
Carbon Tetrachlonde
Cosubstrates
Toluene
Methane
Air
Percent of total
20
20
7.0
6.2
24
2.3
58
96
7.2
17
12
95
Water
Percent of total
72
62
90
93
60
97
42
40
66
29
78
5
Solids
Percent of total
73
18
31
07
15
11
0.08
trivial
27
53
10
trivial
TABLE 6.4. KINETICS OF DEPLETION OF NATURAL HYDROCARBONS IN UNSATURATED SOIL AND
SUBSURFACE MATERIAL
Soil
type (Ref)
Sand (a)
Sand (a)
Loam (a)
Sandy Loam (b)
Loamy Clay (c)
Sandy Silly Loam (c)
Sandy Clay Loam (c)
Sand (d)
No Data (e)
Muck (0
Methane Propane and Benzene Toluene Ethyl- o-Xylene
Butanes benzene
43.6 10.1 3.6 25
19.6 4.0 2.0 0.8
19.3 44 1.8 I.I
319 146
104
157
173
1.400
150 288
163
a. Miller and Canter. 1991. b. English. 1991: c Bender and Conrad. 1992; d. Hoeks. 1972.
c Anonymous; f. Kampbell et al., 1987
6-7
-------
The hypothetical representative subsurface material described in Table 6.3 has 0.1
1 of soil water per kg. If the data in Table 6.1 on typical concentrations of chlorinated
solvents in ground water are divided through by ten, they can be used to estimate the
mass of contaminant in mg per kg of typical aquifer material. Typical concentrations of
chlorinated solvents are near 1 mg/kg. The rates of removal of chlorinated hydrocarbon
presented in Tables 6.5 and 6.6 indicate that the major fraction of chlorinated solvent
contamination could be removed in a few hours to a few days.
TABLE 6J>. MINERALIZATION OF VAPORS OF CHLORINATED SOLVENTS IN SOIL ACCLIMATED TO
DEGRADE VAPORS OF NATURAL HYDROCARBONS
Refer- Soil
ence Type
Substrate Initial Solvent Cone
and
Nutrients mglkg soil fig/1 air
Mineralization
Rate
(mg/kg soil/day)
With Without
Substrate Substrate
Tnchlororethvlene
(a) Sandy
clay
(b) Organic
nch
muck
(c) no
data
(c) no
data
(d) Coarse
sand
Chloroform
(a) Sandy
clay
1 .2-Dichloromethane
(a) Sand
(a) Silly
loam
(a) Sandy
clay
(a) Sandy
clay
2.5% 50 20.000
methane
N.P.K.S
0.2% 3.2 1,100
propane
butanes
N.P.S 1.030
01% 1.030
toluene
N.P.S
20% 1.5 260
methane
applied
4 times
23% 100 40,000
methane
N.P.K.S
1 3% 100 40,000
methane
N.P.S.K
2.2% 100 40.000
methane
N.P.S.K
1.1% 100 40,000
methane
N.P.S.K
4.9% 100 40.000
methane
0 39 0.35
27
-------
TABLE 6.6. REMOVAL OF VAPORS OF TRICHLOROETHYLENE AND VINYL CHLORIDE IN
SUBSURFACE MATERIAL UNDER OPTIMAL CONDITIONS"
Source
Trich loroeth v lene
Tucson,
Arizona
St. Joseph.
Michigan
Racine,
Wisconsin
Vmvl Chloride
Racine,
Wisconsin
Tucson.
Arizona
So,l
Type
Sandy
Sand
Organic
Rich
Loam
Loam
Sand
Substrate Initial Solvent Cone.
and
Nutrients mglkg soil Ugll air
0.6% gasoline 17 4.200
vapors
N.P.K
0.6% 17 4.200
gasoline
vapors
N.P.K
0.6% 17 4.200
gasoline
vapors
N.P.K
4.0 1.000
0.6%
gasoline
vapors
0.6% gasoline 4.0 1 .000
vapors
N.P.K
Mineraliyition
Rale
(mg/kg soil/day)
1.47
2.00
0.53
1.75
1.35
a Kampbell and Wilson, 1993.
6.9. PROBLEMS ENCOUNTERED WITH THE TECHNOLOGY
If concentrations are high, in the milligrams per liter range, bioventing can
economically remove a major fraction of contaminant mass prior to polishing with
activated carbon to meet regulatory endpoints. If concentrations are low, within one or
perhaps two orders of magnitude of the goal, bioventing can meet acceptable
concentrations with a relatively few applications of substrate or oxygen. However,
bioventing to degrade chlorinated organics through cooxidation should not be expected to
reduce the average concentrations presented in Table 6.1 down to drinking water MCLs
(Maximum Contaminant Levels).
6.10. RELEVANT EXPERIENCE WITH SYSTEM DESIGN
S.C. Johnson & Son, Inc., operates a facility in Racine, Wisconsin, that packages
wax products in aerosol cans. A mixture of propane and butanes is used as the
propellant gas. To treat propellant gas that escapes during the filling process, they
constructed a soil bed bioreactor that is 184 feet long, 159 feet wide and 4 feet deep. Air
6-9
-------
containing the propane and butanes is injected at the bottom of the bed, and works its way
to the surface (Figure 6.2). The waste stream at their facility is very similar to air that
would be used for the intentional cooxidation of chlorinated solvents. Soil acclimated to
degrade propellant gas also degrades trichloroethylene. The performance of their soil
bed reactor is the best model available for bioventing of chlorinated solvents at field scale.
16" Manifold From Aerosol
Production Line (Gas House)
1/8" Holes, 30° Off Bottom of
MPE,8"Apart
Slotted Drain to
Process Sewer
Figure 6.2. Soil bed constructed by S.C. Johnson & Son, Inc. in Racine, Wisconsin, to treat an
airstream containing 2,000 to 3»500 ppm of propellant gas (a mixture of propane and
butanes).
The performance of the soil bed reactor was related to the air loading rate. The
removal efficiency for propellant gas is greater than 90% when the flow of air is less than
3 cubic feet per minute per 100 square feet(cfm/100 sq. feet) of bed surface area. Removal
drops to less than 50% at flows greater than 6 dm/100 sq. feet (personal communication,
Thomas T. Stocksdale, Safety and Environmental Affairs Manager, S.C. Johnson & Son,
Inc., Racine, Wisconsin, see Figure 6.3 for data). At a bed depth of four feet, and an
estimated air-filled porosity of 0.2, 3 scf/100 sq. feet corresponds to an average residence
time for air of 30 minutes.
6-10
-------
100%
I 90%
I 80%
o
~ca
u
£
40%
0 1 2 3 4 5 6 7 8 9 10 II 12 13 14
AirFlow(cfm/IOOsqft)
1990
1991
• Estimated Curve
Figure 6^. Effect of flow rate on removal efficiency of propellent gas hydrocarbons in a soil
bed bioreactor.
If this soil bioreactor were used to treat air containing 1,100 jig/1 of
trichloroethylene, 0.22 mg of trichlorethylene vapor would be exposed to each kg of soil for
30 minutes. If it performed according to the laboratory study of Kampbell et al. (1987, see
Table 6.5), this reactor could remove 0.56 mg/kg of trichloroethylene in 30 minutes under
zero-order kinetics. If this behavior is general, subsurface material optimized to remove
the primary substrate will remove vapors of chlorinated solvents to low concentrations
where first-order kinetics apply.
There was extensive lateral migration of air injected into their bed. To prevent
migration of injected air under their facility, they installed sheet piling along three sides
of the soil bed. "Hot spots" developed in their bed that had high flow rates of air and poor
removal of propellant gas hydrocarbons. To correct "hot spots" they mechanically deep
till and regrade the soil.
6-11
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REFERENCES
Anonymous. Roy F. Weston, Inc. 1990. Final Report Task Order 8: Biotreatment of
Gaseous-Phase Volatile Organic Compounds. Contract DAAA 15-88-D-0010.
United States Army Toxic and Hazardous Materials Agency.
Bender, M., and R. Conrad. 1992. Kinetics of CH4 oxidation in oxic soils exposed to
ambient air or high CH4 mixing ratios. FEMS Microbiol. Ecol. 101(1992):261-270.
Broholm, K., T.H. Christensen, and B.K. Jensen. 1991. Laboratory feasibility studies on
biological in-situ treatment of a sandy soil contaminated with chlorinated
aliphatics. Environ. Technol. 12:279-289.
Broholm, K., B.K. Jensen, T.H. Christensen. and L. Olsen. 1990. Toxicity of 1,1,1-
trichloroethane and trichloroethene on a mixed culture of methane-oxidizing
bacteria. Appl. Environ. Microbiol. 56(8):2488-2493.
Davis, J.W., and C.L. Carpenter. 1990. Aerobic biodegradation of vinyl chloride in
groundwater samples. Appl. Environ. Microbiol. 56(12):3868-3880.
English, C.W. 1991. Removal of Organic Vapors in Unsaturated Soil. Dissertation
presented to the Faculty of the Graduate School of the University of Texas at
Austin. May 1991.
Henson, J.M, M.V. Yates, J.W. Cochran, and D.L. Shackleford. 1988. Microbial
removal of halogenated methanes, ethanes, and ethylenes in an aerobic soil
exposed to methane. FEMS Microbiol. Ecol. 52(3-4): 193-201.
Hoeks, J. 1972. Changes in composition of soil air near leaks in natural gas mains. Soil
Science. 113:46.
Janssen, D.B., A. Scheper, L.Dijkhuizen, and B. Witholt. 1985. Degradation of
halogenated aliphatic compounds by Xanthobacter autotrophicus GJ10. Appl.
Environ. Microbiol. 49(2):673-677.
Kampbell, D.H. and B.H. Wilson. 1993. Bioremediation of chlorinated solvents in the
vadose zone. In: In Situ and On-Site Bioreclamation. The Second International
Symposium. April 5-8,1993. San Diego, California. In Press.
Kampbell, D.H., J.T. Wilson, H.W. Read, and T.T. Stocksdale. 1987. Removal of volatile
aliphatic hydrocarbons in a soil bioreactor. J. Air Pollut. Cont. Assoc. 37(10): 1236-
1240.
Miller, D.E., and L.W. Canter. 1991. Control of aromatic waste air streams by soil
bioreactors. Environ. Progress. 10(4):300-306.
Plumb, R.H., and A.M. Pitchford. 1985. Volatile organic scans: Implications for
ground water monitoring. In: Proceedings of The Petroleum Hydrocarbons and
Organic Chemicals in Ground Water • Prevention, Detection, and Restoration.
National Water Well Association. Dublin, OH. pp. 207-221.
6-12
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Rittmann, B.E., and P.L. McCarty. 1980. Utilization of dichloromethane by suspended
and fixed-film bacteria. Appl. Environ. Microbiol. 39(6): 1225-1226.
Speitel, G.E., and F.B. Closmann. 1991. Chlorinated solvent biodegradation by
methanotrophs in unsaturated soils. J. Environ. Eng. 117(5):541-558.
Wackett, L.P., and D.T. Gibson. 1988. Degradation of trichloroethylene by toluene
dioxygenase in whole-cell studies with Pseudomonas putida Fl. Appl. Env.
Microbiol. 54(7): 1703-1708.
Wilson, B.H., and M.V. White. 1986. A fixed-film bioreactor to treat trichloroethylene-
laden waters from interdiction wells. In: Proceedings of The Sixth National
Symposium and Exposition on Aquifer Restoration and Groundwater Monitoring.
National Water Well Association. Dublin, OH. pp. 425-435.
6-13
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SECTION?
IN-SITUBIOREMEDIATION TECHNOLOGIES FOR PETROLEUM-DERIVED
HYDROCARBONS BASED ON ALTERNATE ELECTRON ACCEPTORS (OTHER THAN
MOLECULAR OXYGEN)
Martin Reinhard
Stanford University
Department of Civil Engineering
Stanford, California 943054020
Telephone: (415)7230308
Fax: (415)725-8662
7.1. FUNDAMENTAL PRINCIPLES
As currently practiced, conventional in-situ biorestoration of petroleum-
contaminated soils, aquifer solids, and ground water relies on the supply of oxygen to the
subsurface to enhance natural aerobic processes to remediate the contaminants.
However, anaerobic microbial processes can be significant in oxygen-depleted subsurface
environments that are contaminated with petroleum-based compounds and/or
chlorinated solvents. The purpose of this chapter is to discuss anaerobic
biotransformation of petroleum-derived ground-water contaminants, to discuss both
laboratory and field evaluation of the process, and to discuss important site conditions
that would influence a successful bioremediation.
7.1.1. Comparison of Oxygen and Alternate Electron Acceptor Based In-Situ
Bioremediation Technologies
In-situ bioremediation technology for the decontamination of soil and ground
water contaminated with petroleum-derived hydrocarbons involves the stimulation of
naturally occurring microorganisms that are capable of degrading the organic
contaminants (Atlas, 1981; Lee et al., 1988). Biostimulation consists of adding those
nutrients and/or electron acceptors that limit bacterial growth to the contaminated zone.
Bioaugmentation, on the other hand, involves introduction of adapted or genetically
engineered microorganisms into the contaminated aquifer. Using contaminants as
substrates for energy and growth, microorganisms convert the contaminants into
harmless products, principally COg, cell mass, inorganic salts, and water. When oxygen
is consumed, anaerobic microorganisms may grow using alternate electron acceptors.
Anaerobic degradation of aromatic hydrocarbons has initially been identified at
field sites (Reinhard et al. 1984) and in microcosm studies (Wilson etal., 1987) and has
now been demonstrated in the laboratory under a number of redox conditions, including
reduction of nitrate, iron(III) and manganese(IV) oxides, sulfate, and carbon dioxide.
In contrast to aromatic hydrocarbons, aliphatic hydrocarbon degradation without oxygen
has not been reported. In aquifers contaminated with biodegradable organic compounds,
electron acceptors tend to be used successively in order of decreasing free energy yield.
Oxygen is the preferred electron acceptor, followed by nitrate, manganese(IV) and
iron(III) oxides (MnO2 and FeOOH, respectively), sulfate, and carbon dioxide. This
sequence applies to pH 7 and should be valid for most field conditions where the
appropriate microorganisms occur.
7-1
-------
The conventional approach to hydrocarbon bioremediation is based on aerobic
processes. Anaerobic bioremediation has been tested only in a very few cases and is still
considered experimental. For instance, in a review of 17 sites contaminated with
hydrocarbon fuels and oils (Staps, 1990), hydrogen peroxide was used as the electron
acceptor at seven sites, air at five, combinations of nitrate-ozone and nitrate-air at one
site each, and nitrate alone was used only at three sites. Much available information has
been developed in laboratory studies; however, the applicability of these results to field
conditions remains to be studied. Anaerobic transformation rates can be slow and lag
times long and unpredictable, except for transformation in denitrifying systems which
can be fast. In spite of slow rates, anaerobic bioremediation could play a significant role
in the future mainly because the principal factor limiting aerobic bioremediation, the
difficulty of supplying oxygen to the subsurface, is circumvented.
Although aerobic biodegradation of refined petroleum products is relatively rapid
and complete under ideal growth conditions, application of anaerobic processes may be
preferable in ground waters because ideal aerobic growth conditions are difficult to
maintain in an aquifer. Rapid aerobic degradation requires ample supply of nutrients
and oxygen, good mixing, and a high microbial mass, conditions that are difficult to
maintain in aquifers (Wilson, B. et al., 1986; Lee et al., 1988). Furthermore, at many sites
there may be a very high abiotic oxygen demand due to hydrogen sulfide, Fe2* or other
readily oxidizable compounds, making it difficult to increase the reduction potential into
the aerobic range (> 0.82 V, see Table 8.5, Section 8). The advantages of aerobic
bioremediation may become inconsequential if the overall degradation rate is controlled
by slow dissolution, dispersion and/or desorption. Mass transfer limitations are
especially severe at sites where petroleum-derived hydrocarbons are present as
nonaqueous phase liquids (NAPLs). Under such conditions, natural or passive biological
degradation (aerobic or anaerobic) may be sufficiently fast for removing hydrocarbons
that are slowly released into the ground water.
At contaminated sites, a range of other, site-specific factors can limit
biotransformation, such as the occurrence of metals or other toxics, accumulation of
toxic intermediates and suboptimal temperatures. Nearby NAPL concentrations may
reach toxic levels, thereby limiting biological activity. It is unknown whether aerobic or
anaerobic processes are more readily inhibited by such factors. Thus, the decision to use
either aerobic or anaerobic processes may depend on site-specific conditions.
Since the consumption of 02 is relatively fast and the rate of Ojj supply is slow due
to low 62 solubility in water, expansion of the aerobic zone is limited by the rate of Oz
supply to the aquifer. Anaerobic conditions are expected to persist within aerobically
treated aquifers, especially in relatively impermeable zones and zones further away from
the injection wells. Water is an inefficient mass transfer medium for O2 due to the low
water solubility of 0%. For the degradation of relatively small amounts of hydrocarbons,
large amounts of water need to come in contact with the aquifer solids. The complete
oxidation of 1 mg hydrocarbon compounds requires 3.1 mg Oz (Hutchins and Wilson,
1991). Thus, for the bioremediation of 1 kg of aquifer material containing 10 g/kg
hydrocarbon compounds, a minimum of 3.1 m3 of oxygenated water containing 10 mg/1
02 must be supplied. Potentially, the overall degradation efficiency can be increased by
using alternate electron acceptors that are more water soluble. The ratio of feed water to
contaminant mass degraded is higher if the electron acceptor concentration in the feed is
increased. Nitrate salts are much more water soluble (92 g/1 or 1.33 M as sodium nitrate)
than 62 (10 mg/1 or 0.31 mM). Comparing the water solubilities and the half-reactions for
Q2 and nitrate reduction (Table 8.5, Section 8),
7-2
-------
2NO3- + 12H+ + 10 e- _^ N2 + 6H,,0 (1)
(2)
it is evident that the reducing equivalents that can be introduced into an aquifer using
saturated sodium nitrate solution is approximately 50 times higher than with a
saturated oxygen solution.
The use and potential benefits of anaerobic in-situ remediation technology are the
subject of this review. Nitrate, sulfate, iron(III) oxide and carbon dioxide and, to a very
limited extent, mixed electron acceptor systems are considered. However, only nitrate
and nitrate in combination with oxygen sources have been tested in field applications.
The general approach to in-situ bioremediation for ground-water cleanup has
been summarized by Sims et al. (1992). It is applicable to both aerobic and anaerobic
processes and consists of the following basic tasks:
(1) Site investigation to determine the distribution, mobility and fate of the
contaminants under site specific conditions.
(2) Performance of treatability studies to determine the potential for
bioremediation and to define the required operating and management
practices.
(3) Development of a bioremediation plan based on fundamental engineering
principles.
(4) Establishment of a monitoring program to evaluate performance of the
remediation effort.
The principal design considerations of the aerobic and anaerobic bioremediation
processes are the same because water serves as the nutrient feed solution in both cases.
The principal factors that must be considered include all aspects affecting mass transfer,
substrate retardation, and bioavailability. For anaerobic bioremediation, special
attention must be given to the adaptation status and growth condition of the indigenous
anaerobic bacteria, and the initial redox status of the aquifer. Potential advantages of
anaerobic bioremediation include:
(1) Alternate electron acceptors (except ferric iron) are more water soluble
and, consequently, require lower volumes of nutrient solution to be supplied
to the contaminated zone.
(2) Reduced plugging problems because of lower biomass yields of anaerobic
bacteria and a lesser tendency for iron precipitation.
7.1.2. Hydrocarbon Transformation Based on Alternate Electron Acceptors
7.1.2.1. Laboratory Studies
Grbic-Galic (1989, 1990) summarized the laboratory research on anaerobic
hydrocarbon transformation involving microcosms and enrichment cultures. Most
laboratory studies have attempted to (1) demonstrate biotransformation and
mineralization of the substrate by obtaining carbon mass balances, (2) identify the
electron acceptor by determination of the reaction stoichiometry, and (3) determine
7-3
-------
optimal growth conditions. Most studies have been conducted with microcosms or
enrichment cultures; although, in a few cases, isolation of pure strains has been
reported. For example, Dolfinget al. (1990) and Lovley and Lonergan (1990), respectively,
reported isolation of pure nitrate- and iron(III)-reducing strains capable of using
aromatic hydrocarbons as a sole carbon source. However, growth parameters and the
effect of environmental conditions on aromatic transformation have been investigated
only in a very few studies.
Under denitrifying conditions, oxidation of monoaromatic compounds has been
demonstrated in a number of systems (e.g., Kuhn etal., 1988; Mihelcic and Luthy, 1988;
Altenschmidt and Fuchs, 1991; Ball et al., 1991; Evans et al., 199 la; Evans et al., 1991b;
Flyybjerg etal., 1991; Hutchins et al., 1991a; Evans et al., 1992). The stoichiometry of the
denitrification reaction of toluene assuming no cell growth with NO3' reduced completely
to N2 and toluene completely oxidized to CO2 is
C^Hg + T^H^T^NOg-—^ 3.6 N2 + 7.6 H20 + 7 CO2 (3)
When biodegradation of benzene, toluene, ethylbenzene and xylenes (BTEX)
mixtures was tested under denitrifying conditions, degradation tended to be sequential,
with toluene being the first substrate to be degraded, followed by the degradation of p- and
m-xylene, ethylbenzene and o-xylene. Mihelcic and Luthy (1988) reported the
degradation of naphthalene. Benzene does not seem to be degraded (Kuhn et al., 1988;
Evans et al., 1991a; Evans et al., 1991b; Hutchins et al., 199 la) although in one study
Major et al. (1988) reported removal under conditions thought to be denitrifying.
Hutchins et al. (199 la) reported longer lag times and slower degradation rates in core
material contaminated with JP-4 aviation fuel than in uncontaminated core material.
Using an enrichment culture and ethylbenzene as the substrate, Ball et al. (1991)
have shown that single aromatic substrates can be degraded rapidly (within hours) and
that nitrate reduction to nitrogen gas proceeds through nitrite. Similar findings were
reported by Evans et al. (199 la, 1991b) for toluene. Ball et al. (1991) also demonstrated that
composition and preparation of the growth medium can affect the observed
transformation rates.
Ball et al. (1991) tested inocula from different sources for the potential to degrade
BTEX compounds. They found that microorganisms with the ability to degrade aromatic
hydrocarbons are not ubiquitous. Sewage seed that contains a diverse population of
microorganisms, for instance, did not adapt to the aromatic compounds tested. Much
work remains to be done before such experimental data can be interpreted with
confidence and the biodegradation potential under field conditions can be predicted.
Systems
Bioremediation using sulfate as the electron acceptor involves oxidation of
aromatic hydrocarbons by sulfidogenic organisms coupled with reduction of sulfate to
hydrogen sulfide (Edwards et al., 1991; Haag et al., 1991; Beller et al., 1992; Edwards et
al., 1992). For toluene, the stoichiometry of this reaction, assuming no cell growth, may
be written as (Beller et al., 1992)-
C7H8 + 4.5 SO42- + 3 HjjO — *-
2.25 HS-+ 2.25 H 28+ 7HCO3+ 0.25 H+ (4)
7-4
-------
As in some denitrifying systems, degradation under sulfate-reducing conditions
is also sequential with toluene being the preferred substrate, followed byp-xylene and
with o-xylene degraded last (Edwards et al., 1991,1992). Ethylbenzene and benzene were
not degraded under the conditions of the experiment. In a follow-up study, Edwards and
Grbic-Galic (1992) observed benzene degradation in the absence of all other aromatic
substrates. After a lag time of 30 days under strictly anaerobic conditions, these authors
observed mineralization of benzene and suspected sulfate to be the electron acceptor.
Accumulation of HS" may inhibit the process, however, and is a problem that remains to
be resolved.
'ir'n
Lovley and Lonergan (1990) have isolated an iron-reducing bacterium capable of
degrading toluene, p-cresol and phenol. For toluene, the stoichiometry of the process is
C7H8 + 36 Fe3+ + 21 H2O —^ 36 Fe2* + 7 HCOs + 43 H+
(5)
whereby 36 moles of Fe(III) are required to oxidize one mole of toluene. Relative to other
anaerobic processes, Fe(III) reduction has a very unfavorable substrate to electron
acceptor ratio. Lovley et al. (1989) found that toluene was transformed into CQz and Fe2*
at a ratio which agreed with the above stoichiometry.
Transport of the dissolved Fe(II) from the aquifer could cause secondary problems
such as clogging and fouling of the aquifer. Furthermore, the supply of large amounts
of colloidal iron(III) oxide or soluble Fe(III) citrate (Lovley et al., 1989) to an aquifer has
not been tested. To develop bioremediation strategies based on iron reduction, a better
understanding of occurrence, nutritional requirements, growth conditions and
metabolism of iron-reducing bacteria must be developed.
Fermentativc/Carbon Dioxide-Reducing Systgnjff
Under methanogenic/fermentative conditions, several aromatic hydrocarbon
compounds, including benzene and toluene, have been shown to transform into CO 2 and
methane (Grbic-Galic and Vogel, 1987). The culture originated from sewage seed and
was enriched under methanogenic conditions using ferulic acid as the only carbon
source. Assuming no cell growth, the stoichiometry for the transformation reaction is
C/He+SHzO —>> 2.5 CO 2 +4.5 CH4 (6)
Biotransformation was studied with toluene or benzene as the only carbon source.
Biotransformation began after a three month lag time and was complete after 60 days of
incubation. Since this ground-breaking study, several other aromatic substrates have
been shown to be degraded under methanogenic conditions, including styrene,
naphthalene and acenaphthalene (Grbic-Galic, 1990), as well as benzothiophene, a
sulfur-containing heterocyclic compound (Godsy and Grbic-Galic, 1989).
Fermentation/methanogenic degradation could be used as a passive
bioremediation technology (Section 9) and is likely to be an ongoing process at many sites
where the geochemical conditions have evolved naturally, without human intervention.
Reliable assessment of the process is difficult under field conditions since mass balances
are difficult to establish. An indication that the process is occurring is the detection of
methane in combination with characteristic intermediates such as aromatic acids
(Reinhard et al., 1984; Wilson et al., 1987; Baedecker and Cozzarelli, 1991).
7-5
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Acceptor Svste"is
In aquifer segments augmented by electron acceptors, different electron acceptors
are likely to co-occur either within the same aquifer compartment, or spatially separated
into adjacent compartments. Few laboratory studies have examined mixed electron
acceptor systems, although they are likely to be common at field sites. For instance, at
the sites where denitrifying conditions were investigated, 0% was frequently present in
the nitrate feed water. Both electron acceptors were consumed, but the effect of the
oxygen on the overall process was not determined.
Different electron acceptors and products of aromatic degradation processes can
react with each other in a number of biological and chemical reactions. Seller et al.
(1992) have studied the link between dissimilatory sulfate reduction to sulfide and
iron(III) reduction to iron(II) by a sulfate-reducing enrichment culture. Ferric iron
appeared to reoxidize hydrogen sulfide in an abiotic process and/or lower the inhibitory
effect of hydrogen sulfide. Toluene was the sole carbon and energy source, but other
substrates were not tested.
7.1.2.2. Large Scale Bioremediation Studies Using Nitrate
Nitrate is the only alternate electron acceptor with demonstrated potential for use
in large scale in-situ bioremediation applications involving petroleum-derived
hydrocarbons (Table 7.1). The other possible alternate electron acceptors (iron(III),
sulfate, and CC^) have been found in systems that may be classified as passive
bioremediation, such as in landfill leachate plumes (Reinhard et al., 1984) and at spill
sites (Lovleyet al., 1989).
Table 7.1 summarizes results of selected field studies, where denitrification was
tested as a means to remove aromatic hydrocarbon contamination. The Traverse City
and the Rhine River Valley studies involved actual contamination sites. The Borden
experiment involved injection into the aquifer of a mixture containing benzene, toluene,
and xylene isomers (BTX) in one experiment and gasoline in another. In general, BTX
compounds were found to disappear within the nitrate amended zone. Interpretation of
these data, however, was complicated by a number of factors, especially the
co-occurrence of nitrate and 0%, and the lack of complete characterization of the organic
substrates. Nitrate removal exceeded the expected amount based on the substrates
analyzed, and this was attributed to the dissolved organic carbon in ground water
(Berry-Spark et al., 1988). Werner (1985) proposed that if O2 and nitrate are present
simultaneously, O2 is used for the first oxidation step to produce partially oxygenated
products and nitrate is then used for mineralization of the oxidation products.
7.1.3. Maturity of the Technology
Anaerobic transformation of aromatic hydrocarbon compounds is a very recent
discovery, and it is too early to predict an impact on bioremediation technology. The
technology to use nitrate has been tested in a limited number of cases and is still highly
immature, although available data are very promising. Research on aromatic
transformation under all other reducing conditions is still at very early stages.
Currently, it is not possible to predict the site conditions under which biostimulation
using nitrate or sulfate will be successful and define optimal operating conditions for a
given site.
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TABLE 7.1. FIELD STUDIES WHERE DENTTRIFICATION HAS BEEN EVALUATED
STUDY SITE AND
AUTHORS
CONTAMINATION AND
CONDITIONS
MAJOR IMPLICATION FOR IN-SITU
BlOREMEDIATlON
Traverse City, MI,
Hutchins and
Wilson, 1991
Borden, Ontario,
Berry-Spark et al.,
1988
Seal Beach,
Reinhard et al., 1991
Rhine Valley, FRG,
Werner, 1985
JP-4 fuel;
NaNO3:62 mg/1;
Oa: 0.5 to 1 mg/1
Gasoline and
BTX; Oxygen
and nitrate
Gasoline
contaminated
ground water feed;
N03-(6mg/l)
Fuel oil (?);
Aerated water (02);
NO3- (>300 mg/1);
P04(>0.3mg/l);
NH4* (>1.0 mg/1)
(1) removal of benzene, toluene,
m,p-xylene; recalcitrance of o-xylene.
(2) nitrate removed exceeded
stoichiometric amount of BTEX
removal.
(3) partitioning of compounds into the
water phase appears to be a major
factor determining compound removal.
(1) BTX transform more slowly when
gasoline is present than in systems
where BTX are the only substrate.
(2) In systems containing both O? and
nitrate, aerobic and (facultative)
denitrifying organisms appear to
cooperate.
90% nitrate removal in mixed
nitrate/sulfate system, aromatics removal
toluene>p-xylene>o -xylenobenzene
(1) Removal was fastest for benzene,
slower for toluene and slowest for
p-xylene.
(2) Oxygen suspected to be electron
acceptor initiating the transformation.
7.1.4. Primary Repository rf Expertise
Application of the technology requires a broad range of expertise including
microbiology, ground-water hydrology, geochemistry and engineering. Such expertise is
available only at a few U.S. governmental laboratories (Environmental Protection Agency
and Geological Survey) and U.S. and European universities. Anaerobic microbiology of
pollutant transformation is being studied by a growing number of academic and
governmental research laboratories.
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CONTAMINATION THAT IS SUBJECT TO TREATMENT
7.2.1. Chemical Nature
Denitrification has been shown to be effective only for monocyclic and polycyclic
aromatic hydrocarbons. Aliphatic hydrocarbon compounds appear to be nondegradable
in anaerobic systems.
7.2.2. Range cf Concentration
The upper and lower concentration limits for which the technology applies have
not yet been defined with certainty. Berry-Spark et al. (1988) were not able to identify a
lower limit of BTX removal in a field study where BTX compounds were artificially
injected. Most laboratory studies using single substrates or substrate mixtures were
conducted with compound concentrations in the mg/1 range.
73. REQUIREMENTS FOR SITE CHARACTERIZATION AND IMPLEMENTATION
OFTHE TECHNOLOGY
For designing and implementing the bioremediation plan, the site has to be
characterized with respect to physical (hydrologic), chemical, and biological
characteristics. Most of these characteristics are generic to all bioremediation
applications.
Physical
-------
3) Limiting factors, including limiting nutrients.
7.4. FAVORABLE SITE CHARACTERISTICS
Site characteristics that are generally favorable for bioremediation include
shallow and permeable aquifers, which can readily be supplied with nutrients without
excessive pumping costs. However, these characteristics are also favorable for other
remediation technologies such as soil gas extraction, or excavation. Anaerobic
bioremediation technology, passive or active, should be of greater advantage at sites that
are not amenable to treatment with conventional technologies, such as deep or otherwise
inaccessible sites.
7.5. UNFAVORABLE SITE CHARACTERISTICS
7.5.1. Chemical and Physical Nature of the Contamination
1) NAPLs may act as a long-term source for contaminants and may be toxic
for microorganisms.
2) Mixed wastes may inhibit adaptation of native organisms.
3) Inhibition by toxic metals. Bollag and Barabasz (1979) and Werner (1985)
report that heavy metals such as cadmium, copper, lead and zinc impair
denitrifying activity.
4) Water quality characteristics, i.e. pH or salt content, are incompatible with
bacterial physiology.
7.5.2. Site Hydrogeology and Source Characteristics
1) Aquifer heterogeneity and clay lenses may make contaminated pockets
inaccessible for treatment.
2) Source boundaries are unknown.
3) Dissolution of minerals may lead to secondary water quality problems.
7.5.3. Infrastructure and Institutional Issues
1) Efficiency of technology is unproven.
2) Potentially harmful by-products or end-products may be formed.
7.6. OPTIMAL SITE CONDITIONS
Optimal growth condition for anaerobic bacteria growing on aromatic
hydrocarbons have yet to be determined. Of specific interest are biomass, biomass
diversity, biomass yields, temperature, bioavailability, nutrients, formation of inhibiting
intermediates, end- or by-products (Texas Research Institute, 1982) and factors
specifically related to ground-water conditions such as surface attachment and mobility.
Sims et al. (1992) and Wilson, J. et al. (1986) list the following issues:
7-9
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1) The concentration of required nutrients in the mobile phase.
2) The advective flow in the mobile phases or the steepness of concentration
gradients within the phases.
3) Opportunity for colonization of microbially active organisms capable of
contaminant degradation.
4) Co-occurrence of waste materials that may inhibit biotransformation.
7.7. PROBLEMS ENCOUNTERED WITH THE TECHNOLOGY
Data to evaluate problems with the technology are very sparse. Only data on
denitrifying systems have been reported in the literature and only for very few sites. The
main concern of these studies has been the efficacy of the process and not potential
problems of the technology. Because failed attempts to remediate a site based on
denitrification have not been documented, it is difficult to state whether application of the
technology is limited by some inherent problems, such as the formation of intermediates
and products of health concern. Werner (1991) states without elaboration that the
technology "cannot be transferred to the general practice of bioremediation."
Some specific limitations have been reported by Hutchins et al. (1989; 1991b):
Only aromatics are subject to treatment.
Higher molecular weight compounds, which were sorbed more strongly,
were not degraded.
Leaching of the more soluble compounds by the nutrient feed water
appeared to be a major removal mechanism and may pose secondary
disposal problems.
7.8. PROPERTIES OF SITE AND CONTAMINANTS DETERMINING THE COST OF
REMEDIATION
Since supply of nutrient solutions is expected in all systems, hardware
installation costs of anaerobic systems are expected to be the same as for aerobic systems.
Pumping costs should be lower, however, due to higher electron acceptor concentrations
and, consequently, lower nutrient solution volumes. This advantage could be diminished
if cleanup times caused by slow growth rates of anaerobes are very long. In any case,
source characterization, contaminant distribution and determination of the hydrologic
conditions, i.e., site heterogeneity and co-occurrence of mixed wastes, are major cost
factors and are the same as those for conventional pump-and-treat and aerobic
bioremediation. Thus, anaerobic processes are most likely of greatest advantage in
passive remediation schemes where the electron acceptors naturally present can be
utilized and the costs of external nutrient supply can be avoided.
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7.9. PREVIOUS EXPERIENCE WITH COST OF IMPLEMENTING THE
TECHNOLOGY
Hutchins et al. (1989, 199Ib) evaluated treatment costs of the Traverse City
bioremediation project, which used nitrate as the electron acceptor. They calculated unit
costs for the remediation with respect to (1) volume of JP-4 fuel removed, (2) volume of
contaminated aquifer material treated, and (3) total aquifer material treated and
considered costs for construction, labor, chemicals, and electrical service. The unit costs
for the remediation were $22 per liter JP-4, $200 per m3 aquifer material contaminated
with JP-4, and $17 per m3 of aquifer material down to the confining aquifer. Of course, to
assess the viability of the technology, the costs of this technology must be compared with
other technologies including conventional and passive treatment technologies.
7.10. FACTORS DETERMINING REGULATORY ACCEPTANCE OF THE
TECHNOLOGY
Regulatory acceptance of the technology is limited by:
1) Inadequate understanding of the underlying science, including the
chemical, biological and hydrologic factors that guarantee the success of
the technology.
2) Lack of successfully completed and documented field demonstration
projects.
3) Unpredictable transformation rates leading to uncertain cleanup times
and poor process control.
4) Uncertain environmental impact, potential formation of harmful by- and
end products.
5) Lack of treatment objectives commensurate with the capabilities of the
treatment technologies.
6) Nitrate is regulated with respect to the National Drinking Water Standard.
The maximum contaminant level for nitrate is 10.0 mg/1 measured as
nitrogen.
These potentially unfavorable factors may be compensated for by reduced
environmental impacts associated with conventional remediation technologies (e.g.,
excavation and off-site disposal).
7.11. PRIMARY KNOWLEDGE GAPS AND RESEARCH OPPORTUNITIES
Because anaerobic biotransformation of aromatic hydrocarbon compounds has
been reported only recently, details of the process such as characteristics of the
organisms and microbial communities, and factors which effect rates and adaptation
times, are not yet sufficiently understood. As the published reports are getting more and
more detailed, increasingly specific questions can be asked. The questions listed below
have been excerpted from the cited reports.
7-11
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7.11.1. Degradation Under Ideal Conditions:
1) In what habitats do we find microorganisms capable of transforming
aromatic hydrocarbon compounds, specifically,
- to what extent is pre-exposure of the inoculum to hydrocarbons and exact
reproduction of the site-geochemical conditions a prerequisite for aromatic
hydrocarbon transformation capability?
- what are the geochemical characteristics of these habitats?
2) What are the nutritional requirements of the microorganisms that are
capable of anaerobically degrading aromatic hydrocarbons?
3) What are typical lag-times, specifically
- how do lag times depend on factors such as substrate structure,
co-occurrence and concentration of other more easily degradable substrates
and environmental factors?
4) What are the degradation rates under ideal conditions and how are they
influenced by environmental factors such as the presence of oxidants,
temperature, pH?
5) What are the pathways by which aromatic hydrocarbon compounds are
degraded and what intermediates and dead-end products are formed?
6) What products and intermediates are toxic?
7.11.2. Degradation Under Ground-water and Soil Conditions
1) What degradation rates can be expected in porous, sorbing media?
2) How do redox-active solids such as iron(III) oxides and iron sulfides
influence the process?
3) What abiotic processes are linked to the biodegradation?
4) How can nutrient conditions and environmental factors be improved?
5) Are humic and fulvic acids serving as electron acceptors?
7.11.3. Degradation at Sites
1) How can favorable growth conditions be stimulated under site conditions?
2) What is the minimum aquifer permeability?
3) What is the effect of aquifer heterogeneity?
4) What is the effect of nonaqueous phase liquid hydrocarbons?
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7.11.4. Degradation at the Hydrocarbon/Water Interface and Within the Nonaqueous
Phase
1) Are organisms growing at the oil/water interface and/or within the
nonaqueous phase?
2) What factors are limiting the growth of these bacteria?
3) What is the significance of these organisms in bioremediation schemes?
7.11.5. Methods for Monitoring Performance of Bioremediation Process
1) What in-situ methods are suitable for monitoring the process efficiency?
7.11.6. Design of Optimal Nutrient and Electron Acceptor Systems
1) What are the best nutrient feed systems considering local hydrogeological
and geochemical factors?
ACKNOWLEDGEMENT
I thank Harry A. Ball and Harry R. Beller for helpful dicussions of the
manuscript.
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Edwards, E.A., and D. Grbic-Galic. 1992. Complete mineralization of benzene by aquifer
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Evans, P.J., D.T. Mang, K.S. Kim, and L.Y. Young. 1991b. Anaerobic degradation of
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Evans, P.J., W. Ling, B. Goldschmidt, E.R. Ritter, and L.Y. Young. 1992. Metabolites
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Flyvberg, J., E. Arvin, B.K. Jensen, and S.K. Olson. 1991. Bioremediation of oil- and
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of alkylated benzenes in denitrifying laboratory aquifer columns. Appl. Environ.
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SECTION 8
BIOREMEDIAT1ON OF CHLORINATED SOLVENTS USING
ALTERNATE ELECTRON ACCEPTORS
Edward J. Bouwer
Department of Geography and Environmental Engineering
The Johns Hopkins University
Baltimore, Maryland 21218
Telephone: (410)516-7437
Fax: (410)516-8996
8.1. INTRODUCTION
The contamination of ground water and soils with chlorinated solvents, such as
trichloroethene (TCE), tetrachloroethene (PCE), carbon tetrachloride (CT),
1,1,1-trichloroethane (1,1,1-TCA), and chloroform (CF), is widespread (Pye et al., 1983).
Their extensive production and use makes these compounds among the most prevalent
contaminants in ground water at waste disposal sites. Over 270 million metric tons of
the fifty most widely used chemicals were produced in 1988 (Chem. Eng. News, 1989).
Synthetic organic compounds including chlorinated solvents accounted for
approximately one-third of the chemical production. Many of these chlorinated
compounds are known or potential threats to public health and the environment, so there
is an urgent need to understand their fate in the environment and develop effective
control methods. Flushing the subsurface with water so that chlorinated solvents can
dissolve and be pumped to the surface for aboveground treatment is used most frequently
for remediation. Because the subsurface is geologically complex and chlorinated
solvents tend to sorb to soils, they are not readily leached from the soil, and such pump-
and-treat systems are generally inefficient and slow. Furthermore, most aboveground
treatment technologies involve physical/chemical processes (e.g., air stripping and
carbon adsorption) that simply sequester the contaminants or transfer them to another
environmental medium.
Biological processes offer the prospect of converting organic contaminants to
harmless products. This cleanup approach, termed bioremediation, stimulates the
growth of indigenous or introduced microorganisms in regions of subsurface
contamination and, thus, provides direct contact between microorganisms and the
dissolved and sorbed contaminants for biotransformation. The process typically entails
perfusion of nutrients and one or more electron donors or electron acceptors through the
contaminated soil. Certain chlorinated solvents are biotransformed by methanotrophic
bacteria under aerobic conditions, and such aerobic bioremediation has been
demonstrated to be successful on a small scale in the field (Section 5). However, the
aerobic methanotrophic bacteria cometabolize the chlorinated solvents while using
methane as a primary substrate. The limited solubility of methane and oxygen in water
and competition between methane and the chlorinated solvent for the initial enzyme can
restrict the growth of microorganisms and biodegradation of contaminants in the vicinity
of a spill, thus reducing the success of aerobic bioremediation. Hydrogen peroxide can be
used to increase the oxidant capacity, but this also has disadvantages, including its
toxicity to microorganisms and its reactivity with aquifer materials.
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Anaerobic bioremediation where electron acceptors other than oxygen are used is
potentially advantageous for overcoming the difficulty in supplying oxygen for aerobic
processes. Nitrate, sulfate, and carbon dioxide are attractive alternatives to oxygen as an
electron acceptor because they are very soluble in water, inexpensive, and nontoxic to
microorganisms. Their high aqueous solubility and low reactivity relative to oxygen
make them easier to distribute throughout a contaminated zone. Fe(III) and Mn(IV)
might be present in the mineral phases of aquifer solids and could serve as alternate
electron acceptors for iron- and manganese-reducing bacteria, respectively. Exploiting
anaerobic microbial processes for bioremediation of chlorinated solvents is in its infancy.
Demonstration of this technology in the field is limited; therefore, the use of alternate
electron acceptors for bioremediation of chlorinated solvents must be viewed as a
developing treatment technology. Establishing the utility of anaerobic bioremediation for
chlorinated solvents is an important scientific and engineering challenge.
This section addresses some important issues concerning the transformation of
chlorinated solvents in the absence of oxygen that can be applied to the problem of
environmental contamination as well as to the development of engineered treatment
processes for subsurface cleanup. These include a discussion of metabolism and
biotransformation of chlorinated solvents with alternate electron acceptors, approaches
for treatment, reaction stoichiometry, biotransformation rates, and limitations.
&2. METABOLISM AND ALTERNATE ELECTRON ACCEPTORS
Biotransformations are driven by the ultimate goal of increasing the size and
mass of microbial populations. Microorganisms must transform environmentally
available nutrients to forms that are useful for incorporation into cells and synthesis of
cell polymers. In general, cells utilize reduced forms of nutrients for these synthesis
reactions. Reducing nutrients requires energy and a source of electrons. An electron
donor is essential for growing cells; energy is made available for cell growth when the
electron donor transfers its electrons to a terminal electron acceptor. Following is an
example of a biotransformation in which an organic contaminant typified as benzene
(CeHe) serves as electron donor and is oxidized to innocuous compounds and supports
microbial growth:
7.502 — ^6CO2+3H2O (la)
C6Hg + 1.5 HCO3' + 1.5 NH4* — »» 1.5 C^Cv^N + 1.5 H2O (Ib)
In Reaction la, the transfer of electrons between benzene (electron donor) and QZ
(electron acceptor) provides energy for synthesis of cellular material (CsHvOgN) from the
benzene carbon (Reaction Ib). By this process, a portion of an organic contaminant
serves as a primary energy source that is converted to end products, and a portion of the
contaminant carbon is synthesized into biomass.
The terminal electron acceptor used during metabolism is important for
establishing the redox conditions and the chemical speciation in the vicinity of the cell.
Common terminal electron acceptors include oxygen under aerobic conditions, and
nitrate, Mn(IV), Fe(III), sulfate, and carbon dioxide under anaerobic conditions.
Microorganisms preferentially utilize electron acceptors that provide the maximum free
energy during respiration. Of the common electron acceptors used by microorganisms,
oxygen has the highest redox potential and provides the most free energy to
microorganisms during electron transfer (Figure 8.1). The redox potentials of nitrate,
Mn(IV), Fe(III), sulfate, and carbon dioxide are lower (Figure 8.1). Consequently, they
8-2
-------
yield less energy during substrate oxidation and electron transfer according to the order
listed in Figure 8.1. These latter compounds comprise the alternate electron acceptors
available for development of anaerobic bioremediation technologies. The importance of
microbial reactions involving Mn(IV) and Fe(III) to organic contaminant
biotransformations is unknown. Therefore, this section will focus on microbial systems
involving nitrate (denitrification), sulfate (sulfate reduction), and carbon dioxide
(methanogenesis) as electron acceptors.
1.0
Aerobic
(Oxygen as /
Electron Acceptor) l
*
> 0.5
ii
I
g
o
1
OC
Typical Primary .
Substrates I =
(Electron Donors)
-0.5'
+ 4e--»2H2O
2NO3 + 12H* + 10e* -» N2 + 6H20
Mn02(s) + HCOj + 3H* + 2e~ -»
MnCO3(s)+2H2O
FeOOH(s) + HCO3 + 2H* + e- -»
FeC03(s)+2H20
SO; + 9H+ + 8e~ -> HS~ + 4H2O
CO2 + 8H* + 8e- -4 CH4 + 2H2O
2CO2 + 8H* + 8e- -» CH3COOH + 2H2O
in
e>
*SI
a
•o
Ul
?
M
i
i
Figure 8.1. Important electron donors and acceptors in biotransformation processes. Redox
potentials were obtained from Stumm and Morgan (1981).
&3. BIOTRANSFORMATION OF CHLORINATED SOLVENTS IN THE PRESENCE
OF ALTERNATE ELECTRON ACCEPTORS
The redox environment is an important factor affecting microbial respiration and
biotransformation of organic contaminants. Some compounds are only transformed
under aerobic conditions; others require strongly reducing conditions; and still others
are transformed in both aerobic and anaerobic environments. Reviews of aerobic and
anaerobic biotransformations of petroleum hydrocarbons and aerobic biotransformation
of chlorinated solvents are presented in other sections within this volume. Examples of
-------
how metabolism with alternate electron acceptors influences the biotransformations of
some chlorinated solvents of concern are described in this section. This knowledge
coupled with the spatial distribution of electron acceptors and other redox species within
a region of subsurface contamination is important for identifying zones conducive to
biotransformation of a particular chlorinated solvent. The coupling of redox conditions
and chlorinated solvent biotransformation is also important in establishing how to
chemically manipulate the medium to achieve a desired biotransformation.
In the absence of molecular oxygen, microbial reduction reactions involving
organic contaminants increase in significance as environmental conditions become
more reducing. Nearly 25 years ago, Castro and Belser (1968) found that the soil
fumigants ethylene dibromide (1,2-dibromoethane), l,2-dibromo-3-chloropropane, and
2,3-dibromobutane were transformed in soil slurries via a reductive dehalogenation
reaction. In reductive dehalogenation, the halogenated compound becomes an electron
acceptor; and in this process, a halogen is removed and is replaced with a hydrogen
atom. About a decade later, investigations were initiated to evaluate the fate of
chlorinated solvents (mainly chloromethanes and chloroethenes) in anaerobic
environments. The results of investigations with microcosms and enrichments from
environmental samples under anaerobic conditions are summarized in Table 8.1.
Anaerobic biotransformation of chlorinated solvents has been observed in field studies
(Roberts et al., 1982), in continuous-flow fixed-film reactors (Bouwer and McCarty, 1983b;
Vogel and McCarty, 1985, 1987; Bouwer and Wright, 1988), and in soil (Kloepfer et al.,
1985), sediment (Barrio-Lage et al., 1986), and aquifer microcosms (Wilson, B. et al., 1986)
under conditions of denitrification, sulfate reduction, or methanogenesis. The initial
step in the anaerobic biotransformation was generally reductive dechlorination. For
example, CF was produced from CT, and 1,1-dichloroethane (1,1-DCA) was produced
from 1,1,1-TCA (Table 8.1).
The transformations of PCE and TCE have been studied most intensely. General
agreement exists that transformation of these two compounds under anaerobic
conditions proceeds by sequential reductive dechlorination to dichloroethene (DCE) and
vinyl chloride (VC); and in some instances, there is total dechlorination to ethene or
ethane. Of the three possible DCE isomers, 1,1-DCE is the least significant intermediate;
several studies have reported that cis-l,2-DCE predominates over trans-l,2-DCE
(Barrio-Lage et al., 1986; Parsons et al., 1984; Parsons and Lage, 1985). CT, CF, 1,2-DCA,
1,1,1-TCA, and PCE were partially converted to carbon dioxide during the anaerobic
biotransformations. Reductive dechlorination of 1,1,1-TCA and PCE occurred first prior
to mineralization to carbon dioxide. Most of the experiments were conducted under
methanogenic conditions. Several of the chlorinated compounds were also transformed
by similar pathways under conditions of denitrification and sulfate reduction (Table 8.1).
Several studies provide evidence for anaerobic transformation of chlorinated
solvents by pure cultures of bacteria (Table 8.2). The bacteria involved ranged from strict
anaerobic microorganisms, such as methanogens, sulfate-reducers, and clostridia to
facultative anaerobes such as Escherichia coli or Pseudomonas putida. Reductive
dechlorination was the predominant reaction pathway. Consequently, the chlorinated
solvent biotransformation studies with environmental samples (mixed microbial
cultures) and pure bacterial cultures indicate that a broad variety of bacteria possess the
enzymatic capability to reductively dechlorinate the compounds. An electron donor, such
as low molecular weight organic compounds (lactate, acetate, methanol, glucose, etc.) or
H2, must be available to provide reducing equivalents for reductive dechlorination.
Toluene was recently found to be a suitable electron donor for the reductive
dechlorination of PCE to DCE in anaerobic aquifer microcosms (Sewell and Gibson, 1991).
-------
TABUE 8.1. ANAEROBIC TRANSFORMATION OF SELECTED CHLORINATED SOLVENTS IN
MICROCOSMS AND ENRICHMENT CULTURES UNDER DIFFERENT REDOX CONDITIONS
Chlorinated
solvent0
CT
CF
1,2-DCA
1,1.1-TCA
1,1.2,2-TeCA
HCA
1,1-DCE
C1S-1.2-DCE
rrans -1,2-
DCE
TCE
PCE
Redox
condition^
dn
sr
me
dn
sr
me
me
dn
sr
me
me
ae
dn
sr
me
me
me
me
me
sr
me
Transfor-
mation c
+
+
+
..
.-
+
•*•
..
+
+
+
4-
+
+
+
+•
•f
+
*•
+•
+
Intermediate1*
CF
CF
CF
_.
..
nd
nd
..
1,1-DCA
1,1-DCA
-•
..
nd
nd
nd
VC
VC
CA + VC
as + trans- 1.2-DCE
1,2-DCE
TCE
TCE
End
product
nd
n d
COz
__
..
C02
CO2
_.
CA
CO2
1,1.2-TCA
PCE
nd
nd
nd
nd
nd
n d
n d
nd
cis- 1,2-DCE
COj
ethene
cis + trans-
1,2-DCE
c
-------
TABU: &2. REDUCTIVE DEHALOGENATION REACTIONS CATALYZED BY PURE CULTURES OF
BACTERIA
Halogenated
compound?
Bacteria
Products
Refer-
enceW
CT
CF
1,2-DCA
1,1,1-TCA
Methanobacterium
thermoautotrophicum
Methanosarcina barkeri
Desulfobacterium autotrophicum
Acetobacterium woodii
Clostridium tkermoaceticum
Clostridium sp
Escherichia coli
two Methansarcina sp.
several methanogens
Methanobacterium
thermoautotrophicum
Desulfobacterium autotrophicum
Acetobacterium woodii
Clostridium sp.
CF- DCM + C02
CF- DCM
CF- DCM + C02
CF- DCM - CM + CO2
CF - DCM - CM + C02
CF — DCM + unidentified
CF
DCM- CM
ethene
1,1-DCA
1,1-DCA + acetate + unidentified
c,e
i
c,d,e
d,e
d
h
b
J
a.d
c
c
d
h
BA
1,2-DBA
PCE
1,2-DBE
several methanogens
several methanogens
several methanogens
Desulfomomle tiedjei
Acetobacterium woodii
several methanogens
ethane
ethane
TCE
acetylene
a
a
c,f,g
g
d
a
Abbreviations stand for: CT = carbon tetrachlonde; CF = chloroform; DCA = dichloroethane;
TCA = trichloroethane; BA = bromoethane; DBA = dibromoethane; DBE = dibromoethene; CM =
chloromethane; DCM = dichloromethane; TCE = trichloroethene; PCE = tetrachloroethene.
a = Belay and Daniels, 1987; b = Criddle et al., 1990a; c = Egli et al., 1987; d = Egli et a]., 1988; e =
Egli et al., 1990; f = Fathepure and Boyd, 1988; g = Fathepure et al., 1987; h = Galli and McCarty,
1989; i = Krone et al., 1989; j = Mikesell and Boyd, 1990
Conversion of the chlorinated aliphatic compounds to less chlorinated alkenes and
alkanes via reductive dechlorination is of little or no benefit in the context of anaerobic
bioremediation. The intermediates commonly observed, such as cis-l,2-DCE,
trans- 1,2-DCE, VC, 1,1-DCA, and CF, also pose a threat to public health. The possible
formation of toxic metabolites has been the major impediment to the development of
practical anaerobic bioremediation in the field for cleanup of chlorinated solvent
contamination. For anaerobic bioremediation to be useful, chlorinated solvents must be
-------
biotransformed to nonchlorinated, environmentally acceptable products. Some recent
laboratory studies have demonstrated that this is possible and help provide impetus to
further develop anaerobic biological processes for bioremediation.
8.3.1. Carbon TetrachlorideBiotransformation
A denitrifying Pseudomonas sp. (strain KG) capable of rapid and complete
biotransformation of CT was isolated from aquifer material (Griddle et al., 1990a). This
bacterium was able to completely convert CT to carbon dioxide and an unidentified
water-soluble fraction without simultaneous production of CF. Denitrification was
confirmed by consumption of nitrate and acetate as primary substrate and production of
protein. This CT biotransformation could potentially be exploited in anaerobic
bioremediation of contaminated ground water. The use of denitrifying organisms would
be advantageous because nitrate is highly soluble in water and easily added. When
reduced iron and cobalt were provided in the growth medium, CT biotransformation to
mineralized products was inhibited. Consequently, careful attention must be paid to the
trace-metal availability in the engineering of systems that direct CT to nonhazardous end
products.
8.3.2. Tetrachloroethene and Trichloroethene Biotransformation
Reductive dechlorination of PCE, via TCE, cis-l,2-DCE, and VC to ethene was
observed at 20°C in a laboratory-scale fixed-bed column packed with a mixture (3:1) of
anaerobic sediment from the river Rhine and anaerobic granular sludge (de Bruin et al.,
1992). In the presence of lactate (1 mM) as an electron donor, 9 uM PCE was
dechlorinated to ethene. Ethene was further reduced to ethane. Nearly complete
conversion (95 - 98%) of PCE occurred in the bioactive anaerobic column with no
chlorinated products remaining (< 0.5 ug/1). A novel bacterium, designated "PER-K23,"
was enriched from the anaerobic column that catalyzed the dechlorination of PCE via
TCE to cis-l,2-DCE and coupled this reductive dechlorination to growth (Holliger, 1992).
Ha/COg or formate were the only electron donors that supported growth with PCE or TCE
as electron acceptors. PER-K23 did not grow in the absence of PCE or TCE. PCE or TCE
could not be replaced with oxygen, nitrate, nitrite, sulfate, sulfite, thiosulfate, sulfur,
fumarate, or carbon dioxide as electron acceptor with H2 as electron donor. All electrons
derived from HZ or formate consumption could be recovered in dechlorination products
and biomass formed. This dependence on a chlorinated hydrocarbon as an electron
acceptor is an important step in reducing the chemical requirements (electron donor)
and increasing the reaction rate of anaerobic bacterial reductive dehalogenation. The
isolation of this novel organism is an important initial step in the development of stable
cultures for converting chlorinated solvents to harmless products.
Anaerobic enrichment cultures obtained from wastewater digested sludge which
support methanogenesis were capable of completely dechlorinating PCE and TCE via
1,2-DCEs and VC to ethene without significant conversion to CO 2 or CHj (Freedman and
Gossett, 1989). The rate-limiting step in the transformation sequence appeared to be
conversion of VC to ethene. It was necessary to supply an electron donor, such as
methanol, hydrogen, formate, acetate, or glucose, to sustain reductive dechlorination of
PCE and TCE. Additional studies with these enrichment cultures yielded an anaerobic
microbial system capable of dechlorinating PCE as high as 550 uM to 80% ethene and
20% VC within 2 days at 35°C (DiStefano et al., 1991). Methanol was required for this
conversion of PCE to ethene at a concentration level approximately twice that needed for
complete dechlorination of PCE to ethene. When the incubation was allowed to proceed
for as long as 4 days, virtually complete conversion of PCE to ethene resulted, with <1% of
the initial 550 uM PCE (250 ug/1) persisting as VC. These findings are encouraging for
8-7
-------
anaerobic bioremediation of PCE-contaminated sites because a high initial volumetric
PCE dechlorination rate was observed at 35°C (275 uM/day) and a relatively large fraction
(about one-third) of the supplied electron donor was used for dechlorination. Whether
such rates and stoichiometry are possible at lower temperatures typical for the
subsurface remain unknown.
8.4. APPROACHES FOR TREATMENT
The previous section has indicated that under certain environmental conditions
with alternate electron acceptors, CT is mineralized to carbon dioxide and PCE is
completely dechlorinated to ethene and ethane. These favorable reactions could be
exploited in subsurface bioremediation by involving the following steps: (1)
characterization of site hydrogeology and contamination, (2) removal of any separate
immiscible phase, (3) assessment of biotransformation, (4) system design and operation,
and (5) monitoring of system performance (Thomas and Ward, 1989). These steps are
developed fully in earlier sections within this volume; only an overview is presented here.
Information regarding site geology and hydrology must be defined in order to properly
determine the eventual location of the treatment system. Geological considerations
should include stratigraphic effects such as horizontal extent of the aquifer and
heterogeneity of the soil. Hydrogeological data include porosity, permeability, and
ground-water velocity, direction, and recharge/discharge (Freeze and Cherry, 1979). In
addition, hydraulic connection between aquifers, potential recharge/discharge areas,
and water table fluctuations must be considered. It is important to initially identify the
contaminants present and their concentrations, since the microbial systems capable of
biotransformation and rates are compound specific.
Bioremediation with alternate electron acceptors will involve the stimulation of
microbial growth by perfusion of electron donor, electron acceptor, and nutrients
through the formation. The process is most attractive when indigenous bacteria are
used, as this avoids the significant problem of injecting and distributing a population of
bacteria acclimated to the contaminants. The frequent occurrence of reductive
dehalogenation reactions in anaerobic ground waters suggests that microorganisms
involved are frequently present in the subsurface. Formations with hydraulic
conductivities of 1(M cm/sec or greater are most amenable to bioremediation. Other
factors that are considered favorable for applying biotransformation in a cleanup
operation are listed in Table 8.3. For comparison, unfavorable conditions for in-situ
bioremediation are also included in Table 8.3.
Feasibility studies for the biotransformation of the contaminants are usually
conducted in the laboratory using subsurface material collected from the site prior to the
design and operation of a full-scale system. These experiments are conducted to
establish the presence of microorganisms capable of biotransforming the organic
contaminant(s), their nutrient and electron acceptor requirements, and the range of
contaminant concentrations that are not completely inhibitory to the microorganisms.
Sorption studies should be conducted to determine the extent and rates of partitioning of
contaminants onto the aquifer solids. The greater the degree to which the contaminants
are sorbed, the more difficult it is for the contaminant to come into contact with
microorganisms, and the longer the time for cleanup.
-------
TABLE &3. FAVORABLE AND UNFAVORABLE CHEMICAL AND HYDROGEOLOGICAL Srre CONDITIONS
FOR IMPLEMENTATION OF IN-SITU BIOREMEDIATION"
FAVORABLE FACTORS
CHEMICAL CHARACTERISTICS
small number of organic contaminants
nontoxic concentrations
diverse microbial populations
suitable electron acceptor condition
pH6to8
HYDROGEOLOGICAL CHARACTERISTICS
granular porous media
high permeability (K > 10* cm/sec)
uniform mineralogy
homogeneous media
saturated media
UNFAVORABLE FACTORS
CHEMICAL CHARACTERISTICS
numerous contaminants
complex mixture of inorganic and organic compounds
toxic concentrations
sparse microbial activity
absence of appropriate electron acceptors
pH extremes
HYDROGEOLOGICAL CHARACTERISTICS
fractured rock
low permeability (K < 1(H cm/sec)
complex mineralogy
heterogeneous media
unsaturated-saturated conditions
aWagner et al., 1986
The use of alternate electron acceptors for control of chlorinated solvents is likely
to be accomplished in situ. The advantage of in-situ treatment is that the contaminated
water or aquifer material do not have to be pumped or transported. It might be possible to
establish anaerobic conditions for the treatment of soil and ground water near the land
surface by using infiltration galleys that allow substrates and nutrient laden water to
percolate through the soil. When contamination is located at greater depths, the
approach to in-situ anaerobic bioremediation is accomplished by infiltrating or injecting
electron donor, electron acceptor, and nutrients into the contaminated subsurface to
stimulate anaerobic microorganisms in the contaminant plume. Alternatively, the
necessary growth factors could be injected downgradient from the contaminant plume to
establish an anaerobic biological treatment zone. Here, biotransformation of the
8-9
-------
chlorinated solvent occurs as the plume percolates through the zone of microbial activity.
Chemotactic bacteria move in response to chemical agents. When present, chemotactic
bacteria may slowly move upgradient in the direction of increasing organic contaminant
concentrations. This will expand the zone of biotransformation and allow contact with
sorbed and immiscible compounds. A dynamic system that includes injection and
extraction wells and equipment for the addition and mixing of nutrients could be used to
better control flow and movement of electron donor, electron acceptor, and nutrients and
contaminants. The objective is to stimulate anaerobic microorganisms to transform a
portion of the desorbed chlorinated compound with each pass of water laden with
growth-supporting chemicals. Since the microorganisms colonize the soil surfaces,
chlorinated compounds could be biotransformed as they desorb from the aquifer solids.
Biotransformation reduces the solution concentration, thus enhancing the rate of
desorption or dissolution of an immiscible phase. Periodic sampling of the soil and
ground water is essential for determining the progress of the bioremediation.
&5. FIELD EXPERIENCE
Bioremediation of chlorinated solvents using alternate electron acceptors is a
developing treatment technology that is mostly being investigated at the laboratory scale.
Limited field experience exists on stimulation of anaerobic biotransformation for control
of chlorinated solvents. One field study demonstrating this technology was conducted at
the Moffett Field Naval Air Station, Mountain View, California (Semprini et al., 1991).
This site was used earlier to study in-situ restoration of chlorinated aliphatics by
methanotrophic bacteria (Roberts etal., 1990). Reducing conditions were promoted in the
field in a 2-m test zone by stimulating a consortium of denitrifying bacteria, and perhaps
sulfate-reducing bacteria, through the addition of acetate as primary substrate (25 mg/1).
The aquifer contained both nitrate (25 mg/1) and sulfate (700 mg/1). CT was continuously
injected at a concentration of 40 ug/1, and between 95 and 97 percent CT
biotransformation was observed in the 2-m test zone with stimulated anaerobic growth.
CF was an intermediate product and represented 30 to 60 percent of the CT transformed.
Other halogenated aliphatics were biotransformed, but at slower rates and lower extents
of removal. Removals achieved for Freon-11, Freon-113, and 1,1,1-TCA ranged between
65 to 75 percent, ID to 30 percent, and 11 to 19 percent, respectively.
A second field demonstration was conducted at a chemical transfer facility in
North Toronto (Major et al., 1991). The aquifer at this site was contaminated with
organic solvents (methanol, methyl ethyl ketone, vinyl and ethyl acetate, and butyl
acrylate) and PCE. Samples of the aquifer material were amended with PCE plus
acetate/methanol. Over a 145-day incubation period in the laboratory, PCE was
dechlorinated to TCE, then cts-l,2-DCE, VC, and in many instances, to ethene. From
these results the investigators hypothesized that the presence of methanol in the
contaminated site serves as primary substrate for complete dechlorination of PCE by
anaerobic microorganisms. In-situ anaerobic bioremediation appears to be occurring at
the site without addition of chemicals.
8.6. SEQUENTIAL ANAEROBIC/AEROBIC TRANSFORMATIONS OF
CHLORINATED SOLVENTS
The combination of an anaerobic process followed by an aerobic process has
promise for bioremediation of highly chlorinated organic contaminants to innocuous
products. Generally, anaerobic microorganisms can reduce the number of chlorines on
a chlorinated compound via reductive dechlorination as described in a previous section.
8-10
-------
Chlorinated compounds are relatively oxidized by the presence of chlorine substituents.
The susceptibility to reduction reactions increases as the number of chlorine substituents
increases. Conversely, as the number of chlorine substituents decreases on a given
organic compound, reductive dechlorination reactions become less rapid and are less
likely to occur. Therefore, it is difficult to achieve complete loss of the chlorine
substituents by reductive dechlorination under anaerobic conditions. Mono- and
dichlorinated compounds tend to accumulate from the transformation of polychlorinated
organic compounds under reducing microbial conditions. Only in specialized cases has
complete dechlorination been observed via reductive microbial processes.
However, aerobic microorganisms are capable of transforming some chlorinated
compounds, especially those with fewer chlorine substituents. This oxidative process
often results in complete mineralization to carbon dioxide and is mediated by three
general mechanisms: incorporation of oxygen in the carbon-hydrogen bond, oxidation of
a halogen substituent, and oxidation of a carbon-carbon double bond via epoxidation.
With fewer chlorine substituents, the more reduced the compound, and the more
susceptible it is to oxidation. With removal of chlorines, oxidation becomes more
favorable than does reductive dechlorination. Therefore, the combination of anaerobic
and aerobic processes has potential utility as a control technology for chlorinated solvent
contamination. The approach is to first stimulate anaerobic bacteria followed by
creating oxic conditions for methanotrophs. In such a sequence the products of an
incomplete anaerobic dechlorination could be oxidized by cometabolic reactions involving
methanotrophs.
Anaerobic dechlorination and aerobic biodegradation have not been shown to
occur in sequence within the same natural system. However, this combination of
anaerobic and aerobic reactions for treatment has been tested in the laboratory. PCE and
TCE were transformed to DCE under methanogenic conditions in a 23-liter laboratory
aquifer simulator containing contaminated soil and ground water (Dooley-Danna et al.,
1989). A recirculation flow of glucose and nutrients was used to maintain methanogenic
conditions. Oxygen was then introduced and the oxidation of DCE by methanotrophic
bacteria was initiated. The sequential anaerobic/aerobic manipulations resulted in
complete biotransformation of the PCE and TCE. Hexachlorobenzene, PCE, and CT were
dechlorinated to at least the dichlorinated products in a methanogenic biofilm column
reactor fed acetate as the primary substrate (Vogel et al., 1989). All of the reductive
dechlorination products in the effluent of the methanogenic biofilm reactor were fed to an
aerobic biofilm reactor seeded with settled sewage. The mono- and dichlorinated
compounds were effectively utilized by the aerobic biofilm. Although sequential
anaerobic/aerobic treatment is a promising alternative to overcome the possible
accumulation of partially dechlorinated intermediates under anaerobic conditions,
alternating reducing and oxidizing conditions will be difficult to achieve in the field.
8.7. PERFORMANCE
8.7.1. Physical/Chemical Properties
An important factor in the success of subsurface biotransformation is the
availability of the contaminant for microbial reactions. Important physical chemical
properties influencing contaminant availability include density, water solubility, Henry's
constant (H), and n-octanol/water partition coefficient (Kow). Such physical chemical
data are summarized in Table 8.4 for some chlorinated solvents commonly encountered
in ground-water contamination problems.
8-11
-------
TABLE 8.4. PHYSICAL CHEMICAL PROPERTIES OF CHLORINATED SOLVENTS (COMMON TO
GROUND-WATER CONTAMINATION8
Compound
trichloroethylene
tetrachloroethylene
chloroform
1 , 1-dichloroethylene
1,1,1-trichloroethane
vinyl chloride
carbon tetrachloride
Density,
glml
1.4
1.63
1.49
1.013
1.435
gas
1.59
Solubility,
mgll
1,100
200
8,200
250
480
1,100
800
Henry's constant,
atm
550
1,100
170
1,400
860
35,500
1,200
logioKow
2.29
2.88
1.95
0.73
2.49
0.60
2.64
8 Data obtained from Verschueren (1983), U.S. EPA (1985), and Lyman etal. (1982).
The n-octanol/water partition coefficient (Kow), which characterizes the
hydrophobia nature of the compound, indicates the tendency for the compound to
partition (sorb) into soil organic matter. Compounds with low solubility and high K^
tend to sorb strongly to aquifer solids, which retard their movement and decrease their
availability for biotransformation. Conversely, contaminants with high water solubility
and low KOW are quite mobile and can be transported great distances with ground-water
flow. Chlorinated solvents fall into the latter class of compounds. Typical values of
logioKow for chlorinated solvents range between 0.6 and 3.0 (Table 8.4). Chlorinated
solvents migrate at rates 10% to nearly 100% of the velocity of ground water (Mackay et
al., 1985). On a relative basis, chlorinated solvents sorb less strongly than aromatic
hydrocarbons common to petroleum mixtures. This is a favorable property for
bioremediation. Also, the aqueous solubilities of chlorinated solvents are high, making
them readily available as substrates for microorganisms. However, chlorinated solvent
sorption is significant enough that effects of desorption coupled with geologic complexity
often make extraction problematic in pump-and-treat remediation (Mackay and Cherry,
1989). Furthermore, the high aqueous solubility can lead to inhibitory concentrations in
the vicinity of a spill.
The Henry's constants for chlorinated solvents are high (>100 atm), making
volatilization an important loss process in open systems. Volatilization can occur in the
vadose zone or during soil excavation but is not significant under saturated flow
conditions necessary to achieve anaerobic conditions and utilize alternate electron
acceptors.
8.7.2. Concentration Range
Chlorinated solvents are inhibitory to anaerobic microorganisms, which restricts
the range of concentrations appropriate for treatment. Belay and Daniels (1987) reported
that nearly complete inhibition of pure methanogenic cultures occurred for DCA, DCE,
8-12
-------
and TCE at exposure concentrations in the range of 50 to 150 mg/1. Partial inhibition (20 -
50%) was observed for exposure concentrations in the range of 10 to 50 mg/1. CT toxicity
studies with methanogens conducted by Yang and Speece (1986) showed similar
findings. Inhibition of unacclimated cultures was noted at 0.5 mg/1; but with
acclimation, 15 mg/1 could be tolerated. The dechlorinating culture reported by DiStefano
et al. (1991) functioned effectively with PCE at 550 uM (91 mg/1). Consequently, the
maximum allowable concentration for treatment depends on the specific chlorinated
compound and appears to range between 10 and 100 mg/1. Many of the chlorinated
solvent biotransformation studies described in Tables 8.1 and 8.2 were performed with
initial concentrations less than 1000 ug/1. Inhibitory effects were not observed at these
low concentration levels typical of many contaminated ground waters.
Nearly complete removal of the parent chlorinated compound can occur during
anaerobic biotransformation. For many of the studies described in Tables 8.1 and 8.2,
residual concentrations of the starting chlorinated compound were in the low ug/1 or
even <1 ug/1 in some cases. However, the repeatedly observed incomplete reductive
dechlorination results in accumulation of lesser chlorinated compounds and ineffective
treatment. The concentrations of chlorinated intermediates and products remaining
after anaerobic biotransformation of chlorinated solvents, and consequently the degree of
success, are critically linked to how complete the reductive dechlorination reactions have
taken place. Concentrations below typical health based standards of 5 to 10 ug/1 were not
achieved in the two field trials presented on page 8-10. However, the anaerobic microbial
systems showing conversion of PCE to ethene (DiStefano etal., 1991; de Bruin etal., 1992)
and CT to CC>2 (Criddle et al., 1990b) are especially encouraging for development of this
treatment technology. Residual concentrations below 5 ug/1 were observed in these
laboratory microcosms. Consequently, it appears that optimized systems for anaerobic
biotransformation can meet relevant regulatory endpoints.
8.7.3. Favorable RedoK Conditions
There are several possible electron acceptors for anaerobic biotransformations. In
many subsurface systems, the redox state is governed by microbial activity. Energetic
considerations can be used to obtain insight into which chlorinated compounds are most
susceptible to reductive dechlorination in the presence of different electron acceptors
being used by the microorganisms. For assessing whether a given compound may in
principle undergo a redox reaction like reductive dechlorination, one needs to know the
(standard) reduction potentials of the half-reactions involving the compound of interest
and its oxidized or reduced transformation product, as well as the reduction potentials of
the natural oxidant(s) or reductant(s) present in a given system. The (standard)
reduction potentials at pH 7 of some important microbial electron acceptors are given in
Table 8.5 together with the reduction potentials of some half-reactions involving
chlorinated solvents. The half-reactions are ordered in decreasing redox potential values
expressed in volts. A favorable thermodynamic redox reaction is obtained by coupling
any given reduced species as electron donor with an electron acceptor of higher redox
potential (listed above the electron donor half-reaction listed in Table 8.5). For example,
acetate as electron donor can be coupled with all of the other half-reactions listed above it
to yield a thermodynamically favorable reaction. The conversion of CT to CF, PCE to
TCE, 1,1,1-TCA to 1,1-DCA, CF to methylene chloride, TCE to trans-1,2-DCE, and
Irons-1,2-DCE to VC appear to be energetically favorable under sulfate-reducing (sulfate
as electron acceptor) and methanogenic (CO2 as electron acceptor) conditions. Reduction
of ethylene dibromide to ethene and hexachloroethane to PCE even appears
thermodynamically possible under aerobic respiration and denitrification. Several of the
reductive dechlorination reactions appear to be thermodynamically possible with Fe(III)
and Mn(IV) as electron acceptors.
8-13
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TABLE 85. STANDARD REDUCTION POTENTIALS AT 25*C AND pH 7 FOR SOME REDOX
COUPLES THAT ARE IMPORTANT ELECTRON ACCEPTORS IN MlCROBIAL
RESPDUTION AND FOR SOME HALF-REACTIONS INVOLVING CHLORINATED
SOLVENTS
HALF REACTION"
Oxidized species
C13C-CC13 -i- 2e-
Q, + 4H+ + 4e
2NO3 +12H++10e
CCl4+H+ + 2e-
Cl2C=CCl2 + H* + 2e-
Cl3C-CH3+H++2e-
CHC13 + H* + 2e-
HClC=CCl2 + H* + 2e-
MnO2(s) + HCO3 + 3H* + 2e'
t-HClC=C!H + H+ + 2e-
FeOOH(s) + HCO3 +2H++e
SD4a+9H+ + 8e-
CO2+gH* + 8e'
2C02+8H+ + 8e-
2H+ * 2e-
Reduced species
= C12C=CC12 + 2C1-
=2H2O
= N2 + ^O
= CHCl3-i-C]-
= HC1C=CC12 + C1-
= HC12C-CH3+C1-
= CH2C12 + C1-
= t-HC!C=ClH + Cl-
= MnCO3(s)+2H2O
= H2C=CHC1+C1-
= FeCO3(s) + 2H2O
=HS +4H2O
= CH4 + 2H2O
= CH3COOH + 2H2O
= H2
E'
(volts)b
+1.13
•K).82
+0.74
+0.67
+0.58
+0.57
+0.56
+0.54
+0.52
+0.37
-0.05
•O22
•O24
-0.40
-0.41
a Half-reactions in bold are common electron acceptors in microbial respiration.
bData from Stumm and Morgan (1981) and Thauer et al. (1977). Values are for aqueous
solution with pH = 7, [HCO3'] = 0.001 M, and [CM = 0.001 M.
8-14
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8.8. BIOTRANSFORMATION STOICHIOMETRY
In order to maintain reducing conditions for anaerobic microbial reactions that
may be employed for bioremediation of chlorinated solvents, an electron donor must be
available along with the appropriate alternate electron acceptors). Furthermore, the
presence of a suitable electron donor is often necessary to prevent accumulation of
chlorinated intermediates. These chemicals along with other major growth nutrients
are often not among the chemical constituents available in the contaminated region and
growth is limited without them. In order to stimulate anaerobic microbial growth and
engineer a microbial treatment system for organic contaminant control, the chemical
needs of the microorganisms must be defined and are given by the reaction
stoichiometry.
Studies with methanogenic bacteria that biotransform chlorinated aliphatic
compounds such as PCE, CF, and 1,1,1-TCA, using acetate as primary electron donor
and carbon source, indicated the ratio of acetate mass used to the mass of chlorinated
compounds transformed varied between 100/1 and 1000/1 (Bouwer and McCarty, 1983a).
Recently, de Bruin et al. (1992) determined that 1 mM lactate was used for the complete
dechlorination of 9 uM PCE to ethene. This amount of lactate is 150 times the minimum
reducing equivalents necessary for a complete reduction of PCE to ethane. Consequently,
large quantities of an electron donor like acetate or lactate may need to be injected into the
contaminated soil system for anaerobic treatment of even a relatively small amount of
chlorinated solvent contamination. However, de Bruin et al. (1992) did not attempt to
optimize the lactate/PCE stoichiometry. The amount of lactate needed is likely to be
markedly less (W. de Bruin, personal communication, 1992).
The appropriate amounts of electron donor, electron acceptor, and nutrients that
must be supplied for growth of the anaerobic bacteria and the amounts of biomass and
other products that will be formed can be estimated using the thermodynamic model
reported by McCarty (1971). In this model, electrons from the electron donor can be
coupled with the electron acceptor to generate energy or can be used to synthesize
biomass. The relative amounts of the electron donor being oxidized for energy and being
converted to biomass is established with an energy balance. The amount of energy
released during oxidation of the electron donor must balance the amount of energy
required to synthesize the cell material.
The balanced equations for the methanogenic system capable of complete
reductive dechlorination of PCE to ethene as described by de Bruin et al. (1992) are given
in Table 8.6. Such stoichiometric relationships established with the model can be used to
determine the appropriate solution of lactate (electron donor) and nutrients to flush
throughout the zone of contamination. The application of this stoichiometry for PCE soil
contamination appears in Figure 8.2. When organic contaminants enter the subsurface,
the residual amount retained on the soil after drainage has been found to range between
0.5% to 4%by volume depending on the soil type (Wilson, J. et al., 1986). Medium to fine
sand typically retains Ito 2% by volume of the organic contaminant. For this sand with
bulk density of 2000 kg/m3, about 16.3 kg of residual PCE (10 liters) will remain per m3 of
soil after free product recovery by pumping, and normal drainage leaves a residual of 1%
by volume (Figure 8.2). According to the first set of reactions in Table 8.6, 970 kg of lactate
and 17.4 kg of ammonia nitrogen would be required per m3 of soil in order for the
reductive dechlorination of PCE to ethene to proceed properly and be in balance. Other
nutrients that are required in lesser quantities for bacterial growth are not included in
the balanced equations. However, the phosphorus requirement is about one-sixth of that
for nitrogen. Thus, this PCE biotransformation would require 2.9 kg of phosphorus per
m3 of soil. Bacterial biomass is represented by the empirical formula CgH-jOjN and 140
8-15
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kg of cells would be formed per m3 of soil during these anaerobic reactions. In this
example, most of the reducing equivalents supplied as lactate evolve as methane (221 kg
or 317 m3 of gas). The amounts of the innocuous products ethene and HC1 formed from
PCE are shown in Figure 8.2.
Chemical Inputs
970 kg lactate
17.4 kg ammonia-N
2.9 kg P
Products of Anaerobic
Bioremediation
1 m
140kg biomass
221 kg methane'
2.75 kg ethene -
14.3 kg HC1
317m3 gas
2.4 m3 gas
Residual PCE
16.3 kg or 10 liters
Estimated Time - 3 years
Figure 8.2. Chemical requirements and products of anaerobic bioremediation for one cubic
meter soil contaminated with PCE using microbial system reported by de Bruin et
al. (1992).
As a result of the lactate addition for PCE biotransformation in the above example,
140 kg of anaerobic biomass would be formed. The microbial growth forms a larger mass
than that of the original PCE retained by the soil. This biomass growth is likely to reduce
the soil permeability and interfere with water flow during injection of chemicals. About
80% of the biomass will be biodegradable, and if additional oxygen or nitrate is supplied,
eventually the biomass will decay to about 20% of the amounts given in Figure 8.2.
The mass balance relationships given above for lactate in a potential anaerobic
bioremediation illustrate the enormous quantities of chemicals required and point to the
urgent need to develop microbial systems with tighter coupling between the reducing
equivalents from the electron donor and the reductive dechlorination reactions of the
chlorinated solvent. A great improvement in the stoichiometry was recently reported by
DiStefano et al. (1991) in a methanogenic system using methanol as primary electron
donor. Nearly one-third of the methanol consumption was used for dechlorination of
PCE to ethene. The resulting stoichiometry appears in Table 8.6. The corresponding
chemical inputs and products per m3 of soil for the hypothetical PCE contamination
introduced above appear in Figure 8.3. The amounts of methanol (13.6 kg) and nutrients
(0.077 kg N and 0.013 kg P) necessary appear reasonable for field-scale applications. The
amount of biomass formed (0.62 kg) per m3 of soil is not likely to cause problems with
clogging near the injection well or permeability reduction in the formation.
8-16
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Chemical Inputs
I3.6kg melhannl
0.077 kg ammoma-N
0.013 kg P
Products of Anaerobic
Bioremediation
0 62 kg bumiass
3.3 kg methane = 5.0 m3 gas
2.75kgelhene = 2.4m3 gas
l43kgHCl
Residual PCE
16 3 kg or 10 liters
Estimated Time = I year
Figure &3. Chemical requirements and products of anaerobic bioremediation for one cubic
meter soil contaminated with PCE using microbial system reported by DiStefano
etal. (1991).
TABLE 8.6. STOICHIOMETRIC RELATIONSHIPS FOR POSSIBLE BIOREMEDIATION REACTIONS
INVOLVING COMPLETEDECHLORINATION OF PCE TOETHENE"
Stoichiometrv for the Microbial System Reorted bv de Bruin et al. ( 1992)
PCE(C2C14) + 0.671actate(C3H5O3-) + 2.33 H2O
4HC1 + 1.33 CO2 + 0.67 HCO3'
HOlactate + 12.6 NH4+ + 60rLjO _ ^.
12.6 biomass (C^OgN) + 134 CH4 + 97.9 HCCy + 36.3 CO2
Stoichiometrv for the Microbial System Reported bv DiStefano et al. (1991)^
PCE (C2C14) + 1.33 methanol (CH3OH) + 1.33 H2O — *-
+ 4HC1 + 1.33 CO
3.0 methanol -f- 2.25 HC03"
2.25 acetate (CHgCCO') + 0.75 CO2 + 3.75 H2O
2.25 acetate + 0.056 NH4+ + 202H2O + 0.083 CO2
0.056 biomass + 2.11 CH4 + 2.19HC03'
8 All compounds were considered in aqueous phase except C02, CH4, and ethene were taken as gaseous Free
energy of formation values for the organic compounds were obtained from Handbook of Organic Chemistry
(Dean, 1987)
b The acetate produced from acetogenesis of methanol was assumed to undergo methane fermentation
8-17
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Another favorable biotransformation with an alternate electron acceptor is the
conversion of CT to carbon dioxide under denitrification conditions as described in Table
8.1 (Griddle et al., 1990a). However, the relationship between acetate and nitrate
consumption and CT biotransformation was not determined, thus further work is needed
to clarify the stoichiometry.
8,9. BIOTRANSFORMATION RATES
Knowledge of biotransformation rates is useful in a bioremediation process for
determining the length of time required for meeting a treatment objective.
Biotransformation half-lives observed for reductive dechlorination of chlorinated solvents
in both environmental samples and the field range from weeks to months. For example,
CF and other trihalomethanes were removed from the Palo Alto, California, Baylands
Aquifer during injection of reclaimed municipal wastewater with half-lives of 3 to 6
weeks (Roberts et al., 1982). A much slower decline occurred in the concentrations of
chlorinated ethanes and ethenes, yielding half-lives of 5 to 9 months. Similar anaerobic
biotransformation rates for chlorinated solvents have been observed in sediment, aquifer
microcosms, and in laboratory-scale batch and continuous-flow systems. Consequently,
reductive dechlorination rates by indigenous microorganisms appear to be quite slow.
One of the objectives in a bioremediation scheme is to increase the numbers of
desired microorganisms. Bouwer and Wright (1988) illustrate that elevation of the
biomass concentration by one to two orders of magnitude by the addition of growth
substrates and nutrients can correspondingly decrease the half-lives to between days and
weeks. The lactate addition in the methanogenic system of de Bruin et al. (1992) yielded a
high dechlorination rate of PCE to ethene (89 umol/1/day). At this volumetric
dechlorination rate, the anaerobic bioremediation of PCE illustrated in Figure 8.2 would
require about 1,100 days or 3 years to complete. An even higher initial volumetric rate of
PCE dechlorination (275 umol/1/day) was obtained at 35°C with methanol as electron
donor in the anaerobic system of DiStefano et al., (1991). At this volumetric
dechlorination rate, the anaerobic bioremediation of PCE would be shortened to about 350
days or 1 year (illustrated in Figure 8.3).
In most studies of anaerobic biotransformation of chlorinated solvents, only 0.005
to 1.6% of the electrons available from the primary substrate (electron donor) were used
for dechlorination. The transformations are believed to occur by cometabolism, and these
cometabolic reactions are slow and inefficient due to the uncoupling of contaminant
transformation and microbial growth. In cometabolism, enzymes produced by the
microorganism to metabolize the primary substrate can interact with an organic
contaminant and bring about its transformation in a fortuitous manner. The
cometabolite does not provide energy for growth or maintenance, so the microorganism
does not benefit from cometabolic transformations. Limited evidence exists that a
halogenated compound can serve as sole energy and carbon source for anaerobic
bacteria. Recently a homoacetogen has been isolated which used methyl chloride for
growth (Traunecker et al., 1991). Another example is the novel bacterium PER-K23
already described (p. 8-7) which carries out respiration with Hg as electron acceptor and
PCE as electron donor (Holliger, 1992). The PCE reductive dechlorination rate for
PER-K23 is several orders of magnitude higher than reaction rates observed for
cometabolic reductive dechlorinations (Table 8.7). The coupling between reductive
dechlorination and respiration is encouraging for developing anaerobic microbial
systems with faster dechlorination rates that could be exploited for the bioremediation of
sites contaminated with chlorinated solvents.
8-18
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TABLE 8.7. PCE DECHLORINATION RATES BY DIFFERENT ANAEROBIC BACTERIA
Organism Reaction Product Dechlormation Refer-
Rate ence"
(timolldaylmg
protein)
PER-K23 respiration cis-l,2-DCE 475 a
Acetobacterium woodu cometabolism TCE 0.086 b
Methanosarctna sp. cometabolism TCE 0.00084 c
Methanosarcina mazei cometabolism TCE 0.00048 c
Desulfomonde tiedjei cometabolism TCE 0.0023 c
"a = Holliger (1992); b = Egli et al. (1988); c = Fathepure et al (1987)
Limited information is available on rates of CT biotransformation with nitrate as
electron acceptor. CT was nearly completely biotransformed in batch microcosms after 3
weeks of incubation under denitrification conditions (Bouwer and McCarty, 1983b). CT
was assimilated into cell mass, mineralized to CO2, and reductively dechlorinated to CF,
indicating simultaneous oxidative and reductive reactions. CT disappeared in a 10-day
period with concomitant CF production in a denitrifying enrichment sample from
Moffett Field, California (Criddle et al., 1990a). A field test of anaerobic bioremediation
demonstrated that CT was transformed at rapid rates with half-lives on the order of
hours to days through the addition of acetate in the presence of nitrate and sulfate
(Semprini et al., 1991). CF was observed as an intermediate of the CT biotransformation.
Although the rates of CT biotransformation reported in these studies are reasonably fast,
the formation of CF as an intermediate product is objectionable. The denitrifying
Pseudomonas sp. described on p. 8-7 was capable of complete biotransformation of CT
without production of CF in 2 days. Further work is needed to evaluate if this reaction
can be deployed in anaerobic bioremediation.
8.10. LIMITATIONS
The formation of chlorinated intermediates, biotransformation stoichiometry and
rates, influence of mass transfer, and water quality changes under anaerobic conditions
are major concerns and possible impediments to practical implementation of anaerobic
bioremediation technology.
8-19
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Incomplete reductive dechlorination of chlorinated solvents is often encountered
under anaerobic conditions, which results in the formation and accumulation of lesser
chlorinated aliphatic compounds. This reaction mechanism is probably the major
obstacle to widespread deployment of anaerobic processes for in-situ bioremediation. The
chlorinated compounds formed are objectionable and pose a threat to public health. An
electron donor compound, such as lactate, methanol, or H,, must be supplied to
stimulate growth of the anaerobic microorganisms involved in the reductive
dechlorination reactions. For this cometabolism, often the mass of primary substrate
(electron donor) to mass of chlorinated solvent biotransformed ranges between 100/1 and
1000/1. Consequently, a large quantity of electron donor is needed along with additional
nutrients like nitrogen and phosphorus. Proper delivery of such large masses of
chemicals is an engineering challenge. The high levels of chemicals needed are
converted to large amounts of end products, such as methane gas, carbon dioxide, and
biomass. How to control these by-products, particularly in heterogeneous environments,
is an important question yet to be resolved. The biomass growth is likely to fill up the pore
space, causing marked permeability reduction in the formation. This plugging of the
formation will in turn interfere with proper delivery of the chemicals required. Slow
reaction rate is a final concern for the anaerobic microbial systems involved in reductive
dechlorination. Half-lives for anaerobic reductive dechlorination are typically on the
order of months in the field. Extrapolation of optimal rates presently observed in the
laboratory suggests cleanup times of years in the field. The slow rates are likely to be
problematic with most regulatory timetables.
The influence of mass transfer on the availability of chlorinated solvents for in-
situ bioremediation is a potential drawback. Although the chlorinated solvents tend to
weakly sorb to aquifer solids, as indicated by their small K. values, slow desorption can
be the rate limiting step and control the bioavailability of the compounds. The practical
effect of slow diffusion from within soil aggregates and other kinetic limitations to
desorption is to decrease the rate of removal of the chlorinated solvents from the aquifer,
thereby increasing the time required to achieve cleanup and the amount of chemicals
that must be added to sustain anaerobic microbial activity. The ability to deliver electron
donors, electron acceptors, and nutrients to the microorganisms is a second mass
transfer problem. The effects of geologic complexity, such as strata of gravel, sand, silt
and clay, and fractured layers, along with the difficulty of locating the sources of
subsurface contamination, can severely hamper the supply of chemicals throughout the
zone of contamination.
The intended use of the aquifer after treatment could create a problem with
stimulating anaerobic processes. If the aquifer is to be a source of drinking water, then a
number of negative water quality issues arise due to making the aquifer anaerobic. As
conditions switch from oxic to anoxic, some metals will be solubilized, particularly iron
and manganese. These metals cause taste and odors and stain materials that come in
contact with the water (e.g. pipes, bathtubs, toilet bowls, sinks, and clothes). Metabolites
excreted by the anaerobic biomass increase the organic matter content of the water.
Disinfectants added to the water to control pathogens react with this organic matter to
form disinfection by-products. Regulations under the Safe Drinking Water Act of 1986
are currently aimed at limiting the formation of such disinfection by-products. The
addition of nitrate to aquifers to stimulate denitrifying conditions may be of concern
because nitrate is regulated under the National Drinking Water Standards with a
maximum contaminant level of 10.0 mg/1 measured as nitrogen. Production of
microbial metabolites can also solubilize cadmium, copper, lead, and zinc oxides
(Francis and Dodge, 1988) which may facilitate their passage into the distribution
system.
8-20
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8.11. RESEARCHNEEDS
The positive attributes about bioremediation of chlorinated solvents under
anaerobic conditions are that the compounds can be rendered harmless if reductive
dechlorination is complete and costs are likely to be lower than other technologies.
Additional research is needed to overcome the many limitations detailed above.
Research needs specific for biotransformation of chlorinated solvents with alternate
electron acceptors appear below:
1. There is need to determine the environmental factors and metabolic
requirements necessary for either the complete reductive dechlorination of
chlorinated solvent compounds or mineralization of certain compounds like
CT to carbon dioxide. Achieving this goal requires further investigation into
the physiology and metabolism of the various anaerobic microorganisms,
focusing on both enriched mixed cultures and pure cultures. The aim of this
research is to be able to specify completely and unambiguously the optimum
conditions for growth and cometabolic transformation so that the
biotransformation can be reliably controlled in the field.
2. It is essential to develop ways to minimize chemical requirements,
particularly the need for an electron donor, and ways to increase reaction
rates for the cometabolism of chlorinated solvents under anaerobic conditions.
The kinetics of individual transformation processes must be thoroughly
understood to enable confident prediction of intermediate and product
formation and to permit design and scale-up of processes in the field.
3. The development of stable anaerobic microbial consortia that use reductive
dechlorination for respiration rather than cometabolism would be beneficial
for improving the success of anaerobic processes for in-situ bioremediation.
4 There is need to evaluate the feasibility of sequentially creating anaerobic
followed by aerobic conditions in the subsurface in order to biotransform
chlorinated solvents in a two step process. The impacts of the changing redox
conditions on ground-water quality need to be assessed.
5 It is necessary to continue research to determine factors that govern the
influence of sorption/desorption of chlorinated solvents on the performance of
their bioremediation. If desorption is limiting, then ways must be devised to
increase their availability for microbial transformation.
6. Optimal findings from the laboratory efforts mentioned above need to be
demonstrated in the field by conducting small-scale studies at
well-instrumented sites. Specific tasks to be conducted include evaluation of
methods for delivery of chemicals (electron donors, electron acceptors, and
nutrients) and their efficacy for stimulating anaerobic microorganisms,
characterization of the extent of the biological active zone, and evaluation of
contaminant residuals and times required for the microbial reactions.
8.12. CONCLUDING REMARKS
Chlorinated solvents are difficult to control in the environment; their widespread
usage, uncontrolled disposal, and chemical/physical properties make them common
8-21
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ground-water contaminants. The transformations of chlorinated solvents in the
presence of alternate electron acceptors are important reactions that can affect their fate
and can be applied to the development of treatment technology. This section addressed
some of the important factors that affect these biotransformations in the subsurface
environment and that can be applied to the development of bioremediation technology.
Many different anaerobic bacterial species are capable of catalyzing the reductive
dechlorination of chlorinated solvents. The conversion of CT to carbon dioxide under
denitrification is another important anaerobic transformation. An electron donor
(primary substrate) is required to supply energy for bacterial growth and for activation of
enzymes necessary for the transformation. The possible conversion of chlorinated
solvents to harmless products under anaerobic conditions has lead to an interest in using
in-situ techniques for biotransformation of these contaminants as an alternative to
aboveground treatment systems that generally involve physical/chemical processes that
do not destroy the contaminants. The objective of bioremediation is to stimulate the
growth of indigenous or introduced microorganisms in regions of subsurface
contamination, and thus, provide direct contact between microorganisms and the
dissolved and sorbed contaminants for biotransformation. The anaerobic process will
typically require the perfusion of an electron donor, electron acceptor, and nutrients
through the contaminated soil. However, the supply of these chemicals can be difficult in
tight and heterogeneous soils. The formation of chlorinated intermediates, the large
amounts of electron donor necessary for the cometabolic reductive dechlorination
reactions, slow rates of desorption, and negative water quality changes are additional
major concerns and possible impediments to practical implementation of anaerobic
bioremediation technology. Both basic laboratory studies and well-controlled field
experiments are needed to establish feasibility and overcome the present limitations with
exploiting anaerobic processes for bioremediation of chlorinated solvents.
ACKNOWLEDGEMENT
The author thanks Dr. Gosse Schraa, Department of Microbiology, Agricultural
University, Wageningen, The Netherlands, for fruitful discussions on this topic.
8-22
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Fathepure, B.Z., J.P. Nengu, and S.A. Boyd. 1987. Anaerobic bacteria that dechlorinate
perchloroethene. Appl. Environ. Microbiol. 53(11):2671-2674.
Fathepure, B.Z., and S.A. Boyd. 1988. Dependence of tetrachloroethylene dechlorination
on methanogenic substrate consumption by Methanosarcina sp. strain DCM.
Appl. Environ. Microbiol. 54<12):2976-2980.
Francis, A.J., and C.J. Dodge. 1988. Anaerobic microbial dissolution of transition and
heavy metal oxides. Appl. Environ. Microbiol. 54<4):1009-1014.
Freedman, D.L., and J.M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic
conditions. Appl. Environ. Microbiol. 55(9):2144-2151.
Freeze, R.A., and J.A. Cherry. 1979. Ground Water. Englewood Cliffs, New Jersey.
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Galli, R., and P.L. McCarty. 1989. Biotransformation of 1,1,1-trichloroethane,
trichloromethane, and tetrachloromethane by a Clostridium sp. Appl. Environ.
Microbiol. 55(4):837-844.
Holliger, C. 1992. Reductive Dehalogenation by Anaerobic Bacteria. Ph.D. Thesis.
Department of Microbiology, Agricultural University. Wagenmgen, The
Netherlands.
Kastner, M. 1991. Reductive dechlorination of tri- and tetrachloroethylenes depends on
transition from aerobic to anaerobic conditions. Appl. Environ. Microbiol.
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Klecka, G.M., S.J. Gonisor, and D.A. Markham. 1990. Biological transformation' of
1,1,1-trichloroethane in subsurface soils and ground water. Environ. Toxicol.
Chem. 9:1437-1451.
Kloepfer, R.D., D.M. Easley, B.B. Haas, Jr., T.G. Deihl, D.E. Jackson, and C.J.
Wurrey. 1985. Anaerobic degradation of trichloroethylene in soil. Environ. Sci.
Technol. 19(3):277-280.
Krone, U.E., K. Laufer, R.K. Thauer, and H.P.C. Hogenkamp. 1989. Coenzyme F.^ as a
possible catalyst for the reductive dehalogenation of chlorinated Cj hydrocarbon in
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Lyman, W.J., W.F. Reehl, and D.H. Rosenblatt. 1982. Handbook of Chemical Property
Estimation Methods. McGraw-Hill. New York, New York.
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in groundwater. Environ. Sci. Technol. 19(5):384-392.
Mackay, D.M., and J.A. Cherry. 1989. Groundwater contamination: Pump-and-treat
remediation. Environ. Sci. Technol. 23(6):630-636.
Major, D.W., E W. Hodgins, and B.J. Butler. 1991. Field and laboratory evidence of in
situ biotransformation of tetrachloroethene to ethene and ethane at a chemical
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Mikesell, M.D., and S.A. Boyd. 1990. Dechlorination of chloroform by Methanosarcina
strains. Appl. Environ. Microbiol. 56(4): 1198-1201.
Parsons, F., P.R. Wood, and J. DeMarco. 1984. Transformation of tetrachloroethene and
trichloroethene in microcosms and groundwater. J. Amer. Water Works Assoc.
72(2):56-59.
Parsons, F., and G.B. Lage. 1985. Chlorinated organics in simulated groundwater
environments. J. Amer. Water Works Assoc. 77(5):52-59.
Parsons, F., G.B. Lage, and R. Rice. 1985. Biotransformation of chlorinated organic
solvents in static microcosms. Environ. Toxicol. Chem. 4:739-742.
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States. University of Pennsylvania Press. Philadelphia, Pennsylvania.
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changes during ground water recharge in the Palo Alto Baylands. Water Res.
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in-situ biodegradation of chlorinated ethenes: Part 1, Methodology and field site
characterization. Ground Water. 28(4):591-604.
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Scholtz-Muramatsu, H., R. Szewzyk, U. Szewzyk, and S. Gaiser. 1990.
Tetrachloroethylene as electron acceptor for the anaerobic degradation of
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biotransformation of carbon tetrachloride, freon-113, freon-11, and 1,1,1-TCA
under anoxic conditions. In: On-Site Bioreclamation Processes for Xenobiotic and
Hydrocarbon Treatment. Eds., R.E. Hinchee and R.F. Olfenbuttel.
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SECTION 9
NATURAL BIOREMEDIATION OF HYDROCARBON-CONTAMINATED
GROUND WATER
Robert C. Borden
Civil Engineering Department
North Carolina State University
Raleigh, North Carolina 27695-7908
Telephone: (919)515-2331
Fax: (919)515-7908
9.1. GENERAL CONCEPT OF NATURAL BIOREMEDIATION
The basic concept behind "Natural Bioremediation" is to allow naturally occurring
microorganisms to degrade contaminants that have been released into the subsurface
and at the same time minimize risks to public health and the environment. Use of this
approach will require an assessment of those factors that influence the biodegradation
capacity of an aquifer and the potential human and environmental risks. Ongoing
research has shown that an aquifer's assimilative capacity depends on the metabolic
capabilities of the native microorganisms, the aquifer hydrogeology and geochemistry,
and the contaminants involved.
Natural bioremediation is not a "No Action" alternative. In most cases, natural
bioremediation is used to supplement other conventional remediation techniques. The
type and extent of conventional remediation techniques used depend on the
environmental conditions in the aquifer, the extent of contamination, and the risk to the
public and environment. In some cases, only removal of the primary source (e.g.,
leaking tanks, contaminated soil) may be necessary. In other situations, conventional
ground-water remediation by pump and treat may be used to reduce the concentrations
within the aquifer. Once contaminant concentrations are reduced below some defined
level, the pump and treat system may be terminated and natural bioremediation used to
complete the cleanup.
Implementation of a natural remediation system differs from conventional
techniques, in that a portion of the aquifer is allowed to remain contaminated.
Depending on site specific conditions, use of natural bioremediation may require a
variance from existing regulations, may involve questions of third party liability and
property rights, or require public hearings and review by elected officials. Natural
bioremediation is less predictable than conventional pump and treat or excavation.
Consequently, some type of risk evaluation will usually be required whenever natural
bioremediation is considered.
The purpose of this chapter is to discuss the potential for natural bioremediation to
be incorporated into an overall remedial design at a hazardous waste site. The various
biological processes using oxygen, nitrate, ferric iron, sulfate, and carbonate as electron
acceptors will be addressed. Considered also are the effect of environmental conditions
on biodegradation, site characterization needed for natural bioremediation, necessary
parameters to be monitored, performance and prediction of natural bioremediation, and
issues that may affect the costs associated with the technology. Well-documented case
studies of natural bioremediation at former wood preserving facilities and petroleum
releases are also presented.
9-1
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At present, there is almost no operating history to judge the effectiveness of
natural bioremediation. Early attempts at aquifer remediation focused on using
conventional remediation techniques to remove or permanently immobilize
contaminants at the highest priority sites. While at many low priority sites, regulators
may have assumed that natural bioremediation would be adequate to control the
migration of dissolved contaminants, typically these sites have not been monitored
sufficiently to determine if this approach is actually effective or to identify those factors
that influence the efficiency of natural bioremediation. At present, there are no well-
documented full scale demonstrations of natural bioremediation, although there has
been some limited research into the processes that control the natural biodegradation of
dissolved hydrocarbon plumes (Borden et al., 1986; Barker et al., 1987; Franks, 1987; Hult,
1987a; Chiang et al., 1989; Wilson et al, 1993). At present, the primary repositories of
expertise on natural bioremediation are in universities, in industry, the U.S.
Environmental Protection Agency (U.S. EPA) and the U.S. Geological Survey (U.S.G.S.).
Use of natural bioremediation is typically much less expensive than other remediation
technologies. Consequently, there has been less incentive for private consultants and
service companies to invest funds developing this technique.
9.2. HYDROCARBON DISTRIBUTION, TRANSPORT AND BIODEGRADATION IN
THE SUBSURFACE
Dissolved hydrocarbons are among the most common ground-water contaminants
and can originate from spilled fuels (gasoline, diesel, jet fuel, heating oil), solvents, wood
preservatives, and coal gasification wastes. Many of these wastes are initially present as
nonaqueous phase liquids (NAPLs), which contain many different components.
Gasoline contains primarily the lighter, lower boiling point compounds like pentane and
benzene; while creosote and coal tars contain more of the higher boiling point
compounds.
Ground water that comes in contact with the residual hydrocarbons will dissolve a
portion of the NAPL. The amount of each individual component that dissolves in water
can be roughly estimated as the aqueous solubility times the mole fraction of the
individual component in the oily phase. Ground waters that have come in contact with
petroleum fuels typically become contaminated with BTEX compounds (benzene, toluene,
ethylbenzene, and xylenes) because these compounds are the most water soluble.
Ground water in contact with gasoline will typically contain more BTEX than ground
water in contact with heating oil and other heavier hydrocarbons, because gasoline
contains a higher percentage of BTEX. These ground waters may also contain high
concentrations of fuel additives since many of these additives are highly soluble in water
and are present in relatively high concentrations in some gasolines.
Once an individual hydrocarbon constituent is dissolved, it may be transported by
moving ground water. While the movement of most petroleum constituents will be
retarded to some extent by sorption onto aquifer material, the more soluble compounds
are usually not sorbed to a large extent, except in aquifers with a high organic carbon
content. The primary mechanisms that will limit the subsurface transport of dissolved
hydrocarbons are biodegradation and, to a lesser extent, volatilization. Volatilization
results in a transfer of the lower boiling point, more volatile compounds from the ground
water to the soil gas within the unsaturated zone. At present, the significance of this
process is unknown, although it is expected that the relative importance of volatilization
will be much less for large spills where a larger portion of the plume is present at
significant depth below the water table. Nonbiological (abiotic) reactions such as
9-2
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hydrolysis are of lesser importance because many hydrocarbons are relatively stable
under the environmental conditions found in most aquifers.
9.2.1. Petroleum Hydrocarbon Biodegradation
Hydrocarbon biodegradation can be represented by the chemical reaction
Hydrocarbon + electron acceptors + microorganisms + nutrients •»
-• carbon dioxide + water + microorganisms + waste products
The rate and extent of hydrocarbon biodegradation in the subsurface will depend
on several factors, including (1) the quantity and quality of nutrients and electron
acceptors; (2) the type, number and metabolic capability of the microorganisms; and (3)
composition and amount of the hydrocarbons. While virtually all petroleum
hydrocarbons are biodegradable, the rate and extent of biodegradation can be highly
variable. Depending on environmental conditions, biodegradation may be very rapid or
very slow.
9.2.2. Subsurface Microorganisms
Recent studies have shown that an active diverse microbial population exists in
the subsurface, often at great depth. The organisms present appear to be predominantly
bacteria, but some fungi and protozoa have been identified (Ghiorse and Wilson, 1988).
The native organisms appear to be well adapted to low nutrient conditions. Many of the
organisms identified grow very poorly or not at all under high nutrient conditions, yet
thrive at low levels of organic carbon (Ghiorse and Balkwill, 1985). Biochemical analyses
have indicated the presence of storage granules, allowing survival during extended
starvation periods (White et al., 1983).
Most of the organisms identified are aerobes, but strict anaerobes have been
identified from a few sites (Ghiorse and Wilson, 1988). Microbially mediated
denitrification was observed in a sand and gravel aquifer contaminated with treated
sewage (Smith and Duff, 1988). Anaerobic bacteria were identified by van Beelen and
Fleuren-Kemila (1989) from two sandy aquifers, a saturated peat soil and a river
sediment. Chapelle et al. (1987) identified methanogenic and sulfate-reducing bacteria
from sediments collected 20 to 180 m below grade in the Maryland coastal plain. Recent
work by Jones et al. (1989) has shown that methanogens are present in the subsurface at
over 300 m below grade in sediments near Aiken, SC. Although the microbial
community was dominated by aerobic microorganisms, sulfate-reducing and
methanogenic organisms could be identified from most sediments throughout the depth
profile. In most cases, the total number of methanogens were very low, but the
organisms present were capable of degrading a wide variety of organic substrates, e.g.,
benzoate, phenol, lactate, formate, acetate.
The ability of microorganisms to degrade a wide variety of hydrocarbons is well
known. In an early review, Zobell (1946) identified over 100 microbial species from 30
genera that could degrade some type of hydrocarbon. In a more recent study, Ridgeway
et al. (1990) identified 309 gasoline-degrading bacteria from a shallow coastal aquifer
contaminated with unleaded gasoline. Hydrocarbon-degrading microorganisms are
widespread in the environment and occur in fresh and salt water, soil, and ground
water. Litchfield and Clark (1973) analyzed ground-water samples from 12 different
aquifers throughout the United States that were contaminated with hydrocarbons. These
workers found hydrocarbon- utilizing bacteria in all samples at densities up to 1.0 x 106
cells per ml. After a gasoline spill in Southern California, McKee et al. (1972) found
50,000 hydrocarbon-degrading bacteria per ml or higher in samples from wells
9-3
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containing traces of gasoline, while a noncontaminated well had only 200 organisms per
ml.
9.2.3. Use of Different Electron Acceptors For Biodegradation
Hydrocarbon biodegradation is essentially an oxidation-reduction reaction where
the hydrocarbon is oxidized (donates electrons) and an electron acceptor (e.g., oxygen,
nitrate, etc.) is reduced (accepts electrons). There are a number of different compounds
that can act as electron acceptors including oxygen (O^), nitrate (NOaO, iron oxide (e.g.
Fe(OH)3), sulfate (804-), and carbon dioxide (CO2). Aerobic bacteria can only use
molecular oxygen (Og) as an electron acceptor. Anaerobic bacteria use other compounds
such as NO3-, S04=, Fe(OH)3, or CO2 as electron acceptors. Oxygen is the most preferred
electron acceptor because microorganisms gain more energy from aerobic reactions.
Sulfate and carbon dioxide are the least preferred because microorganisms gain the least
energy from these reactions.
9.2.3.1. Aerobic Biodegradation
Almost all petroleum hydrocarbons are biodegradable under aerobic conditions.
Oxygen is a cosubstrate for the only known enzyme that can initiate the metabolism of
hydrocarbon (Young, 1984) and is later used as an electron acceptor for energy
generation. Under ideal conditions, biodegradation rates for low molecular weight
aliphatic, olefinic and aromatic compounds can be very high. Alvarez and Vogel (1991)
observed essentially complete removal of mixtures of benzene, toluene, and p-xylene in
aquifer slurries and pure cultures after 3 to 13 days incubation. Using aquifer material
from a gas plant facility in Michigan, Chiang et al. (1989) found between 80 and 100%
removal of BTX (120-16,000 ppb) in microcosms with sufficient oxygen. Half-lives for
biodegradation varied between 5 and 20 days. In a field test of aerobic biodegradation at a
former wood preserving facility, the total polynuclear aromatic hydrocarbon
concentration dropped by over 90% within 24 hours of the start of the test when sufficient
oxygen was available (Borden et al., 1989).
The ease of biodegradation will depend somewhat on the type of hydrocarbon.
Moderate to lower molecular weight hydrocarbons (C10 to C24 alkanes, single ring
aromatics) appear to be the most easily degradable hydrocarbons (Atlas, 1988). As the
molecular weight increases, so does the resistance to biodegradation. Gasoline contains
primarily the low to moderate molecular weight compounds while diesel and coal tars
contain more of the higher molecular weight compounds. Jamison et al. (1975) found
that the vast majority of gasoline components were readily degraded by a mixed
microbial population obtained from a gasoline contaminated aquifer. Many of the
individual gasoline components would not support microbial growth as a sole carbon
source but did disappear when gasoline dissolved in water was used as the substrate.
This suggested that a mixed microbial population may be necessary for complete
degradation. In a study of the catabolic activity of bacteria from an aquifer contaminated
with unleaded gasoline, Ridgeway etal. (1990) found that most isolates were very specific
in their ability to degrade hydrocarbons. Although all of the 15 hydrocarbons tested were
degraded by at least one isolate, most organisms were able to degrade only one of several
closely related compounds. Toluene, p-xylene, ethylbenzene, and 1,2,4-trimethylbenzene
were most frequently utilized, whereas cyclic and branched alkanes were least utilized.
In many cases, the major limitation on aerobic biodegradation in the subsurface
is the low solubility of oxygen in water. For example, aerobic toluene biodegradation can
be represented by the theoretical reaction:
9-4
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C6H5-CH3 + 9 02 bacteria^ 7 CQ2 + 4 H2O + Energy
(1)
Water saturated with air contains from 6 to 12 mg/1 of dissolved oxygen. Complete
conversion of toluene (and many other hydrocarbons) to carbon dioxide and water
requires approximately 3 grams of oxygen per gram of hydrocarbon. Using this ratio,
the oxygen present in water could result in the biodegradation of 2 to 4 mg/1 of dissolved
hydrocarbon by strictly aerobic processes. If the hydrocarbon concentration is greater
than this, biodegradation may be incomplete or may occur via slower anaerobic
processes.
9233,. Biodegradation via Nitrate Reduction
When the oxygen supply is depleted and nitrate is present (or other oxidized forms
of nitrogen), some facultatively anaerobic microorganisms will utilize nitrate (NO3-) as a
terminal electron acceptor instead of oxygen. For toluene, this process can be
approximated by the theoretical reaction:
C6H5-CH3 + 6 NO3- bacteria^ 7 CQ2 + 4 H2O + 3 N2 + Energy
(2)
Over the past decade, researchers have found that toluene, ethylbenzene, m-, p-,
and o-xylene, naphthalene and a variety of other compounds can be biodegraded using
nitrate as the terminal electron acceptor (Kuhn et al., 1985; Zeyer et al., 1986; Kuhn et al.,
1988; Hutchins et al., 1991; Mihelcic and Luthy, 1991). At this time, there is some
question about the biodegradability of benzene under denitrifying conditions. Several
investigators have reported benzene to be recalcitrant (not biodegradable) under
denitrifying conditions (Kuhn et al., 1988; Zeyer et al.; 1990; Hutchins et al., 1991)
whereas other studies indicate that benzene is degraded (Major et al., 1988; Kukor and
Olsen, 1989).
9.2.3.3. Biodegradation Using Ferric Iron
Once the available oxygen and nitrate are depleted, subsurface microorganisms
may use oxidized ferric iron [Fe(III)] as an electron acceptor. Microorganisms have
been identified that can couple the reduction of ferric iron with the oxidation of aromatic
compounds including toluene, phenol, p-cresol and benzoate (Lovley and Lonergan, 1990;
Lovley et al., 1989). Large amounts of ferric iron are present in the sediments of most
aquifers and could potentially provide a large reservoir of electron acceptor for
hydrocarbon biodegradation. This iron may be present in both crystalline and
amorphous forms. The forms that are most easily reduced are amorphous and poorly
crystalline Fe(III) hydroxides, Fe(III) oxyhydroxides, and Fe(III) oxides (Lovley, 1991).
A possible reaction coupling the oxidation of toluene to the reduction of Fe(III) in ferric
hydroxide [Fe(OH)3] can be approximated as:
C6H5-CH3+ 36Fe(OH)3 bacteria ^
7 C02 + 36 Fe+2 + 72 OH' + 22 H20 + Energy (3)
9-5
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The reduction of Fe(III) may result in high concentrations (10 to 100 mg/1) of
dissolved Fe(II) in contaminated aquifers. Lovley et al. (1989) found that in an aquifer
contaminated by a crude oil spill, the selective removal of benzene, toluene and xylenes
from the plume was accompanied by an accumulation of dissolved Fe(II) and depletion of
Fe(III) oxides in the contaminated sediments. Although the exact mechanism of
microbial ferric iron reduction is poorly understood, the available evidence suggests that
iron reduction is an important mechanism in the subsurface biodegradation of dissolved
hydrocarbons.
9.2&4. Biodegradation via Sulfate Reduction and Methanogenesis
Past research has shown that a wide variety of problem organics may be
biodegraded by sulfate-reducing and/or methanogenic (methane generating)
microorganisms (Grbic-Galic, 1990). These compounds include creosol isomers
(Smolenski and Suflita, 1987), homocyclic and heterocyclic aromatics (Berry et al., 1987),
alkylbenzenes (Grbic-Galic and Vogel, 1987; Seller et al. 1991), and unsaturated
hydrocarbons (Schink, 1985). Sulfate reducers could potentially biodegrade toluene using
sulfate in the following theoretical reaction (Seller et al., 1991):
C6H5-CH3 + 4.5 SO4= + 3 H20 bacteria ^
2.25 H2S + 2.25 HS' + 7 HCOa' + 0.25 H+ + Energy (4)
Methanogenic consortia (groups of microorganisms which generate methane) could
potentially biodegrade toluene using water as an electron acceptor (Vogel and Grbic-
Galic, 1986) in the following theoretical reaction:
C6H5-CH3 + 5 H2O bacteria^ 4.5 CH4 + 2.5 CO2 + Energy
(5)
At this time, little is known about the effect of sulfate reduction and methanogenic
biodegradation on the fate of dissolved hydrocarbons in the subsurface. While there are
well-documented reports of toluene biodegradation via sulfate reduction (Seller et al.,
1991) and methanogenesis (Grbic-Galic and Vogel, 1987), the extent and significance of
hydrocarbon biodegradation using these electron acceptors is poorly understood. This
may be partially due to the characteristics of these microorganisms. Sulfate-reducing
and methanogenic consortia are known to be very sensitive to a variety of environmental
conditions including temperature, inorganic nutrients (nitrogen, phosphorus, trace
metals), toxicants, and pH (Zehnder, 1978). An imbalance in any of these factors could
significantly reduce the rate and extent of biodegradation.
9.2.4. Effect of Environmental Conditions on Biodegradation
In most cases, the environmental factor that has the greatest influence on the rate
and extent of biodegradation is the availability of suitable electron acceptors (oxygen,
nitrate, etc.). In addition to electron acceptor concentration, temperature, pH, and
nutrients can influence biodegradation.
The optimum temperature for growth of most microorganisms present in shallow
aquifers is between 25°C and 40°C. In northern portions of the continental U.S., shallow
ground-water temperatures can be as low as 3°C and could significantly reduce the
-------
growth rate of subsurface microorganisms. This lower growth rate may be somewhat
offset by the higher solubility of oxygen in water at lower temperatures. In the central
and southern U.S., ground-water temperatures are higher (15°Cto 25°C)and should not
significantly impair biodegradation.
The optimum pH for microbial growth is dependent on the specific
microorganisms and their respiration pathways. Aerobic microorganisms often tolerate
a wider range in pH, whereas many anaerobes are sensitive to pH and operate efficiently
only in a narrow pH range. Denitrification and methanogenic biodegradation rates are
usually optimum between pH 7 and 8, and may drop off rapidly below pH of 6 (van den
Berg, 1974; U.S. EPA, 1975). The pH of most water supply aquifers is between 6.0 and 8.5,
although waters having lower pH are not uncommon (Hem, 1989).
The primary nutrients required for microbial growth are nitrogen, phosphorus,
sulfur and low levels of various minerals (Fe, Mn, etc.). Dissolution of the parent rock
typically releases some minerals and hence it is unlikely that these nutrients would be
completely absent (McNabb and Dunlap, 1975). Depending on the extent of microbial
growth, one or more of these nutrients may become limiting. In enhanced
bioremediation projects, nitrogen and phosphorus are frequently added to allow
maximum growth (Lee et al., 1988). In passive remediation systems, the extent of
microbial natural growth will be much lower and nutrient limitations will probably be
less severe. Lee and Ward (1984) found that addition of nitrogen, phosphorus and trace
minerals increased bacterial growth in creosote contaminated ground water but did not
increase the extent of contaminant removal.
9.3. NATURAL BIOREMEDIATION OF A HYDROCARBON PLUME
While there are no truly typical sites, it may be useful to consider a hypothetical
site where the aquifer hydrogeology and geochemistry are reasonably well defined. For
this hypothetical case, assume that a small release of gasoline has occurred from an
underground storage tank (UST). The soils immediately below the tank are
contaminated with moderate levels of residual hydrocarbon. A simple schematic of this
site is shown in Figure 9.1.
Rainfall infiltrating through the hydrocarbon-contaminated soils will leach out
some of the more soluble hydrocarbon components, probably benzene, toluene,
ethylbenzene and xylenes (BTEX); fuel additives, methyl tertiary butyl ether (MTBE) and
ethylene dibromide (EDB); and a smaller portion of the less soluble constituents (aliphatic
hydrocarbons and higher molecular weight aromatics). As the hydrocarbon-
contaminated water migrates downward through the unsaturated zone, a portion of the
dissolved hydrocarbons may biodegrade. The extent of biodegradation will be controlled
by the size of the spill and the rate of downward movement. For larger spills, the
available oxygen will be consumed and aerobic biodegradation will not continue.
Dissolved hydrocarbons that do not completely biodegrade within the unsaturated
zone will be carried downward, enter the saturated zone and be transported
downgradient within the water table aquifer. Figure 9.2 shows a simple schematic plan
view of a dissolved hydrocarbon plume undergoing biodegradation. Near the source
area, dissolved hydrocarbons enter the saturated zone and flow downgradient. Native
microorganisms will use the available oxygen in the source area to biodegrade a portion
of the hydrocarbon. Dissolved hydrocarbons that are not biodegraded will then be carried
downgradient in a plume of anaerobic contaminated water. In this region, the extent of
biodegradation will probably be limited by the available oxygen supply. Because the
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solubility of oxygen in water is relatively low, only a small amount of hydrocarbon may be
biodegraded aerobically in the source area.
Aerobic - Unconiammated Ground Water
Figure 9.1. Profile of a typical hydrocarbon plume undergoing natural bioremediation.
Oxygenated-Uncon laminated
Ground Water
Flow
—• • • -Aerobic Margin' •
Flow
Oxygenated-Uncon laminated
Ground Water
Figure 9.2. Plan view of a typical hydrocarbon plume undergoing natural
bioremediation.
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As the plume migrates downgradient, dispersion will mix the anaerobic
hydrocarbon-contaminated water with clean oxygenated water at the plume fringes.
This is the region where most aerobic biodegradation occurs. After an acclimation
period, a population of aerobic hydrocarbon-degrading bacteria will develop in the
sediments of this fringe area. As oxygenated water mixes with hydrocarbon
contaminated water, the attached bacteria will consume both the hydrocarbon and
oxygen, preventing the further spread of the contaminant plume. This is the reason why
many dissolved hydrocarbon plumes appear long and narrow. As the dissolved
hydrocarbons disperse outward, they come in contact with oxygenated ground water and
biodegrade. If this process is allowed to continue indefinitely, the dissolved hydrocarbon
plume will reach a quasi-steady state condition where the long-term rate of hydrocarbon
dissolution is equal to the rate of biodegradation. The major limitations to this process
are the amount of oxygen present and the extent of mixing. Recent field studies
(Freyberg, 1986; Moltyaner and Killey, 1988) have shown that mixing (dispersion) in most
aquifers is very limited and consequently, the overall rate of aerobic biodegradation may
be slow.
Hydrocarbon biodegradation using nitrate will probably follow the same general
pattern as aerobic biodegradation. Nitrate present in the uncontaminated ground water
will mix with hydrocarbon at the plume fringes and increase the amount of
biodegradation. While oxygen is preferred to nitrate by most microorganisms, nitrate is
still a good electron acceptor and many hydrocarbons can be biodegraded in this manner
(e.g. toluene, ethylbenzene, xylenes). Also, any hydrocarbons biodegraded using nitrate
will not express a demand for oxygen, which will allow further aerobic biodegradation of
other hydrocarbons that require oxygen.
In the core of the plume, conditions may become highly reducing and other
anaerobic biodegradation reactions may occur. Certain hydrocarbons and many
bacterial waste products may be biodegraded by iron reduction, sulfate reduction or
methanogenic biodegradation. While little is known about these processes, it is clear that
they do occur. Field monitoring has shown that in the core of some hydrocarbon plumes,
sulfate concentrations are reduced and dissolved iron and methane concentrations are
elevated (Wilson et al., 1990). It is not yet known whether these conditions result from
direct anaerobic attack by bacteria on the hydrocarbon molecule or anaerobic
biodegradation of bacterial waste products. From a practical perspective, it may not
matter. Any organic carbon (hydrocarbons or waste products) biodegraded by an
anaerobic pathway will reduce the total oxygen demand on the aquifer. By reducing the
overall oxygen demand, more oxygen will be available for those compounds that can only
be biodegraded aerobically.
9.4. CASE STUDIES OF NATURAL BIOREMEDIATION
Over the past ten years, there have been a number of well-documented studies
which have demonstrated that plumes of dissolved hydrocarbons will biodegrade in the
subsurface without human intervention. These studies have included former wood
preserving facilities and petroleum releases.
One of the earliest studies of natural bioremediation was conducted at the United
Creosoting Company site in Conroe, Texas, by a team of researchers from the R.S. Kerr
Environmental Research Laboratory (U.S. EPA) and the National Center for Ground
Water Research. Early work (Lee and Ward, 1984; Wilson et al., 1985) demonstrated that
an adapted population of creosote-degrading microorganisms was present within the
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contaminated zone but not in the uncontaminated regions of the aquifer. Later studies
correlated creosote biodegradation with the availability of dissolved oxygen (Lee and
Ward, 1984). These results were used to develop and calibrate the computer model,
BIOPLUME, to simulate hydrocarbon transport and aerobic biodegradation within the
aquifer (Borden and Bedient, 1986; Borden et al., 1986). Model results indicated that
removal of the contaminant source would be sufficient to contain the hydrocarbon plume
and active remediation by pump and treat would not be required.
Microbiologists from the U.S. Geological Survey have studied two different
creosote-contaminated aquifers where methanogenic degradation of organic compounds
has been observed. Field studies at a contaminated aquifer in St. Louis Park, Minnesota,
showed that methane production was occurring in zones within the aquifer that had
been contaminated with creosote (Godsyet al., 1983). Later studies demonstrated that the
presence of anaerobes (denitrifiers, iron reducers, sulfate reducers and methanogens)
was highly correlated with the presence of creosote. More recent work at an abandoned
creosote plant in Pensacola, Florida, has shown a wide variety of organic compounds
present in the aquifer were undergoing methanogenic biodegradation and that transport
distances in the aquifer could be correlated with biodegradation rates observed in
laboratory microcosms (Troutman et al., 1984; Goerlitz et al., 1985).
Monitoring at petroleum contamination sites suggests that methanogenic
biotransformation of petroleum related compounds may be more common than has
generally been assumed. Ehrlich et al. (1985) observed elevated numbers of sulfate-
reducing and methanogenic bacteria in a jet fuel contaminated aquifer. Evans and
Thompson (1986) and Marrin (1987) monitored methane concentrations in soil gas to map
subsurface hydrocarbon contamination. In a study of soil gas concentrations near
underground storage tanks, Payne and Durgin (1988) found elevated methane
concentrations at over 20% of the 36 sites surveyed. Methane gas production can be so
rapid that safety problems occur at some sites. Hayman et al. (1988) had to develop a
special apparatus to remove the large quantities of methane generated from a fuel spill at
the Miami, Florida, airport.
Hult (1987b) observed the production of large volumes of methane in the
unsaturated zone immediately below a crude oil spill at the U.S. Geological Survey
research site in Bemidji, Minnesota. At this same site, Eganhouse et al. (1987) observed
a two order of magnitude decrease in alkylbenzene concentration over a downgradient
travel distance of 150 m. This decrease was accompanied by elevated concentrations of
aliphatic and aromatic acids in the ground water (Baedecker et al., 1987). The acids
included benzoic, methylbenzoic, trimethylbenzoic, toluic, cyclohexanoic, and
dimethylcyclohexanoic. These are the same acids identified by Grbic-Galic and Vogel
(1987) as intermediates in anaerobic degradation of alkylbenzenes. Ground-water and
sediment analyses demonstrated that methanogenic biodegradation caused a drop in pH
and a rise in bicarbonate concentrations in the ground water. The actual drop in ground-
water pH appears to have been limited by dissolution of carbonate minerals (and possibly
aluminosilicates) (Siegel, 1987).
9.5. SITE CHARACTERIZATION FOR NATURAL BIOREMEDIATION
The first step in evaluating a site for potential application of natural
bioremediation is to complete a conventional site characterization. This characterization
should include: (1) detailed description of the subsurface hydrology and geology; (2)
delineation of the contaminant source area and any mobile NAPLs; (3) delineation of the
horizontal and vertical extent of the contaminant plume; and (4) identification of any
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downgradient receptors (wells or surface discharges) that could potentially be affected.
In some cases it may be appropriate to model ground-water flow and/or transport at the
site to gain a better understanding of the hydrologic system and contaminant transport
pathways.
In addition to the basic data required inmost remedial investigations, information
will be needed to evaluate the ability of the aquifer to assimilate wastes and the potential
risks if the system does not perform as expected. Specific questions that should be
addressed are described below.
9.5.1. Is the Contaminant Biodegradable?
The first major question to be addressed is whether the contaminants are
biodegradable by microorganisms present at the site. The level of detail required to
answer this question will depend on the type of contaminant and general site conditions.
The most common dissolved hydrocarbons (benzene, toluene, ethylbenzene and
xylenes) released from gasoline spills are known to be readily biodegradable under
aerobic conditions (Jamison et al., 1975; Gibson and Subramaman, 1984; Thomas et al.,
1990; Alvarez and Vogel, 1991). In addition, aerobic hydrocarbon-degrading
microorganisms are very common in nature and have been recovered from virtually all
petroleum-contaminated sites that have been studied (Litchfield and Clark, 1973). For
most petroleum sites, extensive studies to confirm the presence of BTEX-degrading
microorganisms are probably not necessary. In contrast to BTEX, there is much less
information available on the biodegradability of many fuel additives such as methyl
tertiary butyl ether (MTBE), 1,2-dibromoethane (EDB), or 1,2-dichloroethane (EDO. If
persistence of fuel additives is a concern, site specific studies may be needed to confirm
the presence of microorganisms capable of degrading these compounds and to estimate
biodegradation rates.
Ground water contaminated with creosote, coal tar and heavier petroleum
products often contains higher molecular weight aromatic compounds (fluorene,
phenanthrene, dibenzofuran, etc.). These compounds often biodegrade much more
slowly and may persist for long time periods even under ideal conditions (Lee, 1986;
Borden et al., 1989). Site specific laboratory studies may be needed to determine if these
compounds are biodegradable by subsurface microorganisms and if the rates of
biodegradation are sufficient to contain the contaminant plume.
9.5.2. Is Biodegradation Occurring in the Aquifer?
Probably the most important question to address is whether the compounds of
concern are actually biodegrading in the aquifer. The simplest way to answer this
question is to examine the ground-water monitoring data and determine if there is a
significant decline in the total mass of the contaminant as the plume migrates
downgradient. Unfortunately, it is often difficult to evaluate changes in total mass
without an extensive monitoring well network. Comparison of dissolved hydrocarbon
concentrations at individual points is not sufficient to prove biodegradation, since
dispersion will reduce the point concentrations even if there is no biodegradation. To
overcome these problems, other parameters are often used as secondary indicators of
biodegradation.
One very useful method for assessing the extent of biodegradation is to monitor
changes in the concentration of inorganic compounds within the aquifer.
Biodegradation of dissolved hydrocarbons will result in the removal of electron acceptors
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(oxygen, nitrate, and sometimes sulfate) and release of waste products (carbon dioxide
and sometimes reduced iron and methane) in areas where microorganisms are most
active. If Held monitoring indicates that oxygen, nitrate and/or sulfate is being depleted
(or carbon dioxide, soluble iron, or methane is being produced) within the plume, this is
a good indication that one or more of the contaminants are being biodegraded. The major
limitation of this approach is that it is not possible to determine which specific
compounds are being degraded.
A second method that can be used to determine if individual compounds are being
biodegraded is to examine changes in the ratio of different contaminants along the flow
path. If one contaminant declines more rapidly than another, this suggests that some
process is removing that contaminant. Field monitoring at several hydrocarbon plumes
(Jasiorkowski and Robbins, 1991) and a sanitary landfill leachate plume (Barker et al.,
1986) has shown a more rapid downgradient decline in o-xylene concentrations than m-
or p-xylene. Since all of the xylene isomers should sorb to the aquifer equally, the only
explanation for this pattern would be biodegradation of the o-xylene.
9.5.3. Are Environmental Conditions Appropriate for Biodegradation?
As previously discussed, virtually all hydrocarbons are biodegradable. Yet
extensive plumes of dissolved hydrocarbons persist in some aquifers. Why does this
apparent contradiction occur? The answer lies with the environmental conditions in the
specific aquifer.
Virtually all hydrocarbons biodegrade more rapidly in the presence of dissolved
oxygen. If dissolved oxygen concentrations are low in a specific aquifer, the rate of
natural biodegradation will be lower. Also, the pH of the aquifer should be near
neutrality, adequate inorganic nutrients should be present (nitrogen, phosphorus, and
trace minerals), and no toxicants should be present that could inhibit microbial growth.
In most cases, it is not necessary to perform extensive investigations to precisely
determine the concentrations of nitrogen, phosphorus, trace minerals, and potential
toxicants. Past studies have shown that most aquifers do not contain toxicants and do
contain adequate levels of inorganic nutrients to support moderate levels of microbial
growth (Lee, 1986). If field monitoring indicates that biodegradation is occurring, it can
reasonably be assumed that aquifer conditions are appropriate for microbial growth.
Where field monitoring data suggest that biodegradation is being inhibited, additional
laboratory studies may be needed to identify those factors that are limiting
biodegradation. When performing laboratory studies, it is very important to design the
experiment to simulate actual conditions within the aquifer. For example, if the oxygen
supply in the aquifer is limiting, laboratory studies conducted with an excess of oxygen
(or nitrogen, phosphorus, etc.) will overestimate the actual extent of biodegradation and
lead to erroneous conclusions.
9.5.4. If the Waste Doesn't Completely Biodegrade, Where Will It Go?
Natural bioremediation, like other available techniques, is not foolproof.
Instances arise where for some unforeseen reason, the contaminant plume does not
biodegrade as expected. In order to adequately manage a natural remediation system, it
is first necessary to evaluate the consequences of a system failure. In most cases, the
primary consequences of a failure will be: (1) contamination of water supply wells; or (2)
contamination of surface water. Appropriate controls should be incorporated into a
natural remediation system to identify a failure and eliminate it.
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9.6. MONITORING NATURAL BIOREMEDIATION SYSTEMS
One of the most important factors to consider in planning a natural
bioremediation system is monitoring system performance. The monitoring system
typically includes: (1) interior wells to monitor the actual plume distribution and
indicator parameters; and (2) guardian wells at the outside edge of the area of
contamination to monitor potential offsite migration and determine if additional
remedial measures are required.
Interior wells may be monitored to evaluate the overall system performance.
Parameters to be monitored typically include: (1) individual hydrocarbon components; (2)
dissolved oxygen; (3) nitrate; (4) dissolved iron; (5) redox potential; (6) carbon dioxide; (7)
pH; and (8) total organic carbon. Monitoring of individual hydrocarbon components can
be performed using standard techniques and provides an indication of the treatment
effectiveness.
Dissolved oxygen is monitored to determine if one or more of the organics are
biodegrading and as an aid in defining the contaminant plume. Typically, both
dissolved oxygen and hydrocarbon concentrations will be reduced at the margins of the
plume. Dissolved oxygen can be measured in the field using electrodes or field test kits.
Collection of accurate data on dissolved oxygen concentrations in ground water is
difficult because of problems with aerating the samples during collection. One to two
mg/1 of oxygen may be added to the sample during collection unless special precautions
are taken to prevent aeration. The extent of aeration can be reduced by using special
pumps and filling the well casing with argon gas, but in most cases aeration cannot be
completely eliminated.
Nitrate and iron may be monitored to determine the extent of anaerobic
biodegradation of the hydrocarbons and any bacterial waste products. Nitrate can be
monitored by collecting samples using conventional techniques and then transporting to
the laboratory for analysis. Collection of samples for iron analysis is more difficult
because of problems with iron present in suspended solids and precipitation of dissolved
iron during transport. One method that may be used is to filter samples in the field
during collection, preserve them with a concentrated acid and then analyze for total iron.
While this procedure does not differentiate between dissolved ferric and ferrous iron, in
most cases essentially all iron in excess of 0.5 mg/1 will be in the reduced ferrous form
(Hem, 1989).
Measurement of redox potential is relatively simple and can provide a good
qualitative indicator of the overall oxidation-reduction status of the aquifer. Redox
potential can be measured using a platinum electrode and a standard pH meter. In
locations where the redox potential is negative, the ground water is strongly reduced,
indicating significant bacterial decomposition. In areas where the redox potential is
positive, the ground water is oxidizing, indicating that the contaminant plume has not
reached this point or that bacterial degradation has not occurred. In most cases, redox
potentials should not be used for precise calculations but as a qualitative indicator of
environmental conditions within and outside the contaminant plume (Barcelona et al.,
1989).
Carbon dioxide and pH can be monitored to evaluate the extent of bacterial
respiration and determine if conditions are suitable for biodegradation. If the pH falls
outside of a specified range (typically 5 to 9), biodegradation may be inhibited.
Accumulation of carbon dioxide within and adjoining the contaminant plume is
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indicative of bacterial respiration. Direct interpretation of carbon dioxide concentrations
is sometimes difficult because of shifts in the dominant form of inorganic carbon with pH
and release of inorganic carbon during dissolution of certain minerals.
Individual hydrocarbon components can be monitored to determine the extent of
the contaminant plume and any organic waste products produced during biodegradation
of the dissolved hydrocarbons. In some cases, dissolved hydrocarbons will not be
completely biodegraded but will be converted to nontoxic organic waste products.
Monitoring total organic carbon (TOG) will provide some indication of the total oxygen
demand exerted by the contaminant plume.
Guardian wells may be installed at the outside edge of the contamination area to
monitor system performance, evaluate the potential for offsite migration and determine
if additional remedial measures are required. In most cases, these wells are used for
regulatory purposes and are only monitored for compounds of regulatory concern. These
wells may also be monitored occasionally for indicator parameters (oxygen, nitrate, etc.)
to confirm that the wells do in fact intercept the 'plume' of ground water that has
undergone biodegradation.
9.7. PERFORMANCE OF NATURAL BIOREMEDIATION SYSTEMS
Under optimal conditions, natural bioremediation should be capable of completely
containing a dissolved hydrocarbon plume. While there are few well-documented cases
where this has occurred, there is a great deal of anecdotal evidence that suggests that
natural bioremediation can be effective in containing dissolved hydrocarbon plumes.
Typically greater than 90% of all underground tanks are used to store gasoline and other
petroleum fuels. Yet a study by the California Department of Health Services (Hadley
and Armstrong, 1991) found that by far the most common ground-water contaminants
were chlorinated solvents, not petroleum constituents. These results suggest that the
petroleum contaminants are being removed to below detection limits before reaching
water supply wells.
In many aquifers, conditions will not be perfect for natural bioremediation and
less than optimal biodegradation will occur. The extent of aerobic biodegradation will be
controlled by the amount of contamination released, the rate of oxygen transfer into the
subsurface, and the background oxygen content of the aquifer. When large amounts of
contamination enter the subsurface, they overwhelm the capacity of an aquifer to
assimilate them. As a result, extensive contamination may persist for long distances.
When hydrogeologic conditions such as clayey, confining layers or naturally occurring
organic deposits reduce the rate of oxygen transfer into the subsurface, the assimilative
capacity of the aquifer will be lower. Anaerobic biodegradation may be inhibited by low
pH, low buffering capacity, or absence of appropriate electron acceptors (nitrate, iron,
etc.). Heterogeneous conditions within the aquifer may prevent mixing and allow a
portion of the plume to migrate rapidly. If this occurs, the extent of biodegradation may
be less than would be expected for more uniform conditions.
9.8. PREDICTING THE EXTENT OF NATURAL BIOREMEDIATION
One of the most frequently asked questions is "How far will the plume migrate
before it biodegrades?" Unfortunately, this is a very difficult question to answer.
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To predict the maximum extent of plume migration, it is necessary to estimate1 (1)
the rate of migration; and (2) the rate of biodegradation. The rate of contaminant
migration can be estimated by measuring the hydraulic gradient and permeability of the
aquifer. Accurate estimation of the biodegradation rate within an aquifer is much more
difficult. Results from laboratory studies may significantly over- or underestimate
biodegradation rates if environmental conditions in the laboratory differ from conditions
in the field.
Computer models may be used to combine the results of field and laboratory
investigations and to predict the actual extent of biodegradation in an aquifer. At
present, there are a number of computer models that have been developed to simulate
contaminant biodegradation (Molz et al., 1986; Odencrantz et al., 1989; MacQuarrie et
al., 1990). Two of the most commonly used models for simulating hydrocarbon
biodegradation are: (1) first-order decay models; or (2) BIOPLUME II.
Kemblowski et al. (1987) describe the use of a first order decay model to simulate
hydrocarbon biodegradation at several sites. Their results show that this simple
approach can adequately match the observed hydrocarbon distribution in the aquifers
studied. The major limitation of this method is in estimating the first-order decay rate
before extensive data are collected. Once the contaminant plume is properly delineated
and shown to be biodegrading, it is possible to match the field data to a first-order decay
equation and estimate the decay rate.
Hydrocarbon biodegradation may also be simulated using the computer model
BIOPLUME II (Rifai et al., 1989). BIOPLUME II is based on the U.S.G.S. Method of
Characteristics model (Konikow and Bredehoeft, 1978) and includes advection,
dispersion, oxygen-limited biodegradation, and first-order decay in a two-dimensional
aquifer. Oxygen-limited biodegradation is simulated as an instantaneous reaction
between oxygen and hydrocarbon. Calibration of BIOPLUME II is relatively simple
because the only data required are the aquifer hydrogeology, background oxygen
concentrations and contaminant source concentrations.
The major limitations of BIOPLUME II are the inability to accurately simulate
dissolution of residual hydrocarbons and anaerobic biodegradation of hydrocarbons or
bacterial waste products. BIOPLUME II assumes that all contaminants are converted
directly to carbon dioxide and water using 3 mg of oxygen for every mg of hydrocarbon
degraded. In many cases, this significantly underestimates the amount of
biodegradation (Chiang et al., 1989) and leads to a conservative prediction. This error is
presumably due to anaerobic degradation of bacterial waste products and certain
hydrocarbons. Anaerobic decay can be simulated in BIOPLUME II using a first-order
decay rate, but this approach suffers from the same limitation as the simple first-order
decay models. There are no accurate methods available to estimate these decay rates
without first collecting extensive field data.
In summary, there are no good methods available at this time for predicting the
extent of hydrocarbon biodegradation without first characterizing the contaminant
plume. Once the contaminant plume is defined, there are several methods that can be
used to analyze the available data and evaluate the effect of different alternatives on
contaminant migration. As additional field data becomes available from different sites,
it may become possible to estimate the decay rate by extrapolating results from similar
aquifers and avoid extensive field data collection.
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9.9. ISSUES THAT MAY AFFECT THE COSTS OF THIS TECHNOLOGY
One of the major factors controlling the costs of natural bioremediation is
acceptance of this approach by regulators, environmental groups and the public. At sites
where natural bioremediation is strongly opposed, the costs of implementation may
actually be higher than conventional remediation technologies (e.g. pump and treat). In
North Carolina, regulations have been in place for five years that allow responsible
parties to request a reclassification of contaminated ground water to a nonwater supply
use. Once reclassified, the responsible party would not be required to actively remediate
the site. At present there are over 2,000 sites under investigation, with over 100 pump-
and-treat systems in operation. Yet no one has ever filed a request for reclassification.
The apparent cause is a perception by the responsible parties that the legal,
administrative and site characterization costs for reclassification would be excessive,
and the probability of success would be low.
The second major issue limiting application of natural bioremediation is third
party liability. A hydrocarbon plume that is left in place to naturally biodegrade may
migrate under an adjoining property, posing a potential risk to public health and the
environment. Even when public health is not at risk, adjoining property owners may
have strong concerns about a contaminant plume migrating under their property and
the potential impact on property values. In such cases, natural bioremediation could be
coupled with active plume management technology, such as purge wells, to prevent
undesirable impact to third parties.
9.10. KNOWLEDGE GAPS AND RESEARCH OPPORTUNITIES
Currently, there are no reliable methods for predicting the effectiveness of natural
bioremediation without first conducting extensive field work. Existing mathematical
models cannot be used in a predictive mode because they either: (1) require extensive field
data for calibration; or (2) greatly underestimate the extent of anaerobic biodegradation.
This is often the primary reason why natural bioremediation is not seriously considered
when evaluating remedial alternatives. Without some reasonable assurance of success,
responsible parties are not willing to risk the large sums of money required for legal,
administrative and site characterization costs.
Over the next several years, there is potential to dramatically improve our ability
to predict the extent of natural bioremediation. Several organizations [U.S. EPA,
American Petroleum Institute (API), Electric Power Research Institute (EPRI)] are
funding extensive field studies to characterize dissolved hydrocarbon plumes undergoing
natural bioremediation. These studies will generate an extensive database that will be
used to improve our understanding of the basic processes that control natural
biodegradation and to develop more accurate models for predicting the extent of natural
bioremediation. In order to use this database effectively, additional research is needed in
two general areas: (1) anaerobic hydrocarbon biodegradation; and (2) biodegradation
modeling.
We now know that many hydrocarbons can be biodegraded under anaerobic
conditions using nitrate, iron, sulfate, water and carbon dioxide as terminal electron
acceptors. What we do not know is what factors control the rate of anaerobic hydrocarbon
biodegradation and why anaerobic hydrocarbon biodegradation occurs in some locations
and not in others. Detailed laboratory studies are needed to resolve these questions.
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Primary emphasis should be placed on coordinating these laboratory studies with the
ongoing field work to maximize benefits.
Existing models of hydrocarbon biodegradation do not adequately represent
anaerobic biodegradation. Consequently, these models grossly under-predict the extent
of biodegradation at many sites. Until this problem is resolved, natural bioremediation
will not be seriously considered at many sites where it is a reasonable alternative. The
extensive field database being collected by EPA, API and EPRI provides an outstanding
opportunity to resolve this problem. By coordinating model development with the field
data collection, in the next few years we can significantly improve our ability to predict
the extent of natural bioremediation.
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9-23
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SECTION 10
NATURAL BIOREMEDIATION OF CHLORINATED SOLVENTS
Timothy M. Vogel
Environmental and Water Resources Engineering
Department of Civil and Environmental Engineering
The University of Michigan
Ann Arbor, MI 48109-2125
10.1. SUMMARY
Halogenated solvents, as some of the most mobile constituents of hazardous
wastes, can pose a threat to subsurface drinking water supplies. Fifteen years ago,
many of these highly chlorinated organic compounds were considered recalcitrant to
biological degradation in the environment. Since then, researchers have shown that
these compounds undergo chemical reactions with half-lives ranging from days to years.
Their transformation products may contain no halogens, as a result of hydrolysis, or still
be partially halogenated alkenes, as a result of elimination of hydrogen halide.
Elimination reactions, rather than hydrolysis reactions, will predominate with greater
halogenation. Microbially-mediated reactions of chlorinated solvents usually involve
oxidation or reduction reactions. Oxidation reactions are generally slower with highly
halogenated compounds than with compounds containing fewer halogen substituents,
while the opposite is true for reduction reactions. Oxidation reactions do not
dehalogenate in the first rate-limiting step, but in subsequent steps. Reduction reactions
normally include the dehalogenation of these solvents, producing less halogenated
homologues. The dechlorination occurs under anaerobic conditions and results in less
chlorinated, and often aerobically degradable, products. Engineered systems, or in-situ
bioremediation, can effectively employ either aerobic alone or sequential
anaerobic/aerobic microbial processes to biodegrade chlorinated solvents. The natural
bioremediation of chlorinated solvents depends on the appropriate subsurface
environmental conditions. These conditions must promote growth of either anaerobic or
aerobic microorganisms and allow for the contact between chlorinated solvent and these
microbes. The rate of natural bioremediation may be slow enough that analyses of
subsurface chemical constituents would indicate the presence of several chlorinated
products of original chlorinated solvents. Enhanced bioremediation requires improving
microbial growth conditions, reducing mass transfer limitations, and controlling
movement of subsurface chlorinated solvents.
10.2. FUNDAMENTAL PRINCIPLES
Hazardous wastes contain many different classes of compounds, including
metals, polyaromatic hydrocarbons (PAH), polychlorinated biphenyls (PCB), aromatic
hydrocarbons (e.g., benzene), and halogenated solvents. Large volumes of both
chlorinated aliphatic and aromatic hydrocarbons are produced each year for a variety of
domestic and commercial purposes (Merian and Zander, 1982; Pearson, 1982). Table 10.1
contains a list of some halogenated solvents and their annual production rates. They
10-1
-------
have become widely distributed in the environment as a result of discharges from
industrial and municipal wastewaters, urban and agricultural runoff, leachates from
landfills, and leaking underground tanks and pipes. Sediments beneath some industrial
sites contain chlorinated hydrocarbons in excess of 1,000 ppm (Phelps et al., 1990).
Several examples of well-documented contaminant plumes in ground water exist both in
the United States and Canada (Mackay and Cherry, 1989). Most of these plumes are
composed of large quantities (0.4 x 109 to 5.7 x 109 liters) of water contaminated by
chlorinated industrial solvents, such as tetrachloroethylene (PCE), trichloroethylene
(TCE), and 1,1,1-trichloroethane (TCA), or aromatic hydrocarbons, such as benzene,
toluene, and xylene (Mackay and Cherry, 1989). Such plumes pose problems regarding
containment and remediation. Halogenated solvents are relatively mobile in the
environment, being both highly volatile and generally less retarded in ground water than
many other constituents of hazardous wastes. Due to their relatively high mobility,
halogenated solvents can be transported from hazardous waste sites to drinking water
wells by ground water. Even a spill of relatively small volume can contaminate millions
of gallons of drinking water. A survey conducted by the U.S. Environmental Protection
Agency documented that 22% of approximately 466 randomly sampled subsurface
drinking water sources contain mixtures of volatile organic chemicals at detectable
levels (Symons et al.,1975; Westrick etal., 1984). Some halogenated solvents potentially
pose significant human health hazards, which is in part reflected in the drinking water
maximum contaminant levels (MCLs) set by the U.S. Environmental Protection Agency.
These MCLs range from near 1 microgram per liter (jig/1) for compounds such as vinyl
chloride to 200 ug/1 for TCA (Table 10.1). The ultimate fate of these halogenated aliphatic
compounds in ground-water supplies is often controlled by their chemical and biological
reactivity. Although chlorinated organic solvents have been released into the
environment for decades, they have only come under intense international scrutiny in
the last 10 years. Investigations of the fate of these compounds in the environment,
including volatilization into the atmosphere, sorption onto sediments, bioaccumulation
and concentration in aquatic and terrestrial organisms, and dissolution in surface and
ground waters have led to increased understanding of the movement of these
compounds. Yet, none of these processes actually degrades these compounds.
TABLE 10.1. PRODUCTION,PROPOSED MAXIMUM CONTAMINANT LEVELS, AND TOXICITY
RATINGS OF COMMON HALOGENATED ALIPHATIC COMPOUNDS8
Compound
Trihalomethanes
Vinyl chloride
1,1-Dichloroethylene
trans -1,2 Dichloroethylene
Trichloroethylene
Tetrachloroethylene
1 , 1 -Dichloroethane
1,2-Dichloroethane
1 , 1. 1-Tnchloroethane
Production* (million
Iblyr)
7000
200
<0001
200
550
<0001
12.000
eoo
MCLe
(mg 11 or ppm)
100
1
7
--
5
..
..
5
200
Carcino-
genicity*
1
3
-.
3
..
..
2
3
Vogel et al, 1987
Federal Register, 1985
Maximum contaminant level, Van Nostrand Remhold Co, 1984
Carcmogenicity 1 = chemical is carcinogenic. 2 = chemical probably is carcinogenic,
3= chemical cannot be classified
10-2
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Indeed, these compounds were considered resistant to chemical and biological
degradation, although medical studies had shown that some chlorinated compounds are
metabolized by rat liver cells (Anders, 1983). One reason for the apparent recalcitrance of
highly chlorinated compounds is their oxidized nature. The reactivities of chlorinated
solvents are controlled mainly by the degree of halogenation. Chlorine substituents tend
to make the chlorinated solvent fairly oxidized, unlike most hydrocarbons composed of
only carbon and hydrogen. The specific chemistry of these solvents controls the types of
reactions they undergo. Due to the variety of the reactions that halogenated solvents
undergo, a diverse range of transformation products may be found at contaminated sites.
The products often have fewer chlorine substituents than the original compound, and
may be more or less toxic. In some cases, the halogenated solvents are completely
mineralized to carbon dioxide (Vogel et al., 1987). An understanding of the changes that
chlorinated compounds are subject to in the environment provides the basis for
understanding their natural bioremediation and for designing of remediation systems.
Conditions can be induced that stimulate the transformations of hazardous compounds
to the least harmful products possible. A crucial point in the complete destruction of
chlorinated hydrocarbons, as will be discussed, is the removal of chlorine substituents
from the molecule.
10.3. CHEMICAL REACTIONS
Halogenated solvents generally undergo either or both substitution and
dehydrohalogenation reactions in water (Vogel et al., 1987). Substitution reactions of
halogenated solvents in water (hydrolysis) involves the replacement of a halogen
substituent with a hydroxy (-OH) group, forming an alcohol. For example, chloroethane
is hydrolyzed to ethanol (Vogel and McCarty, 1987a). If the halogenated solvent has more
than one halogen, further substitution reactions can occur. For example, TCA is
partially hydrolyzed through a series of substitution reactions to acetic acid (Mabey et al.,
1983); and 1,2-dibromoethane (EDB) is partially hydrolyzed to ethylene glycol (Weintraub
et al., 1986). Dehydrohalogenation reactions of halogenated solvents in water usually
involve the elimination of hydrogen halide from an alkane and the formation of an
alkene. Some halogenated solvents undergo both hydrolysis and dehydrohalogenation in
water. For example, the two compounds described to undergo hydrolysis above, TCA and
EDB, also form 1,1-dichloroethylene (1,1-DCE) (Vogel and McCarty, 1987b) and vinyl
bromide (bromoethylene) (Vogel and Reinhard, 1986), respectively, as a result of
dehydrohalogenation. The likelihood that a halogenated solvent will undergo either
hydrolysis or dehalogenation depends in part on the number of halogen substituents.
More halogen substituents on a compound tend to increase the chance of
dehydrohalogenation reactions occurring. The opposite is true for hydrolysis reactions.
Bromine substituents are generally more reactive than chlorine substituents. So, less of
these substituents than chlorine substituents are needed to cause the halogenated solvent
to undergo dehydrohalogenation. Location of the halogen substituents on the carbon
skeleton also has some effect on both type and rate of reaction.
The rates of substitution and dehydrohalogenation reactions are also dependent on
the degree of halogenation. Substitution reactions generally decrease in rate with
increasing halogenation. Monohalogenated alkanes have half-lives at 25°C on the order
of days to months. Polychlorinated alkanes have half-lives that range up to thousands of
years for carbon tetrachloride. Dehydrohalogenation rates for halogenated solvents
increase with increasing halogenation. Unfortunately, many reported environmentally
significant chlorinated solvent half-lives are the result of extrapolation from experiments
performed at elevated temperatures (Mabey and Mill, 1978). Some chlorinated solvent
reaction rates are so slow as to make experiments run at environmental temperatures
-------
impractical. However, in order to accurately extrapolate rate values from elevated
temperatures, experiments need to be conducted at several different temperatures. The
data can be extrapolated using the Arrhenius equation in a manner that includes
statistical evaluation (Vogel and Reinhard, 1986). The resulting range of values for
chlorinated solvent half-lives can be used to estimate lifetimes of chemicals in ground
water. However, other processes such as sorption will influence these estimates.
Typically, the activation energies for chlorinated solvent hydrolysis and
dehydrohalogenation reactions are approximately 100 KJoules/mole, which results in a
factor of 3.5 change in reaction rate and half-life with each 10°C change in temperature.
So, values listed at 25°C, such as three years for TCA would be 10.5 years at 15°C. Beyond
the impact of the initial chemicals themselves, the approximate half-lives and the
potential products of chlorinated solvent chemical reactions in water tend to suggest that
compounds that undergo dehydrohalogenation can cause the most significant health
hazard. In addition, natural bioremediation can also act on both the original compound
and its chemical (or biological) products, leading in some cases to a mixture of
chlorinated compounds in the subsurface.
10.4. MICROBIOLOGICAL REACTIONS
Chlorinated solvents can undergo both substitution and elimination reactions
similar to those described above, yet mediated by microorganisms. However, the most
common reactions are those involving the transfer of electrons to or away from the
chlorinated solvent. These reactions, oxidation for removal of electrons from chlorinated
solvents and reductions for addition of electrons to chlorinated solvents, are dependent to
some extent on the degree of chlorination of the chlorinated solvent and upon the redox
conditions of the microorganisms (Vogel et al., 1987). The more highly chlorinated
solvents are more highly oxidized and, therefore, are more likely to undergo reduction
reactions. The less chlorinated solvents are less oxidized and are more likely to undergo
oxidation reactions. Redox conditions vary from the most oxidative (+800 mv), where
oxygen is present and aerobic microbes grow, to the most reduced, where neither oxygen,
nitrate nor sulfate exists, and methanogens grow (~ -350 mv). Hence, aerobic conditions
should be more suitable to the oxidation of less chlorinated solvents and methanogenic
conditions should be more suitable to the reduction of highly chlorinated solvents.
Oxidations of chlorinated solvents usually involve the addition of oxygen without the
removal of halogen in the first and rate-limiting step (Vogel et al., 1987). Subsequent
release of chlorine substituents might not actually be associated with oxidation reactions,
but with substitution reactions.
10.4.1. Aerobic
Most research to date has described the microbial oxidations of mono- or
dihalogenated aliphatic compounds. The major exception to this is the work done on the
oxidation of trichloroethylene (TCE). Several different microbes or microbial
enrichments have been shown to be capable of TCE oxidation (Fogel et al., 1986; Nelson et
al., 1986; Little et al., 1988) and chloroform oxidation (Strand and Shippert, 1986).
Apparently, the ease of oxidation increases with decreasing number of halogens. Hence,
dichloroethylene would be oxidized faster than TCE. Unfortunately, due to the nature of
contaminant release in the environment, mass balances are difficult to achieve, and no
strong evidence for the oxidation of halogenated solvents has been derived from actual
hazardous waste sites.
Highly chlorinated organic compounds are much more oxidized than many
natural organics. As such, these compounds do not provide much energy upon further
104
-------
oxidation in aerobic environments. Most aerobic biodegradation processes start with' a
step that involves the insertion of oxygen into a bond on the molecule. Due to the
electrophilic nature of that oxygen insertion, other electrophilic substituents (e.g.,
chlorine) hinder the reaction. Hence, the observation that increasing chlorination
within a homologous series often leads to a decrease in aerobic (oxidative) biodegradation
(Vogel et al., 1987).
Studies of the aerobic biodegradation of chlorinated compounds have illustrated
several major pathways of oxidation. These pathways resemble those for the
nonchlorinated homologues. For example, the oxidation of chlorinated ethylenes
involves the formation of a chlorinated epoxide similar to that for ethylene:
Example:
Cl H Cl O H
\ / \/\/
c = c —^ c-c
/ \ / \
H Cl H Cl (1)
The epoxide degrades rapidly in water. In both of these cases, the microbe that degrades
these compounds might require a natural nonchlorinated compound for growth and
energy. The enzymes produced for degradation of that "normal" substrate are also
capable of degrading the pollutant (cometabolism). The possibility that these microbes
would adapt to the use of chlorinated compounds as sources of energy and carbon exists,
but might have limited engineering applications. Selective pressure in natural
environments will not be great if pollutant concentrations are relatively low from a
microbial adaptation point of view, even if these concentrations are high from a
regulatory point of view.
In other aerobic degradation of lightly chlorinated compounds, microbes have
been shown to grow on the pollutant when it exists in sufficiently high concentration.
Most of these compounds are mono- or dichlorinated organics. A common reaction is the
microbially mediated substitution reaction where a hydroxyl group replaces a chlorine
(Brunneret al., 1980).
Example:
H
T T ? T
H-C -C-H > H-C-C-H
II II
H H H H (2)
After which, the compound is further oxidized and the metabolites enter the
anabolic and catabolic pathways of the microbe.
Although progress has been made in culturing aerobes capable of degrading
organic compounds with higher and higher degrees of chlorination, many highly
chlorinated compounds remain resistant due to their highly oxidized state. Examples of
aerobically recalcitrant chlorinated organics include tetrachloroethylene, which has not
-------
been observed to undergo epoxidation, and hexachlorobenzene, which have all carbons
occupied with chlorine substituents, allowing no site for hydroxylation. These highly
chlorinated organic compounds are not, however, resistant to anaerobic biodegradation
(Vogel and McCarty, 1985; Gibson and Suflita, 1986; Tiedje et al., 1987; Vogel and
McCarty, 1987a; Vogel, 1988; Freedman and Gossett, 1989; Bagley and Gossett, 1990; Nies
and Vogel, 1990; Bhatnagar and Fathepure, 1991).
10.4.2. Anaerobic
Under different anaerobic conditions, both in laboratory studies and in the
environment, highly chlorinated organics, such as tetrachloroethylene (Vogel and
McCarty, 1985), hexachlorobenzene (Gibson and Suflita, 1986), and polychlorinated
biphenyls (Nies and Vogel, 1990), have been shown to undergo reductive dechlorination.
Reductions of chlorinated solvents normally involve the removal of a chlorine substituent
and either its replacement with a hydrogen or removal of a second chlorine substituent
from alkanes and formation of an alkene. The first mechanism, commonly called
reductive dechlorination, can occur with both alkanes and alkenes. Reductive
dechlorination has been described for the sequence of ethylenes from tetrachloroethylene
to vinyl chloride (Vogel and McCarty, 1985) (Figure 10.1) and to ethylene (Freedman and
Gossett, 1989), and for TCA to chloroethane (Vogel and McCarty, 1987a) (Figure 10.2)
under methanogenic conditions in laboratory studies. In the case of TCA, potential
products are complicated by the chemical reactions (denoted by A) that co-occur with
biological reactions. Relative rate studies on the reductive dechlorination of various
chlorinated ethanes and ethenes have shown a general decrease in rate with decreasing
number of chlorine substituents, opposite to the trend shown for oxidation reactions
under aerobic conditions. The relative rates of reduction under methanogenic conditions
have been quantified in two cases (Table 10.2) (Bouwer and McCarty, 1988; Vogel, 1988).
From these data and that for the chemical reactivity described above, the disappearance
of an initial chlorinated solvent and the appearance of its products under favorable
anaerobic conditions might be derived, as will be discussed later.
1.2-DCE I
Cl H
CC12=CCI2—»»CHC1°CCI2 . CH2=CHCI ~»* CH^CH2 —+* 2CO2 •«• HCI
|PCE| |TCE| X H H f \Vmy\ Chlonde| | Ethylene [
Cl Cl
I 1.2-DCE~1
Figure 10.1. PCE anaerobic transformations.
1O6
-------
COj
Figure 10.2 Abiotic (A) and biotic (B) transformations of 1,1,1-trichloroethane.
TABLE 10.2. RELATIVE RATES OF DEGRADATION BY METHANOGENIC CULTURES
Compound
Suspended Growth"
Attached Growth11
Carbon Tetrachloride
Chloroform
Trichloroethylene
Acetate
1, 1-Dichloroethylene
1,1,1-Trichloroethane
1, 1 ,2-Trichloroe thane
1,2-Dichloroethane
1,1-Dichloroethane
Chloroe thane
Vinyl Chlonde
..
..
1.04
1.00
0.55
0.45
0.43
005
0.04
002
0.005
33
1.3
-
1.00
-
0.73
-
--
-
..
~~
« Vogel, 1988.
b Bouwer and McCarty, 1988
Possibly the observed reductive dechlorination of chlorinated solvents results from
the reduction-oxidation reaction between the highly oxidized chlorinated compound and
reduced compounds in the microbe. Bacteria contain a number of metal-organic
compounds, which are involved in electron exchange reactions. Many of these
metal-organic compounds, when in their reduced form, have been shown to dechlorinate
10-7
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chlorinated organics (Gantzer and Wackett, 1991; Assaf-Anid et al., 1992). The
metal-organic compound becomes oxidized as a result of this reaction, but might be
reduced again in the normal metabolic processes of the cell. In this case, the ultimate
source of electrons would be the organic substrate (e.g., acetate) that the anaerobe is
using for energy and carbon. The thermodynamics are very favorable for the oxidation of
metal-organics and concomitant reduction of chlorinated compounds. The fortuitous
nature of this reaction and the requirement for an ultimate external electron donor have
significant implications for bioremediation of these compounds.
On the other hand, anaerobic microorganisms might adapt to the use of a
chlorinated compound as an alternative terminal electron acceptor. If this occurs, the
chlorinated compound degradation would be directly tied to energy production in the cell.
Although this type of anaerobic dechlorination mechanism has been observed, it might
not be very common, due to typically low concentrations of dissolved chlorinated organics
in the environment.
10.5. PREDICTIONS OF PRODUCT DISTRIBUTION
The natural biodegradation of chlorinated solvents discussed above leads to
predictions of the natural bioremediation of these compounds. The next section describes
the outcome of natural anaerobic reductive dechlorination of the chlorinated solvents,
TCE and TCA, based on relative rates (Table 10.2). One of the major unknown aspects
regarding the microbial transformations of halogenated solvents in ground water is the
number and activity of appropriate microbes. Research is needed for understanding the
diversity of occurrence of these microorganisms. The simplest case is the addition of
PCE to an anaerobic aquifer. Assuming a particular microbial activity, PCE is
transformed to TCE, and then to 1,2-DCE, which is subsequently transformed to vinyl
chloride (VC) as described previously. A simple model using relative rates determined
in lab studies (Vogel, 1988, Table 10.2) can show the distribution of products over time or
distance in an active anaerobic subsurface zone (Figure 10.3). A more active microbial
population would increase the rate of transformations. For the case where TCA has
contaminated an anaerobic aquifer, the transformations are complicated by the potential
simultaneous chemical and microbial reactions. The simplest TCA case is where no
microbial activity exists at all. In this case, only acetic acid and 1,1-DCE are formed due
to hydrolysis and dehydrohalogenation, respectively (Figure 10.4). When some
methanogenic activity exists, TCA is partially reduced to 1,1-dichloroethane (DCA),
which is subsequently reduced to chloroethane, which can be, in some cases,
mineralized to carbon dioxide (Figure 10.5). Furthermore, the 1,1-DCE from the
dehydrohalogenation of TCA is reduced to vinyl chloride (VC) (Figure 10.5). Acetic acid
is also microbially mineralized partially to carbon dioxide by methanogens. If the
microbial activity increases, then the pathways are dominated by the microbial reactions,
not the chemical reactions. In this case, DCA, VC, and carbon dioxide (C02)
predominate after time (Figure 10.6). The most complicated case presented here is when
both TCE and TCA contaminate an anaerobic aquifer. When appropriate microbial
activity occurs, dichloroethylene isomers (1,2-DCE from TCE and 1,1-DCE from TCA) are
reduced to VC (Figure 10.7). TCA is reduced to DCA. Note that at one point in time or
space, the major products are DCA and the DCE isomers. Remember that if the ground
water should carry the mixture of chlorinated solvents out of an active microbial
(methanogenic) area, then most of the reductive dechlorination reactions would stop.
The ratios of chemical compounds would not change, except due to the influence of other
processes, such as aerobic degradation.
103
-------
Cd
8 10 12
Time (Years)
14
16
18
20
Figure 10.3. Reductive dechlorination of trichloroethylene (TCE) under hypothesized anaerobic
field or laboratory conditions.
0 2 4 6 8 10 12
Time (Years)
Figure 10.4. Chemical degradation of 1,1,1-trichloroethane (TCA).
10-9
-------
14
16
i
18
0 2 4 6 8 10 12
Time (Years)
Figure 10.5. Chemical and microbial degradation of TCA Gower microbial activity).
20
0
14
16
18
8 10 12
Time (Years)
Figure 10.6. Chemical and microbial degradation of TCA (higher microbial activity).
20
10-10
-------
lOOq
8 10 12
Time (Years)
14
I
16
18
20
Figure 10.7. Chemical and microbial degradation of both TCE and TCA,
10.6. RATIONALE FOR TECHNOLOGY
Many examples of the transformation of halogenated compounds under anaerobic
conditions by reductive dechlorination (Vogel and McCarty, 1985; Vogel and McCarty,
1987a; Vogel et al., 1987; Freedman and Gossett, 1989; Bagley and Gossett, 1990) have
been published, supporting the effectiveness of this first step for chlorinated solvent
degradation. Reductive dechlorination as described above is relatively rapid for
chemicals with a higher number of chlorine substituents, such as highly chlorinated
PCBs, hexachlorobenzene (HCB), perchloroethylene (PCE), trichlorethylene (TCE),
carbon tetrachloride (CT), chloroform (CF) and 1,1,1-trichloroethane (TCA) when
compared with their less chlorinated homologues (Tiedje et al., 1987; Vogel and McCarty,
1987a; Vogel et al., 1987; Bouwer and Wright, 1988; Fathepure et al., 1988). Upon
reduction, these polychlorinated compounds lose chlorine, and the resulting products
are usually more susceptible to hydrolytic and oxidative processes and less susceptible to
further reduction. These lower chlorinated compounds have been shown to be
successfully degraded by aerobic bacteria (Kuhn et al., 1985; de Bont et al., 1986; Schraa et
al., 1986; Strand and Shippert, 1986; Spain and Nishino, 1987; van der Meer et al, 1987;
Vogel et al, 1987; Henson et al., 1988). Therefore, the anaerobic/aerobic sequential
biodegradation of highly chlorinated compounds by indigenous microbes could occur and
should be encouraged.
In order for compounds to undergo natural anaerobic/aerobic sequential
environmental conditions, compounds would have to diffuse or flow from anaerobic
zones to aerobic zones. This could occur near sites that contain easily degradable
reduced organics, thus consuming oxygen near the source of contamination.
10-11
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This scheme requires the establishment of anaerobic conditions, followed by
aerobic conditions, which is not the normal ecological trend. This sequence can, in some
cases, be physically modeled spatially in a flow-through system (e.g., Figure 10.8) or in
chronological order with the same material.
Chloroform
Tetrachloroethylene
Hexachlorobenzene
o
. vH
PQ
u
s
CF
PCE
HCB
Dichloro-
benzene
Dichloro-
methane
Dichloro-
ethylene
CO2
Figure 10.8. Schematic illustrating the reductive dechlorination of polychlorinated compounds
in an anaerobic biofilm and subsequent mineralization of the products of
anaerobic treatment in an aerobic biofilm.
In most systems described, the anaerobic microbial conditions resulted in
dechlorination of the highly chlorinated organic compounds, although none of the
compounds were completely dechlorinated, with the exception of vinyl chloride going to
ethylene (Freedman and Gossett, 1989). Typically, mono-, di-, and trichlorinated
compounds remained after the anaerobic phase was complete (Table 10.3). In a physical
model, relative amounts of mono-, di-, and trichlorinated dechlorination products
varied depending on the organic substrate and nutrients fed to the system and the
hydraulic residence time, among other parameters (Fathepure and Vogel, 1991).
10-12
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TABLE 10.3. PRODUCTS OF ANAEROBIC DECHLORINATION
Initial Compound Major Products
Hexachlorobenzene -» di- and trichlorobenzene
Tetrachloroethylene -» 1,2-dichloroethylenes
Trichloromethane -» dichloromethane
Hexachlorobiphenyl -» trichlorobiphenyl
Aerobic degradation of mono- and dichlorinated organic compounds was fairly
rapid in several systems tested (Brunner et ah, 1980; Fogel et al., 1986; Nelson et al., 1986;
Henson et al., 1988; Walton and Anderson, 1990; Fathepure and Vogel, 1991). When
carbon-14 radiolabelled compounds were used to track the carbon, considerable amounts
of carbon-14 labelled carbon dioxide were produced (Table 10.4) (Walton and Anderson,
1990; Fathepure and Vogel, 1991). The most recalcitrant chlorinated solvents of those
that underwent aerobic degradation were the trichlorinated compounds. In some cases,
these compounds are (as were the mono- and dichlorinated compounds) the result of
reductive dechlorination of more chlorinated compounds under anaerobic conditions.
TABLE 10.4. PRODUCTS OF AEROBIC DEGRADATION*
Initial Compound Product(s)
Dichlorobenzene CO 2
Dichloroethylenes CO 2
Dichloromethane CO 2
a Fathepure and Vogel, 1991.
Anaerobic reductive dechlorination in these studies was dependent on organic
substrate and nutrients. The solution shown in Table 10.5 represents an example of
nutrients for anaerobic bacteria. In many subsurface environments, these nutrients
may already exist dissolved in ground water. Another critical addition is the reduced
organic that will supply energy and carbon for the growth of anaerobes. Further, it will
be the ultimate source of electrons for the reductive dechlorination of the chlorinated
organic compounds. In addition, its degradation by aerobes will also deplete oxygen,
keeping the region anaerobic. Products (such as methane) of the anaerobic degradation
of these substrates might provide substrates for aerobic microbes later.
10-13
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TABU: 10.5. PROPOSED NUTRIENTS PORBIOREMEDIATION
Compound Concentration (milligrams per liter)
(NH4)2HP04 80
NH4C1 1,000
K2HP04 200
NaCl 10
CaCl2 10
MgClo 50
CoCl2,6H2O 1.5
CuCl2,2H20 0.2
Na2MoO4, H2O 0.23
ZnCl2 0.19
NiS04,6H2O 0.2
FeS04,7H2O 1.0
A1C13,6H2O 0.4
HgBOg 0.38
Several important implications can be derived from these laboratory studies. The
sequential anaerobic/aerobic biodegradation of chlorinated organic compounds might be
a viable treatment technology, either naturally or induced. The anaerobic phase requires
the induction of active anaerobic metabolism (e.g., methane production) by a consortium
of anaerobes, not necessarily one specific microbe. This anaerobic consortium must be
supported by nutrients and organic substrate(s). If these compounds already exist in
solution, they need not be added. Due to the apparent lack of dependence of the anaerobes
on the chlorinated compounds for growth in many cases, the concentration of these
chlorinated solvents can theoretically be reduced to zero.
For the aerobic phase, oxygen (possibly in the form of hydrogen peroxide) must be
added or mixed in naturally in order to oxidize the anaerobic conditions. Viable aerobes
capable of degrading chlorinated compounds must either be present or be added. In
cases where the compounds are sufficiently high in concentration, these aerobes will use
the chlorinated compound as a source of energy and carbon. Hence, the theoretical
minimum concentration achievable would be the lowest concentration capable of
supporting the microbial community. Possibly, other substrates that will not
out-compete the chlorinated compound, but will aid in supporting the microbial
community, can be added or already exist. Additional substrates are required when the
chlorinated compounds are too low in concentration to induce enzymes or support
aerobic microbial growth. The sequential anaerobic/aerobic biodegradation of
chlorinated organic compounds in laboratory studies provides a sound, but not complete,
basis for field application testing.
Many cases of natural anaerobic reductive dechlorination of chlorinated solvents
in ground waters have been observed by consulting engineers, who question the apparent
production of less chlorinated products. In some cases, these products seem to disappear
once they reach aerobic zones. The research site at St. Joseph, Michigan, might provide
10-14
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evidence of this natural combination of anaerobic (McCarty and Wilson, 1992) and aerobic
processes. The plume contains high concentrations of methane and relatively low
concentrations of chlorinated solvents at the time it discharges to surface water. Natural
cooxidation of the chlorinated compounds during aerobic metabolism of the methane in
the oxygenated benthic sediments at the interface between the plume and surface water
is likely. However, the phenomenon has not been carefully documented.
To date, no organism has been found which can use the anaerobic metabolism of
halogenated solvents as the sole source of carbon and energy. Instead, microorganisms
generally require a primary source of carbon and energy. Although the reduction of a
halogenated solvent (the secondary substrate) is usually energetically favorable at
standard state, the organism may or may not benefit from this reduction due to inability
to control energy enzymatically or due to low concentrations leading to relatively little
available energy. The relationship may be fortuitous. A proposed mechanism for the
reductive dechlorination of halogenated solvents by methanogens, the group of
organisms most commonly cited as being responsible for reductions, involves the
transfer of electrons through the cell membrane during the anaerobic metabolism of
primary substrate (Zeikus et al., 1985). In this case, the halogenated solvent may divert
some of these electrons and use them for dechlorination. Hence, it is conceivable that
microbes might use the chlorinated solvent as an electron acceptor.
10.7. PRACTICAL IMPLICATIONS
The implication of a microbial degradation process that is dependent on the
primary substrate concentration, but not on the halogenated solvent concentration, is
that there is no lower limit to the final concentration of the halogenated solvent. If the
chlorinated compound served as a primary substrate, then fewer organisms would
survive when its supply became low, and further dechlorination would cease (McCarty,
1984). As a secondary substrate, the halogenated solvent can continue to be reduced by a
large, healthy population of bacteria grown on the primary substrate until the
halogenated solvent has been completely degraded.
Actual contamination is often a mixture of complex chemicals. Therefore, future
technologies aimed at bioremediation of halogenated solvents should achieve complete
destruction of all hazardous chemicals. Based on metabolic and kinetic limitations of
anaerobic and aerobic bacteria, a two-stage biological process consisting of an initial
anaerobic dechlorination of highly chlorinated chemicals followed by aerobic degradation
of the partially dechlorinated metabolites may effectively treat wastes containing complex
mixtures of chlorinated hydrocarbons.
In a given environment, any or all of the types of reactions discussed above may
occur. The conditions present, as well as the structure of the chlorinated compound,
dictate to a large extent the transformations that will predominate, and the expected
products. The only chemical reactions that have a significant effect on degradation
products are the hydrolysis of monochlorinated solvents and the dehydrohalogenation of
chlorinated polyalkanes. However, biological reactions can achieve rates with half-lives
as low as a few days, and may be significantly influenced by controlling environmental
conditions.
An important factor influencing biological degradation is whether the necessary
organisms are present. This should be determined before a full-scale remediation
scheme is begun by sampling the aquifer material in the area of the contaminated
ground water to be treated, and running laboratory-scale treatability studies.
10-15
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Microbial substrates and electron acceptors are the next factors to be considered.
In a given location, the biological transformations will be oxidation reactions, mediated
by aerobic organisms, as long as oxygen is present. Once oxygen has been depleted,
alternative electron acceptors such as nitrate and sulfate will be used. Finally, anaerobic
(even methanogenic) microorganisms will dominate and reduction reactions will take
place. The types of reactions that actually occur, however, may be dictated by the type
and amounts of substrates added to the ground water. Thus degradation conditions can
occur or be imposed according to the products desired.
In remediation of ground water, the choice of a treatment method should be based
on which type of reaction will be the most rapid and whether the expected products are
less hazardous than the original halogenated solvent. For example, reduction of PCE
leads to the production of vinyl chloride, which is a known carcinogen. However, the
above discussion suggests an efficient means of converting halogenated solvents to
nonhazardous compounds: sequencing anaerobic and aerobic treatments. PCE would be
reduced to TCE and DCE under anaerobic conditions, then an aerobic environment
would be provided in which the TCE and DCE would be oxidized to carbon dioxide.
Overall degradation rates would be maximized, and no vinyl chloride would be produced
(Fathepure and Vogel, 1991).
10.8. SPECIAL REQUIREMENTS FOR SITE CHARACTERIZATION
Considerable site characterization of parameters directly related to in-situ
biological activity, in addition to other site characteristics, is required (Table 10.6). As
natural bioremediation of chlorinated compounds is controlled in part by the natural
redox of the ground water, sites amenable to anaerobic reductive dechlorination require
information regarding the ability of indigenous microbes to undergo anaerobioses. An
example of a specific characteristic is the dissolved oxygen concentration in the ground
water at the site. This information is critical to understanding which type of microbial
community is dominant. Clearly, measurement of chemical parameters directly related
to microbial activity and not just analyses of priority pollutants is critical for evaluating
the likely success of natural bioremediation. For example, as listed in Table 10.6, pH,
temperature, ionic strength, presence or absence of heavy metals and of potential
electron acceptors all play important roles in determining the type and extent of
microbial activity. Laboratory tests that evaluate microbial activity and potential toxicity
for a given site will aid in determining potential for natural bioremediation. These
methods might measure C-14 carbon dioxide or methane production or, in the case of
anaerobic conditions, the production of dechlorinated products.
10.9. FAVORABLE SITE CHARACTERISTICS
Since highly chlorinated solvents (e.g., tetrachloroethylene: PCE) do not appear to
undergo aerobic degradation, the degree of chlorination of the solvent is critical for
differentiating between whether aerobic (less chlorinated solvents) or anaerobic
conditions will be effective. The degree of chlorination will also, to some extent, control
the sorption onto organic matter in the aquifer and, thus, the retardation of the solvent
through the aquifer or soil. The relative retention of the solvent affects the practicality of
potential natural bioremediation, as will be discussed below.
10-16
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TABLE 10.6. SOME INFORMATION NEEDED FOR PREDICTION OF ORGANIC CONTAMINANT
MOVEMENT AND TRANSFORMATION IN GROUND WATER"
Hydraulic
Contaminant Source Wells
location
amount
rate of release
location
amount
depth
pump rates
Hydrogeologic Environment
extent of aquifer and
aquitard
characteristics of aquifer
hydraulic gradient
ground-water flow rate
Sorption
Chemical
Biological
Distribution
Coefficient
characteristic of
concentration
Ground-water
Characteristics
ionic strength
pH
temperature
toxicants
Ground-water
Characteristics
ionic strength
pH
temperature
nutrients
substrate
macro (P, S, N)
trace
organism
concentration
distribution
type
Characteristics of
the Aquifer Solid
organic carbon content
clay content
Aquifer
Characteristics
potential catalysts:
metals, clays
Aquifer
Characteristics
grain size
active bacteria -
number
Monod rate - constants
Contaminant
Characteristics
octanol/water partition
coefficient
solubility
Contaminant
Characteristics
potential products
concentration
Contaminant
Characteristics
potential products
toxicity
concentration
a Vogel, 1988
10-17
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The characteristics of the contaminated site also affect the relative mobility of the
chlorinated solvents and the potential for natural bioremediation. Hydrologic and
geologic conditions that allow for microbes, chlorinated solvents, and necessary
nutrients to occur in the same locations at the site will improve the probability of success.
The strong sorption of solvents on organic-rich aquifer or soil material will make these
compounds less available to microbial activity and, thereby, possibly slow rates. On the
other hand, porous subsurface materials, such as sands, are better for ground-water
flow but often have less associated microbial activity than organic-rich biodegradation
materials and might not be conducive to anaerobic conditions.
The interplay between environmental conditions in the subsurface, proposed
remedial actions, and regulatory requirements renders the natural bioremediation of
halogenated solvents difficult to pursue. Yet, the possibility for implementing aerobic or
anaerobic/aerobic sequential systems successfully are enhanced by relatively porous
subsurface material, no heavy metal toxicity, easy access to potential nutrients, ability to
reinject recovered contaminated ground water, and hydrogeologic control over the
contaminated plume.
10.10. UNFAVORABLE SITE CHARACTERISTICS
Site characteristics that would make the natural bioremediation of halogenated
solvents difficult are those that hinder the initiation of appropriate microbial conditions
and activities, and those characteristics that prevent the commingling of pollutant with
active microbes. As an example, although heavily chlorinated solvents tend to sorb more
strongly than less chlorinated solvents, they still undergo reductive dechlorination more
rapidly than the less chlorinated solvents under active methanogenesis. For lightly
chlorinated solvents, sorption is less important. They tend to undergo degradation under
aerobic versus anaerobic conditions (Figure 10.9). Toxic heavy metals, high sulfide
concentrations, and lack of appropriate nutrients are chemical characteristics that will
negatively affect the natural bioremediation of chlorinated solvents. Hydrologic and
geologic characteristics that would not benefit the natural bioremediation of chlorinated
solvents include fractured rock systems where small microbial populations exist.
Increasing these populations can be difficult, although development of subsurface
biofilms with continuous recycling of ground water with appropriate nutrients added
might be effective. As mentioned above, physical systems or regulations that prevent
mixing of chlorinated solvents and nutrients will hinder or prevent natural
bioremediation of chlorinated solvents.
Given the potential difficulties and lack of information regarding site
characteristics, the natural bioremediation of chlorinated solvents, even highly
chlorinated compounds such as PCE, is still likely at many sites. Indeed, throughout the
USA, sites that were initially contaminated with only PCE, TCE, or TCA have shown
active dechlorination patterns near the center of contaminated plumes forming the
reductive dechlorination products. In some cases, these products appear to be somewhat
degraded, as they migrate into aerobic zones. The limitations on these unaided natural
processes are currently unknown. Several possibilities exist, such as low microbial
activity, for which nutrient addition might be increased.
As mentioned earlier, the extent of degradation in using this approach can
theoretically be complete. Since, in most cases, the microbes are not utilizing the
chlorinated solvents as food or substrate, microbes could degrade the last molecule,
assuming contact exists. The microbes need to be supported on the appropriate
substrates and nutrients and be given sufficient opportunity for contact with the
10-18
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chlorinated solvent. Cometabolism defined in this way provides tremendous potential for
eventually degrading pollutants to levels below detection.
C8
V)
o
o
C/3
cs
at
00
a
Sorption
Reductive
Dechlnnnanon
Rale
Degree of Chlorination
Monochlorinated
0.25
Polychlorinated
Figure 10.9. Relationships between degree of chlorination and anaerobic reductive
dechlorination, aerobic degradation and sorption onto subsurface materiaL
One critical aspect of the natural bioremediation of chlorinated solvents is the rate
at which they are degraded. As was mentioned before, the rate is dependent on which
types of chlorinated solvent and microbial community interact. Reductive dechlorination
rates (estimated as first-order constants or k/ks in Monod terms) suggest that the
dechlorination of highly chlorinated compounds might occur in active anaerobic
conditions within one year (Vogel, 1988). This rapid microbial dechlorination rate would
undoubtedly be controlled by other factors such as sorption/desorption, availability of
nutrients, temperature, etc.
A major difficulty with this approach is the lack of predictability of the activity of
in-situ microbial communities. In laboratory studies, mainly unreported, some active
anaerobic communities have little ability to dechlorinate these solvents. However, in
other studies, anaerobic conditions induced in soil and aquifer material samples
exhibited reductive dechlorination. Our lack of understanding of the important members
of anaerobic microbial communities for reductive dechlorination of chlorinated solvents
lends uncertainty to this technology. For aerobic conditions, research has also shown
that not all aerobes are active toward lightly chlorinated solvents. An example of this
was the lack of TCE-oxidizing ability in all microbes that oxidized aromatic compounds
(Nelson et al., 1987; Nelson et al., 1988). Two remedies exist: the first would involve
changing the subsurface environmental conditions to activate and grow the appropriate
microbes; the second is the addition of selected microbes to the subsurface.
10-19
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10.11. COST EVALUATIONS
The difficulties in assessing cost and factors that affect costs include the lack of
demonstration sites of these processes. Natural unaided bioremediation of chlorinated
solvents imposes little or no costs other than the time for the natural processes to
proceed. The facilitated natural bioremediation requires control of the subsurface
hydrological regime sufficiently to mix nutrients, solvents, and microbes. In addition,
the costs of nutrients like those listed in Table 10.5, probably will not be significant, in
most cases, although electron donors (e.g. methanol) and acceptors (e.g. hydrogen
peroxide) might have significant costs associated with them.
As described previously, this natural bioremediation has the potential to reduce
the concentrations of chlorinated solvents below detection levels, but would probably
require either permission to reinject/recirculate contaminated water or to leave the site to
slowly proceed without intervention, assuming adequate anaerobic and aerobic zones.
10.12. KNOWLEDGE GAPS
Unfortunately, without expanding our knowledge regarding the general ability of
anaerobes to reductively dechlorinate those solvents, the effects of temperature, pH, ionic
strength, soil or aquifer type on these subsurface microbial communities and the
environmental distribution of the appropriate microbes, further development of this
technology will be limited to sites where intensive treatability studies have shown it to be
appropriate. Indigenous anaerobic microbes' dechlorination abilities and kinetic
coefficients need to be evaluated. Implementation of this technology at study sites where
conditions seem appropriate will aid in developing experience.
Clearly, research to date has demonstrated the potential for both anaerobic and
aerobic degradation or transformation of chlorinated solvents in laboratory studies.
However, research that addresses the problems associated with moving a technology
from well-controlled environments into nature is lacking for these processes. Some of
these studies must be undertaken by microbial ecologists who can evaluate both the
biological potential and spatial distribution in nature. Other studies need to be performed
by hydrogeologists who can evaluate the ability to induce anaerobic conditions throughout
a contaminated site. Finally, risk associated with the natural bioremediation, both the
process and possible products, needs to be evaluated.
10.13. CONCLUSION
When a chlorinated solvent is introduced into the environment, it may be
transformed by chemical and biological reactions into a variety of products. These may
be more or less hazardous than the original chlorinated solvent. Although chemical
transformations may be quite slow, biological reactions often proceed quickly. The types
of microbial conversions and the resulting products will depend on the chlorinated
solvent and environmental conditions. An understanding of these transformations
provides an insight into the natural processes and methods for producing conditions that
will maximize degradation rates and lead to the conversion of chlorinated solvents to
compounds that are not hazardous to human health. One such process might be
anaerobic/aerobic degradation of chlorinated solvents.
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REFERENCES
Anders, M.W. 1983. Bioactivation of halogenated hydrocarbons. J. Toxicol.-Clin.
Toxicol. 19:699-706.
Assaf-Anid, N. L. Nies, and T.M. Vogel. 1992. Reductive dechlorination of a
polychlorinated biphenyl congener and hexachlorobenzene by vitamin B12. Appl.
Environ. Microbiol. 58(3):1057-1060.
Bagley, D.M., and J.M. Gossett. 1990. Tetrachloroethene transformation to
trichloroethene and cis-l,2-dichloroethene by sulfate-reducing enrichment
cultures. Appl. Environ. Microbiol. 56(8):2511-2516.
Bhatnagar, L., and B.Z. Fathepure. 1991. Mixed cultures in detoxification of hazardous
wastes. In: Mixed Cultures in Biotechnology. Eds., G. Zeikus and E.A. Johnson.
McGraw-Hill, Inc. New York. pp. 293-340.
Bouwer, E.J., and J.P. Wright. 1988. Transformations of trace halogenated aliphatics in
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compounds in acetate-grown biofilms. Biotech. Bioengr. 27:1564-1571.
Brunner, W., D. Staub, and T. Leisinger. 1980. Bacterial degradation of dichloroethane.
Appl. Environ. Microbiol. 40(5):950-958.
de Bont, J.A.M., M.J.A.W. Vorage, S. Hartmans, and W.J.J. van den Tweel. 1986.
Microbial degradation of 1,3-dichlorobenzene. Appl. Environ. Microbiol. 52(4):
677-680.
Fathepure, B.Z., J.M. Tiedje, and S.A. Boyd. 1988. Reductive dechlorination of
hexachlorobenzene to tri- and dichlorobenzenes in anaerobic sewage sludge.
Appl. Environ. Microbiol. 53(2):330-347.
Fathepure, B.Z., and T.M. Vogel. 1991. Complete degradation of polychlorinated
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Fogel, M.M., A.R. Taddeo, and S. Fogel. 1986. Biodegradation of chlorinated ethenes by
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Freedman, D.L., and J.M. Gossett. 1989. Biological reductive dechlorination of
tetrachloroethylene and trichloroethylene to ethylene under methanogenic
conditions. Appl. Environ. Microbiol. 55(9):2144-2151.
Gantzer, C.J., and L.P. Wackett. 1991. Reductive dechlorination catalyzed by bacterial
transition-metal coenzymes. Environ. Sci. Technol. 25(4):715-722.
Gibson, S.A., and Suflita, J.M. 1986. Extrapolation of biodegradation results to
groundwater aquifers: Reductive dehalogenation of aromatic compounds. Appl.
Environ.Microbiol. 52(4):681-688.
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Henson, J.M., M.V. Yates, J.W. Cochran, and D.L. Shackleford. 1988. Microbial
removal of halogenated methanes, ethanes, and ethylenes in aerobic soil exposed
to methane. FEMS Microbiol. Ecol. 53:193-201.
Kuhn, E.P., P.C. Colberg, J.L. Schnoor, O. Wanner, A.J.B. Zehnder, and R.P.
Schwarzenbach. 1985. Microbial transformations of substituted benzenes during
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Sci. Technol. 19(10):961-968.
Little, C.D., A.V. Palumbo, S.E. Herbes, M.E. Lidstrom, R.L. Tyndall, and P.J. Gilmer.
1988. Trichloroethylene biodegradation by a methane-oxidizing bacterium. Appl.
Environ. Microbiol. 54(4):951-956.
Mabey, W.R., V. Barich, and T. Mill. 1983. Hydrolysis of polychlorinated alkanes,
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Environmental Chemistry. Ed., O. Hutzinger. Vol. 3, Part B. Springer-Verlag.
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Phelps, T.J., J.J. Niedzielski, R.M. Schram, S.E. Herbes, and D.C. White. 1990.
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ethenes and ethanes by anaerobic microorganisms. Amer. Chem. Soc. Ann.
Mtg., Extended Abstract. 344-346.
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SECTION 11
INTRODUCED ORGANISMS FOR SUBSURFACE BIOREMEDIATION
J. M. Thomas and C. H. Ward
Rice University
National Center for Ground Water Research
Department of Environmental Science and Engineering
Houston, Texas 77251
Telephone: (713)527-4086
Fax: (713)285-5203
11.1. FUNDAMENTAL PRINCIPLES OF THE TECHNOLOGY
Inocula of microorganisms have been widely used for bioremediation of
hazardous waste sites. There is little documentation of the efficacy of this process, and
important questions still persist about the environmental responsibility of adding
nonindigenous microorganisms. This section reviews the properties of the subsurface
and the properties of microorganisms that influence their transport through geological
material, their survival, and their capacity to degrade contaminants.
11.1.1. Review of the Development of the Technology
The concept of microbial movement through the subsurface was first addressed as
early as the mid-1920s for microbial enhanced oil recovery (MEOR). At that time,
Beckmann (1926) suggested that microorganisms that produce emulsifiers or
surfactants could be transported into an oil-bearing formation to recover oil that remains
after a well has stopped flowing. The addition of microorganisms to oil-bearing
formations to enhance oil recovery by biosurfactant or biogas production has since been
investigated and appears promising (Bubela, 1978). At about the same time, research on
the transport of microorganisms through the subsurface environment was being
conducted to determine the effectiveness of on-site wastewater disposal systems (i.e. pit
latrines, septic tanks, land disposal of sewage) in removing pathogens (Caldwell, 1937,
1938). More recently, the concept of transporting microorganisms with specialized
metabolic capabilities for subsurface bioremediation has been proposed (Leeet al., 1988;
Thomas and Ward, 1989).
The addition of microorganisms to the subsurface in remedial operations would be
beneficial when contaminants resist biodegradation by the indigenous microflora, where
evidence of toxicity exists, or when the subsurface has been sterilized by the
contamination event. Seed microorganisms have been added to the subsurface to aid in
contaminant biodegradation; however, the role of the added microorganisms has never
been differentiated from that of the indigenous microflora (Lee et al., 1988; Thomas and
Ward, 1989). Operations in which seed organisms are added to enhance contaminant
biodegradation in the subsurface usually involve treating contaminated ground water in
a closed-loop system by withdrawal and treatment in an aboveground bioreactor or by
physical methods, after which the treated ground water is reinjected into the subsurface.
The treated ground water that is reinjected contains adapted microorganisms from the
bioreactor or is amended with contaminant-degrading organisms to enhance
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biodegradation in situ (Ohneck and Gardner, 1982; Quince and Gardner, 1982a, b;
Winegardner and Quince, 1984; Flathman and Githens, 1985; Flathman and Caplan,
1985, 1986; Flathman et al., 1985; Quince et al., 1985).
For added microorganisms to be effective in contaminant degradation, they must
be transported to the zone of contamination, attach to the subsurface matrix, survive,
grow, and maintain their degradative capabilities (Thomas and Ward, 1989). When
injected into a nonsterile formation, the added organisms must compete with the
indigenous microflora for limiting nutrients and escape predation. Transport will
depend upon complex interactions between the subsurface and the microorganism.
Physical phenomena related to the composition of the subsurface formation that affect
transport include filtration and adsorption (Gerba and Bitton, 1984). Passage through
the subsurface will depend on grain size and related values of hydraulic conductivity (K)
and channels made by cracks and fissures. However, transport by channeling probably
will result in uneven seeding of the formation. Obviously, organisms that are larger
than the average pore size cannot move with ground water and will be retained by aquifer
solids. In addition, microbial cells may be removed from solution by sorption in
sediments high in clay and organic matter.
One of the first studies that addressed microbial transport through subsurface
materials for the purpose of contaminant degradation was published by Raymond et al.
(1977). These investigators reported that heterotrophs and hydrocarbon-degrading
bacteria penetrated and were detected in the effluent of 1.45 x31 cm columns packed with
unconsolidated sands having effective hydraulic conductivity (K) values ranging from
3.38 x 1O3 to 1.9 x 1CH cm/sec, which were run at a flow rate of about 30 ml/h (Darcy flow
18/cm/hr). Microorganisms also penetrated and were detected in the effluent of 3.8 x 10
cm sandstone (consolidated) cores, with hydraulic conductivities ranging from 1.8 x 10-5
to 7.2 x 10-5 cm/sec, through which water was passed under unknown pressure. In a
separate experiment, it was determined that the added microorganisms were utilizing
the gasoline.
11.1.2. Matrix Properties That Affect Transport
Bioremediation of the subsurface is usually limited in subsurface material with
hydraulic conductivities less than 104 cm/sec, because of the difficulty in pumping fluids
through material with lower K values. Hence, for practical purposes, microbial
transport through the subsurface to enhance bioremediation will probably be limited to
material with hydraulic conductivities of 10-4 cm/sec or greater. Laboratory studies
using materials that have been screened or sieved has produced a distorted view of
microbial transport in geological materials; the transport of microbial cells with the flow
of ground water has been underestimated. In-situ geological materials have a more
heterogeneous distribution of pore sizes than laboratory simulations. As a practical
consequence, microbial inocula can move readily through the larger pores in many
subsurface environments.
Several studies have been conducted to determine the effect of hydraulic
conductivity on microbial transport for selective plugging of subsurface materials for
MEOR. Hartet al. (1960) reported that the plugging effect of injecting water containing
1.2 x 106 dead bacteria/ml through 2.5 x 8 cm sandstone (consolidated) cores maintained
at an input pressure of 40 psi were not different at hydraulic conductivities ranging from
1.2 to 2.9 x KM cm/sec. Kalish et al. (1964) found that when 1 x 106 dead cells/ml of a
Pseudomonas aeruginosa strain was transported through 2.54 x 5.08 to 10.16 cm
sandstone cores at constant flow rate, core plugging was inversely related to K at
hydraulic conductivities ranging from 3 x 10-5 to 2.8 x 10-4 cm/sec. Bubela (1978) reviewed
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several papers on MEOR and found that secondary recovery in formations with K values
of 9.66 x 10'5 cm/sec or greater could be enhanced using microbiological techniques;
however, recovery declined or was insignificant in less permeable formations.
Jenneman et al. (1985) found that the rate of penetration of a motile Bacillus sp. through
nutrient-saturated sandstone cores under static conditions was independent of hydraulic
conductivity at values above 9.66 x 1O6 cm/sec but rapidly decreased for cores with K
values below it.
More recent studies have been conducted to determine the effect of K on microbial
transport for enhancement of subsurface bioremediation. Marlow et al. (1991) reported
that extent of transport of a yeast, Rhodotorula sp., after 10 pore volumes through sand
columns with hydraulic conductivity values of 5.59 x 1O2 and 1.37 x 1O1 cm/sec was about
2 and 50%, respectively, of the initial number of cells added (1 to 2 x 105 cells/ml). Fontes
et al. (1991) investigated the effects of grain size, bacterial cell size, ionic strength of the
transporting fluid, and heterogeneities of the medium on microbial transport and found
that grain size was the most important variable. When transporting a gram-negative
coccus in a low ionic strength fluid at a flow rate of 88 ml/h (Darcy flow 4.9 cm/hr)
through sand packed in a 4.8 x 14 cm column to achieve K values of 2.0 cm/sec and 0.37
cm/sec, 88.35 and 14.50% of the cells, respectively, were recovered in the effluent.
The mineralogy of the matrix may also affect transport. Scholl et al. (1990) found
that bacterial attachment, and hence removal from solution, may be affected by the
surface charge on the minerals present. These authors reported that attachment of
bacteria (negatively charged) isolated from ground water was greater for limestone, iron
hydroxide coated quartz and iron hydroxide coated muscovite (positively charged) than
for clean quartz and clean muscovite (negatively charged) in batch experiments.
Attachment to coated muscovite was greater than that to coated quartz whereas
attachment to clean muscovite and quartz was not different. When bacterial cells (1.77 x
109 cells/ml) were transported through columns (2 x 20 cm) packed with coated or
uncoated quartz at a flow rate of 21 ml/h (Darcy flow 6.7 cm/hr), 99.9 and 97.4% of the
cells, respectively, were retained in the columns.
Another matrix property that affects transport is sediment structure. Smith et al.
(1983) reported that Escherichia coli was transported to a greater extent through intact
cores than through cores of disturbed or structureless soil. In addition, movement
through intact cores appeared to be related to the presence of macropores and
channeling. For intact cores, there was no relationship between clay and organic matter
content and the extent of transport. These authors suggested that the use of studies in
which the porous material is sieved and then packed homogeneously in columns to
determine microbial transport will not be predictive of transport in situ because the
natural pores and channels will be destroyed.
Madsen and Alexander (1982) reported that vertical transport of Rhizobium
japonicum and Ps. putida was facilitated by percolating water, plant roots and
percolating water, and a burrowing earthworm. Neither species of bacteria was
transported further than 2.7 cm below the surface without facilitation. Transport
through channels was thought to be the most important mechanism for microbial
movement.
To summarize, matrix properties that will affect transport include hydraulic
conductivity, mineralogy, and sediment structure. Hydraulic conductivity was the most
studied parameter affecting transport through porous media; however, the results of
laboratory studies in which samples of porous media were packed to homogeneity may
produce underestimates of microbial transport. The use of intact cores will provide the
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full range of pore sizes present in situ for microbial transport through available
macropores.
11.1.3. Properties of Organisms that Affect Transport
Properties of organisms that affect transport include size, shape, stickiness,
condition, and motility and chemotaxis. In general, no one property has a dominant
influence on transport of microbial cells. The relative influence of organismal properties
is less important than the properties of the geological matrix, or operational factors such
as cell density, or chemical properties of the ground water.
Kalish et al. (1964) investigated the effect of cell size and aggregation tendencies of
Ps. aeruginosa (0.5 x 1.5 um; no aggregation), Micrococcus roseus (0.8 am; occurs
singly or as aggregates), and B. cereus (1x6 um; occurs singly or as long chains) on
their ability to plug sandstone cores with high (2.6 to 3.3 x 104 cm/sec) and low (2.1 to 2.9
x 10-5 cm/sec) hydraulic conductivity at initial cell densities of 1 x 105 and 1 x 10* dead
cells/ml, respectively. The authors found that the aggregation tendency of the cells was
more important than cell size in causing a reduction in hydraulic conductivity. Ps.
aeruginosa, which is intermediate in size, caused the least amount of plugging; M.
roseus which is the smallest and occurs as single cocci or as aggregates, caused more
plugging, while B. cereus, which is the largest and occurs as single rods or in chains,
caused the most plugging. However, when Ps. aeruginosa and another nonaggregating
but larger bacterium, Proteus vulgaris (0.5 to 1.0 x 1 to 3 um) were transported through
sandstone cores with similar hydraulic conductivity, the larger organism, P. vulgaris,
caused the most plugging.
Gannon et al. (1991) investigated the transport of the rod-shaped organisms
Enterobacter, Pseudomonas, Bacillus, Achromobacter, Flavobacterium, and
Arthrobacter strains through 10 x 5 cm columns packed with a loam soil at a flow rate of
2.5 cm/h; cells with lengths less than 1 um were transported to a greater extent than
larger cells. The presence of capsules and the hydrophobic nature and the net surface
electrostatic charge of the cell, properties that may affect sorption, did not influence
transport.
Fontes et al. (1991) transported a gram-negative coccus (approximately 0.75 um in
diameter) and gram-negative rod (approximately 0.75 x 1.8 um) with similar
hydrophobicities through columns packed with unconsolidated sand with hydraulic
conductivities of 0.37 and 2.0 cm/sec at a flow rate of 88 ml/h; the coccus was transported
to a greater extent than the rod. Jang et al. (1983) transported Ps. putida, Clostridium
acetobutylicum spores, and vegetative cells and spores of B. subtilis through 2.54 x 7.62
cm sandstone cores with a hydraulic conductivity of 3.9 x 103 cm/sec at a flow rate of 40
ml/h (Darcy flow 7.9 cm/hr); Cl. acetobutylicum spores were transported to the greatest
extent. In addition, B. subtilis spores were transported to a greater extent than were the
vegetative cells. Another property that may affect transport is the condition of the cell.
MacLeod et al. (1988) reported that starved cultures of Klebsiella pneumoniae, which
were smaller and less sticky than vegetative cells, were transported through artificial
(glass beads) rock cores to a greater extent than the vegetative cells.
Motility and chemotaxis (the ability of a cell to detect and move with substrate
gradients) may be important in the movement of microorganisms to contaminants
localized in the subsurface. Jenneman et al. (1985) compared two taxonomically similar
strains, En. aerogenes, which is motile, and K. pneumonia, which is nonmotile. The
motile strain penetrated nutrient-saturated sandstone cores of similar length and
hydraulic conductivities (4.5 to 6.1 x 10~* cm/sec) 3 to 8 times faster than the nonmotile
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strain under nonflow conditions. In addition, penetration of either strain was not related
to hydraulic conductivity within the ranges tested.
Using isogenic strains of E. coli under anaerobic conditions, researchers from
this same laboratory found that motility, but not chemotaxis, may be important in the
penetration of cells through unconsolidated porous media (Reynolds et al., 1989). Mutant
nonchemotactic motile strains penetrated 2.01 x 8 cm nutrient-saturated sand cores
faster than a chemotactic motile strain (wildtype) and nonmotile mutants, under static
conditions. In addition, a motile strain that was chemotactic toward but unable to utilize
galactose, penetrated sand cores at the same rate in the presence or absence of a
galactose gradient. Furthermore, nonmotile strains of E. coli which produced gas
penetrated nutrient-saturated cores about 5 to 6 times faster than mutant nongas-
producing strains, suggesting that gas production is important for movement of
nonmotile cells. For both motile and nonmotile strains, penetration rates were directly
related to growth.
To summarize, the properties of the microorganisms that may affect transport
include size, aggregating tendencies, shape, condition, and motility and chemotaxis.
The results of studies designed to investigate which cell characteristics are most
important are mixed. Cell size is important in that transport will be limited or prevented
for cells that are bigger than the average pore size; however, cells that tend to aggregate,
even if they are small, will not be good candidates for transport. For microorganisms
that form spores, the spore, which is smaller than the vegetative stage, may be
transported more efficiently. Microorganisms that are in a starved state usually are
smaller and produce less extracellular polysaccharide, which allows the organism to
attach to surfaces. Thus the reduced size and stickiness of the cells should enhance
transport. Finally, motility may enhance transport.
11.1.4. Operational Factors that Affect Transport
The most important operational factor is the ionic strength of the water used to
introduce the microorganisms into the subsurface. Transport is greatly facilitated in
water with low ionic strength. The concentration of microbial cells or spores may also
affect the rate and extent of transport through subsurface materials. At high cell
densities, filtration of cells can significantly reduce hydraulic conductivity. At low cell
density, the cells sorb to aquifer matrix materials to a greater extent. Transport is also
related to the rate of flow of water. A greater proportion of cells are transported in water
that is moving rapidly.
Hart et al. (1960) reported on the effect of injection concentration of cells on
hydraulic conductivity of consolidated sandstone cores (2.5 x 8cm; K= 9.66 x 1O5 cm/sec)
maintained at an input pressure of 40 psi. Hydraulic conductivity decreased as the
injection concentration of cells increased from 1.2 x lOSto 1.2 x 107 cells/ml. Kalish et al.
(1964) also found that the injection concentration of dead cells of Ps. aeruginosa and M.
luteus at concentrations ranging from 1 x 106 to 20 x 106 and 1 x 105 to 1 x 106 cells/ml,
respectively, was inversely related to the final hydraulic conductivity of sandstone cores
through which the cells were transported at constant flow rate (initial K = 3 x 1O4
cm/sec). The same trend of decreasing K with increasing cell concentration was
observed when Ps. aeruginosa was transported through high (3 x 10-* cm/sec) and low
(1.5 x 10-5 cm/sec) permeability sandstone. At influent concentrations of 5 x 107 but not 1
x 106 cells/ml, Jang et al. (1983) observed formation of a filter cake at the inlet and a
pressure drop along the core when Ps. putida was injected under nongrowth conditions
through 2.54 x 7.62 cm sandstone cores (K = 3.9 x 10-3 cm/sec; flow rate of 40 ml/h). A
filter cake did not form at influent concentrations of 1 x 106 cells/ml
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In contrast to studies that indicate an inverse relationship between transport and
cell density, Smith et al. (1983) found that cell densities ranging from 105 to 108 cells/ml
had no effect on the extent of transport of E. coli through several different intact cores.
These authors speculated that these relatively large concentrations of cells could not
saturate the adsorption and filtration sites of these soils. Reynolds et al. (1989) found that
the penetration rate of an E. coli strain through 2.01 x 8 cm sand cores increased in
proportion to the logarithm of cell concentration. Bacterial concentrations between 101 to
107 cells/ml were tested. These authors speculated that factors that may inhibit cell
movement through porous media on a finite basis, such as sorption, may be negated by
increasing the initial number of cells added.
Gannon et al (1991) also found a direct relationship between transport and cell
density. Bacterial cells were transported in deionized water or a 0.01 M NaCl solution
through columns (5 x 30 cm) packed with sandy aquifer material at a Darcy flow rate of
lO-4 cm/sec. An increase in injected cell density from 10s to 109 cells/ml increased the
total recovery of cells transported in deionized water and the salt solution from 44 to 57%
and 1.5 to 44%, respectively. The authors suggested that the small increase in recovery of
cells injected in deionized water resulted from the lack of appreciable adsorption under
conditions of low ionic strength. However, the large difference in recovery of cells
transported in the salt solution, a condition of high ionic strength which favors sorption,
was the result of a smaller percentage of cells at the higher density that was retained by a
finite number of sorption sites.
Although injection concentration may physically retard or enhance transport,
injection concentration or inoculum size will be important in initiation and maintenance
of biodegradation. Ramadan et al. (1990) found that inoculum size affected the
biodegradation potential of bacteria inoculated into lake water. Inoculation of Ps. cepacia
into lake water resulted in mineralization of 1 ug/ml p-nitrophenol at concentrations of
3.3 x 10« and 3.6 x 105 cells/ml but not at 330 cells/ml. The absence of biodegradation at
low cell concentrations was a result of protozoa that grazed the population to
nondetectable levels. Grazing by protozoa could significantly reduce the number of
naturally occurring and/or introduced contaminant-degrading organisms and affect the
rate and extent of bioremediation. Sinclair (1991) reported that 100 or less eucaryotes
were detected in samples from two uncontaminated sites; however, large numbers of
protozoa (2.66 x lOVgram dry weight) were detected in samples contaminated with jet
fuel, aviation fuel, and creosote in which sufficient organic carbon was present to
support high numbers of bacteria (Sinclair, 1991; Madsen et al., 1991).
Flow rate is another factor that may affect transport of microorganisms through
the subsurface. In all published reports, an increase in flow rate increased transport.
Kalish et al. (1964) found that reductions in hydraulic conductivity in sandstone as a
result of plugging by suspensions of dead cells of P. vulgaris could be partially reversed
by increasing the flow rate, which concomitantly increased the pressure differential
across the core. Smith et al. (1983) reported a direct relationship between flow rate and
the extent of transport of E. coli through intact cores. By increasing the flow rate from 0.5
to 4 cm/h, the extent of transport in a silt loam increased six times. Marlow et al. (1991)
reported that transport of Rhodococcus sp. through sand packs with K = 1.37 x 10-1
cm/sec was facilitated by increasing the injection rate; increasing the flow rate by a
factor of two nearly doubled the number of cells transported through the column.
Gannon et al. (1991) transported bacterial cells (108 cells/ml) in deionized water or a 0.01
M NaCl solution through a column (5 x30 cm) packed with sandy aquifer material. Flow
rates ranged from 1 x 10-2 to 2 x 10-2 cm/sec. Doubling the rate of flow increased the total
11-6
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recovery of cells transported in deionized water and on 0.01 M NaCl solution,
respectively, from 60 to 77% and from 1.5 to 3.9%.
In summary, the operational factors that will affect microbial transport include
cell concentration, flow rate, and the ionic strength of the transporting fluid. The results
of studies designed to investigate the effects of cell density on transport have been mixed.
The effects of cell density on transport may be organism- and site-specific. Microbial
filtration and clogging of the matrix will be of concern. The direct relationship between
flow rate and microbial transport always has been found. Finally, there will be an
inverse relationship between the ionic strength of the transporting fluid and microbial
transport. Microorganisms tend to sorb to surfaces under conditions of high ionic
strength and to a less extent under condition of low ionic strength; hence more cells will
be transported in a fluid of low ionic strength.
11.1.5. Environmental Factors that Affect Survivability of Added Organisms
As mentioned previously, transported organisms must not only reach the zone of
contamination but must compete with the indigenous microflora for nutrients, escape
predation, retain their biodegradative capabilities, and often tolerate extremes in pH,
temperature, and other environmental variables. Hardly anything is known about
environmental factors and survivability in the subsurface environment. By extrapolation
from experience with surface water and soil, predation will probably be the most
important factor limiting the survival and activity of introduced microorganisms.
Goldstein et al. (1985) reported that the success of adding nonindigenous
microorganisms to the environment may be dependent on the concentration of the target
compound, the presence of toxicants or predators, the preferential use of alternate
substrates, or the mobility of the introduced organisms. In one experiment, these
authors found that mixing enhanced the mineralization of 5 ug/g p-nitrophenol (PNP) in
sterile soil inoculated with a PNP-degrading organism, suggesting that mixing was
required to move the organisms through the soil to effectively degrade the PNP. Zaidi et
al. (1988) reported that pH and substrate concentration affected the survival and
biodegradation capabilities of introduced organisms in lake water. These authors found
that an increase in pH from 7 to 8 inhibited the mineralization of PNP in sterile and
nonsterile lake water inoculated with a Pseudomonas sp.
The presence of predators and inhibitors may also affect the survival and
biodegradation potential of inoculants. Zaidi et al. (1989) found that the addition of a
eucaroytic inhibitor to lake water inoculated with a Corynebacterium sp. increased the
extent of mineralization of 26 ng/ml PNP, but did not increase mineralization of higher
concentrations of PNP; the authors suggested that the organisms were not able to replace
those cells cropped by eucaryotic grazing at the lower concentrations of PNP.
In summary, the same factors that affect survival of microorganisms in the
surface soil and water environments will affect the survivability in the subsurface.
These factors include substrate concentrations, pH, temperature, and the presence of
toxicants, predators, and alternate substrates. However, little information is available
concerning the survivability of introduced microorganisms in the subsurface.
11.1.6. Field Demonstrations of Microbial Transport
Field demonstrations have documented the transport of introduced
microorganisms through the subsurface. In one demonstration, bacteria native to the
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aquifer actually moved faster than the bulk flow of ground water, perhaps due to size
exclusion chromatography.
In 1978, Hagedorn et al. (1978) transported antibiotic-resistant fecal bacteria
through subsoil under saturated conditions at depths of 30 to 60 cm below the surface to
investigate the potential for ground-water contamination by septic tank discharge. When
inocula at concentrations ranging from 3 to 5 x 10s cells/ml were added, bacteria were
detected in sampling wells located 50 cm from the injection point after 1 day. In some
wells, cells were detected as far as 1,500 cm from injection after 8 and 12 days. Microbial
numbers peaked in sampling wells after rainfall, suggesting that transport was
associated with rainfall patterns. Researchers from this same laboratory investigated
the transport of antibiotic-resistant E. coli through different horizons of hillslope soils
under saturated conditions (Rahe etal., 1978a, 1978b). Inocula at a concentration of 1.4 x
109 cells/ml were injected into the subsoil at depths ranging from 12 to 80 cm, and their
numbers monitored downslope at distances of 2.5 to 20 m from the injection point at
depths ranging from 12 to 200 cm. Irrespective of inoculation depth, the cells moved
downslope to zones of high permeability and then through macropores. In addition,
transport was faster in a subsoil with greater slope and hydraulic conductivity than that
with a lesser gradient and hydraulic conductivity.
In a similar study, McCoy and Hagedorn (1980) investigated transport under
saturated conditions of antibiotic-resistant strains of E. coli through a concave hillslope,
which was located in a transition area between two soil series. Inocula at 4 x 107 cells/ml
were injected into horizontal injection lines located at depths of 12,35, and 70 cm, and the
numbers of transported bacteria were monitored at 2.5, 5.0,10.0, and 15.0 m downslope at
depths ranging from 12 to 200 cm. Bacterial transport varied with depth of injection in
the upper soil series but not in the transition zone where flow paths converged. Flow in
this zone resulted more from channeling rather than matrix flow as the water moved
upward into more transmissive layers because of the hydraulic gradient and a
nontransmissive clay layer.
Harvey et al. (1989) investigated the transport of bacteria and microspheres
through a sandy aquifer (K = 0.1 cm/sec) in natural and forced gradient tracer
experiments. The bacteria to be transported were cultured from ground water collected
at the site and stained with a DNA-specific fluorochrome. A conservative tracer (Cl- or
Br-), microspheres of different diameters (0.2 to 1.3 urn) and surface charges, and the
indigenous bacteria (0.2 to 1.6 jim in length) were then injected into a well screened at 10
to 11 m below the surface and their transport was monitored at multilevel wells placed
1.7 and 3.2 m downgradient of the injection point.
In the forced gradient experiment, both bacteria and carboxylated microspheres
were injected. Breakthrough of bacteria occurred somewhat earlier than that of
bromide. The microspheres were retained by aquifer sediments to a greater extent than
bacteria. Transport of microspheres was directly related to size.
In the natural gradient experiment, carboxylated microspheres of diameters
ranging from 0.2 to 1.3 urn, uncharged microspheres with a diameter of 0.6 urn, and
microspheres with a diameter of 0.8 |im and containing carbonyl surface groups were
injected. Transport of the carboxylated microspheres was directly related to size. For the
microspheres with different surface characteristics, increasing breakthrough times
were observed for uncharged, carbonyl containing and carboxylated particles,
respectively.
11-8
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The results of these field demonstrations suggest that microorganisms can 'be
transported significantly through subsurface materials. These data are contradictory to
many laboratory experiments in which subsurface material was packed to achieve
homogeneity and eliminate any macropores and channels that could have facilitated
transport.
11.1.7. Inoculation to Enhance Biodegradation of Hydrocarbons
Microorganisms have been added to samples of soil and water in the laboratory
and field to enhance biodegradation of hydrocarbons; however, the results of these
studies have been mixed. Atlas (1977) stated in a review on stimulated petroleum
biodegradation that seeding will not be necessary in most environments because of the
ubiquity of hydrocarbon-degrading organisms. Although hydrocarbon-degrading
organisms may be ubiquitous, the problem with natural bioremediation of these
compounds is that the rate of biodegradation is often too slow. Nutrient addition and
agents that render the compounds more bioavailable may enhance these rates. However,
inoculation may be important in environments in which the population of hydrocarbon-
degrading organisms is too low or absent, or the environment is too harsh. In the latter
case, the added organisms must be able to tolerate the extreme conditions. In addition,
inoculation may be beneficial in the biodegradation of the high-molecular-weight
polycyclic aromatic hydrocarbons, which are recalcitrant (Bossert and Bartha, 1986). If
seeding is considered as a method for hydrocarbon remediation, a mixture of
microorganisms will be required. Zajic and Daugulis (1975) found that multiple species
were required to degrade the complex composition of crude oil.
Several investigators have studied the effects of inoculants on hydrocarbon
degradation. Schwendinger (1968) investigated the effect of adding nitrogen, phosphorus
and a bacterial seed (Cellulomonas sp.) to reclaim soil contaminated with oil. Seeding
did not enhance bioreclamation in soil amended with 25 ml/kg oil and inorganic
nutrients; however, seeding did enhance bioreclamation in soil amended with 62 and 100
ml/kg oil and inorganic nutrients. Similarly, Jobson et al. (1974) also reported that a
mixed population of hydrocarbon-degrading organisms slightly stimulated the
degradation of the n-alkanes with chain lengths of CM to C& but had no effect on other
components in soil amended with crude oil. Lehtomaki and Niemela (1975) reported that
the addition of a mixture of hydrocarbon-degrading microorganisms to soil amended
with 0.5% light fuel oil or heavy waste oil had no effect on oil decomposition. Westlake et
al. (1978) added hydrocarbon-degrading bacteria to field plots amended with oil in the
boreal region of the Northwest Territories and found that seeding did not enhance
biodegradation above those plots which received fertilizer.
Several investigators have isolated organisms that can degrade the recalcitrant
high-molecular-weight polycyclic aromatic hydrocarbons. Mueller et al. (1990) isolated a
strain of Ps. paucimobilis from a creosote waste site that can metabolize several PAHs
when its enzymes are induced by growth on fluoranthrene. The organism uses
fluoranthrene, 2,3-dimethylnaphthalene, and phenanthrene, and to a lesser extent
anthracene, benzo[b]fluorene, naphthalene, 1-methylnaphthalene, and 2-
methylnaphthalene, as sole sources of carbon and energy. Washed cells of a
fluoranthrene-grown culture were active against these compounds and biphenyl,
anthraquinone, pyrene, and chrysene as well. The authors speculated that this
organism may be effective in treating mixtures of PAHs, which are characteristic of
creosote waste sites.
Heitkamp and Cerniglia (1988) isolated a Mycobacterium sp. from sediments
exposed to petroleum hydrocarbons which was able to mineralize naphthalene,
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phenanthrene, fluoranthrene, pyrene, 1-nitropyrene, 3-methylcholanthrene, and 6-
nitrochrysene when grown in the presence of peptone, yeast extract, and starch.
Inoculation of soil and water mixtures with fluoranthrene-induced cells enhanced the
mineralization of fluoranthrene by 93% over uninoculated samples (Kelley and
Cerniglia, 1990). In addition, peptone, yeast extract, and starch amendments greatly
enhanced the fluoranthrene-mineralization capability of the inoculant.
Microorganisms subjected to genetic manipulation may loose traits that allow
them to survive and express the contaminant-degrading genes under conditions that
prevail in the subsurface environment. A unique inoculation experiment involved the
introduction of genetically modified ground-water bacteria, which harbored plasmids for
toluene/xylene metabolism (TOL) and antibiotic resistance (RK2), into microcosms
containing uncontaminated and artificially contaminated aquifer material (Jain etal.,
1987). Although the inoculant was stably maintained for 8 weeks of incubation;
inoculation did not enhance the biodegradation of toluene and chlorobenzene in the
artificially contaminated microcosms above that of uninoculated microcosms.
In summary, inoculation to enhance biodegradation of most hydrocarbons usually
is not necessary because of the ubiquity of hydrocarbon-degrading microorganisms.
Microorganisms have coevolved with hydrocarbons, and many have metabolic capability
to degrade the compounds. However, inoculation may be beneficial in environments in
which there are harsh conditions or high molecular-weight polycyclic aromatic
compounds.
11.1.8. Inoculation to Enhance Biodegradation of Chlorinated Compounds
In contrast to the situation with naturally occurring organics, inoculation to
enhance the biodegradation of chlorinated compounds may be beneficial. This is
particularly true for pentachlorophenol.
As early as 1965, MacRae and Alexander (1964) inoculated alfalfa seeds with a
strain of Flavobacterium sp. that degraded 4-(2,4-dichlorophenoxy)- butyric acid, in order
to protect the developing plants in herbicide-treated soils. Seeds were planted in
nonsterile and sterile soils. Inoculation afforded protection in sterile soil but not in the
presence of the indigenous microflora. Edgehill and Finn (1983) reported that the
addition of 106 cells per g dry soil of a pentachlorophenol (PCP)-degrading strain of
Arthrobacter reduced the half-life of 20 ug PCP/g soil from 2 weeks to 1 day in laboratory
experiments; the bacterium used PCP as the sole source of carbon and energy. In
addition, PCP biodegradation was directly related to inoculum size; PCP was reduced by
90% after 24,40 and 100 h, after the addition of 10^, IQS and 10* cells/g soil, respectively.
The results of a field experiment conducted using soil in an outdoor shed indicated that
inoculation and mixing enhanced PCP degradation. After 12 days, 25% was removed in
uninoculated plots, 50% was removed in inoculated but unmixed plots, and 85% was
removed in inoculated and mixed plots.
Martinson etal. (1984) inoculated samples of river water with 106 cells/ml of a
PCP-degrading strain of Flavobacterium and found that 90% of the PCP (1 ppm) was
removed within 48 h, whereas none was removed in uninoculated samples.
Investigators from this same laboratory investigated mineralization of PCP from
contaminated soils by inoculation with the same PCP-degrading Flavobacterium
(Crawford and Mohn, 1985). When samples of loam, clay and sand were amended with
100 ppm PCP and inoculated, initial rates of mineralization were initially fastest in the
loam and slowest in the sand. However, about 60% of the PCP was mineralized in all
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soils after 6 days of incubation. PCP was not mineralized in uninoculated samples after
10 days. In another experiment in which samples were incubated for a longer period of
time, significant mineralization of 100 ppm PCP was detected in uninoculated as well as
inoculated samples, indicating that the indigenous microflora had acclimated to degrade
PCP. After 40 days of incubation, PCP had been mineralized to the same extent (50%) in
inoculated and uninoculated samples.
In one waste-dump soil contaminated with 298 ppm PCP, however,
mineralization was detected in the inoculated samples only. In another waste-dump soil
contaminated with 321 ppm PCP, the extent of removal was similar in inoculated and
uninoculated samples. These authors speculated that enhancing PCP degradation by
the indigenous microflora may be advantageous at low contaminant concentration while
inoculation may be beneficial at high contaminant concentrations.
Brunner et al. (1985) investigated the effect of inoculation to enhance degradation
of polychlorinated biphenyls in samples of soil. Soil (100 g) was adjusted to 50% water
holding capacity, amended with 100 mg Aroclor 1242/kg, and inoculated with 105 or 109
cells/ml of a PCB-degrading Acinetobacter. After incubation aerobically for 70 days,
inoculation did not greatly enhance mineralization; however, inoculation coupled with
analog enrichment with biphenyl significantly enhanced mineralization above that
observed in samples receiving biphenyl only.
Stormo and Crawford (1992) have developed a unique method for transporting a
chlorophenol-degrading Flauobacterium sp. by encapsulating the cells into polymeric
beads of alginate, agarose, or polyurethane. The cells are encapsulated into beads with
diameters ranging from 2 to 50 urn. The catabolic activity of free and encapsulated cells
is not different. Mineralization of PCP by free or encapsulated cells in 20 cm-columns
containing native aquifer material has been assessed (Keith E. Stormo, personal
communication). Preliminary results indicate that rates of PCP mineralization at
concentrations as high as 200 mg/kg by free or encapsulated cells are not significantly
different; however, encapsulation may enhance long-term survivability. PCP
mineralization by the indigenous aquifer microflora was not observed.
To summarize, inoculation to enhance the biodegradation of chlorinated
compounds may be beneficial in some instances. In contrast to the coevolution of
microorganisms and hydrocarbons, the coexistence of microorganisms and chlorinated
compounds has been relatively short. Many microorganisms cannot degrade these
compounds, or the period required to adapt to degrade the compounds may be long.
When the presence of chlorinated contaminants is posing environmental and health
risks and little or no biodegradation of these compounds is detected, inoculation with
contaminant-degrading microorganisms may be warranted.
11.2. MATURITY OF THE TECHNOLOGY
Inoculation or bioaugmentation has been widely used to stimulate bioremediation
of subsurface material contaminated with petroleum hydrocarbons. A variety of cultures
and formulations are commercially available. The practice of inoculation is based on the
assumption that contamination has persisted in the subsurface because competent
microorganisms were not available and that biodegradation is limited by active biomass
The practice further assumes that biodegradation of the contaminant is not limited by the
supply of the substances required for metabolism of the contaminant, such as oxygen or
mineral nutrients. Because adequate field evaluations have not been done, there is no
way to determine whether perceived benefits were provided by the introduced organisms
11-11
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rather than indigenous organisms, or whether the effective agent was the introduced
organism, or a mineral nutrient, surfactant, or biosurfactant provided along with the
culture.
The inocula are relatively inexpensive compared to other remedial activities, and
they are often used to insure the presence of a competent microbial community. If the
inoculant is the only remedy, the contaminated site should be characterized to
demonstrate that there are adequate supplies of electron acceptors and mineral nutrients
to permit complete destruction of the contaminant.
The use of microorganisms with specialized capabilities to enhance
bioremediation in the subsurface is not an established technology. However, research
has been conducted to determine the potential for microbial transport through
subsurface materials for public health and microbial enhanced oil recovery.
11.3. PRIMARY REPOSITORIES OF EXPERTISE
Institutions where microbial transport research is being conducted include
Cornell University (M. Alexander), Mississippi State University (C. Hagedorn), Rice
University (C. H. Ward), U. S. Geological Survey (R. W. Harvey), University of Arizona
(C. Gerba), University of Calgary (J. W. Costerton), University of Idaho (R. L. Crawford),
University of Oklahoma (G. £. Jenneman; M. J. Mclnerney), and University of Virginia
(A. L. Mills). Current addresses and telephone numbers are listed below.
Martin Alexander
Department of Agronomy
Bradfield Hall
Cornell University
Ithaca. NY 14853
Phone « (607)225-1717
FAX «• (607)255-2106
J W Costerton
Montana State University
Engineering Department
Center for Biofilm Engineering
409 Cobleigh Hall
Bozeman, MT 59717
Phone ff (4061994-4770
FAX* (406)994-6098
Ronald L Crawford
Food Research Center, Room 202
University of Idaho
Moscow, Idaho 83843
Phone 0 (208)885-6580
FAX « (208)885-5741
Charles Gerfaa
Department of Soil and Water Science
University of Arizona
TusconAZ 85721
Phone «• (602)621-6906
FAX« (602)621-1647
Charles Hagedorn
Department of Plant Pathology,
Physiology, and Weed Science
Price Hall
Virginia Polytechnic Institute
Blacksburg.VA 24061
Phone*. (703)231-6361
FAX «: (703)231-7477
Ronald W Harvey
U.S Geological Survey
Water Resources Division
Box 25046. MS 458, Boulder Office
Denver, CO 80225
Phone ff (303)541-3034
FAX# (303)447-2505
G E Jenneman
Phillips Petroleum Company
Bartlesville. OK 74005
Phone ff (918)661-8797
FAX ff (918)662-2047
Michael J Mclnerney
Dept of Botany and Microbiology
University of Oklahoma
Norman, OK 73019
Phone ff (405)325-6050
FAXff (405)325-7619
Aaron Mills
Department of Environmental Sciences
Clark Hall
University of Virginia
Charlottesville. VA 22903
Phone # (804)924-7761
FAX # (804)982-2137
C H Ward
Department of Environmental
Science and Engineering
Rice University
Houston, IX 77251
Phone # (713)527-4086
FAXff. (713)285-5203
11.4. OTHER FACTORS CONCERNING APPLICATION
Although specialized microorganisms that have been cultured using selective
enrichment techniques can be used in environmental applications, those developed using
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genetic engineering techniques cannot be released into the environment for commercial
purposes without prior government approval (Pimentel et al., 1989). Genetically
engineered microorganisms for use in such operations as MEOR, bioremediation of
Superfund sites, extraction and concentration of metals, and production of specialty
chemicals may be regulated under the Environmental Protection Agency's Toxic
Substances Control Action Section 5 (Clark, 1992).
11.5. &TATE OF THE ART OF TRANSPORT OF MICROORGANISMS WITH
SPECIALIZED METABOLIC CAPABILITIES AND RESEARCH
OPPORTUNITIES
Since the study published by Raymond et al. (1977) that indicated that
microorganisms can be transported and enhance degradation of hydrocarbons in a
column packed with sand, no one has demonstrated that inoculation of the subsurface
can enhance bioremediation in the laboratory or field. There is a tendency to work with
organisms that are easy to culture and whose genetics are well understood. Little
consideration is given to developing organisms with good transport properties and
survival traits. Provided that microorganisms can be successfully transported through a
specified aquifer and establish themselves, several different possibilities for application
exist (Table 11.1). The best opportunities involve development of inocula that can degrade
mixed wastes, that have increased tolerance to toxicants, and that produce
bioemulsifiers and biosurfactants to increase their access to oily phase contaminants.
TABLE 11.1. POSSIBLE APPLICATIONS OF INTRODUCED MICROORGANISMS
Speciatixed Capability Purpate
Produce biosurfactant/bioemulsifier Mobilize sorbed/entrained contaminants
Degrade multiple compounds Treatment of mixture of compounds
Degrade recalcitrant compounds Inoculation in absence of acclimation by
indigenousorganisms
Tolerate and degrade toxic compound Inoculation in absence of acclimation by
indigenous organisms
Tolerate high concentration of toxicant Inoculation in absence of acclimation by
indigenous organisms
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