EPA/600/R-03/037
September 2002
Application, Performance, and Costs of Biotreatment
Technologies for Contaminated Soils
Contract No. 68-C-00-185
Task Order No. 13
Submitted to
U.S. Environmental Protection Agency
National Risk Management Research Laboratory
26 West Martin Luther King Drive
Cincinnati, Ohio 45268
Jennifer Goetz
Project Officer
Richard C. Brenner
Task Order Manager
BATTELLE
505 King Avenue
Columbus, OH 43201
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CONTENTS
FIGURES iv
TABLES v
ACRONYMS AND ABBREVIATIONS vi
EXECUTIVE SUMMARY x
1.0 INTRODUCTION 1
1.1 Magnitude and Extent of Soils, Sediment, and Groundwater Contamination 3
1.1.1 Common Contaminants of Concern (COCs) 4
1.1.2 Sources of Soil Contamination 7
1.2 Soil Treatment Goals 7
1.2.1 State and Federal Standards 7
.2.1.1 Resource Conservation and Recovery Act (RCRA) Landban Requirement 8
.2.1.2 EPA Region 3 Preliminary Remediation Goals (PRGs) 8
.2.1.3 EPA Region 9 Preliminary Remediation Goals 8
.2.1.4 U.S. Environmental Protection Agency Soil Screening Guidance 9
.2.1.5 Texas Risk-Reduction Rule Guidance 10
1.2.2 Groundwater Standards 10
.2.2.1 Federal Maximum Contaminant Levels (MCLs) 10
.2.2.2 National Recommended Water Quality Criteria (NRWQC) 11
.2.2.3 California Maximum Contaminant Levels 11
.2.2.4 California Ocean Plan Limits 11
1.2.3 Application of Risk-Based Cleanup Criteria 11
.2.3.1 Criteria for Ex-Situ Treatment of Soil 11
.2.3.2 Criteria for In-Situ Treatment of Soil 12
.2.3.3 American Society for Testing and Materials Guide for Risk-Based Corrective
Action 13
2.0 UNDERLYING PRINCIPLES OF BIOREMEDIATION 14
2.1 Biodegradation of PAHs and Petroleum Hydrocarbons 15
2.2 Biodegradation of Explosives 16
2.3 Biotransformation of PCBs 17
2.4 Biotransformation of Chlorinated Aliphatic Hydrocarbons 18
2.5 Biotransformation of Halogenated Aromatic Compounds 20
2.6 Biotransformation of Dioxin Like Compounds 21
3.0 EXISTING BIOREMEDIATION TECHNOLOGIES AND APPLICATIONS 22
3.1 Bioremediation Technologies and Their Applications - Conventional Technologies 24
3.1.1 Land Treatment 24
3. .1.1 Principles of Operation 24
3. .1.2 Target Contaminants 26
3. .1.3 Advantages and Limitations 26
3. .1.4 Technology Cost Drivers 27
3. .1.5 Technology Performance and Cost 27
3.1.2 Biopile and Biocell Treatment 28
3. .2.1 Principles of Operation 28
3. .2.2 Target Contaminants 34
3. .2.3 Advantages and Limitations 35
3. .2.4 Technology Cost Drivers 35
3. .2.5 Technology Performance and Cost 36
11
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3.1.3 Composting 37
3.1.3.1 Principles of Operation 37
3.1.3.2 Target Contaminants 41
3.1.3.3 Advantages and Limitations 42
3.1.3.4 Technology Cost Drivers 42
3.1.3.5 Technology Performance and Cost 42
3.1.4 Ex-Situ Bioreactors 47
3.1.4.1 Principles of Operation 47
3.1.4.2 Target Contaminants 49
3.1.4.3 Advantages and Limitations 50
3.1.4.4 Technology Cost Drivers 51
3.1.4.5 Technology Performance and Cost 51
3.1.5 Conventional Bioventing 57
3.1.5.1 Principles of Operation 58
3.1.5.2 Target Contaminants 59
3.1.5.3 Advantages and Limitations 61
3.1.5.4 Technology Cost Drivers 61
3.1.5.5 Technology Performance and Cost 62
3.1.6 Enhanced Bioventing 68
3.1.6.1 Principles of Operation 68
3.1.6.2 Target Contaminants 69
3.1.6.3 Advantages and Limitations 69
3.1.6.4 Technology Cost Drivers 70
3.1.6.5 Technology Performance and Cost 70
3.1.7 Bioslurping 74
3.2 Bioremediation Technologies and Their Applications - Emerging Technologies 75
3.2.1 Anaerobic Bioventing 76
3.2.2 Phytoremediation 78
3.2.2.1 Advantages and Limitations 80
3.2.2.2 Technology Performance 80
3.2.3 Sequential Anaerobic/Aerobic Treatment 82
3.2.3.1 Advantages and Limitations 82
3.2.3.2 Technology Performance 82
3.2.4 Natural Attenuation 84
3.3 Existing Bioremediation Technologies Summary 86
3.3.1 Applicability of Bioremediation for Contaminated Soils 88
3.3.2 Bioremediation Advantages and Disadvantages 89
4.0 IMPLEMENTATION OF BIOREMEDIATION TECHNOLOGIES 91
4.1 Site Characterization 91
4.2 Technology Selection 94
4.3 Treatability Studies and Pilot-Scale Testing 95
4.4 Full-Scale Design 95
4.5 Process Optimization 96
4.6 Contaminant Degradation Plateaus 96
4.6.1 Preferential Degradation of Easily Degraded Compounds 96
4.6.2 Exhaustion of a Cometabolic Growth Substrate 97
4.6.3 Limited Contaminant Bioavailability 98
4.6.4 Nutrient Limitations 98
4.6.5 Predation of Contaminant-Degrading Bacteria 99
4.6.6 Energetic Limitations and Buildup of Toxic Intermediates 99
in
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5.0 FUTURE DIRECTION OF BIOREMEDIATION TECHNOLOGIES AND BIOREMEDIATION
RESEARCH 100
5.1 Optimization of Technologies that are Proven in the Marketplace 100
5.2 Development of Emerging Technologies to Bring Them to the Market Place 101
5.2.1 Anaerobic Treatment and Anaerobic/Aerobic Sequencing 101
5.2.2 Phytoremediation 102
5.2.3 Bioaugmentation 103
5.2.4 Natural Attenuation 103
5.3 Fundamental Research and Development Needs for Recalcitrant Waste Streams 103
5.4 Summary of Research Needs 104
6.0 REFERENCES 105
FIGURES
Figure 1-1. Treatment Technologies at Bedford MGP Site 3
Figure 1 -2. Frequency of Occurrence of Maj or Contaminant Subgroups at NPL Sites with RODs 6
Figure 1 -3. Frequency of Occurrence of the Most Common Contaminants at Superfund Sites 6
Figure 3-1. Cost Per Cubic Yard for Remedial Actions Based on Cost and Performance Data Presented in
Appendix B 23
Figure 3-2. Land Treatment 25
Figure 3-3. Land Treatment Schematic 25
Figure 3-4. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before and
After Land Treatment 33
Figure 3-5. Biopile schematic 34
Figure 3-6. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before and
After Biopile/Biocell Treatment 40
Figure 3-7. Composting 41
Figure 3-8. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before and
After Composting Treatment 46
Figure 3-9. Bioslurry Reactor 48
Figure 3-10. Bioslurry Reactor Schematic 49
Figure 3-11. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before and
After Bioslurry Treatment 54
Figure 3-12. Schematic of Typical Conventional Bioventing Process 57
Figure 3-13. Conventional Bioventing System 58
Figure 3-14. Relationship Between Organic Physicochemical Properties of Organic Compounds and
Their Potential for Conventional Bioventing 61
Figure 3-15. Box and Whisker Plots Showing the Range of Contaminant Concentrations Before and
After Conventional Bioventing 66
Figure 3-16. Box and Whisker Plots Showing the Range of Contaminant Concentrations Before and
After Enhanced Bioventing 73
Figure 3-17. Bioslurping 74
Figure 3-18. Bioslurping Schematic 75
Figure 3-19. Phytoremediation 78
Figure 3-20. Phytoremediation Schematic 79
Figure 3-21. Groundwater Natural Attenuation Schematic 85
Figure 3-22. Illustration of Decreased Biodegradability of Aged Phenanthrene in Soil 86
Figure 4-1. Contaminant Degradation Plateau 97
IV
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TABLES
Table 1-1. Status of Remediation Sites as of 1996 4
Table 1-2. Media and Contaminant Types at Remediation Sites as of 1996 5
Table 1-3. Summary of Contaminant Types Found at Remediation Sites(a) 5
Table 1-4. Most Common Types of DOD Sites Needing Cleanup 7
Table 1-5. Cleanup Criteria Equations Provided in Texas Regulations 10
Table 1-6. Cleanup Criteria Provided in ASTM Risk-Based Guides 13
Table 2-1. Effectiveness of Treatment Technologies for Contaminant Groups 14
Table 3-1. Technology Maturity 22
Table 3 -2. Costs per Cubic Yard for Remedial Actions: Data used to make Figure 3-1 24
Table 3-3. Summary of Site Characteristics at Land Treatment Installation 29
Table 3-4. Concentrations of Contaminants of Concern Before and After Land Treatment: Data
Used to Generate Figure 3-4 32
Table 3-5. Summary of Site Characteristics at Biopile/Biocell Installations 38
Table 3-6. Concentrations of Contaminants of Concern Before and After Biocell/Biopile Treatment:
Data Used to Generate Figure 3-6 39
Table 3-7. Summary of Site Characteristics at Composting Installations 44
Table 3-8. Concentrations of Contaminants of Concern Before and After Composting Treatment:
Data Used to Generate Figure 3-8 45
Table 3-9. Summary of Site Characteristics atBioslurry Installations 52
Table 3-10. Concentrations of Contaminants of Concern Before and After Bioslurry Treatment: Data
Used to Generate Figure 3-11 53
Table 3-11. Cleanup Goals for Primary Contaminants at the French Site 55
Table 3-12. Efficiency of Bioslurry Reactor for Removing PAHs at the Southeastern Wood
Preserving Site 56
Table 3-13. SABRE Process Effectiveness at Yorktown Site 57
Table 3-14. Summary of Site Characteristics at Conventional Bioventing Installations 64
Table 3-15. Concentrations of Contaminants of Concern Before and After Conventional Bioventing:
Data Used to Generate Figure 3-15 65
Table 3-16. Typical Full-Scale Conventional Bioventing Costs (Downey etal., 1994) 67
Table 3-17. Cost Comparison of In-Situ Bioremediation Technologies Utilized at Fuel Spill Sites 68
Table 3-18. Summary of Site Characteristics at Enhanced Bioventing Installations 72
Table 3-19. Concentrations of Contaminants of Concern Before and After Enhanced Bioventing:
Data Used to Generate Figure 3-16 72
Table 3-20. Primary Equipment Used for Conventional Bioremediation Technologies 87
Table 3-21. Factors That Tend to Increase Costs for Conventional Bioremediation Technologies 88
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ACRONYMS AND ABBREVIATIONS
A acetone
AFB Air Force Base
AFCEE U.S. Air Force Center for Environmental Excellence
ARAR Applicable or Relevant and Appropriate Requirement
ASTM American Society for Testing and Materials
BAP benzo(a)pyrene
BB brominated biphenyl
BDAT Best Demonstrated Available Treatment Technology
BFSS Bioremediation in the Field Search System
BGS Below Ground Surface
BTEX benzene, toluene, ethylbenzene, and xylenes
CA Corrective Action
CAH chlorinated aliphatic hydrocarbon
CB chlorobenzene
CERCLA Comprehensive Environmental Response, Compensation, and Liability Act
CFA Civilian Federal Agencies
CFR Code of Federal Regulations
CPU colony forming units
CO2 carbon dioxide
COC contaminant of concern
CSF cancer slope factor
CSTR continuously stirred tank reactor
CWA Clean Water Act
DAF dilution attenuation factor
DCA dichloroethane
DCE dichloroethylene
ODD dichlorodiphenyldichloroethane
DDE dichlorodiphenyldichloroethylene
DDT dichlorodiphenyltrichloroethane
DGGE denaturing gradient gel electrophoresis
DNAPL dense, nonaqueous-phase liquid
DNT dinitrotoluene
DO dissolved oxygen
DoD United States Department of Defense
DOE United States Department of Energy
DRO diesel range organics
DTSC Department of Toxic Substances Control
EPA U.S. Environmental Protection Agency
ESTCP Environmental Security Technology Certification Program
FAME fatty acid methyl ester
Fe3+ Ferric Iron
FRTR Federal Remediation Technologies Roundtable
VI
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GAC granular activated carbon
GE General Electric
GW groundwater
H2 hydrogen gas
HAC halogenated aliphatic compound
HDPE high-density polyethylene
HEAST Health Effects Assessment Summary Tables
HHEM Human Health Evaluation Manual
HMX octahydro-1,3,5,7-tetranitro-1,3,5,7-tetraaxocine
IAAP Iowa Army Ammunition Plant
IRIS Integrated Risk Information System
K potassium
LAP Joliet Army Ammunition Plant
LNAPL light, non-aqueous-phase liquid
MAFB McClellan Air Force Base
MCAGCC Marine Corps Air Ground Combat Center
MCE methylene chloride extractable
MCL maximum contaminant level
MCLG maximum contaminant level goal
MEK methyl ethyl ketone
MGP manufactured gas plant
MNA Monitored Natural Attenuation
Mn4+ manganese ion
N naphthalene
N2 nitrogen gas
NA not available
NAPL nonaqueous-phase liquid
NCEA National Center for Environmental Assessment
ND non detectable
NFESC Naval Facilities Engineering Service Center
NG no growth
NH4+ ammonium ion
NO3" nitrate ion
NPDES National Pollutant Discharge Elimination System
NPL National Priorities List
NRMRL National Risk Management Research Laboratory
NRWQC National Recommended Water Quality Criteria
NS not specified
NTIS National Technical Information Service
O2
O&M
P
PAH
oxygen gas
operations and maintenance
phosphorus
polycyclic aromatic hydrocarbon
vn
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PCB polychlorinated biphenyl
PCDD polychlorinated dibenzo-p-dioxins
PCDF polychlorinated dibenzo-furan
PCE perchloroethene
PCL Protective Concentration Levels
PCP Pentachlorophenol
PLFA phosphilipids fatty acid
PO4 phosphate ion
ppm parts per million
PRG Preliminary Remediation Goal
PVC polyvinyl chloride
RAGS Risk Assessment Guidance for Superfund
RCRA Resource Conservation and Recovery Act
R&D research and development
RDX cyclotrimethyle netrinitramine
REACH-IT Remediation and Characterization Innovative Technologies
RfD reference dose
RI/FS Remedial Investigation/Feasibility Study
RI radius of influence
RIMS Remediation Information Management System
RMX hexahydro-1,3,5 -triaza-1,3,5 -trinitrocyclohexane
ROD Record of Decision
SABRE Simplot Anaerobic Bioremediation Ex-Situ
scfm standard cubic feet per minute
SMC Stauffer Management Company
SO42" sulfate ion
SPLP Synthetic Precipitation Leaching Procedure
SSL soil screening level
SVE soil vapor extraction
SVOC semivolatile organic compound
SWRCB State Water Resources Control Board
T&E Test and Evaluation
TAC Texas Administrative Code
TCA trichloroethane
TCDD 2,3,7,8-tetrachlorodibenzo-p-dioxin
TCE trichloroethylene
TCLP Toxicity Characteristic Leaching Procedure
TNRCC Texas Natural Resource Conservation Commission
TNT trinitrotoluene
TPH total petroleum hydrocarbon
TRRP Texas Risk Reduction Program
TSD treatment, storage and disposal
UHC underlying hazardous constituents
USCG United States Coast Guard
U.S. EPA United States Environmental Protection Agency
UST underground storage tank
UTS universal treatment standard
Vlll
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VC vinyl chloride
VOC volatile organic compound
IX
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EXECUTIVE SUMMARY
The U.S. Environmental Protection Agency's (EPA's) Office of Research and Development (ORD) has
devoted considerable effort over the last two decades to advancing the understanding of appropriate
applications of bioremediation. Over the years, research direction has transitioned from substantial
emphasis on mechanistic studies to a greater emphasis on evaluation of bioprocesses in the field. The
initial research impetus provided the background information necessary for successful field applications,
and was accomplished collectively through in-house research studies and cooperative research projects
with public and private research institutes. The field efforts are conducted through the Bioremediation in
the Field Program, supported by EPA/ORD, EPA's Office of Solid Waste and Emergency Response
(OSWER), and the EPA Regions through the Superfund Innovative Technology Evaluation (SITE)
Program and Cooperative Research and Development Agreements (CRADAs) with companies. This
two-phase program has resulted in the development of cost-effective technical approaches to site cleanup
that have been validated in the field.
Remedial activities have been conducted on groundwater, soils, sediments, and landfills with a range of
contaminants, including chlorinated solvents, polycyclic aromatic hydrocarbons (PAHs), petroleum
hydrocarbons, oils, and many others. These activities range from catalyzing a shift in the nation's
remedial approaches to groundwater cleanup using bioremediation to employing biotreatment
technologies to remediate the Exxon Valdez oil spill, this country's largest cleanup effort.
As with other treatment strategies, the effectiveness and cost of biotreatment technologies are both site-
and contaminant-specific. Because of the potential advantages offered by bioremediation, there remains a
strong interest in the continued development of biotreatment processes. There are many cases where
bioremediation can be employed with relative confidence. The aerobic degradation of petroleum
hydrocarbons and low-molecular-weight aliphatic and aromatic hydrocarbons is well understood and has
been applied at hundreds of sites using bioventing, biosparging, land treatment, biopile treatment, or
composting. Bioslurry reactors also have been used historically, but tend to be less widely used than
these other alternatives due to their higher capital costs and lower throughput rates. Regulatory approval
for the aerobic biotreatment of these contaminants can be readily obtained, and the above processes can
be applied with confidence to meet treatment goals. For such easily degraded contaminants, treatability
tests can be minimized or even eliminated at most sites.
Whereas the biological treatment of easily degraded contaminants is relatively well understood and
accepted, a large number of contaminants remain for which there are no readily available bioremediation
technologies and for which biotreatment remains challenged. Reports of new and previously
undocumented biotransformation pathways for recalcitrant contaminants continue to appear in the
literature and suggest that new biodegradation pathways and mechanisms will continue to be discovered.
Examples include recent reports of the anaerobic degradation of benzene and PAHs under sulfate-
reducing conditions (Coates et al., 1996, 1997), anaerobic oxidation of dichloroethylene (DCE) and vinyl
chloride (VC) (Bradley and Chapelle, 1996, 1997), the ability to stimulate anaerobic PCB dechlorination
by the addition of surrogate polybrominated biphenyl compounds to soils or sediments (Bedard et al.,
1998), and the complete dechlorination of poly chlorinated biphenyl (PCBs) (Bedard and van Dort, 1998).
These studies and others provide an optimistic future for the biodegradation of environmentally persistent
contaminants, and reflect the need for further research for the development of new and innovative
bioremediation strategies and technologies to address recalcitrant contaminants and increasingly
challenging site conditions.
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1.0 INTRODUCTION
This report is a critical review of biological treatment processes for remediation of contaminated soils.
The focus of this review is on cost and performance of biological treatment technologies demonstrated at
full- or field-scale. Contaminants of concern include primarily organic chemicals and, to a lesser extent,
inorganic chemicals. The report was prepared by Battelle for EPA under Contract 68-C-00-185, Task
Order 13. Primary authors were Dr. Victor S. Magar, Dr. Bruce Alleman, Dr. Andrea Leeson, Mr. James
Abbott, and Ms. Regina Lynch.
Soils may be contaminated with a wide range of organic (e.g., petroleum hydrocarbons, organic solvents,
pesticides and herbicides, dioxins and furans, poly chlorinated biphenyls [PCBs], and energetic
compounds) and inorganic (mostly metals) compounds. Much of the contaminant residue in terrestrial
environments is found in surface soils, vadose-zone soils, and the capillary fringe. Their presence may
threaten human or ecological receptors through a variety of exposure routes including direct contact with
contaminated soil media, transport to the groundwater with further transport to a receptor, and
aboveground volatilization. Conventional physical treatment processes have focused on physical removal
of these contaminants from the vadose zone through excavation or soil vapor extraction (SVE) with
ex-situ vapor treatment. Excavated soils commonly require treatment prior to disposal.
Biological treatment (or biotreatment) has been used to treat contaminated soils at Superfund sites for
many years. As with other treatment strategies, the effectiveness and cost of biotreatment technologies
are both site-specific and contaminant-specific. Because of the potential advantages offered by
bioremediation, there remains a strong interest in the continued development of biotreatment processes.
There are many cases where bioremediation can be employed with relative confidence. The aerobic
degradation of petroleum hydrocarbons and low-molecular-weight aliphatic and aromatic hydrocarbons is
well understood and has been applied at hundreds of sites using bioventing, land treatment, biopile
treatment, or composting.
Bioremediation of some contaminant waste streams has gained preapproval from the EPA as a
"presumptive remedy." Since the enactment of the Comprehensive Environmental Response,
Compensation and Liability Act (CERCLA), the Superfund remedial and removal programs have found
that certain categories of sites have similar characteristics, such as types of contaminants present, disposal
practices performed, or environmental media affected. Based on information acquired from evaluating
and cleaning up these sites, the Superfund Program has taken the initiative to develop presumptive
remedies to accelerate future cleanups at similar sites. The presumptive remedy approach can be used to
streamline remedial decision-making for corrective actions conducted under the Resource Conservation
and Recovery Act (RCRA).
Presumptive remedies are preferred technologies for common categories of sites, based on EPA's
experience and its scientific and engineering evaluation of alternative technologies. The objective of the
presumptive remedy initiative is to use the Superfund Program's experience to streamline site
characterization and expedite the selection of cleanup actions. Over time, presumptive remedies are
expected to ensure consistency in remedy selection and reduce the cost and time required to clean up
similar types of sites. In general, presumptive remedies are expected to be used at all appropriate sites
except where unusual site-specific circumstances are present. Conditions at a site also may justify
considering other technologies along with the presumptive remedy. These potential alternatives may then
be combined with other components of the presumptive remedy to develop a range of alternatives suitable
for site-specific conditions.
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The primary presumptive remedy for treating organic contamination of soils, sediments, and sludges at
wood-treater sites is bioremediation. Bioremediation has been selected as the primary presumptive
remedy for these wastes because the EPA believes that it effectively treats wood-treating wastes at
relatively low costs, and because it has been selected most frequently to address organic contamination at
wood-treater Superfund sites. Bioremediation at wood-treater sites may be accomplished with ex-situ or
in-situ processes. However, at some wood-treater sites, ex-situ bioremediation may be able to achieve
higher performance efficiencies than the in-situ processes due to increased access and contact between
microorganisms, contaminants, nutrients, water, and electron acceptors.
Bioremediation of wood-treater sites is generally inexpensive at $50 to $150/cu yd of treated soil.
Expected removal efficiencies are as follows (EPA, 1995):
Ex situ: 64% - 95% for PAHs; 78% - 98% for chlorophenols
In situ: 51% for PAHs; 72% for pentachlorophenol (PCP).
Efficiency can be limited by lack of indigenous microbes, the presence of toxic metals and/or highly
chlorinated organics, low permeability soils, pH outside of the 4.5-8.5 range, winter weather, and
excessive or insufficient rainfall. Studies on the bioremediation of creosote contamination indicate that
biotreatment processes work well on 2-, 3-, and often 4-ring compounds, but generally not as well on 5-
or 6-ring compounds (EPA, 1995). However, despite these limitations, bioremediation of wood-treater
sites successfully meets the EPA's CERCLA criteria for overall protection of human health and the
environment; long-term effectiveness and performance; reduction of toxicity, mobility, or volume;
short-term effectiveness; implementability; cost; and compliance with Applicable or Relevant and
Appropriate Requirements (ARARs).
While biotreatment is accepted as a presumptive remedy for wood treating wastes, this does not imply its
use as a presumptive remedy for other organic waste streams. Furthermore, the Presumptive Remedy
Program does not specify the type of biotreatment technology that should be used for wood-treater sites;
rather, this determination should be based on historical data and the type of waste stream being
considered. While use of a presumptive remedy helps streamline the technology selection process,
providing significant potential for cost and timesavings, it does not ensure that the process will meet
treatment goals.
For the last decade, EPA's National Risk Management Research Laboratory (NRMRL) in Cincinnati, OH
has evaluated the biotreatment of PAH-contaminated soils using land treatment, bioslurry treatment,
composting treatment, and biopile treatment, and has amassed comprehensive information on the
biodegradation of PAHs in soils. Most of this work has been done at pilot scale at EPA/NRMRL's Test
and Evaluation (T&E) Facility. Other processes, such as the combination of soil washing and
biotreatment for PCP-contaminated soils have been evaluated.
EPA/NRMRL is currently evaluating all of these processes in the field at a former manufactured gas plant
(MGP) facility in Bedford, IN. The goal of this ongoing demonstration project is to evaluate several
bioremediation technologies for meeting PAH cleanup goals as an alternative to conventional thermal and
"dig and haul" strategies.
The primary objective of the Bedford project is to compare the performance of three active
bioremediation treatments with natural attenuation on site soil with moderate PAH concentrations. The
active treatments include phytoremediation, ex-situ land treatment, and ex-situ biopile/composting
treatment. Besides these three active treatment technologies, two other ex-situ technologies, bioslurry
treatment and chemical oxidation with biotreatment, are being evaluated at the site as a secondary
objective to determine their effectiveness for reducing PAH concentrations in heavily contaminated soil.
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Figure 1-1 is an overview picture of the Bedford site, showing several of the bioremediation technologies
in operation. SVE was used at the site early in the study to reduce benzene concentrations to below risk
levels before soil was used for chemical oxidation.
Figure 1-1. Treatment Technologies at Bedford MGP Site
Each of the four primary objective treatment technologies was replicated in nine treatment blocks in a
randomized arrangement for statistical evaluation. Phytoremediation and natural attenuation were each
operated and monitored for 3 years in nine in-situ treatment plots. Ex-situ biopile/composting treatment
and land treatment systems were operated and monitored for three 1-year periods, treating three treatment
plots per year.
Results from this large field-scale study are not yet available. However, the extent of this study alone
demonstrates the dedication and interest of EPA/NRMRL to thoroughly research and advance
bioremediation field efforts.
Despite the research of EPA and others, there is uncertainty regarding how widely bioremediation
technologies are being used for full-scale treatment of contaminated soils, and how effectively they are
being applied. The purpose of this report is to summarize cost and performance data from a wide variety
of sites where bioremediation has been employed at full scale to treat contaminated soils, assess its
effectiveness, compare the cost and performance of various biotreatment technologies, and outline future
directions for additional research.
1.1 MAGNITUDE AND EXTENT OF SOILS, SEDIMENT, AND GROUNDWATER
CONTAMINATION
Biotreatment of contaminated soils is the primary focus of this report. Data are provided in this section to
help the reader become more aware of the overall scope of soil, sediment, and groundwater remediation
efforts in the U.S., and and the general nature of the contaminants involved.
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1.1.1 Common Contaminants of Concern (COCs)
Table 1-1 summarizes the data available as of 1996 indicating the status of sites remaining to be
remediated under major federal and state programs in the United States. These programs are as follows:
CERCLA (i.e., Superfund)
RCRA Corrective Action (CA)
RCRA Underground Storage Tank (UST)
Cleanup activities by Department of Defense (DoD)
Cleanup activities by Department of Energy (DOE)
Cleanup activities by Civilian Federal Agencies (CFA)
Cleanup activities by state agencies.
Table 1-2 shows the relative amounts of the different media and contaminants present at sites for four of
these programs. The media types considered are groundwater (GW), soil, and sediment, and the
contaminant types are volatile organic compounds (VOCs), semivolatile organic compounds (SVOCs),
and metals. As indicated by the table, many of the sites in all of the programs involve both soil and
groundwater contamination and contamination by organics and metals. A more detailed breakout of the
different types of contaminants found at various sites is shown in Table 1-3, and the distribution of
contaminants found most often at Superfund sites with completed Records of Decision (RODs) is
provided in Figures 1-2 and 1-3. Figure 1-2 illustrates the frequency of major contaminant subgroups at
National Priority List (NPL) sites with RODs. Figure 1-3 shows the 12 contaminants most commonly
found needing remediation at NPL sites. As indicated by these tables and figures, there is a wide range of
chemical contaminants present and many of the sites contain mixtures of several different classes of
contaminants.
Table 1-1. Status of Remediation Sites as of 1996
Program
Superfund
RCRACA
RCRA UST
DoD
DOE
CFA
States
Total
Sites Remaining to be
Remediated
547
3,000
165,000
8,336
10,500
>700(a)
29,000(b)
>217,083
Estimated Date of
Completion
Not available
2025
Not available
2015
2070
Varies with agency
Varies with state
Not applicable
Estimated Cost to
Remediate
(Billions, in 1996 $)
7
39
21
29
63
15
13
187
(a) Number of facilities, some of which contain more than one contaminated site.
(b) Number of sites needing attention, but some sites may not require remediation.
Source: U.S. EPA, 1997, EPA/542/R-96/005
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Table 1-2. Media and Contaminant Types at Remediation Sites as of 1996
Program
Superfund
RCRACA
DoD(c)
DOE(d)
Percent of Media Types at Sites in
Program(a)
GW
76
82
71
72
Soil
72
61
67
72(e)
Sediment
22
6
6
NA(e)
Percent of Contaminant Types at Sites
in Program^
VOCs
71
67
65
38ฎ
SVOCs
61
30
69
NA(f)
Metals
65
46
43
55
(a) Media type percentages total to greater than 100 because many sites have more than one type of
contaminated media.
(b) Contaminant type percentages total to greater than 100 because many sites have more than one type
of contaminant.
(c) DoD sites also involve contamination from fuels (22%), explosives (8%), and radionuclides (1%).
(d) DOE sites also involve contamination from radionuclides (90%).
(e) DOE media type data include soil and sediment under soil.
(f) DOE contaminant type data combine VOCs and SVOCs.
Source: U.S. EPA, 1997b, EPA/542/R-96/005
Table 1-3. Summary of Contaminant Types Found at Remediation Sites(a)
s
_
OK
O
PM
Superfund
RCRACA
DOD
Percent of Sites with Specific Contaminant Types
VOCs
o
- J
ฃ "c
5 0
Z
30
32
44
SVOCs
VI
ffi
36
18
16
-------
700
600
500
4A
ฃ 400
It-
ฐ 300
ฃ 200
Z
100
0-
y
601
(b4%)
497
(53%)
VOCS
[TM1 svocs
lyTTj Metals
616
(65%)
360
( iS*, i
287
(30%)
340
(.*%)
227
(24%)
224
(24X1
205
C2%)
Contaminant Subgroups
^ oirttain coe if more of ttiese contemirents .
Source: U.S. EPA. 19ST. EPAS42JR-S9DD5
rbrrtes List sites virthfec^ year 1932throiฃh 1ffl4 Reccfds
fi (RODs). Asrte
Figure 1-2. Frequency of Occurrence of Major Contaminant Subgroups at NPL Sites with RODs
500,
400
300
200
100
473
(50%)
445
(47%)
340
(36%)
272
(29%)
(41%)
352
(37%)
VOCS
SVOCS
Metals
306
(32%)
276
(29%)
273
(29%)
Contaminants
Ncte: Bsed cr d^a arai^le ta 944 NMbr^ P rii( iles Ut t ites with fisc^ year 1SB2 through 1EQ4 Records of Decis br, (R ODs). A s fe
rr^f contancre of more of these ozntemirants.
Sourca:U.E. EPA 1067, EPA/S42/R-OOn05
Figure 1-3. Frequency of Occurrence of the Most Common Contaminants at Superfund Sites
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1.1.2 Sources of Soil Contamination
Soil contamination is primarily the result of inadequate practices for handling and storing hazardous or
toxic solids or liquids (e.g., spills and leaks) and/or managing waste materials. The large number of
contaminated sites related to these practices is illustrated in Table 1-4, which summarizes the main
classifications of DoD sites requiring cleanup. Similarly, contamination remediated under the RCRA
UST program often results from inadequate monitoring and maintenance that allowed stored liquids to
leak from buried storage tanks. Many of the contaminated sites resulted from practices that were legal,
and in some cases, considered to be fully adequate at the time the practice was in use. The failure of these
practices became known only after extensive groundwater and soil contamination was found long after
the waste materials had been placed in or discharged to the site.
Table 1-4. Most Common Types of DoD Sites Needing Cleanup
Classification of Site
USTs
Spill Area
Landfill
Surface Disposal Area
Storage Area
Disposal Pit/Dry Well
Ordnance Area
Fire/Crash Training Area
Surface Impoundment/Lagoon
Other
Total
Number of Sites in
Classification
1,199
1,029
940
700
569
535
496
230
223
2,415
8,336
Percent of Sites in
Classification
14.4
12.3
11.3
8.4
6.8
6.4
5.9
2.8
2.7
29.0
100
Source: U.S. EPA, 1997, EPA/542/R-96/005
National Pollution Discharge Elimination System (NPDES) requirements have decreased the allowable
contaminant levels in individual point source discharges from industrial plants and municipal water
treatment plants, but the number of discharge points can increase in many areas of the country due to
growth. Efforts are also under way to reduce the inputs attributed to surface runoff and combined sewage
overflow, but these reductions also may be offset by growth. The relative contribution of different
contaminant types from point sources compared to those from nonpoint sources is an open question and
undoubtedly varies substantially depending on the industrial and urban land-use mixture in a watershed.
1.2 SOIL TREATMENT GOALS
Soil treatment goals are discussed to provide preliminary estimates of the acceptance criteria for cleanup
using biological treatment methods. Groundwater goals are also briefly discussed to indicate the degree
of cleanup required to protect groundwater from contamination leaching out of soils.
1.2.1 State and Federal Standards
This section provides a general background of the regulatory framework for setting cleanup goals or
guidance for soils and sediment. Groundwater cleanup goals are also discussed because soil treatment
goals may be set based on requirements for groundwater protection. In addition to the standards and
goals discussed in this section, site-specific risk assessment and standards promulgated by local
jurisdictions (e.g., area water boards) often affect soil cleanup criteria. The various standards and goals
overlap, so a site-specific analysis of ARARs is needed to identify the appropriate cleanup criteria for site
-------
soils. Site-specific risk assessment should include a definition of the conceptual model for contaminant
transport and applicable pathways and an evaluation of criteria required under the applicable regulatory
program (e.g., the nine criteria for CERCLA sites).
1.2.1.1 Resource Conservation and Recovery Act (RCRA) Landban Requirement
Landban requirements specify treatment standards that must be achieved prior to land disposal of
hazardous waste (40 CFR 268.40). These requirements consist of treatment standards for each RCRA
waste code and universal treatment standards (UTSs) for the underlying hazardous constituents (UHCs).
The treatment standards for specific waste codes are specified in terms of total concentration, leachable
concentration, or required technology depending on the waste code and the nature of the waste (i.e.,
wastewater or nonwastewater). The UTSs define the maximum allowable total or leachable
concentrations of the underlying hazardous constituents in hazardous waste. Landban treatment standards
were developed based on the performance of the Best Demonstrated Available Treatment Technology
(BDAT) for each waste type and hazardous constituent.
The landban requirements were developed using data collected from the application of specific
technologies applied to specific RCRA wastes. These RCRA wastes are more uniform in physical and
chemical properties than are contaminated soils. The particular challenges associated with the treatment
of soil were recognized and resulted in the development of alternative treatment standards specifically
applicable to land disposal of contaminated soil. The alternative treatment standard for soil requires at
least a 90% reduction from the measured concentrations of UHCs in soils, but levels are not required to
meet goals that would be lower than 10 times the UTS.
1.2.1.2 EPA Region 3 Preliminary Remediation Goals (PRGs)
Region 3 PRGs are risk-based guidelines used to screen sites not yet on the NPL, respond rapidly to
citizen inquiries, and spot-check formal baseline risk assessments (Hubbard, 1999). These PRGs were
developed primarily for screening chemicals during a baseline risk assessment and do not constitute
regulation or guidance. The exposure equations are taken from EPA's Risk Assessment Guidance for
Superfund (RAGS) Human Health Evaluation Manual (HHEM) (U.S. EPA, 1991b, 9285.7-01B) using
exposure factors recommended in RAGS (U.S. EPA, 1991a, 9285.6-03) or supplemental guidance from
the Superfund Program. PRGs are calculated for consumption of tap water, inhalation of ambient air,
consumption offish, and industrial and residential exposure to soil. The target cancer risk is IxlO"6, and
the target hazard quotient is 1.0. Reference doses (RfDs) and cancer slope factors (CSFs) are taken from
the Integrated Risk Information System (IRIS) (U.S. EPA, 1998b, IRIS), Health Effects Assessment
Summary Tables (HEAST) (U.S. EPA, 1997c, HEAST), and provisional values from EPA's National
Center for Environmental Assessment (NCEA). The PRGs have the following important limitations:
Transfers from soil to air and water are not considered.
Cumulative risks from multiple contaminants and media are not calculated.
Dermal risk is not included.
Inhalation risk due to water vapor is calculated using a very simple model (effects of confined
areas or enhanced vaporization [e.g., showering] are not included).
1.2.1.3 EPA Region 9 Preliminary Remediation Goals
The Region 9 PRGs are developed for the following media and pathways:
Groundwater (ingestion from drinking and inhalation of volatiles)
Surface water (ingestion from drinking and inhalation of volatiles)
-------
Residential soil (inhalation of particles, inhalation of volatiles, and dermal absorption)
Industrial soil (inhalation of particles, inhalation of volatiles, and dermal absorption).
Exposure from the direct ingestion of soil is calculated using the method presented in RAGS HHEM
(U.S. EPA, 1991b, 9285.7-01B). Exposure from inhalation of vapors and particulate from soils are
calculated using the revisions to the RAGS HHEM developed for the soil screening guidance document
(U.S. EPA, 1996c, EPA/540/R-95/128). Soil dermal exposure is calculated using chemical-specific
dermal adsorption values for arsenic, cadmium, chlordane, 2,4-D, dichlorodiphenyltrichloroethane
(DDT), lindane, 2,3,7,8-tetrachlorodibenzo-^-dioxin (TCDD), PAHs, polychlorinated biphenyls (PCBs),
and chlorophenols as recommended in the Supplemental Dermal Guidance to RAGS (U.S. EPA, 2001).
A default dermal absorption factor of 10% for semivolatile organic compounds is recommended, but
default dermal absorption values for other chemicals (VOCs and inorganics) are not recommended.
Exposure from groundwater and surface water ingestion and inhalation is calculated based on the RAGS
HHEM. Inhalation of volatile chemicals from water is considered only for chemicals with a Henry's Law
constant greater than or equal to 1 x 10"5 atm-m3/mole and with a molecular weight of less than 200
g/mole. Exposure factors are primarily those recommended in RAGS (U.S. EPA, 1991a, 9285.6-03)
supplemented by more recent guidance from the Superfund Program and the California EPA's
Department of Toxic Substances Control (DTSC). The target cancer risk is IxlO"6, and the target hazard
quotient is 1.0. RfDs and CSFs are taken from IRIS (U.S. EPA, 1998b, IRIS) and HEAST (U.S. EPA,
1997c, HEAST), and provisional values are taken from EPA/NCEA.
Soil screening levels (SSLs) for the protection of groundwater are calculated for 100 of the most common
contaminants at Superfund sites. These generic SSLs were calculated using the default values and
standardized equations presented in the Soil Screening Guidance (U.S. EPA,1996c). SSLs are tabulated
for dilution/attenuation factors (DAFs) of 20 and 1. DAFs are generic estimates to account for typical
natural processes that reduce contaminant concentrations as groundwater migrates from the source zone.
A DAF of 1 would indicate site conditions such as shallow groundwater, fractured media, karst
topography, or source size greater than 30 acres that would typically cause little or no attenuation. In
contrast, a DAF of 20 would account for contaminant dilution and attenuation during transport through
the saturated zone, which would provide an accurate reflection of actual contaminant threat to
groundwater resources. Also included in the PRG table are California EPA PRGs for specific chemicals
where California values may be more restrictive than the federal values.
1.2.1.4 U.S. Environmental Protection Agency Soil Screening Guidance
EPA developed generic SSLs as preliminary screening values to help standardize and accelerate the
evaluation and cleanup of contaminated soils at sites on the NPL with anticipated future residential land-
use scenarios (U.S. EPA, 1996c, EPA/540/R-95/128). Generic SSLs were calculated at 110 for ingestion,
inhalation, and groundwater exposure. Ingestion exposure includes direct ingestion of soil and dermal
exposure for PCP, the only compound with sufficient data to support the calculation. The soil ingestion
calculation uses the methods and data described in the RAGS. Calculation of the inhalation dose
considers both contaminant vaporization from soil and generation and inhalation of contaminated dust.
The inhalation calculation uses a volatilization factor and dispersion modeling approach developed during
preparation of the soil screening guidance document. Calculation of SSLs for the groundwater exposure
pathway considers leaching from the contaminated soil into the groundwater, migration of the
groundwater, and direct ingestion of the groundwater. Groundwater modeling is based on linear
equilibrium partitioning between soil and water and a simple water balance approach to determine a DAF.
The target cancer risk is IxlO"6, and the target hazard quotient is 1.0. RfDs and CSFs are taken primarily
from IRIS (U.S. EPA, 1998b, IRIS) and HEAST (U.S. EPA, 1997c, HEAST) supplemented by a variety
of other sources.
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1.2.1.5 Texas Risk-Reduction Rule Guidance
The Texas Natural Resource Conservation Commission (TNRCC) has developed a risk-based system for
site cleanup under its Texas Risk Reduction Program (TRRP). The required methods for the risk
assessment are described in detail in 30 Texas Administrative Code (TAC) 350 to establish a protocol for
risk-based site cleanup that is acceptable to the TNRCC. The approach follows a three-tiered system
similar to that developed by the American Society for Testing and Materials (ASTM) for risk-based
corrective action guides. The regulations include detailed descriptions of methods to define the
appropriate exposure pathways and to calculate Tier 1 and Tier 2 cleanup goals. Equations are provided
for determining risk-based cleanup criteria for carcinogens and noncarcinogens for the media and
exposure routes shown in Table 1-5. These equations are used to calculate Tier 1 Protective
Concentration Levels (PCLs) based on a target cancer risk is IxlO"6 and the target hazard quotient is 1.0.
The Tier 1 PCLs were calculated for each media and exposure route combination and for residential and
industrial scenarios for exposure to soil along the combined pathways of ingestion, inhalation of vapors
and particulates, and dermal contact, and, for the residential scenario only, ingestion of vegetables.
Table 1-5. Cleanup Criteria Equations Provided in Texas Regulations
Source Media
Groundwater
Groundwater (Class 3)
Groundwater
Groundwater
Surface soil
Surface soil
Surface soil
Surface soil
Subsurface soil
Subsurface soil
Exposure Route
Ingestion
Exposure to Class 3 groundwater
Inhalation of vapors
Discharge to surface water
Inhalation of vapors and particulates
Dermal contact
Ingestion of soil
Ingestion of garden vegetables
grown in contaminated soil
Inhalation of vapors
Leaching to groundwater
Exposure Media
Potable groundwater
Class 3 groundwater
Ambient (outdoor) vapor
Surface water
Ambient (outdoor) air
Surface soil
Surface soil
Vegetables
Ambient (outdoor) air
Groundwater
Source: 31 TAC 335 Subchapters A and S.
1.2.2 Groundwater Standards
This section provides summaries of federal and state standards and guidelines applicable to cleanup of
contaminants in groundwater. These standards and guidelines may influence soil cleanup criteria at sites
where soil cleanup is required to protect groundwater. A common soil cleanup criterion, particularly for
risk-based corrective actions, is ensuring that soil cleanup goals are protective of groundwater. Usually,
determining the leachate concentration from contaminated soils and relating those concentrations to
groundwater cleanup goals identifies the soil cleanup goals.
1.2.2.1 Federal Maximum Contaminant Levels (MCLs)
MCLs are enforceable standards developed under the authority of the Safe Drinking Water Act and define
the maximum permissible level of contaminants in water that is delivered to any user of a public water
system. Maximum contaminant level goals (MCLGs) are distinguished from MCLs in that MCLGs are
nonenforceable goals for drinking water contaminant concentrations. These goals are based on protection
of human health for drinking water, and they allow an adequate margin of safety for public use (U.S.
EPA, 1996d, EPA/822/B-96/002). MCLs are set at levels that also should protect human health but may
have other factors that influence the selection of levels. For example, while MCLs are as close to the
10
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MCLGs as feasible, they also may take into account available treatment technologies and the costs to
large public water systems. MCLGs, on the other hand, are strictly health-based goals.
7.2.2.2 National Recommended Water Quality Criteria (NRWQC)
The NRWQC are nonenforceable guidelines developed and published by EPA as required by Section
304(a)(l) of the Clean Water Act (CWA) (U.S. EPA, 1999a, EPA/822/Z-99/001). These guidelines are
developed to reflect the latest scientific knowledge based solely on data and scientific judgements on the
relationship between contaminant concentrations and adverse effects to human health and the
environment. Cost and feasibility of meeting the guidelines in ambient water are not considered during
the process of setting the guidelines. Guidelines have been set for 157 contaminants based on adverse
effects to human health and toxicity to freshwater and saltwater organisms.
1.2.2.3 California Maximum Contaminant Levels
California MCLs are similar to the federal MCLs, but include lower limits on some chemicals that are
specifically applicable to drinking water in California.
1.2.2.4 California Ocean Plan Limits
The California Ocean Plan sets forth limits or levels of water quality characteristics for ocean water to
ensure reasonable protection of beneficial use (SWRCB, 1997). The total discharge of waste materials
shall not cause violation of these limits. Standards include water quality objectives for ocean water
describing acceptable bacterial characteristics, physical characteristics, chemical characteristics,
biological characteristics, and radioactivity; general requirements for management of waste discharge to
the ocean; and quality requirements for waste discharges to the ocean, including contaminants in
groundwater migrating into the ocean. These numerical discharge limits for chemical contaminants are
based on protection of marine aquatic life or protection of human health.
1.2.3 Application of Risk-Based Cleanup Criteria
There is growing support for the application of risk-based corrective actions to clean up petroleum
contaminants at UST sites EPA, 2002 (http:/www.epa.gov/swerustl/rbdm/rbdmfaq5.htm; last viewed
9/25/02). Risk assessment is required for CERCLA and RCRA CA during site characterization and
remedy evaluation to determine the level of risk (Begley, 1996). In the early history of the CERCLA and
RCRA CA processes, if the risk assessment indicated excessive risk, cleanup criteria were set based on
the site background or practical limits of available technology (Begley, 1996). However, using risk
assessment to set cleanup criteria based on site-specific pathways analysis, considering the planned future
land use for the site, has been gaining acceptance (Begley, 1996).
1.2.3.1 Criteria for Ex-Situ Treatment of Soil
The alternative landban requirements for soil will play a role in setting cleanup criteria for excavated soil
or dredged sediment at many sites. Contaminated soil, once it is excavated, is considered a solid waste
and, therefore, must be evaluated to determine if it is also a hazardous waste. If the soil exhibits a
hazardous waste characteristic or contains a listed hazardous waste, the applicable waste codes will be
applied and the landban standards applicable to those codes and for UHCs must be met prior to land
disposal. RCRA requirements apply directly to hazardous waste soils excavated at non-CERCLA sites or
transferred off site from a CERCLA site. Landban requirements usually are considered as an applicable
requirement for hazardous waste soil treated and managed at a CERCLA site. Additionally, landban
requirements may be applied at a CERCLA site as a relevant and appropriate requirement for soil that is
11
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not classified as RCRA hazardous, but has chemical characteristics that are similar to a RCRA hazardous
waste.
Site-specific risk assessment can be used to set cleanup limits for soil treatment to supplement landban or
when landban provisions do not apply (i.e., nonhazardous soil). Cleanup goals developed by various
regulatory authorities provide a preliminary indication of risk-based goals, but include simplifying
assumptions and do not cover all possible site conditions. For example, a construction worker who is
assumed to work in a trench where groundwater and contaminant vapors accumulate will have a much
shorter period of allowable exposure at higher concentrations compared to the allowable exposure from
the inhalation of vapors emanating from the soil surface. Typically, cleanup criteria are determined by
first calculating PRGs and/or site-specific risk-based limits. The lowest value for each contaminant,
determined by each method, is then selected and used. Risk assessment may result in setting cleanup
criteria that are lower than the landban requirements in special circumstances, such as where multiple
contaminants in the groundwater or multiple pathways of exposure increase the risk above acceptable
levels.
1.2.3.2 Criteria for In-situ Treatment of Soil
When soil is treated in situ, no solid waste is generated so the RCRA hazardous waste rules are not
triggered. Therefore, cleanup criteria are set on a site-specific basis. As with excavated soils, risk
assessment considering site-specific conditions and planned land use is gaining acceptance with
regulatory authorities as an approach for setting cleanup criteria. The cleanup criteria for in-situ soil
should consider the viable pathways for surface soil (e.g., ingestion, volatilization, and dermal exposure)
and groundwater protection for surface and subsurface soil.
Ingestion, volatilization, and dermal exposure are not probable exposure pathways for subsurface soils,
but infiltrating rainwater can leach contaminants out of soil and carry them downward into an aquifer. A
simple approach for setting cleanup criteria for subsurface soil is the use of equilibrium modeling or leach
testing of soil to estimate or measure the leachate concentration; this concentration is then compared to an
accepted groundwater standard or a risk-based standard for groundwater consumption. Leach testing
normally is done using an accepted regulatory test such as the toxicity characteristic leaching procedure
(TCLP) or the synthetic acid precipitation leaching procedure (SPLP). For more complex situations, risk-
based goals may be based on modeling of transport from the source soil to the groundwater consumer. A
detailed transport modeling approach requires considerably more effort than does the simple leachate
estimation approach.
MCLs are generally ARARs at CERCLA sites for cleanup of an aquifer if the groundwater, prior to the
contamination, could have been used at some future date as a drinking water source (U.S. EPA, 1988,
EPA/540/G-89/006). Similarly, RCRA standards require cleanup of groundwater at hazardous waste
treatment, storage, and disposal (TSD) facilities to meet MCLs. A cleanup standard more stringent than
MCLs may be needed in special circumstances, such as where multiple contaminants in the groundwater
or multiple pathways of exposure increase the risk above acceptable levels.
If the aquifer is not suitable for use as a drinking water supply (e.g., low yield and/or high salt content),
cleanup to MCLs may not be required. In this case, cleanup criteria may be set using risk-based standards
alone or in conjunction with other ARARs.
12
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1.2.3.3 American Society for Testing and Materials Guide for Risk-Based Corrective Action
ASTM has developed a standard guide for risk-based corrective action at petroleum release sites (ASTM
E 1739) and a provisional guide for risk-based corrective action applicable to a broad range of
contaminant types (ASTM PS 104-98). Both of these guides describe a three-tiered approach as follows:
Tier 1 evaluation - a risk-based analysis using non-site-specific values for complete and
potentially complete direct and indirect human exposure pathways and qualitative ecological
screening evaluation
Tier 2 evaluation - a risk-based analysis for human exposure pathways using the same methods
applied in the Tier 1 evaluation, but with site-specific analysis of exposure pathways and
qualitative or quantitative analysis of ecological risks
Tier 3 evaluation - a risk-based analysis for human exposure pathways using complex modeling
of contaminant fate and transport and a more quantitative analysis of ecological risks than Tier 2.
Moving from Tier 1 to Tier 3 reduces the degree of conservatism in the cleanup criteria, but involves
increased effort. In particular, Tier 3 calculations typically require a much larger amount of effort than do
Tiers 1 or 2. The guides include detailed descriptions of methods to define the appropriate exposure
pathways and to calculate Tier 1 and Tier 2 cleanup goals based on a wide range of pathways. Equations
are provided for determining risk-based cleanup criteria for carcinogens and noncarcinogens for the
media and exposure routes shown in Table 1-6. These equations are used to calculate Tier 1 cleanup
criteria for some example contaminants; however, the standards are intended to prescribe a method for
calculating risk-based cleanup criteria, not to define the specific chemicals to focus on or the cleanup
values to use.
Table 1-6. Cleanup Criteria Provided in ASTM Risk-Based Guides
Source Media
Air
Groundwater
Groundwater
Groundwater
Surface soil
Subsurface soil
Subsurface soil
Subsurface soil
Exposure Route
Inhalation
Ingestion
Enclosed space vapor inhalation
Vapor inhalation
Ingestion of soil, inhalation of vapors and particulates, and
dermal contact
Vapor inhalation
Enclosed space vapor inhalation
Leaching to groundwater
Exposure Media
Air
Potable
Indoor air
Outdoor air
Soil and air
Outdoor air
Indoor air
Groundwater
Source: ASTME 1739-95
13
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2.0 UNDERLYING PRINCIPLES OF BIOREMEDIATION
Bioremediation may be defined as a process in which a biological, especially microbial, catalyst acts on
pollutant compounds, thereby remedying or eliminating the environmental contamination (Madsen,
1991). Successful bioremediation of soils results from a manipulation of the contaminated system that
encourages biological activity that results in the conversion of the contaminant to a less harmful form
(Turco, 1999). This section focuses on the microbiology of contaminant degradation in soils. As
discussed in Section 1, soils may be contaminated with a wide range of organic (e.g., petroleum
hydrocarbons, organic solvents, pesticides and herbicides, dioxins and furans, PCBs, and energetic
compounds) and inorganic (mostly metals) compounds. Table 2-1 summarizes the effectiveness of
different remedial activities for specific contaminant groups. While Table 2-1 provides a general idea of
how contaminants have been effectively treated in the past, it should be noted that the efficacy of
bioremediation technologies is based on many site-specific factors, and if a technology has successfully
remediated a certain class of contaminants in the past, this does not guarantee future success. Conversely,
if a technology is listed as 'demonstrated ineffective' in Table 2-1, this does not imply that the same
technology will be ineffective at treating the given contaminant at future sites.
Table 2-1. Effectiveness of Treatment Technologies for Contaminant Groups
Treatment Technology
In Situ
Bioventing, Aerobic
Bioventing, Cometabolic /
Enhanced
Monitored Natural
Attenuation
Phytoremediation, Extraction
Phytoremediation,
Mineralization
Slurry-Phase Biotreatment
Ex Situ
Biocell
Biopile
Bioreactor, Aerobic Slurry-
Phase
Bioreactor, with Anaerobic
Slurry-Phase
Composting (mainly
windrows)
Daramend Process
Land Treatment, Active
Cycled Land Treatment (with
non-aerated / anaerobic
phase)
Contaminant Groups
Chlorinated
VOCs/
SVOCs
2
1
1
1
1
2
1
Fuels and
Oils
19
6
2
1
1
6
6
2
1
1
18
2
Creosote,
PAHs
1
1
4
3
2
2
4
1
Metals
1
2
PCBs
1
1
Pesticides
Herbicides
1
1
1
1
Explosives
4
1
1
3
4
2
#
Demonstrated Effective
Somewhat Effective
Demonstrated Ineffective
Number of Sites
14
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It is important to understand the limitations as well as the potential advantages of the different biological
processes for degrading these contaminants when considering bioremediation technologies for
contaminated soils. For example, some contaminants are best degraded aerobically while others are
degraded only under anaerobic conditions. Understanding the biodegradation mechanisms will help in
the proper selection of a bioremediation technology. The degradation characteristics of the soil
contaminants are divided into the following groups: organic compounds that are degraded aerobically or
anaerobically as primary growth substrates (this may include halogenated and non-halogenated
compounds); organic compounds that are degraded cometabolically; and halogenated organic compounds
that are degraded anaerobically via reductive dehalogenation.
2.1 BIODEGRADATION OF PAHs AND PETROLEUM HYDROCARBONS
Hydrocarbons, including PAHs, are introduced into the environment primarily in fuels, oils, creosotes,
coal tars, and other refined petroleum products. These hydrocarbons are weathered in soils through
volatilization, dissolution, sorption, and biodegradation. Lighter-molecular-weight compounds tend to
evaporate into the vadose zone and subsequently into the atmosphere, or dissolve in water that infiltrates
soils. However, for high-molecular-weight compounds, these mechanisms tend to be minor (Prince and
Drake, 1999) and they tend to be much more persistent in the environment. Through adsorption, the
binding of hydrocarbon compounds to soils can make them unavailable for biodegradation (Prince and
Drake, 1999); this is particularly true of the higher-molecular-weight compounds that tend to have a
higher affinity for sorption to soils due to their higher degree of hydrophobicity. However, it is important
to recognize that higher-molecular-weight PAHs are generally more potent toxicants, suggesting that it
may be necessary to treat all PAHs to mitigate risk. Further, weathering of lower-molecular-weight
PAHs and sorption of higher-molecular-weight PAHs may not significantly reduce toxicity.
With respect to the destruction and removal of hydrocarbons from the environment, the most important of
the weathering phenomena is biodegradation. Hydrocarbons are biodegraded by a wide variety of
microorganisms in a broad range of habitats, under aerobic and anaerobic conditions (Prince and Drake,
1999). Aerobically, bacteria that grow on hydrocarbons typically initiate oxidation by incorporating
molecular oxygen into organic compounds by the action of enzymes known as oxygenases (Wackett and
Householder, 1989) that destabilize carbon-carbon bonds and render the organic molecule more
susceptible to degradation. A number of hydrocarbon compounds have been shown to be degraded and
ultimately mineralized to carbon dioxide (CO2) in this manner, including PAHs, such as acenaphthene,
fluorene, dibenzothiophene, anthracene, phenanthrene, fluoranthene, pyrene, chrysene,
benzo[a]anthracene, and benzo[a]pyrene (Prince and Drake, 1999). However, biodegradation of PAHs
and other petroleum hydrocarbons tends to be more difficult with increasing molecular weight, resulting
in reduced degradation rates.
In some cases, the oxygenases show activity for other compounds, a process known as cometabolism.
Because oxygenases tend to be relatively non-specific with respect to the types of organic compounds that
they will react with, the fortuitous oxidation of some contaminants will occur via the degradation of a
primary growth substrate. Methane monoxygenase is the most widely studied non-specific oxygenase,
and is well known for its ability to cometabolically degrade trichloroethylene (TCE) and other chlorinated
aliphatic hydrocarbons (CAH). In addition to cometabolically degrading selective CAH, methane
monooxygenase also has been shown to convert naphthalene to 1- and 2-naphthols (Dalton et al., 1981).
A wide variety of organic compounds can serve as primary growth substrates or as the cometabolically
degraded substrate. While cometabolism has not been exploited as a bioremediation mechanism for PAHs
and other petroleum hydrocarbons to date, it may occur in the environment where petroleum spills result
in significant biological activity, resulting in anaerobic conditions and methane production through
methanogenesis.
15
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Anaerobically, PAHs and other petroleum hydrocarbons are more difficult to degrade and degrade much
more slowly than under aerobic conditions. For some time, it was thought that PAHs could not degrade
anaerobically, especially under extreme sulfate-reducing or other reducing conditions. However,
naphthalene and acenaphthene have been shown to be biodegraded under nitrate-reducing conditions
(Milhelcic and Luthy, 1991; Durant et al., 1995), and naphthalene and phenanthrene have been shown to
be biodegraded under sulfate-reducing conditions (Coates et al., 1996). New degradation pathways are
continuously being discovered, and this list is unlikely to exhaust the potential for anaerobic degradation
of PAHs.
2.2 BIODEGRADATION OF EXPLOSIVES
Explosive compounds include the nitroaromatic compounds (trinitrotoluene [TNT], picrate, and tetryl)
and the nitramines (nitroguanidine, hexahydro-l,3,5-triaza-l,3,5-trinitrocyclohexane [RMX], and
octahydro-l,3,5,7-tetranitro-l,3,5,7-tetraaxocine [HMX]). TNT is historically the most commonly used
of all military explosives. Military grades of TNT also contain as a sum up to 8% 2,4-dinitrotoluene (2,4-
DNT) and 2,6-dinitrotoluene (2,6-DNT) (Major, 1999).
Biological mineralization of the nitroaromatics is difficult due to the electrophilic nature and the
orientation of the nitro groups (Major, 1999), and mineralization by individual bacterial cultures may not
be possible, or practical. The meta spacing of the three nitro groups leaves only two unsubstituted
carbons on the ring, situated meta to each other. Biological aromatic ring cleavage normally requires
placement of phenolic substituents oriented ortho or para to each other. It is likely that the meta
orientation of the nitro groups inhibits the hydroxylation of adjacent carbons on the aromatic ring and,
consequently, ring cleavage, resulting in the persistence of these compounds in the environment (Major,
1999). Thus, the aerobic biodegradation of nitroaromatic compounds tends to favor biotransformation
more than mineralization, resulting in the potential production of a variety of amino-nitro compounds.
TNT has been mineralized in the laboratory. The white rot fungus Phanerochaete chrysosporium is able
to mineralize TNT to CO2 (Fernando et al., 1990; Tudor et al., 1990); however, under strictly aerobic
conditions, reports of mineralization in nature are lacking. Anaerobic reduction of nitro substituents to
amino substituents can occur biologically under reduced conditions (Funk et al., 1993), forming amino-
dinitro-, nitro-diamino-, and triamino-toluene. Anaerobic consortia can be enriched to degrade TNT to
aliphatic end products (Funk et al., 1993, 1994). Presumably, under these conditions the nitro groups of
the nitroaromatic compounds are used as terminal electron acceptors.
The nitramine explosives of environmental interest include the cyclic nitramines
cyclotrimethylenetrinitramine (RDX) and HMX because of their wide use as explosives due to their
explosive power (1.5 to 2 times that of TNT) and rapid detonating velocity (1.3 times that of TNT)
(Major, 1999). Most weapons-grade RDX contains some HMX as an impurity and vice versa. Under
anaerobic conditions, the nitro substituents of RDX are reduced (McCormic et al., 1981, 1985) to nitroso
groups, producing nitrosoamines. This is followed by a series of reactions that result in the cleavage of
the heterocyclic ring to form common byproducts of this degradation process including various
hydrozine, dimethylhydrazine, and dimethylnitrosamine intermediates. Because dimythylnitrosamine is
much more toxic than RDX, accumulation of this compound presumably would require more stringent
remedial action (Major, 1999).
Much less is known about HMX degradation. Under reducing conditions, HMX is known to
biotransform similarly to RDX in which the nitro groups undergo reductions to nitroso groups. However,
HMX reduction proceeds only to the formation of the mono and dinitroso products with retention of the
intact ring system (Major, 1999).
16
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2.3 BIOTRANSFORMATION OF PCBs
Most PCBs in the environment are dispersed at low concentrations in soil, air, water, and sediment;
however, environmental PCB pools remain in soils or sediments at concentrations high enough to pose
environmental or public health risks (Hickey, 1999). The family of PCBs contains 209 theoretically
possible molecular conformations, called congeners. Each congener consists of a biphenyl molecule
substituted with one to ten chlorine substituents. PCBs were produced as mixtures during the
mid-twentieth century and sold for industrial applications under trade names such as Aroclor, Clophen,
Fenclor, or Kanechlor (in the USA, Monsanto Corporation produced Aroclors). Approximately 189 of
the 209 theoretically possible PCBs have been identified in Aroclors and other PCB mixtures (Jones,
1998). Of these, 36 have been identified as the most significant in terms of their toxicity potential or
abundance (McFarland and Clarke, 1989). The PCBs with the greatest toxicity potential are those
substituted in both the para positions and at least two meta positions, because these congeners are
stereochemically similar to 2,5,7,8-tetrachlorodibenzo-^-dioxin (TCDD) (Hickey, 1999; NRC, 2001).
The fate of PCBs is influenced by sorption, volatilization, and degradation, and the extent to which these
processes affect PCB weathering depends strongly on the number and positioning of chlorines on the
PCB molecule. Increased PCB chlorination results in increased hydrophobicity, and increased sorption.
Sorption is the dominant process influencing the fate of PCBs in soils, and is of particular significance
with PCBs because it may attenuate biodegradation. PCBs strongly partition into soil or sediment organic
matter, which has the potential beneficial effect of immobilizing them against leaching into groundwater.
However, in the context of biodegradation, as for PAHs, desorption may be rate limiting by restricting the
release of PCBs into the aqueous phase where they would be more bioavailable (Hickey, 1999).
The primary factors affecting PCB biotransformation are the number and pattern of chlorine substituents
(Hickey, 1999; NRC, 2001). Aerobically, the meta cleavage pathway is the primary method by which
bacteria degrade the biphenyl molecule. Bacteria use biphenyl and monochlorinated biphenyls for
carbon, energy, and growth (Ahmed and Focht, 1973). However, biphenyl degraders are unable to use
PCBs chlorinated on both aromatic rings because they are unable to assimilate halogenated aromatic or
aliphatic acids. Consequently, PCBs substituted on both rings generally fail to support growth (Hickey,
1999). However, these PCBs may be degraded cometabolically, in which degradation of biphenyl or
monochlorinated biphenyls supports bacterial growth, and PCBs chlorinated on both aromatic rings are
degraded fortuitously by the biphenyl dioxygenase and other nonspecific enzymes. Both the number and
location of chlorine substituents strongly affects the ability for PCBs to be degraded cometabolically. In
general, the ability for bacteria to degrade PCBs decreases with increased chlorination, and congeners
with five or more chlorines are relatively recalcitrant to aerobic biodegradation (Furukawa et al., 1983;
Masse et al., 1984; Bedard et al., 1986). The chlorine substituent pattern also affects the metabolic
byproducts of dechlorination and their ability to be degraded further by other bacteria.
Fungi also have been reported to degrade or transform PCBs while growing on a supplemental carbon
source. PCB degradation by white-rot fungi and soil fungi is similar to that of aerobic bacteria in that it is
most extensive for lower-chlorinated congeners. The white-rot fungus Phanerocheate chrysosporium
mineralized PCBs in the laboratory for congeners with four or fewer chlorines (Dietrich et al., 1995), and
the soil fungus Aspergillus niger degraded technical grade PCB mixtures with less than 42% chlorine by
weight (Murado et al., 1976); more chlorinated congeners were not degraded.
Under anaerobic conditions, the primary metabolic pathway is reductive dechlorination, in which chlorine
removal and substitution with hydrogen by bacteria results in a reduced organic compound with fewer
chlorines (Mohn and Tiedje, 1992). Higher chlorinated biphenyls are preferentially dechlorinated over
lower chlorinated congeners, and the step-wise replacement of chlorines with hydrogen atoms results in
17
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the accumulation of mono-, di-, and trichlorobiphenyls (Quensen et al., 1988). In general, reductive
dechlorination preferentially removes chlorines from the meta and para positions and replaces them with
hydrogen atoms, resulting in substantial reductions in carcinogenicity and "dioxin-like" toxicity. In
addition to lowering the overall toxicity of PCB-contaminated materials, the tendency of the PCB mixture
to bioaccumulate is also reduced. For example, 2-chlorobiphenyl and 2,2-bichlorobiphenyl display an
approximate 450-fold decrease in the tendency to bioaccumulate in fish compared with tri- and tetra-
chlorinated PCBs (Abramowicz and Olson, 1995).
Bedard et al. (1998) and Bedard and Van Dort (1998) demonstrated that PCB dechlorination can be
initiated using brominated biphenyl (BB) analogs of chlorobobiphenyl analogs. Anaerobic PCB
dechlorination in Woods Pond (Lenox, Massachusetts) sediments was stimulated using bromophenols.
Mono-, di-, tri-, tetra-, and pentabromophenols were added to sediment microcosms; all were completely
dechlorinated to biphenyl. The PCB dechlorination primed by several brominated biphenyls was nearly
twice as effective as that primed by chlorinated biphenyls (Bedard et al., 1998), where the most effective
primers were 26-BB, 245-BB, 25-3-BB, and 25-4-BB. The 26-BB primed microcosms converted
approximately 75% of the hexa- through nonachlorobiphenyls to tri- and tetrachlorobiphenyls within 100
days, and removed approximately 75% of the PCBs that are more persistent in humans.
2.4 BIOTRANSFORMATION OF CHLORINATED ALIPHATIC HYDROCARBONS
Halogenated aliphatic compounds are frequently used as solvents, degreasers, refrigerants, aerosols, and
pesticides. The popularity of their use has resulted in their frequent release into the environment. The
most common halogenated aliphatic compounds are the CAHs, which include the chlorinated ethenes
(perchloroethene [PCE], TCE, dichloroethylene [DCE] isomers, and vinyl chloride [VC]), the chlorinated
ethanes such as 1,1,1- and 1,1,2-trichloroethane (TCA), and 1,1- and 1,2-dichloroethane (DCA); and
chloromethanes including carbon tetrachloride and methylene chloride. In soils, these compounds exist
as sorbed or dissolved in light non-aqueous phase liquids (LNAPL). Because many of these compounds
form dense NAPLs (DNAPL), they are not commonly present in soils as free-phase liquids, except
possibly trapped in soil pores. The transformation reactions of CAHs in biotic and abiotic systems have
been reviewed extensively (Vogel et al., 1987; Mohn and Tiedje, 1992; Fetzner and Lingans, 1994;
Castro, 1998; and Reinhard et al., 1999). This discussion focuses on microbial transformations of CAHs
under aerobic and anaerobic conditions.
Three primary microbial mechanisms are used for the degradation of CAHs; halorespiration (e.g.,
reductive dechlorination) where the CAH is used as an electron acceptor in the microbial electron
transport chain; direct oxidation where the CAH is used as a substrate for energy and growth; and
cometabolism in which the CAH is oxidized by non-specific enzymes used for bacterial growth on an
alternative primary substrate (Reinhard et al., 1999).
Halorespiration is a reductive process in which the CAH is used as an electron acceptor, and in the
process dehalogenated, resulting in chlorine removal and substitution with a hydrogen atom. For some
compounds that cannot be oxidized under aerobic or anaerobic conditions, such as PCE, carbon
tetrachloride, 1,1,2,2-tetrachloroethane, and some highly chlorinated PCBs, halorespiration is the primary
biological mechanism for their microbial transformation. In the absence of halorespiration, many of these
compounds would persist in the environment indefinitely. Limitations of halorespiration are that this
process requires strict anaerobic conditions, the presence of an electron donor for carbon and energy, and
the presence of bacteria capable of sustaining these reactions. Another limitation is that halorespiration
rates tend to decrease with decreasing chlorination, and the presence of bacteria that can halorespire a
parent compound does not necessarily imply that the dechlorination daughter products also can be
halorespired. This results in the potential accumulation of dechlorination byproducts in the environment,
such as the production of DCE and VC from PCE or TCE dechlorination. The accumulation of
18
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dechlorination byproducts has historically been among the most significant limitations of implementing
this technology in the field, especially for chloroethenes, because the byproduct VC is much more toxic
than its parent compounds PCE, TCE, or DCE. However, as discussed below, recent studies have shown
that VC and DCE can be degraded aerobically, providing a removal mechanism for these compounds
should they accumulate in the environment.
Some CAHs can be oxidized under aerobic or anaerobic conditions where the CAHs are used as growth
substrates (Reinhard et al., 1999). Specifically, evidence has been presented for the aerobic degradation
of monochloromethane (Yokota et al., 1986) or dichloromethane (Rittman and McCarty, 1980); aerobic
degradation of 1,2-DCA (Hage and Hartmans, 1999); and the degradation of VC (Bradley and Chapelle,
1996; Bradley et at., 1998) 1,2-DCE under aerobic (Bradley et al., 1998) and anaerobic (e.g., Fe-reducing,
Mn-reducing, or methanogenic) conditions (Bradley et al., 1998; Bradley and Chapelle, 1996, 1997). The
ability for bacteria to use VC and DCE isomers as growth substrates under aerobic or anaerobic
conditions has opened the possibility for sequential anaerobic/aerobic treatment of chloroethenes, PCE
and TCE could be dechlorinated to DCE and VC under anaerobic conditions, then mineralized to CO2 and
hydrogen chloride (HC1) under aerobic conditions. Thus, under properly controlled conditions where the
complete dechlorination of PCE to ethene is ineffective, partial dechlorination to DCE and VC may be
promoted followed by their degradation to CO2 and HC1. While the engineering application of this
approach has been limited, there is increasing evidence of the sequential anaerobic/aerobic dechlorination
of chloroethenes in nature under natural conditions. Furthermore, this mechanism was promoted as the
optimal condition for complete degradation of chloroethenes in the environment under Monitored Natural
Attenuation (MNA) in the Air Force Center for Environmental Excellence (AFCEE) protocol for
evaluating MNA of chlorinated solvents in groundwater (Wiedemeier et al., 1996).
The aerobic cometabolism of chlorinated ethenes has been studied extensively because of its potential for
site cleanup. Cometabolism relies on the oxidation of a cosubstrate that is affected by oxygenaze
enzymes known as mono- or dioxygenazes. The nonspecific nature of these enzymes means that they are
often used for the oxidation of CAHs or other compounds other than the primary growth substrate. PCE
and carbon tetrachloride are important exceptions to this process, and there is no evidence to date of their
oxidation under aerobic or anaerobic conditions, cometabolically or otherwise.
It is generally accepted that cometabolism of CAHs does not provide energy or carbon for cell growth. In
fact, cometabolism often results in the depletion of stored energy reserves in the cell (Alvarez-Cohen and
McCarty, 1991). This implies that an organic cosubstrate other than the CAHs is required for biological
growth and for the production of the necessary oxygenase enzymes that are used to degrade the CAHs. A
variety of growth substrates have been used for cometabolic CAH degradation, including methane
(Broholm et al., 1993), propane (Wackett et al., 1989), propene (Reij et al., 1995), aromatic compounds
including phenol (Hopkins et al., 1993), toluene (Wackett and Gibson, 1988), and isopropylbenzene
(Dabrock et al., 1992).
Important issues related to the engineering application of cometabolic CAH degradation include CAH
intermediate toxicity (Alvarez-Cohen and McCarty, 1991; Fox et al., 1990) and competitive inhibition.
Because CAHs are degraded by the same nonspecific enzymes responsible for the degradation of the
growth substrate, competition between the CAH and the growth substrate is known to occur.
Intermediate toxicity during CAH degradation and competition can reduce the degradation rates of both
the growth substrate and the CAH. In addition, multiple CAHs can compete (Strand et al., 1990) or create
toxic conditions that can affect CAH degradation rates (Bielefeldt, 1995).
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2.5 BIOTRANSFORMATION OF HALOGENATED AROMATIC COMPOUNDS
Halogenated aromatic compounds (HACs) include all compounds containing halogenated aromatic rings,
including aryl halides such as DDT, dioxins, and dibenzofurans. While the more volatile compounds
such as hexachlorobenzene and pentachloroanisole (Simonich and Kites, 1995) may volatilize after they
are released into the environment and become distributed globally, less volatile compounds such as
p,p'-DDT [p,/?'-dichloro-(bis)-l,l-diphenyl-l,l,l-trichloroethane] and its degradation products/?,/?'-ODD
andp,p '-DDE \p,p '-dichloro-(bis)-l,l-diphenyl-l,l-dichloroethane] tend to remain in place in soils or
sediments to which they are originally bound (Adriaens et al., 1999). Thus, the fate and
biotransformation of HACs is of particular interest for soils where many of these compounds are retained.
Aerobic biodegradation pathways for bacterial and fungal growth on HACs have been reviewed
extensively for halogenated phenols, benzoic acids, benzenes, pesticides, anilines, and herbicides
(Rochkind-Dubinsky et al., 1987; Haggblom, 1992; Neilsen, 1990; Engesser and Fisher, 1991;
Commandeur and Parsons, 1994; Adriaens et al., 1999). While the initial oxidation steps may be carried
out by a variety of enzymes, only a limited number of intermediates are produced; they include
dihydroxylated benzoic acids and substituted catechols. Further degradation of these metabolites via
either ortho or meta ring fission leads to intermediates of central metabolic pathways such as the
tricarboxylic acid cycle (Adriaens et al., 1999). Despite the success of aerobic bacteria to degrade a
variety of HACs, not all HACs are easily degraded aerobically. HACs only serve as sources of carbon
and energy for aerobic or anaerobic bacteria if they can be dehalogenated prior to or after ring fission.
Alternatively, these compounds may be degraded cometabolically, a process during which enzymes or
proteins break the aromatic ring but the bacteria are unable to derive carbon or energy from the HAC.
Under anaerobic conditions, reductive dechlorination represents the most common detoxification method
for HACs (Mohn and Tiedje, 1992). There is increasing evidence that reductive dechlorination is an
energetically favorable reaction in which bacteria use the HACs as electron sinks for energy (Dolfing and
Harrison, 1992; Mohn and Tiedje, 1992, 1991, 1990; Holliger and Schraa, 1994). Alternatively,
dechlorination reactions may be used by bacteria to detoxify contaminants or they may occur fortuitously
cometabolically (Mohn and Tiedje, 1992). However, because HACs in the environment are often present
at trace or ultratrace concentrations, little measurable energetic benefit can be expected to be obtained
from halide respiration (halorespiration) (Adriaens et al., 1999), and transformation of HACs may be
considered a form of secondary metabolism. This makes the engineered application of anaerobic
dechlorination increasingly difficult for contaminated soils where HACs are present at trace
concentrations. Whether microorganisms in soil matrices grow on aryl halides, degrade them via
secondary metabolism, or transform them cometabolically, ultimately the metabolic pathways converge
into one of three possible intermediates, substituted halocatechols or dihydroxybenzoates, which further
degrade and serve as growth substrate for numerous microorganisms. However, in the environment,
dechlorination and mineralization represent only one of a number of possible pathways for aryl halides;
microbially mediated conjugation, polymerization, and reactions with natural humic substances also
determine the fate of the aryl halides. The relative contribution of each of these processes depends largely
on the chemical reactivity of the aryl halide, the physiochemical characteristics of the soil matrix, and the
biological component of the soils (Adriaens et al., 1999).
The susceptibility of the aryl halide compound to oxidative or reductive microbial degradation depends
largely on the oxidation state of the aryl compound. Increased chlorination results in a more oxidized
form of the aryl compound. Thus, not surprisingly, highly chlorinated compounds are more susceptible to
reductive dechlorination while they tend to be difficult or impossible to oxidize aerobically. (The
difficulty for aerobic bacteria to oxidize highly chlorinated aryl halides also may be due to the
unavailability office adjacent carbons on the aromatic ring; it is generally observed that mono- and
dioxygenases will not act on aryl halides with more than two chlorines per ring [Adriaens et al., 1999].)
20
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As Gibbs free energy associated with lesser chlorinated compounds decreases (Dolfmg and Harrison,
1992), oxidative reactions become more favorable (Adriaens et al., 1999). The position of the aryl
halogen also exerts an influence on microbial processes, where certain ortho-, meta-, or/>ara-chlorine
positions are preferentially dechlorinated. For example, chlorobenzene dechlorination under anaerobic
conditions generally requires the presence of adjacent halogens. Isolated halogens on chlorobenzenes,
such as 1,3- or 1,4-dichlorobenzene or 1,3,5-trichlorobenzene are relatively recalcitrant to dechlorination
(Fathepure et al., 1998).
The primary factor that determines which microorganisms will degrade aryl halides in soils is the
availability of electron acceptors and an alternative carbon source. Bacteria will utilize electron acceptors
in order of the most energetically favorable, beginning with oxygen respiration, followed in sequence by
nitrate-, iron(III)-, manganese(IV)-, and sulfate-reduction, ending with methanogenesis. Aryl halide
reductive dechlorination is strongly dependent on the prevailing electron acceptor process and the
availability of an electron donor. Dechlorination of aryl halides is most favorable under methanogenic
and sulfate-reducing conditions, although sulfate reduction also can be inhibitory to reductive
dechlorination (Mohn and Tiedje, 1992; H'ggblom, 1992; H'ggblom and Young, 1995; Adriaens et al.,
1999). Alternatively, the metabolism of aryl halides also varies depending on the primary electron
acceptor process. Hซggblom et al. (1993) investigated the effects of nitrate, sulfate and carbonate
(methanogenesis) as electron acceptors on anaerobic metabolism of monochlorinated phenols and benzoic
acids in freshwater and marine sediments. The respective denitrifying, sufidogenic, and methanogenic
enrichments all were capable of utilizing at least one chlorophenol or chlorobenzoate, but none was
capable of utilizing all six compounds tested. The variety of reports of dechlorination and metabolism of
chlorinated compounds under varying electron acceptor conditions indicates that generalities and
subsequent predictions may be impossible to make at some sites, regarding the degradability of aryl
halogens, without bench-scale testing using site-specific soils and environmental conditions.
2.6 BIOTRANSFORMATION OF DIOXIN-LIKE COMPOUNDS
Dioxins (e.g., polychlorinated dibenzo-p-dioxins [PCDDs], and polychlorinated dibenzo-furans [PCDFs])
are ubiquitous in the environment at subparts-per-million concentrations. In general, these compounds
are unwanted byproducts of combustion and industrial synthesis, and very little is known about their fate
in soils. Under reduced conditions, PCDD have been successfully dechlorinated to mono-, di-, and
trichlorinated dibenzo-p-dioxins (Barkovskii and Adriaens, 1996, 1998). Aerobically, the lesser
chlorinated congeners have been cometabolized to form hydroxylated chlorodiphenylethers and
chlorocatechols (Klecka and Gibson, 1980; Fortnagel et al., 1990). Similarly to the biodegradation
processes of PCBs, the overall fate of PCDD may be dependent on an anaerobic/aerobic sequence of
reactions to completely mineralize these compounds and remove them from the environment (Adriaens et
al., 1999).
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3.0 EXISTING BIOREMEDIATION TECHNOLOGIES AND APPLICATIONS
This section describes existing soil bioremediation technologies. Table 3-1 shows the technologies that
are discussed and categorizes them as conventional or emerging. For the purposes of this report,
conventional technologies are those that have been deployed at full scale and can be implemented readily
without further significant research and development. Emerging technologies are those that have not yet
been significantly implemented at full scale but currently are undergoing or could undergo further
research and development to bring them to the field. An abundance of cost and performance data are
available for soil bioremediation using the conventional technologies. Not surprisingly, cost and
performance data generally are unavailable for the emerging technologies. Furthermore, most of the
conventional technologies may be considered emerging or even developmental in cases where they may
be applied to complex soil conditions and increasingly recalcitrant contaminants.
Table 3-1. Technology Maturity
Technology
Land Treatment
Biopile
Biocell
Composting
Bioslurry Reactors
Aerobic Bioventing
Cometabolic Bioventing
Anaerobic Bioventing
Phytoremediation
Sequential Anaerobic/Aerobic Treatment
Natural Attenuation
Stage of Development
Conventional
S
s
s
s
s
s
Emerging
S
S
^
^
^
Section 3.1 describes the conventional technologies with respect to their principles of operation,
specifically, the biological aspect of the technology, target contaminants, advantages and limitations, and
the technology cost and performance as reported in available literature. Section 3.2 describes emerging
technologies with respect to their principles of operation and future directions. Cost and performance
data are reported for these technologies where available, but in general little such data were found for
these technologies in published reports and documents.
For this report, a review of existing cost and performance data was conducted using on-line searches of
databases including the National Technical Information Service (NTIS), EPA, Federal Remediation
Technologies Roundtable (FRTR), Remediation Information Management System (RIMS), EPA
Remediation and Characterization Innovative Technologies (REACH-IT), Bioremediation in the Field
Search System (BFSS), and Lexis-Nexis. These databases provided cost and performance data for
large-scale demonstrations. However, they did not usually provide extensive technology descriptions or site
information.
As much as possible, full-scale or large-demonstration-scale cost and performance data were obtained and
are reported in this study. A detailed description of how these data were obtained and used for this report is
provided in Appendix A. All applicable cost and performance information was input into a Microsoft
Access database. Site-specific reports generated from the database are provided in Appendix B. Figure 3-1
is a "box and whisker" plot showing the range of costs per cubic yard of the various conventional
22
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technologies identified in the literature and evaluated in this section. Table 3-2 shows the data used to
generate Figure 3-1. In the box and whisker plot, the bottom and top edges of the box are at the 25th and
75th percentiles of the sample set; the horizontal line in the box is at the median (50th percentile); and the
upper and lower whiskers drawn from the box show the most extreme data point within 1.5 interquartile
ranges (i.e., 1.5 times the distance between the 25th and 75th percentiles). The data points with more
extreme values are marked with a star on the plot. Each technology shows a wide range of costs, which
may be influenced by a number of site-specific factors including technology performance; duration of
remedial activities; and requirements for monitoring, pilot testing, and/or design, to name a few. The
highest costs are associated with composting, biopile/biocells, and slurry-phase bioreactors, while the
lowest costs are associated with bioslurping, conventional bioventing, and phytoremediation. The wide
range of costs associated with biopile/biocells and land treatment is caused by application at a wide
variety of sites under highly variable treatment conditions.
10000
1000
u
-------
Table 3-2. Costs per Cubic Yard for Remedial Actions: Data used to make Figure 3-1
Median
25th
Percentile
75th
Percentile
High
Whisker0"
Low
Whisker(a)
Cost per Cubic Yard for Remedial Technologies
Land
Treatment
(30 sites)
$77
$26
$115
$183
$8
Biopile/
Biocell
(17
sites)
$106
$40
$151
$233
$22
Com-
posting
(10 sites)
$281
$202
$332
$465
$166
Slurry-
phase
Bioreactor
(7 sites)
$200
$112
$319
$335
$97
Conventional
Bioventing
(12 sites)
$23
$11
$86
$152
$3
Enhanced
Bioventing
(9 sites)
$73
$56
$106
NA
$10
Bio-
slurping
(2 sites)
$32
$28
$35
NA
NA
Phyto-
remediation
(2 sites)
$16
$9
$23
NA
NA
(a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges, i.e., 1.5 times
the distance between the 25th and 75th percentiles.
3.1 BIOREMEDIATION TECHNOLOGIES AND THEIR APPLICATIONS - CONVENTIONAL
TECHNOLOGIES
This section describes conventional technologies with respect to their principles of operation, the types of
contaminants for which they are appropriate, and cost and performance data reported in the literature.
The technologies discussed in this section include land treatment, biopile/biocell treatment, composting,
bioslurry treatment, conventional bioventing, and enhanced bioventing. The first four technologies are
ex-situ processes, while the next two are in-situ processes. Bioslurping, an in-situ product recovery
technology, is discussed within the bioventing section because the principal biological component of
bioslurping is the venting of the vadose zone, comparable to bioventing. Phytoremediation is an in-situ
technology discussed in Section 3.2.2.
3.1.1 Land Treatment
The term "land treatment" is used to refer to the technology of ex-situ treatment of contaminated soils,
sediments, and sludges in engineered treatment cells or soil lifts designed to provide some level of
process monitoring and control. Land treatment often includes the addition of nutrients, water, and
externally cultured microorganisms, but not large-scale addition of organic materials. Land treatment has
evolved from land farming as more stringent regulations limiting land application of wastes were enacted.
Today, the technology is applied at full-scale to treat a wide range of contaminant types, with most
success seen in treatment of total petroleum hydrocarbons (TPH), although land treatment of PAHs has
shown good results as well.
3.1.1.1 Principles of Operation
Land treatment is typically conducted on a prepared pad that provides some containment and allows for
desired levels of process monitoring and control. The treatment areas range from pads formed from
asphalt or soil covered with impermeable liners to compacted clay. Because the units are exposed to the
weather they are usually equipped with leachate collection and storage systems. Land treatment cells do
not typically include forced aeration systems, and aeration occurs either passively from exposure to the
atmosphere or through surface soil mixing or tilling. Figure 3-2 shows a field application of pilot land
treatment that is hand-tilled, and Figure 3-3 is a schematic of full-scale land treatment.
24
-------
Figure 3-2. Land Treatment
Groundwater
Monitoring Leachate Collection
Wells and Treatment
Contaminated
Soil-
(Optional)
Tilling for Soil Aerationi \Berm
Figure 3-3. Land Treatment Schematic
Laboratory treatability tests are often conducted to determine the biodegradation potential of target
contaminants and the need for amendments to optimize the biodegradation process. Typical amendments
could include moisture, nutrients, pH adjustment, and/or microorganisms. Adding microorganisms,
referred to as bioaugmentation, may increase both the rate and extent of contaminant removal and is being
applied at many demonstration sites.
During land treatment, contaminated soil is excavated, screened to remove rocks and debris as necessary,
mixed with appropriate amendments, placed into lifts in the treatment area, and then allowed to incubate
to affect contaminant destruction. During incubation, the soil is routinely tilled, and the soil temperature,
pH, and moisture are monitored and controlled as necessary. Soil tilling helps promote mixing, aeration,
and bioavailability of contaminants for microbial degradation. Soil-gas oxygen levels can be monitored
to ensure aerobic conditions. If oxygen becomes limiting, lift depths can be reduced and/or tilling
frequency increased.
Monitoring land treatment performance requires collection of grab samples and analyses for target
contaminants. Treatment is complete when the analytical results show that the cleanup goals have been
25
-------
met or when sequential sampling and analyses show that no further treatment is being achieved. If the
cleanup goals are attained, the soils can be used or discarded accordingly. If treatment stops short of
achieving the cleanup goals, the soil may need to be treated further or disposed of according to pertinent
restrictions. The potential uses for the treated soils and the requirements for handling soils that fail to
meet the cleanup criteria must be determined prior to implementation of land treatment.
3.1.1.2 Target Contaminants
Land treatment usually is applied to treat compounds that are directly metabolized and cannot be easily
removed through volatilization. The list of contaminants successfully treated with land treatment include:
TPH
BTEX
Gasoline
Diesel fuel
JP-5 and other jet fuel
Fuel oils
PAHs (higher-molecular-weight compounds are more difficult to degrade)
Creosote
Coke wastes
PCP and other chlorinated phenolics
Non-chlorinated phenolics
Chlorinated benzenes
Certain pesticides
- Dinosep
- 2,4-D
- 2,4,5-T.
3.1.1.3 Advantages and Limitations
The primary advantages of using land treatment include:
The process is destructive with the contaminants being transformed into innocuous end products
The system can be covered and left dormant over winter months during low biological activity
Monitoring allows for control of contaminant migration
The cost of land treatment is usually lower than alternatives such as incineration or hauling and
disposal in a secured landfill
Following remediation, the site can be converted to beneficial uses.
The primary limitations of land treatment include:
Land treatment is land and management intensive
Climatic conditions strongly affect biodegradation
Lift depth is limited by depth of tilling
Volatile emissions and/or dusts can be a nuisance and may pose a health threat
Improper design and/or operation can result in an adverse environmental impact
On-site waste storage is often required
Site selection and permitting may be time consuming and public reaction may be negative.
26
-------
3.1.1.4 Technology Cost Drivers
The major cost drivers for land treatment are as follows:
Land treatment requires large treatment areas
Soil type and composition affect aeration efficacy; permeable, low moisture soils are relatively
easily aerated while silty/clayey soils with high moisture contents are difficult to aerate and
require more extensive and more frequent tilling; presence of large rocks and debris can
interfere with tilling efficiency and may require removal or prescreening
Contaminant type and degradability impact the treatment duration
Operations and maintenance (O&M) considerations include tilling frequency and extent,
periodic water addition, and meeting nutrient requirements
If treated soils cannot be returned to the site, additional post-treatment disposal costs may be
incurred
Volatile or dust emissions may require control measures.
3.1.1.5 Technology Performance and Cost
Land treatment has been applied at full scale at many sites including several Superfund sites. Data from a
number of these case histories are summarized in Table 3-3. The data show that in most cases, land
treatment was successfully applied with the treatment goals achieved and the sites closed. A few of the
case histories indicated that the technology was not effective at achieving remedial goals and further
treatment or controlled disposal of residuals was required.
Range of sites identified Land treatment is applicable at a wide range of sites. It requires excavation of
materials for treatment, and soils can be amended during processing for cell loading. The ability to
economically excavate the soils is key to selecting the technology with depth limitations usually in the
25- to 30-ft below ground surface (bgs) range. More typically, the soils treated by land treatment come
from less than 5 ft bgs. Contaminant type has more of an influence on the selection of this technology
than hydrogeologic and/or soil characteristic constraints. The technology has not been used to treat
inorganic contaminants nor contaminants that require a cosubstrate. Research into these uses is
proceeding, but no successful applications have been reported in the literature.
Technology performance. Table 3-3 presents performance data for land treatment applications for a
number of contaminant classes. The data show that treatment performance including treatment times and
achievable levels are dependent on the contaminant type and concentrations. Other factors that affect
performance include, but are not limited to, soil type, temperature, moisture, waste loading rates,
application frequency, aeration, volatilization, and other site-specific factors.
Figure 3-4 presents box and whisker plots showing the range of starting concentrations (Figure 3-4a) and
ending concentrations after treatment (Figure 3-4b) for various contaminants subjected to land treatment.
The figures depict significant contaminant removals; not all the contaminants identified before treatment
were necessarily monitored after treatment, resulting in fewer contaminants identified in Figure 3-4b than
in Figure 3-4a. Table 3-4 shows the data used to create Figure 3-4.
Technology costs. Land treatment is a low-cost alternative to the more conventional thermal or
physical/chemical treatment technologies with cost estimates typically ranging from $30 to $50/cu yd.
Pretreatment costs include $25K to $50K for laboratory studies and up to $100K for pilot tests. The data
in Table 3-3 show costs as high as $l,754/cu yd, with the higher costs associated with small treatment
volumes and/or highly recalcitrant contaminants. Table 3-3 reflects an inability to gather complete data on
27
-------
all sites either from a reluctance of site owners to impart the information or from missing site data.
Additional site data on land treatment is available in Appendix B.
3.1.2 Biopile and Biocell Treatment
Biopiles and biocells are designed to treat contaminated soil that is excavated, mixed with appropriate soil
amendments, placed in a heap configuration on a pad or prepared surface, and oxygenated through forced
aeration. The main difference between the two treatments is that while a biopile is a free-standing pile of
soil, biocells are contained by walls or sides (e.g., stacked hay bales or large metal dumpsters). The
microbial processes promoted during biopile/biocell treatment are similar to land treatment, but the
incorporation of forced aeration relieves the necessity for tilling and reduces space requirements. Biopiles
and biocells exploit the activity of microorganisms that can thrive using the contaminant as a substrate for
growth and obtaining energy. As such, they are most effective for treating readily degradable
contaminants such as petroleum hydrocarbons. The goal of biopile/biocell treatment is to convert target
contaminants to innocuous products, rendering the soils safe for on-site disposal or other beneficial uses.
The biopile/biocell treatment technology was developed by the Navy to treat petroleum hydrocarbons,
and most of the earliest applications focused on various fuel compositions (see Table 3-3). Recent
applications have expanded the application of biopiles and biocells to treat chlorinated solvents, creosote
compounds, and PAHs. More sophisticated designs for applications that incorporate cometabolic or
anaerobic metabolism are currently in the demonstration stage and have not yet been used for full-scale
application. Such applications require relatively complex control systems to operate and maximize
system performance. Nonetheless, these systems have the potential to expand the contaminant list for this
technology to more effectively treat chlorinated solvents, PAHs, and explosives.
3.1.2.1 Principles of Operation
Biopiles and biocells are used to treat contaminated soil that has been excavated and cannot be placed
back into the ground without contaminant removal and/or soil that must be detoxified prior to off-site
disposal. Biopiles and biocells are constructed by mixing excavated soil with appropriate amendments
(e.g., nutrients, chemical additives to adjust pH, or bulking agents to enhance aeration), then placing the
mixture in a heap configuration on a platform or within a containment system. For small-scale temporary
facilities, the biopiles and biocells can be constructed with a simple high-density polypropylene liner laid
flat or within hay bale 'walls', and with a simple aeration and leachate collection system installed during
construction. The more complex facilities include engineered concrete pads or boxes with built-in
leachate collection and aeration systems.
Aeration systems are required to inject or pull air through the soil. The selection of the operating mode
depends on the volatility of the contaminant and regulatory concern over vapor control. Aeration systems
in cold weather climates must be heat traced to prevent freezing and maintain optimal microbial
conditions, and therefore require knockout systems to remove moisture from the aeration plumbing. The
piles and cells can be covered with plastic to minimize leachate; promote solar heating; and control
runoff, evaporation, and volatilization.
28
-------
Table 3-3. Summary of Site Characteristics at Land Treatment Installations*
Site Name
Gila Indian
Reservation
Rancho Vistoso
Properties
Technical Products,
Inc.
Chevron USA
Products Company
Middle Mountain
Silvex
Popile Superfund Site
Navy Demo Camp
Pendleton, CA
Matagorda Island Air
Force Range
Northern Arizona
Mobil Station 18-566
Burgan oil Field
Kohler Company
manufacturing facility
Site
Status
Closed
Closed
Closed
Closed
Closed
Active
NA
Closed
Closed
Inactive
Active
Closed
Contaminants
of Concern
parathion;
toxaphene
petroleum
(diesel)
1 2-DCE 1 2-
DCA, 1,4-
dichlorobenzen
e, B, methylene
chloride, TCE,
T
BTEX, TPH
2,4,5-
trichloropheno
xyacetic acid,
2,4-
dichlorophenox
yacetic acid
BAP, PCP
TPH
BTEX jet fuel,
TPH
butyl benzyl
phthalate, urea
crystals
diesel/waste oil
Weathered
crude oil
kerosene
(DRO)
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
parathion: 2,000
mg/kg; toxaphene:
30,000 mg/kg
43,000mg/kg
1924 1 4-
dichlorobenzene;
706 1,2-DCE; 612
1,2-DCA; 147 TCE;
236 B; 2273 T; 1120
methylene chloride
26,000 TPH; 8900
BTEX
2,4,5-T, 510mg/kg;
2,4-D, NA
BAP: 21; PCP: 200
29,000
41. 3 mg/kg BTEX
jet fuel; 3400 mg/kg
TPH
phthalate:
38,000mg/kg
660mg/kg
TPH, 67,000; PAHs,
26.5
1600
Unit
Cost
(per cy)
$9
$56
<$113
$25
$35
$85
NA
$87
$180
$38
NA
~$26
Numerical
Treatment Goals
(mg/kg dry weight)
NA
7,000 mg/kg
<1 mg/kg all
contaminants
NA
50mg/kg 2,4,5-T;
NA 2,4-D
PAHs: 3mg/kg; PCP:
88
30mg/kg BTEX jet
fuel; 0.5mg/kg B;
70mg/kg E;
lOOmg/kg T;
lOOmg/kg TPH;
lOOOmg/kg X).
NA
lOOmg/kg
NA
100 DRO
Treatment
Goals
Achieved
NA
Yes
Yes
NA
NA
NA
NA
Yes
NA
Yes
Yes
Yes
Concentration
Achieved
(mg/kg dry
weight)
NA
438 mg/kg
<1 mg/kg all
contaminants
NA
NA
NA
NA
Img/kg BTEX jet
fuel; 0.4mg/kg
benzene,
Ethylbenzene,
Toluene; 60mg/kg
TPH; 10.4mg/kg
Xylene
<90mg/kg phthalate
12mg/kg
TPH, 27,500,
PAHs, 3.45.
ND
Treatment
Effectiveness
Biological treatment
would have been
successful if the
neutralization after
chemical treatment had
been complete.
NA
Contaminant reduction to
acceptable limits occurs
in 6-9 months, some sites
remediated to closure
levels within 3 months.
NA
NA
NA
NA
Tar and asphalt were not
easily consumed;
bioremediation proved
successful in cleaning up
BTEX and TPH
contamination levels.
99% reduction
TPH reduction 59%
PAHs reduction 87%
91% reduction DRO and
100% reduction PVOC
Treatment
Duration
(Months)
24
1
14
NA
NA
180-240
NA
3 mo.
(TPHs); all
others NA
20
0.5
8
13
to
-------
Table 3-3. Summary of Site Characteristics at Land Treatment Installations* (continued)
Site Name
Poly-Carb
Seabury Chevrolet
Tucson Rock and
Sand, Inc.
Fort Greeley UST Soil
Piles
McKesson
Envirosystems
Vandalia Road Site
Bonneville Power
Administration Ross
Complex, Operable
Unit A, Wood Pole
Storage Area
Scott Lumber
Company Superfund
Site
Libby Superfund Site
Brown Wood
Preserving Superfund
Site
Hayford Bridge Road
Toote Mineral
Eastman Chemical
Company
Site
Status
Inactive
Closed
Closed
Closed
Closed
Inactive
Closed
Closed
Active
Closed
Active
Closed
Closed
Contaminants
of Concern
cresol, phenols
diesel/waste oil
diesel/waste oil
Gasoline,
diesel fuel,
BTEX
N,N-dimethyl-
aniline, A,
aniline, BTEX,
methanol,
methylene
chloride, TCE
PAHs
PAHs, PCP
PAHs, BAP
PAHs, PCP
PAHs (primary
constituents in
creosote)
PCB
TPH
Petroleum
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
cresol: 1,000;
phenols: 8,000
36,000mg/kg
8,900mg/kg diesel,
waste oil
3,000 mg/kg gasoline,
1,200 mg/kg diesel, 20
mg/kg BTEX
13,072 mg/kg
methanol; l,830mg/kg
N,N-dimethylaniline;
833mg/kg A;
11.5mg/kg B; 49mg/kg
E; 827mg/kg
methylene chloride; 17
TCE;218mg/kgX.
10,000
150mg/kg PAHs; 62
mg/kg PCP
63,000mg/kg PAHs;
23mg/kg BAP
Carcinogenic PAHs:
7,384;
noncarcinogenic :
26,555; PCP: 2,700.
208 mg/kg TCIC (total
carcinogenic indicator
chemicals)
100,000 mg/kg
500 mg/kg
200 mg/kg BTEX;
1,500 mg/kg TPHs
Unit
Cost
(per
cy)
$544
$8
$40
$76
$115
NA
$470
$1,754
NA
$99
$33
$77
NA
Numerical
Treatment Goals
(mg/kg dry weight)
cresol: 10;
phenols: 20.
7000mg/kg
lOOmg/kg diesel,
waste oil
100 mg/kg DRO; 50
mg/kg GRO; 0.1
mg/kg B; 10 mg/kg
BTEX
10 mg/kg for all
contaminants.
NA
Img/kg PAHs, 8
mg/kg PCP
preferred. If not,
<23mg/kg H
PAHs, <126mg/kg
PCP.
500mg/kg total
PAHs; 14mg/kg
BAP
carcinogenic PAHs:
88; naphthalene: 8;
phenanthrene: 8;
pyrene: 7.3; PCP: 37
TCIC (total
carcinogenic
indicator chemicals)
100 mg/kg within 2
years
25 mg/kg
100 mg/kg
10 mg/kg BTEX;
100 mg/kg TPHs
Treatment
Goals
Achieved
No
Yes
Yes
Yes
Yes
NA
Met tier 2
treatment
goals,
not tier 1 .
Yes
NA
Yes
NA
Yes
Yes
Concentration
Achieved (mg/kg
dry weight)
cresol: 20;
phenols: NA
400
40mg/kg diesel,
waste oil
Nondetectable
levels of GRO
N,N-
dimethylaniline:
4.1; acetone: 1;
aniline: 4.1; BTEX:
0.63; methanol: 1;
methylene chloride:
0 63- TCE- 0 63
NA
6.8 to 21. 8 mg/kg H
PAHs, 6.8 to
20.7mg/kg PCP
130mg/kg PAHs (6
mo.); 8mg/kg BAP
NA
23 mg/kg TCIC
NA
25 mg/kg
1 mg/kg BTEX; 5
mg/kg TPHs
Treatment
Effectiveness
NA
NA
NA
Reduced to below
ADEC level A
cleanup standards in
all but two samples.
NA
reduced total PAHs
by 51%.
Land treatment could
not meet Tier I
cleanup goals for all
soil at the site, so a
gravel cap was
installed.
NA
NA
Cleanup goal
achieved within 1 8
months using land
treatment
NA
NA
NA
Treatment
Duration
(Months)
14
3
3
24
2
4
2.75
6
72
18
24
2
4
OJ
o
-------
Table 3-3. Summary of Site Characteristics at Land Treatment Installations* (continued)
Site Name
Novartis Site
Handy Andy, Inc.
River Site
Whitehorse Airport
Former Golden Eagle
Refinery
Domtar, Inc
Domtar Inc
Hellman Lease
Old Seattle Marketing
Fuel Terminal
Great Falls
International Airport
Burlington Northern
Tie Plant
Idaho Pole Site
BN Somers Site
Treating Plant
Site
Status
Inactive
Closed
Active
Closed
Closed
Closed
Active
Closed
Active
Closed
Closed
Active
Closed
Active
Contaminants
of Concern
Metolachlor
petroleum
(benzene,
BTX)
TPH
TPH, BTEX
TPH,
carcinogenic
TPM
chlorinated
phenols, PAHs,
TPH
PAHs,
naphthalene,
BAP,
benzo(a)anthra
cene,
indeno(l,2,3-
cd)pyrene,
PCP,
phenanthrene,
pyrene.
l)diesel; 2)
kerosene,
gasoline
TPH
jet fuel
PAHs
Dioxins PAHs
pesticides,
CPAHs
PCP
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
170mg/kg
15.8 mg/kg BTEX;
0.056 mg/kg benzene
17,000 mg/kg
TPH: 3,900, BTEX:
480
25,000mg/kg TPH;
20,000mg/kg
carcinogenic TPM
chlorinated phenols:
700; PAHs: 2000;
TPH: 8000.
PCP: 266; PAHs:
11 82; naphthalene:
3.87; phenanthrene:
23.3;
benzo(a)anthracene :
68.6; BAP: 35.7;
indeno(l,2,3-cd)
pyrene: 24.2;
dibenzo(a,h)anthracene
: 9.25; pyrene: 309.
7,000mg/kg diesel;
4,000mg/kg kerosene;
4,000mg/kg gasoline
2,660mg/kg (avg.)
120,000mg/kg
500,000 mg/kg
PCP: 3,800; TCDD:
0.0342
200
70
Unit
Cost
(per cy)
$132
NA
$800
$107
NA
NA
NA
NA
NA
NA
$183
NA
~$20
~$20
Numerical
Treatment Goals
(mg/kg dry weight)
NA
11. 705 mg/kg
BTEX; 0.005
mg/kg benzene
100 mg/kg
TPH: 2000; BTEX:
50
3,000mg/kg TPH,
carcinogenic TPM
chlorinated phenols:
5; PAHs: 100;
TPH: 100
PCP' 5' PAHs
phenanthrene,
pyrene: 100;
naphthalene: 50; all
others' 10
diesel, l,000mg/kg
TPH; kerosene,
lOOmg/kg TPH;
gasoline, lOOmg/kg
TPH.
200 mg/kg TPH
400mg/kg TPH
36 mg/kg
PCDD/PCDF
w/TCDDTE: 0.001;
PCP: 48; B2PAHs:
15;DPAHs: 145.
57 mg/kg
34 mg/kg
Treatment
Goals
Achieved
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
Yes
NA
NA
NA
NA
Concentration
Achieved (mg/kg dry
weight)
1 mg/kg
<11.705mg/kg
BTEX;
<0.005mg/kg B
1 mg/kg
TPH: 9; BTEX: NA
lOOmg/kg
carcinogenic TPM,
NATPH
chlorinated phenols: 1;
PAHs' 35' TPH' 25
NA
diesel, lOmg/kg;
kerosene and
gasoline, 5ug/kg
200mg/kg TPH
130 mg/kg
NA
NA
NA
NA
Treatment
Effectiveness
NA
NA
NA
NA
NA
NA
Reduced all
contaminants to
below treatment
goal.
NA
NA
NA
NA
NA
NA
NA
Treatment
Duration
(Months)
18
18
6
12
NA
6
6
3
36-60
5
24
120
6
3
-------
Table 3-3. Summary of Site Characteristics at Land Treatment Installations* (continued)
Site Name
Alcoa Land Treatment
Unit (LTU) study
Site
Status
Active
Contaminants
of Concern
PAHs, PCBs
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
PAHs, 1,662; PCB,
113
Unit
Cost
(per cy)
NA
Numerical
Treatment Goals
(mg/kg dry weight)
NA
Treatment
Goals
Achieved
NA
Concentration
Achieved (mg/kg
dry weight)
PAHs, 124; PCB,
106 maximum
concentrations
Treatment
Effectiveness
NA
Treatment
Duration
(Months)
Active, 2-
3; Passive,
66+
B = benzene
T = toluene
E = ethylbenzene
X = xylenes
N = naphthalene
MEK = methylethylketone
BAP = benzo(a)pyrene
A = acetone
TCE = trichloroethylene
DCE = dichloroethylene
PAHs = polycyclic aromatic hydrocarbons
PCBs = polychlorinated biphenyls
TPH = total petroleum hydrocarbons
DCA = dichloroethane
CPAHs = carcinogenic polycyclic aromatic
hydrocarbons
CB = chlorobenzene
NA = not available
*Eight sites included in Appendix B were not included in this table due to insufficient information. The excluded sites are Site ID Nos. 03-008,
05-012, 08-012, 08-013, 09-024, 09-025, 10-006, and 10-014.
Table 3-4. Concentrations of Contaminants of Concern Before and After Land Treatment: Data Used to Generate Figure 3-4
Median:
25th
Percentile
75th
Percentile
High
Whisker01
)
Low
Whisker'3
)
Concentrations of Contaminants of Concern
Chlorinated VOC/SVOC
Before
Treatment
(11 sites)
612
144
767
1,120
52
After
Treatment
(9 sites)
1
1.0
1.0
NA
NA
Pesticide, Herbicide
Before
Treatment
(3 sites)
2,000
1,085
16,000
30,000
170
After
Treatment
(24 sites)
18.5
8
147.5
200
1
Fuel / Oil
Before
Treatment
(26 sites)
8,450
2,745
25,750
43,000
500
After
Treatment
(8 sites)
1
0.575
1
NA
0.1
BTEX
Before
Treatment
(12 sites)
125
19
297
480
12
After
Treatment
(8 sites)
28.5
16.25
98.5
130
3
PAH, Creosote
Before
Treatment
(12 sites)
2,500
188
10,095
10,380
11
After
Treatment
(1 site)
106
106
106
NA
NA
OJ
to
(a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges, i.e., 1.5 times the distance between the 25th and 75th
percentiles.
-------
1 ,000,000
100,000
_ ^-v
]S ฃ 10,000
CO i^..
ฅ~ฐ
ซ_ฃฃ 1,000
H^ 100
= =
ฐ.2
T. "ซ
ah 10
= a
.S w
^ a
^ o
0.1
0.01
*
* *
n Pi
_ * i i
* _ nr *
F I
T T
I
IE 1
Chlorinated Pesticide, Fuel / Oil BTEX PAH,
VOC/SVOC Herbicide (26) (12) Creosote
(11) (3) (12)
Contaminant
1 ,000,000
100,000
^_^
'l^s-S
ฃ ^ 10,000
-^a
"a ep 1,000
ซky '
^M
.S M
S S
ซ^ 100
o-2
ซ 2 10
Sa
||
U
0.1
0.01
*
*
f
JL
n ~~
* -
- i , !
n
1
Chlorinated Pesticide, Fuel / Oil BTEX PAH,
VOC/SVOC Herbicide (8) (8) Creosote
(9) (24) (1)
Contaminant
Figure 3-4. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before
and After Land Treatment. (See Appendix B for data used to generate these plots.)
33
-------
The basic underlying principle of biopile/biocell treatment is the promotion of aerobic biodegradation by
microorganisms that are provided with a sufficient supply of oxygen as an electron acceptor and other
amendments (e.g., nutrients) as necessary. The contaminants serve as growth substrates for the
microorganisms that carry out the desired biological reactions. The contaminants and/or their degradation
products can also supply other microorganisms with a needed nutrient/substrate that supports any number
of symbiotic processes. As such, the contaminants must be both aerobically biodegradable and available
to the microorganisms. For this reason, biopiles and biocells must be carefully constructed to optimize
conditions for contaminant biodegradation by maximizing airflow and oxygenation of the pile and
optimizing the distribution of additives such as nutrients.
The excavated soil can be mixed with bulking agents to increase the effective porosity, which will
facilitate airflow. Unlike composting, the bulking agents are not added to supply carbon or nutrients, but
simply to modify the texture of the soil so that the soil mass can be maintained aerobic through forced air
movement. Nutrients, moisture, pH adjustment, and bioaugmentation can be applied during construction
to enhance process performance by reducing the time required for cleanup and/or to lower the achievable
level of treatment. The need for such amendments and the potential benefits that can be realized by
adding them may be determined through laboratory treatability testing.
Biopile and biocell applications for contaminants that require cometabolism or anaerobic degradation are
in the demonstration stage. These systems are more complex than the simple aerobic systems and require
a higher level of monitoring and control. Figure 3-5 is a schematic of a biopile treatment system.
Condensate
Collection
Tank
Blower
Treated Off-Gas
Extraction
Piping
Contaminated Soil
Clean Soil Base
.E/COMPQSTIHSQftQJMICSCDR
Figure 3-5. Biopile schematic
3.1.2.2 Target Contaminants
Currently, biopile/biocell treatment is applicable to compounds that are directly metabolized by the
microorganisms that can survive within the soil matrix. Typically, it is more desirable to stimulate and
promote the activity of indigenous microorganisms, but some success has been achieved using
bioaugmentation with cultures that have known and desired metabolic capabilities. This success has
34
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expanded the list of candidate contaminants, as well as enhanced the performance of the technology.
Biopiles and biocells have been successfully used to treat the following contaminants:
TPH
Gasoline (more volatile constituents tend to volatilize, not biodegrade)
JP-5 and other jet fuels
Diesel fuel
Motor oil
Transformer oil
PCP
TNT
PAHs (higher-molecular-weight compounds are more difficult to degrade)
DCE and VC.
3.1.2.3 Advantages and Limitations
The main advantages associated with biopile and biocell treatment include the following :
Systems are simple and easy to design
Application is not limited by hydrogeologic constraints (beyond depth of excavation) or soil
characteristics that limit in-situ technologies
Contaminants are degraded to innocuous end products; ultimately CO2 and water
Forced aeration allows increased height, requiring less land space than other ex-situ biological
treatment alternatives such as land treatment
Nutrients, water, and microorganisms can easily be added and mixed during construction to
accelerate the biodegradation process
Simple design minimizes operation and maintenance requirements
Biopiles (and, to a lesser extent, biocells) are relatively inexpensive due to low capital and O&M
costs.
The primary limitations for biopile and biocell treatment are:
Excavation is required
Space is required for materials handling, soil preparation, and staging
Controls may be required to prevent volatile emissions to the atmosphere
Concentration reductions greater than 95% and residual concentrations below 0.1 ppm are very
difficult to achieve
Contaminants must be readily biodegradable under aerobic conditions
Biopiles and biocells may not be effective for TPH at concentrations greater than 50,000 mg/kg.
3.1.2.4 Technology Cost Drivers
Major cost drivers of biopile and biocell treatment include:
Biopiles are less land intensive than land treatment, but still require relatively large treatment
areas
Cell size and the volume of soil to be treated will impact the number of cells required; larger cells
will require fewer cell units, but larger cells also are more difficult to construct and maintain
35
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Biopiles require frequent irrigation; leachate collection systems and treatment of leachate (if not
returned to the biopile cell) will impact costs
Volatile or dust emissions may require control measures
Soils with low porosity may require bulking agents to increase the airflow through the pile; soil
screening may be required to remove large rocks, debris, or other bulk objects
O&M considerations include maintaining moisture levels via water amendments and maintaining
nutrient levels and pH via nutrient and buffer amendments
Biopiles require nutrient and/or buffering agents to control pH; such agents are difficult and
expensive to administer once the biopiles are constructed
If treated soils cannot be returned to the site, additional post-treatment disposal costs may be
incurred.
3.1.2.5 Technology Performance and Cost
Biopiles and biocells have been in use for several years. Table 3-5 summarizes available data for biopile
and biocell applications at 19 sites.
Range of sites identified. As the data in Table 3-5 reflect, biopiles and biocells have been used primarily
at petroleum-contaminated sites. Because the soils can be amended during construction, the application
of the technology is not limited by the hydrogeologic constraints that limit in-situ technologies. Typical
criteria given for a successful biopile are:
The contaminant must be biodegradable
TPH values below 50,000 mg/kg can be tolerated; higher concentrations may be toxic
Heterotrophic bacteria should be present at densities of > 1,000 colony forming units (CFU)/gram
of dry soil
Soil pH should be between 6 and 9
Soil moisture should be maintained between 70% and 95% of field capacity.
Most of these variables can be adjusted during construction, and controlled during operation by adding
selected amendments.
Technology performance. The time to achieve treatment depends on a number of factors including
contaminant type, soil physical/chemical characteristics, the level of biological activity that can be
obtained, and the climate for outdoor applications. Typical treatment times range from 3 to 6 months.
Figure 3-6 presents box and whisker plots showing the range of starting concentrations (Figure 3-6a) and
ending concentrations (Figure 3-6b) for various contaminants subjected to biocell or biopile treatment.
The figures depict significant contaminant removals; not all the contaminants identified before treatment
were necessarily monitored after treatment, resulting in fewer contaminants identified in Figure 3-6b than
in Figure 3-6a. Table 3-6 summarizes the data used to create Figure 3-6.
Technology costs. The costs for biopile and biocell treatments are dependent on the biodegradability of
the contaminant, the cleanup goal, the cleanup procedure, other regulatory requirements such as off-gas
treatment, and the volume of soil to be treated. Costs have been estimated to range from $25 to $70 per
ton of contaminated soil treated. The Navy Facilities Engineering Service Center (NFESC) has estimated
the cost of biopile treatment of fuel contaminants to be on the order of $40/cu yd. The costs shown in
Table 3-5 are mostly within this cost range, although some costs range as high as $1500/cu yd. This high
cost is due to the small size of the site; in addition, this particular study was a demonstration site to test
the efficacy of biopile treatment only. Costs for full-scale deployment of this treatment tecnology would
36
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not be as high. The information in Table 3-5 is not as comprehensive as would be preferred due to an
inability to gather complete data on all sites, either from a reluctance of site owners to impart the
information or from missing site data. Additional site data on biocell/biopile treatment is available in
Appendix B.
3.1.3 Composting
Composting is an ex-situ technology designed to treat excavated soils contaminated with a range of
recalcitrant contaminants. The process involves mixing the contaminated soil with bulking agents such as
wood chips, straw, hay or alfalfa, and organic amendments such as cattle and/or chicken manure or
vegetative wastes. The selection of the specific compost ingredients depends on the contaminants to be
treated, the physical/chemical characteristics of the soil, and the availability of low-cost organic
amendments. The goal is to select the proper bulking agents to achieve the desired porosity, and organic
amendments that can provide the proper balance of carbon and nitrogen to promote biological activity in
the compost. For most composting applications, it is necessary to provide amendments that will support
thermophilic microbial activity.
Composting has been used to treat a wide range of contaminants in soils, and the use of the technology
continues to increase as the basic understanding of the fundamentals of the biodegradation pathways of
recalcitrant compounds and of the complex microbial interactions that occur during the composting
process continues to increase. Composting has traditionally been used to treat a variety of solid wastes
including wastewater sludge and the biodegradable fractions of municipal refuse with the primary goal
being volume reduction. Composting contaminated soils differs from these traditional applications in that
detoxification of contaminants sorbed to solid surfaces is desired, while volume reduction is not of
primary concern. Although the ultimate goals are different, the two uses of composting are similar in that
the intensive, usually thermophilic metabolic activity of a diverse and changing microbial population is
promoted to achieve the treatment objective.
3.1.3.1 Principles of Operation
Composting is a biological remediation technology that exploits the intensive activity of a diverse range
of microorganisms to degrade the target contaminants. The biological activity is enhanced through the
addition of readily degradable substrates and a sufficient supply of nutrients. Composting is considered
an aerobic technology because the compost is usually turned, mixed, or aerated to provide oxygen.
However, the high level of microbial activity can deplete the oxygen between turning or aeration
sequences creating anaerobic conditions in portions of the compost material, and microanaerobic zones
can exist in organically rich portions of the compost. These anaerobic processes can be beneficial for
promoting degradation of contaminants such as explosives or chlorinated organic compounds that are
recalcitrant or only partially degrade under strict aerobic or anaerobic conditions. For contaminants that
are readily mineralized under aerobic conditions, the development of anaerobic conditions is not desired,
and turning, mixing, and/or aeration frequencies are adjusted to minimize this potential.
37
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Table 3-5. Summary of Site Characteristics at Biopile/Biocell Installations*
Site Name
BPOil
Indiana Wood
Treating
Hydro-
Quebec/Verdun
Biosites
Hydro-
Quebec/Alma
Boucherville
Electrical Station
Pueblo Chemical
Depot
Dresherbrooke
Chevron Station
#9-
Naval Fuel Depot
Pt Molate
IEI Site
Uran Oil
Complex, Oil and
Natural Gas Corp.
Joliet Army
Ammo Plant
(various)
GRACE Biorem
Technologies
Joliet Army
Ammo Plant
(various)
Institute of Gas
Technology
Mare Island NS
Site
Status
Closed
Closed
Closed
Closed
Closed
Closed
Closed
Closed
Closed
NA
NA
NA
NA
NA
NA
Contaminants of
Concern
Diesel
PAHs
PAHs
PAHs, DRO,
VOCs
PCP
transformer oil
TNT, DNT, RDX
TPH
BTEX, TPH
Diesel, Fuel oil
Fuels
Oil
TNT, Tetryl
TNT, Tetryl
Diesel, Fuel oil
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
2100
4530 total PAHs; 666
carcinogenic PAHs
1000
DRO contamination
<1000;1000-2000,
2000-3000
100
14,000
TNT: 3,800
1,250 TPH
1,200 TPH; 21 B; 19
E; 99 T; 75 X(s)
TPH, 47,000
TPH, 58,000
TPH, 620,000
TNT, 3,000 avg;
Tetryl, 7,500 avg
TNT, 3,000 avg;
Tetryl, 7,500 avg
1,670
Unit Cost
(per CY)
$33
~$222
~$122
NA
~$122
~$150
$110
$30
$22
~$52
~$52
~$52
TNT: $476 Tetryl:
$211
TNT: $l,578Tetryl:
$1,240
NA
Numerical
Treatment Goals
(mg/kg dry
weight)
500 mg/kg
500mg/kg total
PAHs; lOOmg/kg
carcinogenic
PAHs
200mg/kg
DRO levels less
than 250mg/kg
5mg/kg
5,000mg/kg
10 mg/kg
500 mg/kg
lOOmg/kg TPH
TPH < 1,000 ppm
TPH < 1,000 ppm
TPH < 1,000 ppm
TNT, 50 ppm;
Tetryl, 250 ppm
TNT, 50 ppm;
Tetryl, 250 ppm
TPH, 100; BTEX,
.005 (each) mg/kg
Treatment
Goals
Achieved
Yes
Yes
Yes
Yes
Yes
Yes
yes
Yes
Unknown
Yes
Yes
Yes
Yes, for
Tetryl
No
Yes for
some cells
Concentration
Achieved
(mg/kg dry weight)
180
58 total PAHs; 10
carcinogenic PAHs
50
Unknown
1
3,800
nondetectable
NA
NA
230
600
TNT, 90; Tetryl, ND
TNT, 480; Tetryl, 1,875
TPH, from < 100 to <
300
Treatment
Effectiveness
NA
NA
NA
Removes an
average of 93% of
contamination
NA
NA
NA
NA
NA
Treatment has
high success rate
Reduced
contaminants to
well below
treatment goal
concentrations
well below
treatment goal
97% TNT
removal; 100%
Tetryl removal
84% TNT
removal; 75%
Tetryl removal
Varied
Treatment
Duration
(Months)
9
19
4
3
5
11
1.67
3
18
4
2.25
4
4
4
6
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Table 3-5. Summary of Site Characteristics at Biopile/Biocell Installations* (continued)
Site Name
Yorktown Naval
Weapons Station,
Site 6
Decommissioned
Gas Plant
Abandoned tanks
area
Marine Corps Air
Ground Combat
Center
(MCAGCC)
Site
Status
NA
NA
Active
Closed
Contaminants of
Concern
TNT, RDX,
HMX, TCE
Amines and salts
TPH, BTEX,
PAHs
JP-5 jet fuel,
Diesel
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
TNT, 1,329; RDX,
3 19; HMX, 98; TCE,
2,332
Amines, 15,000
TPH, 12,000; BTEX,
<50; PAHs, < 100
683, average
Unit Cost
(per CY)
NA
$34
NA
$40
Numerical
Treatment Goals
(mg/kg dry
weight)
TNT, 15; RDX, 5;
total VOCs 700
None reported
TPH, 500 mg/kg;
BTEX, 50 mg/kg;
PAHs, 60 mg/kg
None reported
Treatment
Goals
Achieved
Yes
Yes
Yes
Considered
successful
Concentration
Achieved
(mg/kg dry weight)
TNT, 2.9; RDX, 13.5;
HMX, ND; TCE, 17
Below detection limit
TPH, 235; PAHs, ND
56 average
Treatment
Effectiveness
After 41 days of
treatment.
continued to 90
days.
Large
concentration
reduction
Below treatment
goals at end of
treatment; 87%
reduction in TPH
On average,
decreased by 92%
Treatment
Duration
(Months)
3
5 bio + 6
leaching
6
<12
PAH = polycyclic aromatic hydrocarbons
DNT = dinitrotoluene
VOC = volatile organic compounds
DRO = diesel range organics
TPH = total petroleum hydrocarbons
TNT = trinitrotoluene
TCE = trichloroethylene
RDX = cyclotrimethylenetrinitramine
HMX = octahydro-l,3,5,7-tetranitro-l,3,5,7-
tetraaxocine
PCP = polychlorophenol
BTEX = benzene, toluene, ethylbenzene, xylenes
*Two sites included in Appendix B were not included in this table due to insufficient information. The excluded sites are Site ID Nos. 00-013 and 05-015.
Table 3-6. Concentrations of Contaminants of Concern Before and After Biocell/Biopile Treatment: Data Used to Generate Figure 3-6
Median:
25th Percentile
75th Percentile
High whisker(a)
Low whisker'3'
Concentrations of Contaminants of Concern
Chlorinated VOC/SVOC (2 sites)
Before
Treatment
1,216
658
1,774
2,332
100
After
Treatment
9
5
13
17
1
Fuel / Oil (10 sites)
Before
Treatment
12,000
1,670
47,000
58,000
683
After
Treatment
233
120
525
1,000
1
PAH, Creosote (3 sites)
Before
Treatment
1,000
550
2,765
4,530
100
After
Treatment
50
26
54
58
1
Explosives (12 sites)
Before
Treatment
3,000
2,582
7,500
NA
98
After
Treatment
285
8
2,081
3,525
1
(a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges (1.5 times the distance between the 25th and 75th
percentiles).
-------
4*
* Sj
'a~Sh
ll
^ ^^
งง
gซ
H U
C^ O
a w>
ฐ
o
U
1,000,000
100,000
10,000
1 ,000
100
10
1
0.1
&
1,000,000
100,000
10,000
1 ,000
100
10
1
0.1
c
V(
*
-,-
-
n i - T
U LJ
ilorinated Fuel / Oil PAH, Explosive
DC/SVOC (9) Creosote 111)
(2) (3)
Contaminant
I
X
T
-
-L
ilorinated Fuel / Oil PAH, Explosive
DC/SVOC (10) Creosote {12)
Contaminant
Figure 3-6. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before
and After Biocell/Biopile Treatment. (See Appendix B for data used to generate these plots.)
40
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The three configurations of composting include static pile, windrow, and in-vessel designs. Static piles
are simple in design and operation and are similar to biopiles. They are often designed with forced
aeration to control oxygen levels and maintain aerobic conditions. Air is pulled through the compost so
that the off-gas can be treated for volatile organic compounds and/or odors as necessary. Windrow
composting entails piling the soil on a containment pad with periodic turning to both keep the pile aerobic
and to provide a high level of mixing. This serves to break up the compost and distribute both the
contaminated soil and the nutrients. Because the piles are turned, windrows do not have forced aeration
systems or vapor control capabilities and may not be appropriate for soils contaminated with VOCs that
have the potential to volatilize to the atmosphere. In-vessel composting involves the use of enclosed
reactors. This allows the operator to control the atmosphere and the mixing with minimal temperature
disturbance. Vessels are typically designed with vapor control and more sophisticated monitoring
systems than the static pile or windrow designs. However, in-vessel composting has a much lower
throughput compared to static pile and windrow composting, which makes it the most expensive
composting method. All composting systems require moisture monitoring and control. Figure 3-7 shows
composting.
Figure 3-7. Composting
3.1.3.2 Target Contaminants
Composting is usually used to treat the more recalcitrant contaminants such as substituted- and
higher-molecular-weight aromatic compounds. Compounds effectively removed by composting include
contaminants associated with explosives such as TNT, RDX, HMX, and nitrocellulose; wood-preserving
chemicals including PAHs and PCP; and certain pesticides. Degradation of these compounds is likely to
benefit from the mixture of anaerobic and aerobic biological processes and the microbial complexity of
composts. Other classes of compounds including petroleum hydrocarbons such as gasoline, diesel fuel,
jet fuel, oil and grease, and chlorinated VOCs have been treated with composting. Often, these
compounds are present when the technology has been designed to treat one of the more recalcitrant
classes of contaminant, but removals nonetheless were substantial. Typically, these petroleum
hydrocarbons and VOCs are better treated with an alternative technology such as biopiles, while
composting is used to treat soils with more recalcitrant compounds that degrade too slowly to make the
alternative technologies effective.
41
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3.1.3.3 Advantages and Limitations
The primary advantages associated with composting include:
Removal of recalcitrant compounds is more rapid than that achieved with biopile treatment
End product can be humus-rich soil appropriate for commercial sale and/or reuse
A wide range of soil textures can be treated through addition and mixing of compost bulking
agents and amendments.
The primary limitations associated with composting include:
Space requirements can be substantial
Handling may require vapor control
Volume of contaminated material can increase substantially due to addition of bulking agents
Process can be susceptible to heavy metal or other toxin concentrations
The contaminated soil in the compost mix is limited to approximately 30% by weight to achieve
thermophilic conditions (EPA530-R-98-008, 1998c).
3.1.3.4 Techn ology Cost Drivers
Factors that drive the cost of composting include:
Composting is less land intensive than land treatment, but still requires relatively large treatment
areas
Cell size will impact the number of cells required; larger cells will require fewer cells, but larger
cells also are more difficult to construct and maintain
Volatile, dust, or odor emissions may require control measures
Soils with low porosity may require bulking agents to increase the airflow through the compost
pile; soil screening may be required to remove large rocks, debris, or other bulk objects
Composting may require nutrient amendments and water, in addition to bulking agents
O&M considerations include turnover frequency, maintaining moisture levels via water
amendments, maintaining nutrient levels and pH via nutrient and buffer amendments; thermophilic
aerobic conditions also must be maintained, which may require amendments with biodegradable
bulking agents.
3.1.3.5 Technology Performance and Cost
Composting of contaminated soil has been operating in the field for a number of years, and limited case
history data are available. Table 3-7 summarizes performance and cost data collected for 10 composting
applications. While these data are not complete in that they do not cover every composting application,
the data contained in the table provide useful information regarding the status of the technology.
Additional information on implementation of composting at full-scale sites can be found in Appendix B.
Range of sites identified. Composting is an ex-situ technology, and its application is not limited by the
same constraints as many in-situ technologies. The material to be treated is mixed with bulking agents
and other compost amendments that can be selected to adjust the porosity, moisture, pH, and other
properties to desired conditions to maximize composting performance. Because the compost is "tailor
made, " the technology can be applied to soils ranging from coarse sands and even gravels to the finest
silts and clays, sediments, and other contaminated solids. The more crucial limiting factors for
composting application are the depth of soil excavation, usually considered to be 20 to 30 ft bgs, and the
42
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presence of toxic levels of contaminants or other compounds that can inhibit biological activity, such as
heavy metals. The effects of such toxins must be determined on a case-by-case basis and is best done
through laboratory treatability studies.
Technology performance. Composting has proven successful for treating soils contaminated with the
explosives TNT, RDX, and HMX, achieving removals of up to 99.7%, 99.8%, and 96.6% in 40 days or
less. Composting also has been shown effective for treating PAHs, PCP, and VOCs. At the Dubose Oil
Products Co. Superfund Site in Florida, PAH, PCP, and VOC concentrations were reduced up to 91.3%,
77.3%, and >69.6%, respectively. Other examples of treatment performance are shown in Table 3-7. The
data illustrate that, for the most part, the technology performed as expected and cleanup goals were
achieved.
Figure 3-8 presents box and whisker plots showing the range of starting concentrations (Figure 3-8a) and
ending concentrations after treatment (Figure 3-8b) for various contaminants subjected to composting
treatment. The figure depicts significant contaminant removals; not all the contaminants identified before
treatment were necessarily monitored after treatment, resulting in fewer contaminants identified in Figure
3-8b than in Figure 3-8a. Table 3-8 summarizes the data used to create Figure 3-8.
Technology costs. The costs of composting are strongly dependent on the volume of soil being treated
and on the composition of the compost mix, primarily the percentage of contaminated soil that can be
included. Typical costs for treating large volumes (> 20,000 cu yd) are reported to be on the order of
$190/cu yd for windrow composting, and between $236 and $290 for aerated static pile and in-vessel
composting. The costs for the examples presented in Table 3-7 are generally within this range or slightly
higher. The information in Table 3-7 may be inadequate due to an inability to gather complete data on all
sites, either from a reluctance of site owners to impart the information or from missing site data.
Composting can be cost competitive with physical/chemical and thermal remedial technologies. For
instance, cost savings at Umatilla Army Depot were estimated at $2.6 million compared to the traditional
treatment method of incineration (EPA530-F-97-045). The end product of the Umatilla effort was a
humus-rich soil that was estimated to be worth $ 10/ton. This represented a total value of $ 150,000. The
Army Corps of Engineers has estimated that the total cost savings of substituting composting for
incineration at the remaining explosives-contaminated sites would be on the order of $200 million (EPA,
1997).
43
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Table 3-7. Summary of Site Characteristics at Composting Installations*
Site Name
Burlington
Northern
Superfund Site
Pueblo
Chemical Depot
Umatilla Army
Depot Activity,
Explosives
Washout
Lagoons,
CERCLA Soils
Operable Unit
Dubose Oil
Products Co.
Superfund Site
Joliet Army
Ammo Plant
MFGOU
Joliet Army
Ammo Plant
(LAP) OU
Stauffer Mgt Co
(SMC)
Tooele Army
Depot, TEAD-
81, SWMU 10
Former paper
mill, Southwest
Fill Area
Site
Status
Inactive
Closed
Closed
Closed
NA
NA
Active
Treatme
nt
NA
NA
Contaminants
of Concern
PAHs, MCE
hydrocarbons
TNT, DNT,
RDX
TNT, RDX,
HMX
VOCs (TCE,
benzene,
toluene, xylene),
PAHs, and PCP
TNT
TNT, DNT,
RDX, Tetryl,
TNB
Pesticides
TNT, DNT,
RDX, HMX
PCBs (mainly
Arochlor 1248)
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
17,871 total PAHs
(1987); 89,000 MCE
hydrocarbons (1993)
TNT, 3,800: TNB, 60;
TND, 38
TNT, RDX 2,000 ;
HMX, 100
38 VOC;2.1 TCE;
69.6 xylene; 367
PAHs; 51 PCP
NA
NA
chlordane, 48; ODD,
243; DDE, 11; DDT,
88; dieldrin, 3;
molinate, 10;
toxaphene, 779
TNT, 2,500; DNT,
5.9; RDX, 1,100;
HMX, 257
16
Unit Cost
(per CY)
NA
$312
$181
~$465
~$332
~$332
$192
$230
~$166
Numerical
Treatment Goals
(mg/kg dry weight)
8,632 total PAHs;
2 1,000 MCE
hydrocarbons
TNT, 3.8; TND, 3.2;
TND, 2. 1
30
<50 PAHs, <1.5
xylenes, <10
benzene, < 0.05 TCE
NA
NA
Chlordane, 2.3; ODD,
12.6; DDE, 8.9;
DDT, 8.9; Dieldrin,
0.19; Toxaphene,
2.75; Molinate, 0.74
TNT, 94; RDX, 34;
HMX, 18,000
None reported
Treatme
nt Goals
Achieved
Yes
(PAHs),
No (MCE
hydrocarb
ons)
yes
Yes
Yes
NA
NA
Yes for
DDE,
DDT,
Dieldrin,
Molinate
NA
NA
Concentration
Achieved
(mg/kg dry weight)
564 PAHs; 22,000
MCE hydrocarbons
<0.5
4 TNT; 2 RDX; 5
HMX
3.3 PAHs; 16.5 PCP;
0.03 xylene; 0.01
TCE
NA
NA
Chlordane, 5; ODD,
23; DDE, 7; DDT, 1;
Dieldrin, ND;
Toxaphene, 29;
Molinate, ND
TNT, 0.1; RDX,
2.96; HMX, 4.45
<3
Treatment
Effectiveness
The concentrations of
total PAHs in the soil
after treatment was less
than the cleanup goal of
8632 mg/kg for all 9
treatment sessions.
However, the cleanup
goal for MCE
hydrocarbons was not
met in any of the 9
treatment sessions.
To below action levels
Windrow composting
performance after 40-
day treatment generally
reduced the levels of
target explosives to
below the cleanup
goals.
Each batch of soil was
treated to less than the
cleanup goals within
14-30 days.
NA
NA
contaminants to below
treatment goals; all by
90%
Considered very
effective
Up to 40%
concentration decrease
overall, most loss in
less-chlorinated
cogeners
Treatment
Duration
(Months)
45
0.5-1
1
7
NA
NA
14
1.33
12
-------
Table 3-7. Summary of Site Characteristics at Composting Installations* (continued)
Site Name
Hawthorne
Army Depot,
Building 101-41
Site
Status
NA
Contaminants
of Concern
TNT, RDX,
HMX,
ammonium
picrate
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
explosives 60,000 -
120,000; PCP in wood
chip amendment, 100
Unit Cost
(perCY)
$250
Numerical
Treatment Goals
(mg/kg dry weight)
Varied: TNB, 4;
ammonium picrate, 7;
RDX, 64; TNT, 233;
HMX, 4,000
Treatme
nt Goals
Achieved
Yes
Concentration
Achieved
(mg/kg dry weight)
TNT, < 5; RDX,
< 25; ammonium
picrate, 0.4; PCP,
ND(<0.1)
Treatment
Effectiveness
Considered very
effective
Treatment
Duration
(Months)
28 days
DDD = dichlorodiphenyldichloroethane
DDE = dichlorodiphenyldichloroethylene
TNB = trinitrobenzene
MCE = methylene chloride extractable
PAH = polycyclic aromatic hydrocarbons
TNT = trinitrotoluene
DNT = dinitrotoluene
RDX = cyclotrimethylenetrinitramine
HMX = octahydro-l,3,5,7-tetranitro-l,3,5,7-
tetraaxocine
VOC = volatile organic compounds
TCE = trichloroethylene
PCP = polychlorophenol
PCB = polychlorinated biphenyl
NA = not available
*One site included in Appendix B was not included in this table due to insufficient information. The excluded site is Site ID No. 05-014.
Table 3-8. Concentrations of Contaminants of Concern Before and After Composting Treatment: Data Used to Generate Figure 3-8
Median:
25th Percentile
75th Percentile
High Whisker'3'
Low Whisker(a)
Concentrations of Contaminants of Concern
Chlorinated
voc/svoc
Before
Treatment
45
29
63
100
2
After
Treatment
0.1
0.1
9
17
NA
Pesticides Herbicides
Before
Treatment
48
11
166
243
3
After
Treatment
5
1
15
29
1
Fuel / Oil
Before
Treatment
89,000
89,000
89,000
NA
NA
After
Treatment
22,000
22,000
22,000
NA
NA
PAH, Creosote
Before
Treatment
9,119
4,743
13,495
NA
NA
After
Treatment
284
143
424
564
3
PCB
Before
Treatment
16
16
16
NA
NA
After
Treatment
3
3
3
NA
NA
Explosives
Before
Treatment
1,100
80
2,250
3,800
6
After
Treatment
3
1
5
NA
0.1
(a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges, i.e., 1.5 times the distance between the 25th and 75th
percentiles.
-------
4*
_ ,
100,000
10,000
, >* 1,000
งง
Is
CJ
C$ O
100
10
0.1
0.01
Chlorinated
VOC/SVOC
(4)
Pesticide,
Herbicide
(7)
Fuel / Oil
(1)
PAH,
Creosote
(2)
PCB
(1)
Explosive
in)
Contaminant
II
OS
"1
o> a
* S
o a
J o
U
100,000
10,000
1,000
100
10
0.1
0.01
I
Chlorinated
VOC/SVOC
(3)
Pesticide,
Herbicide
(7)
Fuel / Oil
(1)
PAH,
Creosote
(2)
PCB
(1)
Explosive
111)
Contaminant
Figure 3-8. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before
and After Composting Treatment. (See Appendix B for data used to generate these plots.)
46
-------
3.1.4 Ex-Situ Bioreactors
Ex-situ bioreactors have a long industrial history and broad applicability. Groundwater pumped from
contaminated aquifers is commonly treated using aboveground bioreactors. Off-gases from soil vapor
extraction systems are often treated using vapor-phase bioreactors prior to discharge to the atmosphere.
Municipal wastewater is most commonly treated using bioreactors, such as activated sludge or trickling
filter systems. Residuals from activated sludge plants are usually treated in anaerobic digesters, a type of
anaerobic bioreactor.
Bioreactors have more limited applicability to the treatment of soils and sediments. This is fundamentally
due to the energy required to mix the media and mass transfer restrictions. Vapor and water require much
less energy to mix than soils and sediments. Mixing typically is required to ensure that reactants or
substrates are readily available to microbes and wastes are diluted and appropriately carried through the
reactor system. The addition of water to soils to create a slurry enhances soil mixing and contaminant and
nutrient mass transfer. The water acts as a lubricant for mixing and as a solvent to dissolve contaminants
and nutrients and to suspend bacteria within the reactor. Bioslurry reactors are the most common ex-situ
reactor configuration for soil remediation. Thus, the literature search for cost and performance data was
restricted to bioslurry reactors.
Bioslurry treatment competes with land treatment and composting in the remediation technology selection
process, and typically is chosen when mass transfer requirements or space limitations drive technology
selection. Bioslurry reactors are used to degrade more recalcitrant compounds for which the reaction
kinetics of a completely mixed system are more beneficial.
3.1.4.1 Principles of Operation
Bioslurry reactors include lagoons or vessels that contain a mixture of contaminated soil and water at a
soil-to-water ratio ranging from 5% to 50% by weight. Slurries are used to accomplish the following
objectives:
Solubilize contaminants
Improve mixing effectiveness
Reduce mixing energy requirements
Homogenize media
Improve mass transfer.
Microbes that are indigenous to the soil or sediments, or exogenous cultures of microorganisms having
desired metabolic capabilities, are used to biodegrade the target contaminants. Reactor design and
operation may include manipulating the media through nutrient addition, aeration, mixing, pH control,
and possibly temperature control to enhance conditions favorable for bacterial growth and enzyme
production and activity.
Bioslurry reactors have been constructed using existing lagoons or ponds or aboveground mixing vessels.
Contaminated soil is excavated from the site and loaded into the lagoon or aboveground vessel that will
serve as the reactor. A bioslurry reactor is designed to apply sufficient mixing energy to suspend the bulk
of solids and prevent excessive sedimentation. When existing lagoons or ponds comprise the reaction
vessel, dredging equipment can be used to lift bottom sediments and achieve mixing.
Bioslurry reactors can be operated in continuous flow mode, like a continuous stirred tank reactor
(CSTR), or in batch mode, depending on the nature of the contaminant and degradation process.
47
-------
Regardless of the mode of reactor operation (continuous flow or batch), the mechanical action of slurry
mixers helps to break up soil and aggregates, enhancing the distribution of air and nutrients and mass
transfer throughout the reactor. Mass transfer refers to the transfer of contaminant mass from soils (e.g.,
sorbed, particulate, or pure phase contaminant in soils) to the aqueous phase, rendering the contaminants
more bioavailable; mass transfer also can pertain to the distribution and transfer of nutrients to the
aqueous phase for microbial bioavailability. The enhanced mass transfer of contaminants and nutrients in
bioslurry reactors are attractive features of this technology that sets it apart from the variety of other ex-
situ treatment alternatives such as land treatment, composting, and biopile/biocell treatment. Mass
transfer is enhanced through mechanical agitation, which results in the breakdown of soil clumps into
smaller particles and results in increased mixing between soil and aqueous phases. These actions increase
the exposure of surface particles to water, from which contaminants can desorb and/or dissolve.
Process options for bioslurry reactors, largely dependent on soil and contaminant type, include:
Batch or continuous flow modes
Inoculation with prepared cultures
Pretreatment (particle size reduction)
Mixing with or without aeration (aerobic or anaerobic treatment)
Nutrient addition
Surfactant addition
Residence time
Solids content
Cometabolite addition.
After treatment, the slurry typically is dewatered to separate solid (sludge) and liquid wastes. The
methods of disposal of the sludge and liquid waste streams depends on their posttreatment characteristics.
Secondary wastewater treatment can add significantly to soil treatment costs. Figure 3-9 shows an
operating bioslurry reactor, and Figure 3-10 is a schematic of a bioslurry reactor.
Figure 3-9. Bioslurry Reactor
48
-------
Exhaust
Air
Air Monitoring
Point
Carbon
(Air)
Nutrients (N, P)
pH Control
Antifoam
Knockout
Drum
Sample
Collection
Port
Pump
Air Diffuser
Air
Figure 3-10. Bioslurry Reactor Schematic
3.1.4.2 Target Contaminants
Bioslurry reactors have a relatively broad applicability. Their primary restriction is to sites impacted with
biodegradable contaminants. Contaminants that have been successfully remediated using bioslurry
reactors include the following:
Wood treating wastes
PAHs
Oil separator sludge
Petroleum hydrocarbons
Munitions and explosives
Pesticides (not including highly chlorinated pesticides)
PCBs
DCE or VC.
Bioslurry reactors tend to be associated with relatively recalcitrant compounds because they tend to be
more costly than other ex-situ treatment alternatives due to their slower throughput rates, more intensive
mechanical requirements, and increased operation and maintenance requirements. Bioslurry reactors tend
to be most commonly applied to sites with the following conditions:
Sites with a high degree of soil and hydrogeologic heterogeneities that confound in-situ system
design may be better suited for ex-situ treatment
Sites with severe treatment time restrictions may benefit from improved treatment rates; the more
rapid treatment kinetics of bioslurry reactors are mostly attributed to their enhanced mass transfer
rates
Sites with space restrictions may benefit from the smaller footprint provided by bioslurry reactors
Sites with relatively recalcitrant compounds may benefit from bioslurry reactors, due to the
potential for enhanced process control using bioslurry reactors.
49
-------
The environmental contaminants for which bioslurry reactors are most applicable are compounds that are
relatively difficult or slow to biodegrade using land treatment, composting, or biopile/biocell treatment,
and that can benefit from enhanced mixing and process control. For example, higher-molecular-weight
PAH compounds may be more degradable with this technology than with the more passive biotreatment
processes. Bioreactors tend to be used most often on residual explosives, which have been shown to
biodegrade only at a very slow rate. This is partly due to mass transfer limitations caused by the physical
form of residual explosives in soil, which tend to occur in clumps and particles. In order to be
biodegraded, individual molecules must interact with enzymes and other reactants. Highly viscous and
semisolid materials such as explosives (e.g., TNT, RDX, and HMX) degrade more slowly, partly because
the bioavailable molecules exist primarily on the outer surface of clumps of contaminants. One
developing strategy for addressing this problem is to first dissolve explosives in a solvent, such as
acetone, and then to treat the homogenized soil/contaminant mixture. This technique has been
demonstrated on a pilot scale, and 2,000 ppm of TNT was remediated adequately within 5 days.
Process control may include operating under anaerobic, sequential anaerobic-aerobic, or strictly aerobic
conditions. Explosives such as TNT, RDX, and HMX have been demonstrated to be susceptible to
anaerobic biodegradation. Unfortunately, many of the degradation products of these biotransformations
are no less toxic than the parent compound(s). Alternating anaerobic with aerobic processes may be
effective at mineralizing these compounds due to the various conditions under which parent and daughter
compounds have been found to degrade. This approach is easier to implement using a bioslurry reactor
than land treatment or biopile/biocell treatment. However, this approach has not yet been widely
implemented and requires further research and development, or at a minimum, pilot testing to
demonstrate its effectiveness.
3.1.4.3 Advantages and Limitations
Bioremediation of contaminated soils, sludges, or sediments using bioslurry reactors offers the following
advantages over many other remediation technologies:
Contaminant bioavailability is enhanced
Process control, including control of pH, temperature, and nutrients, is enhanced
The contaminated solid and liquid fractions are fully contained, greatly enhancing treatment
flexibility
Volatile emissions are controlled (for constructed vessels only)
Space requirements are reduced, particularly compared to land treatment, biopile/biocell
treatment, and composting
Bioslurry reactors may be mounted on trailers and transported for use at multiple sites,
maximizing the utility of fixed costs.
Principal limitations of bioslurry reactors include the following:
The physical nature of soil or sediment slurries make them extremely hard on machine parts
(mixers, pumps, aerators, and other process control equipment incur costly wear and tear,
resulting in increased repair and replacement costs)
Mixing soil slurries is energy intensive, and, similar to activated sludge operations, aeration
expenses can be a major cost constituent
Post-treatment dewatering, secondary wastewater treatment, and solids disposal may be required,
significantly increasing overall treatment costs
50
-------
Bioslurry reactors require more energy per unit soil treated than composting, biopiles, and land
treatment
Bioslurry reactors require more careful monitoring and more intensive O&M than the other land
treatment options.
3.1.4.4 Techn ology Cost Drivers
Major cost drivers of the slurry-phase biotreatment process include:
The physical nature of soil or sediment slurries makes them extremely hard on machine parts
(mixers, pumps, aerators, and other process control equipment incur costly wear and tear, resulting
in increased repair and replacement costs)
Excavation of contaminated media is required, except for lagoon implementation
Post-treatment dewatering, secondary wastewater treatment, and solids disposal may be required,
significantly increasing overall treatment costs
Sizing of materials prior to putting them into the reactor can be difficult and expensive
Heterogeneous soils and clayey soils can create serious materials handling problems. In the case of
free phase contaminants, preventative removal is mandatory
Bioslurry reactors may require careful monitoring, operation, and maintenance
Costs are proportional to throughput, which for bioslurry reactors is relatively slow.
3.1.4.5 Technology Performance and Cost
The use of bioslurry reactors is relatively limited because of the high capital and operating costs of this
technology, especially compared with alternative ex-situ biotreatment technologies. Cost and
performance summaries of these sites are provided in Table 3-9.
Figure 3-11 presents box and whisker plots showing the range of starting concentrations (Figure 3-1 la)
and ending concentrations after treatment (Figure 3-1 Ib) for various contaminants subjected to bioslurry
treatment. The figures depict significant contaminant removals; not all the contaminants identified before
treatment were necessarily monitored after treatment, resulting in fewer contaminants identified in Figure
3-1 Ib than in Figure 3-1 la. Table 3-10 summarizes the data used to create Figure 3-11.
51
-------
Table 3-9. Summary of Site Characteristics at Bioslurry Installations
Site Name
LaCie
Huntsman du
Canada
French Ltd.
Superfund Site
Southeastern
Wood Preserving
Superfund Site
Iowa Army
Ammunition
Plant
Eko Tec Site
Yorktown Naval
Weapons Station
Joliet Army
Ammo Plant
OU; Group 61
Navajo Indian
Reservation
Superfund Site
Bowers Field
Weldon Spring
Ordnance Works
Site
Status
Closed
Closed
Closed
Inactive
Closed
NA
NA
NA
Inactive
Inactive
Contaminants
of Concern
oil and grease
PAHs, BAP,
PCBs, vinyl
chloride, arsenic
PAHs
TNT, RDX
Creosote
TNT, RDX
TNT, DNT,
TNB, RDX,
HMX
Toxaphene
Dinoseb,
Nitroaniline,
other pesticides
and herbicides
TNT
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
13,500
616 PCBs; 750
pentachlorophenol;
400 BAP.
4,000
TNT: 1,500; RDX:
approx. 270.
PAHs, 220
>450
TNT, 6,226; DNT
and TNB, 360; RDX,
310, HMX, 215
4,000
Dinoseb, 34.2
maximum;
Nitroanaline, 13.3
average
TNT, 1,500
Unit Cost
(per CY)
$112
$200
$318
$335
NA
NA
$320
NA
$97
$112
Numerical Treatment
Goals (mg/kg dry weight)
1000
9 BAP; 23 PCBs; 43 vinyl
chloride, 7 arsenic, 14 B
<950 total PAHs; <180
BAP-equivalent carcinogenic
PAHs;
TNT: 196; RDX: 53
total PAHs = 50; BAP and
benzo(a)antracene = 10
TNT, 30; RDX, 100
TNT, 20
NA
Dinoseb contamination
reduced by at least 95%.
TNT contamination reduced
by at least 95%.
Treatment
Goals
Achieved
Yes
Yes
Yes
Yes
Yes
Yes
NA
NA
Yes
Yes
Concentration
Achieved
(mg/kg dry weight)
420
B: ND; BAP: ND
634 total PAHs; 152
BAP equivalent
NA
PAHs, 27; others
<10
TNT < 30; RDX,
<100
TNT, <20 to <50;
DNT<10to<100;
TNB, RDX <10
180
Dinoseb, < 0.03;
Notroaniline, < 0.75;
DDT, malathion,
parathion, < 0.75
TNT, 8.7
Treatment
Effectiveness
NA
NA
Total PAHs efficiency
93%; 67% for BAP-
equivalent
TNT below treatment
levels in 8 weeks.
RDX removal
occurred afterTNT.
All contaminants to
below treatment goal
Considered very
effective
> 99% removal in all
reactors
NA
Removed >99.8%
dinoseb; > 88.6%
parathion; other
herbicides unchanged
99. 4% removal of
TNT
Treatment
Duration
(Months)
3
22
36
2
1
1
2.5
NA
0.75
9
ND = non-detectable TNT = trinitrotoluene HMX = octahydro-l,3,5,7-tetranitro-l, 3,5,7-
NA = not applicable RDX = cyclotrimethylenetrinitramine tetraaxocine
BAP = benzo(a)pyrene TNB = tnmtrobenzene DNT = dmitrotoluene
PAH = polycyclic aromatic hydrocarbon DDT = dichlorodiphenyltrichloroethane
PCB = polychlorinated biphenyl
to
-------
Table 3-10. Concentrations of Contaminants of Concern Before and After Bioslurry Treatment: Data Used to Generate Figure 3-11
Median:
25th Percentile
75th Percentile
High whisker(a)
Low whisker(a)
Concentrations of Contaminants of Concern
Pesticide, Herbicide
Before
Treatment
(3 sites)
34
24
2,017
4,000
13
After
Treatment
(6 sites)
0.8
0.8
0.8
NA
NA
Fuel / Oil
Before
Treatment
(1 site)
13,500
13,500
13,500
NA
NA
After
Treatment
(1 site)
420
420
420
NA
NA
PAH, Creosote
Before
Treatment
(2 sites)
2,110
1,165
3,055
NA
NA
After
Treatment
(2 sites)
331
179
482
NA
NA
PCB
Before
Treatment
(1 site)
616
616
616
NA
NA
After
Treatment
(1 site)
23
23
23
NA
NA
Explosives
Before
Treatment
(9 sites)
360
310
1,500
NA
NA
After
Treatment
(11 sites)
30
10
77
100
9
a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges, i.e., 1.5 times the distance between the 25th and 75th percentiles.
-------
100,000
4*
10,000
I-SP
'S "Sfc
1,000 -
100
ง8
10
S"5
H U
.S w
c? o
1
0.1
0.01 -L
Pesticide,
Herbicide
(3)
Fuel/Oil PAH,
(1) Creosote
(2)
Contaminant
PCB
(1)
Explosive
(9)
V >
100,000
10,000
1,000
100
10
*- -
OB "
o
U
0.1
0.01
Pesticide,
Herbicide
(6)
I
Fuel / Oil
(1)
PAH,
Creosote
(2)
PCB
(1)
Explosive
111)
Contaminant
Figure 3-11. Box and Whisker Plots Showing Concentrations of Contaminants of Concern Before
and After Bioslurry Treatment. (See Appendix B for data used to generate these plots.)
54
-------
The following case studies are reported as examples of bioslurry reactor treatment of soils.
French Limited Superfund Site, Crosby, TX. The French Limited Superfund Site in Crosby, TX (French
site) was an industrial waste disposal facility at which more than 70 million gallons of petrochemical
industry wastes were disposed in an unlined disposal lagoon from 1966 to 1971. This project is the first
reported application of a bioslurry reactor at a Superfund site. The bioslurry reactor was constructed
about and within the existing disposal lagoon. A commercial system (MixFlo) was used to aerate the
system with pure oxygen and control volatile emissions from the lagoon surface. The system was
designed to treat approximately 300,000 tons (2.7 x 108 kg) of subsoil underlying a layer of tarry sludge.
The tarry sludge was removed prior to treatment of the subsoil in the bioslurry reactor.
Performance and factors. Table 3-11 lists the primary contaminants and their respective cleanup goals.
For treatment, the lagoon was subdivided into two smaller lagoons, which were treated separately. The
residence time of liquid and solid media in the lagoons was about the same as the total treatment time (10
and 11 months). The aerator consisted of a 3,400-hp (2.5-megawatt) motor supplying 2,500 Ib (1,100 kg)
of oxygen per hour over the treatment time. The system achieved a mass throughput of approximately 500
tons/day.
Table 3-11. Cleanup Goals for Primary Contaminants at the French Site
Contaminant
Benzo (a) pyrene
Total PCBs
Vinyl chloride
Arsenic
Benzene
Cleanup Goal (mg/kg)
9
23
43
7
14
Superfund Preliminary Site Closeout Report French Limited Site Crosby, Texas,
September 1994. CERCLIS TXD-980514814, p. 6.
The ROD specified bioremediation of the lagoon subsoil, but also listed arsenic (an elemental heavy
metal) as a primary contaminant. The reported concentration results showed that arsenic concentrations
in soil decreased over treatment time, but it is unclear what effect biodegradation processes had on this
contaminant. Nevertheless, the sediments in both lagoons were cleaned to below target levels within 11
months of operation.
Cost range. The costs directly attributable to treatment activities were $26,000,000, resulting in a mean
cost of about $90/ton ($99/1,000 kg) of soil treated. Total costs, which included project management,
pilot studies, and post-treatment activities, amounted to $49,000,000, yielding a total cost of $163/ton
($180/1,000 kg).
Southeastern Wood Preserving Superfund Site, Canton, MS. The Southeastern Wood Preserving
Superfund Site in Canton, MS, evolved from wood-preserving operations from 1928 through 1979. Three
unlined surface impoundments were used for wastewater treatment. Approximately 4,000 mg/kg of
PAHs were found to exist in bottom sediment sludges from the impoundments. The material to be treated
was classified as an RCRA KOOOl-listed hazardous waste. Unlike the French site, engineered slurry
reactors were constructed on site, and media was moved to the reactors for treatment.
Performance and factors. Four bioslurry reactors were operated in batch mode. Each reactor was
circular, with a 38-ft (11.6 m) diameter and a height of 24 ft (7.3 m), resulting in an operating volume of
180,000 gallons (680 m3). Each batch consisted of about 170 cu yd (130 m3) of material. In total, 14,140
55
-------
tons (1.3 x 107 kg) or 10,500 cu yd (8,000 m3) of material were treated. The bioslurry reactors were
aerated at a rate of approximately 350 standard cubic feet per minute (scfm) (165 L/s) with a batch
residence time ranging from 8 to 29 days. Air was used for aeration at the Southeastern Wood Preserving
site, while the French site system used pure oxygen. Table 3-12 lists the average treatment efficiency of
the bioslurry reactors for removal of PAHs. When initial operations revealed that cleanup goals for
specific PAHs (pyrene and phenanthrene) were not being achieved in the design reactor residence time of
30 to 35 days, the cleanup goals were modified to specify total and carcinogenic PAHs, instead of specific
PAHs. This was done by obtaining a variance under 40 CFR 268.44. Hence, cleanup goals were
eventually obtained when the goals were adjusted to meet the performance efficiency of the bioreactors.
Nonetheless, this approach was accepted. The system achieved a mass throughput of approximately 50
tons/day.
Table 3-12. Efficiency of Bioslurry Reactors for Removing PAHs
at the Southeastern Wood Preserving Site
Constituent
Naphthalene
Benzo(a)pyrene
Carcinogenic PAHs
Benzo(a)pyrene
equivalent
Total PAHs
Cleanup Goal
(mg/kg)
NA
NA
NA
180
950
Average Initial
Concentration
(mg/kg)
48
98
1,095
433
8,621
Average Final
Concentration
(mg/kg)
6
79
376
150
655
Average Removal
Efficiency
(%)
88
19
66
65
92
Cost range. The costs directly attributable to treatment activities were $2,400,000, resulting in an
average cost of about $170/ton ($190/1,000 kg) of soil treated. Total costs, which included project
management, pilot studies, and post-treatment activities, amounted to $2,900,000, resulting in a total cost
of $205/ton ($230/1,000 kg).
SABRE Process. Another application of bioslurry reactors is exemplified by the J.R. Simplot
Company's Simplot Anaerobic Bioremediation Ex-Situ (SABRE) Process. This approach includes the
addition of proprietary amendments to the slurry and has been used to remediate explosives in soil. Two
applications are summarized below.
Site descriptions. Soils from the Iowa Army Ammunition Plant (IAAP) were treated in a concrete trench
lined with high-density polyethylene (HDPE). The trench was approximately 50 ft (15 m) long and 8.3 ft
(2.5m) wide and was filled with a 40% soil slurry comprising 40 cu yd of contaminated soil. The
primary contaminants were the explosives TNT and RDX.
Soils impacted with explosives TNT and RDX from Yorktown Naval Weapons Station, VA, also were
treated using the SABRE process. At the Yorktown site, a double-lined bioremediation cell was used
to treat 1,900 cu yd (1,500 m3) of contaminated soil, including preliminary pilot tests followed by
full-scale treatment.
Performance and factors. At the IAAP site, TNT degraded first, followed by RDX. TNT was reduced
from an initial concentration of about 800 mg/kg to below the cleanup standard of 196 mg/kg within
approximately 8 weeks. RDX was reduced from its initial concentration of approximately 260 mg/kg to
below its cleanup standard of 53 mg/kg within approximately 10 weeks.
56
-------
At the Yorktown site, treatment times for both pilot- and full-scale treatment were 30 days. Table 3-13
lists the maximum concentrations of various explosives in soil and the respective cleanup goals achieved
at the Yorktown site.
Cost range. At the IAAP site, treatment costs were projected to be in the range of $300 to $350/cu yd
($390 to $460/m3). The Yorktown soils were treated for a cost of $398/cu yd ($520/m3).
Table 3-13. SABRE Process Effectiveness at Yorktown Site
Explosive
TNT
RDX
HMX
Maximum Concentration
(mg/kg)
35,000
5,400
11,000
Cleanup Goal Achieved
(mg/kg)
30
50
3,900
3.1.5 Conventional Bioventing
Conventional bioventing is the process of aerating soils to stimulate in-situ biological activity and
promote bioremediation. Conventional bioventing typically is applied in situ to the vadose zone and is
applicable to any chemical that can be aerobically biodegraded. To date, it has been implemented
primarily at petroleum-contaminated sites. A typical conventional bioventing system is illustrated in
Figure 3-12. Although bioventing is related to the process of SVE, the primary objectives of these two
bioremediation technologies are different. SVE is designed and operated to maximize the volatilization
of
low-molecular-weight compounds; biodegradation is not typically a design objective. In contrast,
bioventing is designed to maximize biodegradation of aerobically biodegradable compounds, regardless
of their molecular weight, while minimizing volatilization. The major distinction between these
technologies is that the objective of SVE is to optimize removal by volatilization, while the objective of
bioventing is to optimize biodegradation while minimizing volatilization and reducing the capital and
utility costs required for vapor treatment. Although both technologies involve venting of air though the
subsurface, the differences in objectives result in different design and operating conditions for the two
remedial systems.
Low Rate
Air Injection
1
I
I
V {
I
Biodegradation
of Vapors
] 4 1
P finil fiqq I
^ Monitoring t
\
1 E
1 r
*
l (
1
r
Contaminated
Soil
fc
i
.X
PPT/Leeson/102-
Figure 3-12. Schematic of Typical Conventional Bioventing Process
57
-------
3.1.5.1 Principles of Operation
Conventional bioventing is a relatively simple technology involving minimal equipment including a
blower for air injection, vent wells screened throughout the contaminated zone, soil-gas monitoring
points, and associated monitoring equipment. Figure 3-13 shows atypical conventional bioventing
system. Four primary characteristics impact the applicability and/or effectiveness of bioventing. These
include soil-gas permeability, contaminant distribution, zone of oxygen influence, and microbial activity.
Figure 3-13. Conventional Bioventing System
Assuming that contaminants amenable to bioventing are present, geology probably is the most important
site characteristic for a successful conventional bioventing application. Soils must be permeable enough
to allow sufficient soil-gas flow to provide adequate oxygen for biodegradation, on the order of 0.25 to
0.5 vapor pore volumes per day. Soil-gas permeability is a function of soil structure, particle size, and
soil moisture content. Typically, permeability in excess of 0.1 darcy is adequate for sufficient air
exchange. Below this level, bioventing certainly is possible, but field-testing may be required to establish
feasibility. When the soil-gas permeability falls below approximately 0.01 darcy, soil-gas flow is
primarily through either secondary porosity (such as fractures) or through any more permeable strata that
may be present (such as thin sand lenses). Therefore, the feasibility of conventional bioventing in
low-permeability soils is a function of the distribution of flowpaths and diffusion of air to and from the
flowpaths within the contaminated area.
Another important factor affecting the feasibility of conventional bioventing is contaminant distribution
throughout the site. Difficulties in applying bioventing arise when significant quantities of the
58
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contaminant are in the capillary fringe or below the water table due to groundwater fluctuations.
Treatment of the capillary fringe by screening air injection wells below the water table is possible;
however, the ability of bioventing to aerate the capillary fringe and underlying water table has not been
evaluated. Limited oxygenation is expected to occur in saturated soils. If significant contamination exists
below the water table, dewatering should be considered as a means of exposing any contaminated soil to
injected air. Alternatively, a combination of air sparging (air injection beneath the water table) and
bioventing may provide more efficient air delivery to the capillary fringe.
An estimate of the oxygen radius of influence (R:) of air injection wells is an important element of
conventional bioventing design. This parameter is used to design full-scale systems, specifically to space
air injection wells, size blower equipment, and ensure that the entire site receives a supply of oxygen-rich
air to sustain in-situ biodegradation. The radius of oxygen influence is defined as the radius to which
oxygen has to be supplied to sustain maximal biodegradation. This definition of radius of influence is
different than is typically used for SVE, where radius of influence is defined as the maximum distance
from the air extraction or injection well where vacuum or pressure (soil-gas movement) occurs. The
oxygen radius of influence is a function of both air flowrates and oxygen utilization rates, and therefore
depends on site geology, well design, contaminant concentration and microbial activity. As microbial
activity increases, the effective treated area will decrease. Therefore, it is desirable to estimate the
oxygen radius of influence at times of peak microbial activity and to design the bioventing system based
on these measurements.
Finally, conventional bioventing is dependent on providing microorganisms optimal conditions for active
growth. Several factors may affect a microorganism's ability to degrade contaminants; however, those
that impact the bioventing process significantly include availability and type of electron acceptors and
moisture content.
One of the most important factors that influences the biodegradability of a compound is the type and
availability of electron acceptors. For example, following a hydrocarbon spill, anaerobic conditions
typically predominate in the subsurface because of oxygen depletion from microbial activity. While
hydrocarbons may undergo limited biodegradation under anaerobic conditions (Bilbo et al., 1992;
Mormile et al., 1994), in general, aerobic conditions are more suitable for relatively rapid remediation of
petroleum hydrocarbons. Therefore, oxygen supply is critical to the success of a conventional bioventing
system. In field studies, oxygen has been found to be the most important factor in determining the
success of a bioventing system (Leeson and Hinchee, 1996; Miller et al., 1991).
Soil moisture content may impact conventional bioventing by its effect on microorganisms or soil-gas
permeability. Microorganisms require moisture for metabolic processes and for solubilization of energy
and nutrient supplies. Conversely, soil moisture content directly affects soil permeability, with high
moisture contents resulting in poor distribution of oxygen. In practice, soil moisture has been found to
directly limit biodegradation rates only where bioventing has been implemented in very dry desert
environments. A more common influence of moisture is that excess moisture has led to significant
reductions in soil-gas permeability (Leeson and Hinchee, 1996).
A fairly recent improvement to conventional bioventing technology is enhanced, or cometabolic,
bioventing. Cometabolic bioventing, which utilizes the addition of gas-phase additives to the injected
oxygenated air in order to enhance biodegredation, is discussed in detail in Section 3.1.6.
3.1.5.2 Target Contaminants
Any aerobically biodegradable compound can potentially be degraded through bioventing. To date,
conventional bioventing has been applied primarily to petroleum hydrocarbons; however, bioventing of
59
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PAHs (Lund et al, 1991; Hinchee and Ong, 1992; Alleman et al., 1995) and bioventing applied to an
acetone, toluene, and naphthalene mixture (Leeson et al., 1994) have been implemented successfully.
The key to conventional bioventing feasibility in most applications is biodegradability versus volatility of
the compound. If the rate of volatilization greatly exceeds the rate of biodegradation, bioventing is
unlikely to be successful, as removal occurs primarily through volatilization. This will occur most often
in those cases where the contaminant is a fresh, highly volatile fuel. An unsuccessful conventional
bioventing application is unlikely to occur due to a lack of microbial activity. If conventional bioventing
is operated in the injection mode, volatilized contaminants may be biodegraded before reaching the
surface, unlike an extraction mode. Figure 3-14 illustrates the relationship between a compound's
physicochemical properties and its potential for bioventing.
In general, compounds with a low vapor pressure1 cannot be successfully removed by volatilization, but
can be metabolized by microbes if they are aerobically biodegradable. High vapor pressure compounds
are gases at ambient temperatures. These compounds volatilize too rapidly to be easily biodegraded in a
bioventing system, but typically are a small component of fuels and, due to their high volatility, will
attenuate rapidly. Compounds with vapor pressures between 1 and 760 mm Hg may be amenable to
either volatilization or biodegradation. Within this intermediate range lie many of the petroleum
hydrocarbon compounds of greatest regulatory interest, such as benzene, toluene, ethylbenzene, and the
xylenes. As can be seen in Figure 3-14, various petroleum fuels are more or less amenable to
conventional bioventing. Some components of gasoline are too volatile to easily biodegrade, but, as
stated previously, typically are present in low overall concentrations and are attenuated rapidly. Most of
the diesel constituents are sufficiently nonvolatile to preclude volatilization, whereas the constituents of
JP-4 jet fuel are intermediate in volatility.
To be amenable to conventional bioventing, a compound must 1) biodegrade aerobically at a rate
resulting in an oxygen demand greater than the rate of oxygen diffusion from the atmosphere, and
2) biodegrade at a sufficiently high rate to allow in-situ biodegradation before volatilization. Practically,
this means that low vapor pressure compounds need not biodegrade as rapidly as high vapor pressure
compounds for bioventing to be successful. The actual feasibility of bioventing is very site-specific;
therefore, Figure 3-14 should not be used as absolute, but rather as a general guideline.
Of the petroleum hydrocarbons, BTEX generally are the compounds that are regulated most stringently.
Typically, these compounds degrade very rapidly during bioventing, and, at most sites, are degraded to
below detection limits within 1 year of operation of a bioventing system. This trend was illustrated in a
study at Tyndall Air Force Base (AFB) and has been confirmed at numerous bioventing sites (Leeson and
Hinchee, 1996). At Tyndall AFB, two test plots were conducted with initial hydrocarbon concentrations
of 5,100 and 7,700 mg/kg. After 9 months of bioventing, TPH was reduced by 40% from the initial
concentration. However, the low-molecular-weight compounds such as BTEX were reduced by more
than 90%. The low-molecular-weight compounds were preferentially degraded over the heavier fuel
components, which is consistent with previous research (Atlas, 1986).
Bioventing generally is not considered appropriate for treating compounds such as PCBs and chlorinated
hydrocarbons. However, through a cometabolic process, it may be possible to enhance the degradation of
compounds such as TCE through bioventing.
For the purposes of this discussion, compounds with vapor pressures below approximately
1 mm Hg are considered low, and compounds with vapor pressures above approximately
760 mm Hg are considered high.
60
-------
100
;-buty!me hy3ether
Gasoline
Vapor Pressure
Too High to Easily Biovetrt
10'8 10-3 10'1 10
Aqueous Solubility (mmoles/litre)
1000
Figure 3-14. Relationship Between Organic Physicochemical Properties of Organic Compounds
and their Potential for Conventional Bioventing (Leeson and Hinchee, 1996)
3.1.5.3 Advantages and Limitations
The primary advantage of implementing conventional bioventing over other comparable technologies is
the cost savings. Conventional bioventing is a relatively simple technology, and this translates into
significantly reduced costs. In addition, conventional bioventing is an in-situ technology that results in
minimal disturbances to sites. Given that many sites that are amenable to bioventing are in high-traffic
areas, such as gasoline stations, this is a significant benefit, as business can continue uninterrupted once
installation is complete. Finally, the microbial aspect of conventional bioventing results in two benefits:
nonvolatile contaminants can be removed (unlike during SVE), and contaminants are biotransformed to
innocuous byproducts instead of transferring the contaminants to another matrix.
The primary limitation of conventional bioventing is the time involved to complete remediation. While
BTEX components may be removed rapidly, the heavier-molecular-weight compounds may take several
years to be removed. This long time frame may not be acceptable for all sites. Also, bioventing is not
applicable to all sites, depending on contaminant volatility, biodegradability, soil permeability, and site
use restrictions.
3.1.5.4 Technology Cost Drivers
Major conventional bioventing cost drivers include:
Capital equipment costs will increase with increasing site size, including blower size and
capacity, well installation frequency, and well depth
61
-------
Depth of contamination will impact the drilling requirements and the material costs for well
installation; depth also impacts energy costs when air needs to be blown across larger depth
intervals
More permeable soils require less energy to force air through the soil matrix; also, more
permeable soils lead to a more uniform air distribution and consequently more uniform
biodegradation
Off-gas capture generally is not recommended because air flow should be commensurate with
biodegradation rates; however, at high risk sites where off gas could impact human activity,
off-gas capture and treatment may be required, significantly increasing treatment costs.
3.1.5.5 Technology Performance and Cost
Table 3-14 provides a summary of various full-scale conventional bioventing installations, including
information on contaminant type and concentrations, target cleanup levels, and status of site cleanup (See
Appendix B for the data used to generate Table 3-14). Figure 3-15 shows box and whisker plots for
contaminants treated using conventional bioventing, as reported in Appendix B. Figure 3-15a presents
the range of contaminant concentrations before treatment, and Figure 3-15b illustrates their range after
treatment. Table 3-15 illustrates the data used to generate Figure 3-15. Information provided in the tables
and figures in this section is not as comprehensive as desired due to an inability to gather complete data
on all sites, either from a reluctance of site owners to impart the information or from missing site data. Of
the 22 conventional bioventing sites shown, five sites achieved their treatment goals, three sites achieved
some of their treatment goals, two sites did not meet any of their treatment goals, and sufficient
information was not available for 12 of the sites to determine whether or not treatment goals were
achieved.
At the USCG Support Center, BTEX was removed to below detection limits, but TPH remained well
above the treatment goal of 100 mg/kg. Initial TPH concentrations were approximately 2,900 mg/kg, and
the bioventing system was operated for 18 months. In order to meet treatment goals for this site, the
biodegradation rate required may be calculated as follows:
(2,900 mg/kg-100 mg/kg)
-^ A = 5.2 mg/kg/day (1)
30 days
18 months x
month
The actual concentration achieved at the end of 18 months was approximately 1,450 mg/kg, indicating a
much slower average biodegradation rate of approximately 2.7 mg/kg/day using the same equation as
shown above. Both of these biodegradation rates fall within a normal range of 1 to 20 mg/kg/day (Leeson
and Hinchee, 1996), and indicates that the bioventing system was not operated for a long enough time
period to achieve the TPH treatment goals.
A similar situation occurred at Vanier Gas Station, Canada, where treatment goals for benzene, toluene,
and xylene were not achieved although concentrations were reduced. At this site, though, the system was
operated only for 111 days. Most evidence indicates that this would not be sufficient time to achieve
complete removal, even of the more biodegradable BTEX compounds.
At Sorel Gas Station, the conventional bioventing system was operated for 475 days, but only small
amounts of removal were achieved. No information was available on soil type at this site; however, air
was injected below the water table, possibly indicating a significant amount of contamination present in
the capillary fringe. Water-saturated soil would not be amenable to bioventing and this could explain the
lack of degradation at this site. However, without additional site information, it is difficult to determine
the causes for not achieving the treatment goals.
62
-------
Of the remaining 10 sites, information on attaining treatment goals was not available. At four of these
sites, this is due to the fact that the systems are still active at the time when data for these sites were
collected. Four of the sites have been closed indicating that treatment goals were met, but not reported in
an attainable format. The remaining two sites are now inactive (Greenwood Chemical Superfund Site,
and Site ST-20, Eielson AFB). At the Greenwood Chemical Superfund Site, the system was installed as
part of a research study and was not intended to be operated until contaminants were completely removed.
At Site ST-20, the initial system was installed as a research project and was to be expanded to the entire
site once funding was received. Given that this expansion has not occurred, it is probably because the
Base is still awaiting funds to expand operation at this site.
63
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Table 3-14. Summary of Site Characteristics at Conventional Bioventing Installations*
Site Name
George AFB,
OU 3 FT19a
Greenwood
Chemical
Sorel Gas
Station
Vanier Gas
Station
Elmendorf
Air Force
Base
Hill AFB (8
areas at site)
Robins Air
Force Base
BLDG 30,
406, 528, and
a POL Area
ofOffutt AFB
Sites, D-10;
FC-2, S-4,
Kelly AFB
East 15th
Street Service
Station
Site 5,
Savannah
River Site
NASA/Wallo
ps Flight
Facility
Lowry Air
Force Base
Site 280, Hill
AFB
Site
Status
Active
treatment
Inactive
Closed
Inactive
Active
Active
Closed
Active
Active
Active
Active
Closed
Closed
Active
Contaminant
s of Concern
BTEX, TPH
A, T, N, CB,
B, 1,2-DCA
BTEX
BTEX
BTEX, TRPH
BTEX, TRPH
BTEX: TRPH
BTEX; TRPH
BTEX; TRPH
diesel
diesel
Diesel oil,
furnace oil
Heating oil,
BTEX
JP-4 jet fuel,
TPH, BTEX
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
TPH, 65,000; B 22;
T210;E59;X610
A- 47- T- 630- N-
120- CB- 46- B- 94-
1,2-DCA: 2.4
B;41;E: 31; T: 280;
X: 650.
B- 36- E- 76- T- 250'
X: 520.
E: 94. T: 87,
X 430,. TRPH: 5340
B1.2; T: 1150;
E: 5840. X: 17,300;
TRPH: 32,200.
B: 1.3; E: 220;
T: 59; TRPH: 9000;
X:39.
B: 2000; X: 38,000;
T: 7,100; E: 4100
30 B; 884 total
BTEX in all four
sites; 5430 TRPH
5500
100
6970 TPH
14,000 TRPH
5040 TPH; 11,200
Soil-gas TPH
Unit Cost
(per CY)
$14
$27
NA
$82
NA
NA
NA
NA
NA
NA
$4
$146
NA
NA
Numerical
Treatment Goals
(mg/kg dry weight)
< 25 ft: TPH, 10; B,
T, E, 0.005; X,
0.015. Soil 25-100 ft,
TPH 10, BTE 0.3 X
1.0.
NA
B: 5;E: 50; T: 230;
X: 50
B- 5- E- 50- T- 30-
X: 50.
NA
B: 0.2; T: 100; E: 70;
X: 0.36; TRPH: 30.
Site Specific
Site Specific
NA
100
100
50
500 TPH; 500
TRPH; <100 BTEX
NA
Treatment
Goals
Achieved
Expected to
reach by
2003
NA
No
E' yes
others: no.
NA
Yes, except
TRPH
NA
Yes
NA
NA
NA
Yes
NA
NA
Concentration
Achieved
(mg/kg dry weight)
TPH, 14,000; B, 2.9;
T, 68; E, 34; X, 266
A- 0012- T- 0002- N-
0- CB- 0- B- 0- 1 2-
DCA: 0.007
B:14;E:66;T:230;
X:290
B- 6 1- E- 11- T- 68-
X: 80.
E: 52. T: 41; X: 240;
TRPH: 3900
B: 0.64. T: 0.38.
E: 2. X: 5.8.
TRPH: 12,000
B:ND;E: 0.11; TRPH:
1,900; X: 1
B: 0.0042; X: 0.0029;
T: ND; E: 0.019
TRPH: 4.3
T: 8; TRPH: 920;
X: 12
NA
NA
NA
NA
2,600 ppm (Soil-gas
TPH)
Treatment
Effectiveness
Considered effective
A- >99- T- >99- N- >98'
CB- >98- B- >99- 1 2-
DCA: >90
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
Evidence of
hydrocarbon degradation
Treatment
Duration
(Months)
36
15
12.5
4
NA
NA
NA
NA
NA
NA
NA
NA
24+
48+
-------
Table 3-14. Summary of Site Characteristics at Conventional Bioventing Installations* (continued)
Site Name
IRP Site ST-49,
Eglin AFB
USCG Support
Center, Elizabeth
City, NC
Los Angeles Air
Force Base
Site 914, Hill Air
Force Base
Eielson Air Force
Base Source Area
ST20
Fort Bliss, Bldg.
675
Tyndall Air Force
Base
Fort Carson Bldg.
8200
Fort Rucker
SWMU 14
Site Status
Active
Inactive
1&2 Closed,
3 Inactive
Closed
Inactive
Active
treatment
Unknown
Active
treatment
Closed
Contaminants
of Concern
Petroleum;
VOC
TPH (JP-4),
BTEX
TRPH
JP-4 jet fuel,
TPH, BTEX
JP-4 jet fuel,
TPH, BTEX
BTEX
TPH
TPH, BTEX
TPH, BTEX
Maximum
Contaminant
Concentrations
(mg/kg dry weight)
2430
346 BTEX; 2,954
TPH
(1) 1580; (2) 11,800;
(3) 14000
10,000 TPH
1,500 TPH
1,350 avg
> 1,000
TPH, 1,350 avg;
BTEX, 17 avg
TPH, 25,000 avg;
BTEX, 10 avg
Unit Cost
(per CY)
$96
NA
NA
$152
$13
$6
$30
$18
$3
Numerical
Treatment Goals
(mg/kg dry weight)
50 VOC
<100 TPH
Site Specific
38.1 TPH
200 TPH; 21bs/day in
extracted soil gas
NA
NA
NA
NA
Treatment
Goals
Achieved
Yes
No
Yes
Yes
NA
NA
NA
NA
NA
Concentration
Achieved
(mg/kg dry weight)
0
0, BTEX; 1,457, TPH
NA
<6
NA
690 in first year
TPH, < 30
TPH, reduction rate
170 mg/kg/year
TPH, 16.9; BTEX, ND
Treatment
Effectiveness
NA
98% benzene, 60%
TPH.
NA
NA
NA
NA
NA
NA
Effective
Treatment
Duration
(Months)
60
18
NA
15
36+
12
12
12
B = benzene
T = toluene
E = ethylbenzene
X = xylenes
*One site included in Appendix B was not
A = acetone DCA = dichloroethane
CB = chlorobenzene TRPH = total recoverable petroleum hydrocarbons
N = naphthalene VOC = volatile organic compound
TPH = total petroleum hydrocarbons NA = not available
included in this table due to insufficient information. The excluded site is Site ID No. 08-006.
Table 3-15. Concentrations of Contaminants of Concern Before and After Conventional Bioventing: Data Used to Generate Figure 3-15
Median:
25th Percentile
75th Percentile
High whisker(a)
Low whisker3'
Concentrations of Contaminants of Concern
Chlorinated VOC/SVOC
Before
Treatment (2
sites)
4
3
4
5
NA
After Treatment
(2 sites)
0
0
0
NA
NA
BTEX
Before
Treatment (48
sites)
134
41
635
1,350
1
After Treatment
(35 sites)
6
0
59
80
NA
Fuel / Oil
Before
Treatment (31
sites)
6,190
2,055
13,600
25,000
52
After Treatment
(17 sites)
960
30
2,600
3,900
4
PAH, Creosote
Before
Treatment (1
site)
120
120
120
NA
NA
After Treatment
(1 site)
0
0
0
NA
NA
(a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges, i.e., 1.5 times the distance between the 25th and 75th percentiles.
-------
s-^.
.0 +J
CQ
งง
S :-
= a
a 5
S w
^ o
100,000
10,000
-S ฃ 1,000
5-t
= W) 100
.
s s
10
o
U
0.1
0.01
0.001 -i
Chlorinated
VOC/SVOC
(2)
Contaminant
Fuel / Oil
(17)
100,000
10,000
1 ,000
100
10-
1
0.1
0.01
0.001
c
V(
A
I
B
ilorinated Fuel / Oil
DC/SVOC (31)
(2)
*
*
&
*
I
--
BTEX
(48)
BTEX
(35)
Contaminant
Figure 3-15. Box and Whisker Plots Showing the Range of Contaminant Concentrations
Before and After Conventional Bioventing
(See Appendix B for data used to generate these plots)
66
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Overall, conventional bioventing performance has been shown to be highly effective and reliable if
sufficient time is provided to complete removal. Problems with system performance occur when
bioventing systems are viewed as being identical to SVE systems, requiring relatively short time periods
to complete remediation. While BTEX compounds will be removed within this time period, the higher-
molecular-weight compounds are unlikely to be significantly impacted. Further research is necessary on
chlorinated solvent remediation to adequately assess system performance when treating these compounds.
Based on U.S. Air Force and commercial applications of this technology, the total cost ofin-situ soil
remediation using conventional bioventing technology is $10 to $60/cu yd (Downey et al., 1994). At sites
with over 10,000 cu yd of contaminated soil, costs of less than $10/cu yd have been achieved. Costs
greater than $60/cu yd are associated with smaller sites or those using amendments, but bioventing still
can offer significant advantages over more disruptive excavation options. O&M costs are minimal,
particularly when on-site personnel perform the simple system checks and routine maintenance that are
needed. Table 3-16 provides a detailed cost breakdown of remediation of 5,000 cu yd of soil
contaminated with an average concentration of 3,000 mg of JP-4 jet fuel per kg of soil.
Table 3-16. Typical Full-Scale Conventional Bioventing Costs (Downey et al., 1994)
Task Total Cost ($)
Site Visit/Planning 5,000
Work Plan Preparation 6,000
Pilot Testing 27,000
Regulatory Approval 3,000
Full-Scale Construction
Design 7,500
Drilling/Sampling1 15,000
Installation/Startup 4,000
2-Year Monitoring 6,500
2-Year Power 2,800
Soil Sampling at 2 Years 13,500
Total 90,300
Assumes four air injection wells drilled to a depth of 15 ft.
Ward (1992) compared costs of conventional bioventing to other in-situ bioremediation technologies
(Table 3-17). Costs shown in Table 3-17 reflect actual costs for these three technologies at fuel spills at
Traverse City, Michigan. Even though the area treated through bioventing was larger than that treated
with hydrogen peroxide or nitrate, total costs for bioventing were significantly lower than for the other
technologies.
67
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Table 3-17. Cost Comparison ofln-Situ Bioremediation Technologies Utilized at Fuel Spill Sites
(Ward et al., 1992)
Task
Total Costs
Hydrogen Peroxide
Addition
Construction2 45
Labor/Monitoring 72
Chemicals 500
Electricity 24
Total 641
($/m3 of Contaminated
Nitrate Addition
118
96
30
12
256
Earth)
Conventional
Bioventing1
26
40
0.44
6.8
73
1 Values reflect only first 4 months of demonstration.
2Prorated to a 5-year service life on buildings, pumps, and blowers.
3.1.6 Enhanced Bioventing
Enhanced, or cometabolic, bioventing is a variation of conventional bioventing that involves the addition
of a cometabolic substrate within an air stream. Addition of nutrients (N, P, and K) may also be required.
The cometabolic substrate promotes aerobic growth and the subsequent cometabolic degradation of the
contaminant of interest. As an aerobic process, cometabolic bioventing can use compressed air and the
growth substrate is provided as a vapor at low percent levels. As far as physical appearance and
schematic views, cometabolic bioventing looks the same as conventional bioventing. The only difference
is that instead of injecting air alone into the subsurface, a cometabolic substrate (i.e., propane or methane)
is also injected.
3.1.6.1 Principles of Operation
Cometabolism is a biological process that involves the fortuitous degradation of CAHs by bacteria during
the degradation of an organic cosubstrate. For the purposes of this application, this discussion focuses on
the cometabolism of CAHs. Bacteria that grow on hydrocarbons typically initiate oxidation by
incorporating molecular oxygen into organic compounds by the action of enzymes known as oxygenases
(Wackett and Householder, 1989). Two types of oxygenases, monooxygenases and dioxygenases, are
involved in the cometabolic oxidation of CAHs. The mono- and dioxygenases are relatively nonspecific
with respect to the type of organic compounds that they will attack, and bacteria that use these oxygenases
for the degradation of a growth substrate often accidentally attack secondary substrates, such as CAHs.
A variety of growth substrates have been used to stimulate cometabolic CAH degradation, including
aromatic compounds (e.g., toluene and phenol), methane, butane, propane, and propene. Ammonia
oxidation is also known to stimulate cometabolic CAH degradation. Among the organic cosubstrates,
methane, butane, propane, and propene are gaseous carbon compounds under ambient conditions and can
be introduced with air to supply both carbon and oxygen for biological growth. Each of these gaseous
compounds is nontoxic and biodegrades rapidly under aerobic conditions. Because this is an aerobic
process, cometabolic bioventing does not require the displacement of oxygen in the vadose zone. In fact,
the primary differences between cometabolic bioventing and conventional bioventing are the addition of
the cometabolic cosubstrate at low percent levels and subsequent monitoring requirements.
Consequently, cometabolic bioventing can use much of the same equipment and methods as conventional
bioventing. Some exceptions include enhanced safety regulations, such as the use of steel pipe instead of
PVC pipe to carry the mixed gas stream, aboveground monitoring for potentially explosive gases, and
68
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other similar safety measures. The emphasis on safety cannot be overstated. The use of flammable or
potentially flammable gas mixtures in vadose zone soils presents significant, unique safety concerns if
any cometabolic growth substrate has not degraded before air exits the vadose zone.
3.1.6.2 Target Contaminants
Chloroethenes, chloroethanes, and chloromethanes are the most extensively studied compounds for
cometabolism; exceptions include PCE and carbon tetrachloride, neither of which can be oxidized
aerobically. A typical application of cometabolic bioventing could be the addition of methane or propane
to promote the growth of methanotrophs or propane-degrading bacteria, respectively, and the cometabolic
degradation of TCE or lower chlorinated ethenes.
3.1.6.3 Advantages and Limitations
Cometabolic bioventing has several advantages over conventional remediation technologies, combining
the benefits of bioventing with the added value of cometabolic activity. Specific advantages include:
The cometabolic component allows for enhanced biodegradation of CAHs, thus reducing vapor
emissions and potentially increasing the degree of remediation attainable. Cometabolism
achieves mineralization of the CAHs and does not result in the production of toxic byproducts or
the transfer of contaminants to another matrix such as granular activated carbon (GAC), as is the
case with SVE combined with off-gas treatment using GAC.
Conceptually, cometabolic bioventing is a simple process that uses commercial, off-the-shelf
equipment. The bioventing component requires standard air compressors or blowers that can
deliver air to the subsurface efficiently and at relatively low flowrates. To enhance
cometabolism, the process requires the addition of a gaseous cosubstrate (such as methane or
propane) that is widely available and for which no special preparation is needed before use.
Limitations of cometabolic bioventing do exist and require thorough evaluation prior to the wide-scale
implementation of this technology. Some uncertainties about long-term effectiveness remain, as is the
case with any technology, but proper monitoring and evaluation can mitigate these uncertainties. Specific
limitations include:
Similar to most remedial technologies, the effectiveness of cometabolic bioventing can be limited
by soil heterogeneities at a site (i.e., differing permeabilities). In fact, minimal differences in soil
permeabilities may lead to areas within a site that are not treated as rapidly as more permeable
portions of the vadose zone.
Cometabolic bioventing uses potentially explosive gas mixtures that require extra safety
precautions to protect workers and the public.
The use of cometabolic growth substrate should be tested in the laboratory before field testing to
ensure that a bacterial population is present to catalyze the desired reactions for this process.
Very high CAH concentrations could be inhibitory to the growth of bacteria.
Off-gas control may be required when injecting air into the subsurface to ensure that CAHs are
not volatilized and released into the atmosphere.
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Although the above challenges can be overcome technically, the driving force will be the economics of
the process and the potential to develop a cost-effective process that can compete with conventional
technologies such as excavation and SVE.
3.1.6.4 Technology Cost Drivers
Major cometabolic bioventing cost drivers are similar to conventional bioventing cost drivers, except for
the addition of a cometabolic growth substrate. Major cometabolic bioventing cost drivers in addition to
bioventing cost drivers include:
The cost of the cometabolic growth substrate and additional plumbing and capital costs to support
the addition of the growth substrate
Pilot tests (as needed) to demonstrate the efficacy of the cometabolic process
Explosive growth substrates (e.g., methane, propane, or butane) require unique safety features to
prevent explosive conditions; these may include flame or spark resistant materials, preventative
measures to ensure that explosive conditions to not exist above ground,and automatic shutdown
controls in the event of a system failure, to name but a few.
3.1.6.5 Technology Performance and Cost
EPA/NRMRL, the Environmental Security Technology Certification Program (ESTCP), and Battelle are
testing the use of cometabolic bioventing using propane as the primary growth cosubstrate to promote the
biodegradation of TCE in vadose zone soils at Hill AFB, UT. Performance and cost data are not yet
available for this site.
ESTCP, Battelle, and Oregon State University employed cometabolism to treat TCE-contaminated
groundwater at McClellan AFB (MAFB), CA. Although this site involved the treatment of groundwater
and saturated soils, it is presented here as an example case study of the treatment of CAHs using
cometabolism (Lynch et al., 2001). Furthermore, because propane and CAH contaminants enteredhe
vadose zone during the sparging process, cometabolic degradation in the vadose zone also hadthe
potential to enhance CAH removal from the site. The MAFB pilot demonstration usedpropane and
methane as growth substrates to treat TCE-contaminated groundwater. Performance of this technology at
MAFB is summarized below:
For the first 500 days of treatment, propane was added as the cometabolic substrate. However,
there was no sign of propane degradation or cometabolic CAH degradation in the vadose zone
throughout the propane feed period, despite the presence of active propane degradation and
cometabolism in the saturated zone.
Methane was used as the substrate from approximately day 500 to day 550. In several monitoring
points, propane, TCE, and c-DCE were depleted concurrently with methane, and in one well TCE
and c-DCE were reduced to below detection limits, indicating degradation of these compounds by
cometabolism.
Microcosm studies demonstrated that propane-degrading bacteria was present at the site. The fact
that these bacteria were present and cometabolism did not occur in the vadose zone suggests the
following observations:
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o The presence of propane-, methane-, or other substrate-degrading bacteria does not
necessarily imply that cometabolic CAH degradation will ensue. CAH degradation rates
varied widely in microcosms using soils from different sites despite the presence of
active propane degradation in all microcosms.
o Microcosm and other pilot-scale studies are integral to successful implementation of
cometabolic bioventing due to the need to establish the presence and activity of
indigenous, substrate-degrading bacteria.
While these results show that cometabolism can be used for site restoration, it also reinforces the fact that
site-specific microcosm testing is required to verify the presence of microorganisms and microbial
processes of interest. Until this technology becomes widely used, field pilot testing is warranted to
adequately demonstrate the process viability for any given site.
Table 3-18 provides a summary of various full-scale enhanced bioventing installations, including
information on contaminant type and concentrations, target cleanup levels, and status of site cleanup (See
Appendix B for the data used to generate Table 3-18). Figure 3-16 shows box and whisker plots for
contaminants treated using enhanced bioventing, as reported in Appendix B. Figure 16a presents the
range of contaminants before treatment, while Figure 16b illustrates their range after treatment. Table 3-
19 illustrates the data used to generate Figure 3-16. Information provided in the tables and figures in this
section reflects an inability to gather complete data on all sites, either from a reluctance of site owners to
impart the information or from missing site data.
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Table 3-18. Summary of Site Characteristics at Enhanced Bioventing Installations
to
Site Name
Former Bulk
Terminal
NY State
Dept.
Northrop-
Hawthorne
BNRR
Fueling Pump
House
Oakland
Chinatown
Turtle Wax
Car Wash
US Coast
Guard Air
Station
van Oss
Site
Status
Closed
Active
Closed
Closed
Closed
Closed
Closed
Closed
Contaminants
of Concern
Diesel
DCE, A, MEK,
TCE
TRPH
No. 2 diesel
fuel, TPH,
BTEX
BTEX; TPH
B, BTEX
Gasoline, JP-4
jet fuel
Mineral oil
Maximum
Contaminant
Concentrations
(mg/kg dry
weight)
33,000
20 TCE; 4.8 DCE;
7.4 MEK; ISA
20,000
52,000 TRPH
BTEX: 100; TPH:
5000
29,393 BTEX; 181
B
NA
< 5,000
Unit Cost
(per CY)
$56
$390
$18
$10
$90
$60
$73
$106
Numerical
Treatment Goals
(mg/kg dry weight)
100 TPH
TCE: 1.5, DCE: 0.6;
MEK: 0.6; A: 0.2
100
NA
0.05 BTEX; 100
TPH
16 BTEX; 0.25 B
NA
900
Treatment
Goals
Achieved
Yes
Yes
NA
Yes
Yes
Yes
NA
Yes
Concentration
Achieved
(mg/kg dry weight)
10
< 0.005 TCE and DCE;
< 0.05 MEK and A
NA
5,000 TRPH; 73 total
BTEX (benzene ND)
50 TPH
14BTEX;0.16B
<0.1 JP-4;<0.2
gasoline
<490
Treatment
Effectiveness
NA
NA
NA
60%
NA
NA
60% fuel removed in 4
months.
all below goal.
Treatment
Duration
(Months)
24
5
NA
24
18
48
4
10
B = benzene
T = toluene
E = ethylbenzene
X = xylenes
A=acetone
MEK = methyl ethyl ketone
DCE = dichloroethylene
TCE = trichloroethylene
TRPH = total recoverable petroleum hydrocarbons
TPH = total petroleum hydrocarbons
Table 3-19. Concentrations of Contaminants of Concern Before and After Enhanced Bioventing: Data Used to Generate Figure 3-16
Median:
25th Percentile
75th Percentile
High whisker(a)
Low whisker(a)
Concentrations of Contaminants of Concern
Chlorinated VOC/SVOC
Before Treat-
ment (5 sites)
20
7
250
NA
5
After Treat-
ment (5 sites)
0.25
0.01
0.25
NA
NA
BTEX
Before Treat-
ment (3 sites)
181
104
14,787
29,393
100
After Treat-
ment (4 sites)
7
0.2
29
NA
NA
Fuel / Oil
Before Treat-
ment (5 sites)
20,000
5,000
33,000
52,000
NA
After Treat-
ment (6 sites)
30
3
380
490
NA
PAH, Creosote
Before Treat-
ment (1 site)
7
7
7
NA
NA
After Treat-
ment (1 site)
0
0
0
NA
NA
a) High and low whiskers represent the most extreme data point within 1.5 interquartile ranges, i.e., 1.5 times the distance between the 25th and 75th percentiles
-------
r-v
.0 +J
CQ
งง
s -
is
.S w
^ o
100,000
10,000
1,000
100
10
0.1
0.01
0.001
I
Chlorinated
VOC/SVOC
(5)
100,000
10,000
^ 1,000
ki
"a wo 100
-
.
s s
10
o
U
0.1
0.01
0.001 ^
T
Chlorinated
VOC/SVOC
(5)
Fuel / Oil
(5)
BTEX
(3)
PAH,
Creosote
(1)
Contaminant
I
Fuel / Oil
(6)
BTEX
(4)
PAH,
Creosote
(1)
Contaminant
Figure 3-16. Box and Whisker Plots Showing the Range of Contaminant Concentrations
Before and After Enhanced Bioventing
(See Appendix B for data used to generate these plots).
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3.1.7 Bioslurping
Bioslurping is primarily used for rapid and effective removal of LNAPL floating on the water table and is
applicable to most hydrocarbon releases that result in a significant layer of free-phase petroleum product.
The technology consists of applying a strong vacuum capable of supporting multiphase extraction from
aquifer sediments. Free product, groundwater, and soil vapor are extracted together through a manifold
system and separated above ground. Free product is typically recycled, and water and vapor streams
usually require treatment prior to discharge. An operating bioslurping system is shown in Figure 3-17,
and Figure 3-18 is a schematic illustrating how bioslurpers work.
Figure 3-17. Bioslurping
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Light, Nonaqueous-Phase
Liquid Saturation
Figure 3-18. Bioslurping Schematic
Bioslurping is described here because the soil vapor convection imposed by a bioslurping system creates
an oxygen-rich environment in the vadose zone concomitant with free product removal and performs as a
bioventing system above the water table. Indeed the "bio" component of bioslurping is bioventing. (The
"slurping" moniker is derived from the liquid extraction process that typically occurs in slugs, much like
slurping liquid from a straw in an almost-empty glass.) After recoverable free product has been removed
from the water table, a bioslurping system is easily transformed into a standard bioventing system to
bioremediate residual soil contamination. While bioslurping systems are designed with free product
removal as the primary objective, life cycle designs incorporate the likely eventual conversion of the
system to bioventing only.
3.2 BIOREMEDIATION TECHNOLOGIES AND THEIR APPLICATIONS - EMERGING
TECHNOLOGIES
This section describes emerging technologies with respect to their principles of operation and the types of
contaminants for which they may be appropriate. General cost and performance data reported in the
literature are provided, if available, but because these technologies have not yet evolved to significant
full-scale use, full-scale cost and performance data are generally not available in the literature. The
technologies discussed in this section include anaerobic bioventing, phytoremediation, sequential
anaerobic/aerobic treatment, and natural attenuation.
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3.2.1 Anaerobic Bioventing
Anaerobic bioventing is the injection of anaerobic (i.e., oxygen-free) gases into the vadose zone to
establish an anaerobic environment by displacing the oxygen-rich soil gas. Common gases used for this
purpose include nitrogen (N2) and argon (Ar). The anaerobic gas is infused with hydrogen (H2) gas to
supply an electron donor and promote reductive dechlorination of the chlorinated hydrocarbons. Practical
applications of anaerobic bioventing include remediating vadose zone soils contaminated with PCE or
TCE, such as those commonly found at dry-cleaning facilities.
Technology Description. Anaerobic bioventing is comparable to aerobic bioventing with the exception
that a gaseous blend that promotes anaerobic conditions is injected into the vadose zone in lieu of air or
pure oxygen (O2). As with conventional bioventing, gaseous uptake must be monitored. In the case of
anaerobic bioventing, H2 takes the place of oxygen and H2 utilization rates are measured in place of
oxygen utilization rates. The goals of anaerobic bioventing are to reduce O2 in the soil in the pore spaces
of the contaminated vadoze zone through displacement with an inert gas, and to lower the redox level and
promote reductive dechlorination by providing an electron donor in the form of H2.
If the electron donor is molecular hydrogen, the sequential reactions are as follows for PCE
dechlorination to DCE:
PCE + H2 -ป TCE + HC1 (2)
TCE + H2 -ป DCE + HC1 (3)
Similar dechlorination reactions can be expected for other halogenated hydrocarbons including aliphatic
and aromatic compounds. Further dechlorination of DCE to VC and ethene may be possible, but are
likely to involve reduced dechlorination rates and consequently longer bioventing requirements. Longer
acclimation periods also could be expected for DCE and VC dechlorination than for PCE and TCE
dechlorination. Alternatively, DCE and VC could be degraded aerobically or cometabolically. Although
this approach has not yet been tested, aerobic degradation of the PCE and TCE dechlorination byproducts
could be promoted by exchanging the anaerobic gas for air or pure O2 after the reductive dechlorination
process becomes rate limiting.
Advantages and Limitations. Anaerobic bioventing has several advantages over conventional
remediation technologies, combining the benefits of bioventing with the added value of anaerobic
dechlorination. Specific advantages are described in the following paragraphs:
Reductive dechlorination of CAHs should reduce vapor emissions and potentially increase the
degree of remediation attainable.
Aerobically recalcitrant compounds can be treated in situ, without relying on extraction and off-
gas treatment and disposal. Contaminants can be reduced to nontoxic byproducts through careful
application of the anaerobic process to ensure complete dechlorination, or through
anaerobic/aerobic sequencing to promote the dechlorination of higher-chlorinated compounds and
the subsequent aerobic degradation of the dechlorination byproducts.
This is an emerging technology that has the potential to offer a cost-effective means of cleaning
up chlorinated hydrocarbons in vadose zone soils. Because this approach has many similarities to
that of aerobic bioventing, conventional bioventing equipment and knowledge can be used to
develop this technology and apply it in the field. The exception is that extra safety precautions
76
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must be taken when working with compressed H2 gas in the field. Additional work is needed to
develop a more economical way to scrub O2 from air.
Limitations of anaerobic bioventing exist and require thorough evaluation prior to the wide-scale
implementation of this technology. Some uncertainties about long-term effectiveness remain, as is the
case with any technology, but proper monitoring and evaluation can mitigate these uncertainties.
Specific, disadvantages include the following:
Similar to most remedial technologies, the effectiveness of anaerobic bioventing can be limited
by soil heterogeneities at a site (i.e., differing permeabilities). In fact, minimal differences in soil
permeabilities may result in non-uniform distribution of injected gases. This nonuniform
distribution may lead to areas within a site that are not treated as rapidly as more permeable
portions of the vadose zone, or in zones that remain aerobic.
Reductive dechlorination of CAHs generally requires an acclimation period for bacterial growth
or to otherwise stimulate the dechlorination process. Sequential acclimation periods may be
required for parent compounds and their daughter products. The length of an acclimation period
may be able to be predicted using laboratory microcosms, but differences in the field should be
anticipated.
Incomplete dechlorination can result in the production of undesirable byproducts; in the case of
PCE treatment, dechlorination can result in the production of DCE isomers and VC.
Anaerobic bioventing uses potentially explosive gas mixtures that require extra safety precautions
to protect workers and the public.
Because anaerobic bioventing is in a relatively young stage of development, its potential for cost
savings remains unknown.
The use of anaerobic bioventing requires further field testing to demonstrate its effectiveness.
Site-specific pilot testing also may be required.
Very high CAH concentrations could be inhibitory to the growth of anaerobic bacteria.
Off-gas treatment and control may be required when injecting gases into the subsurface to ensure
that CAHs are not volatilized and released into the atmosphere. Controlling the rates of gas
injection will minimize the need for off-gas treatment and control.
Although the above challenges can be overcome technically, the driving force will be the economics of
the process and the potential to develop a cost-effective alternative that can compete with conventional
technologies such as excavation and SVE.
Performance. EPA NRMRL and Battelle are testing anaerobic bioventing in the field at Salina, KS, and
at Hill AFB, UT. There are strong indications that the injection of a N2/H2 gas mixture into the vadose
zones of these sites displaced the O2 in the soil gas and lowered redox potentials. As of the writing of this
report, additional performance data were unavailable.
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3.2.2 Phytoremediation
Phytoremediation is an emerging treatment technology that utilizes plants, plant microbial systems, soil
amendments, and agronomic techniques to remove, stabilize, and/or degrade environmental contaminants.
Figure 3-19 shows a phytoremediation plot in the field at an MGP site in Bedford, IN. The technology
takes advantage of the natural hydraulic and metabolic processes of plants, resulting in a technology that
is passive and driven by solar power. Phytoremediation may be employed exclusively or in tandem with
conventional treatment technologies to remediate soil and groundwater contaminated with heavy metals,
pesticides, chlorinated solvents, explosives, crude oil, PAHs, and landfill leachate.
:
^K^f,ซ<
Figure 3-19. Phytoremediation
Phytoremediation is a broad term that describes a number of mechanisms that are defined by plant
systems. These mechanisms described below are illustrated schematically in Figure 3-20.
Phytoremediation may involve plant uptake of contaminants, or it can exploit enhanced biological activity
associated with the rhizosphere. In general, these two treatment modes are governed by the physical
characteristics of the contaminant(s) of concern. The water solubility and soil sorption capacity are two
major chemical characteristics that influence the phytoremediation of pollutants. Water-soluble inorganic
pollutants and those organic pollutants with intermediate log Kows (approximately 1-4 L/g) are taken up
by roots and are considered to be good targets for phytoremediation. Organic pollutants that fall outside
this range are not readily taken up by plants and are targets for extracellular plant enzymes or microbiota
associated with the rhizosphere. The following phytoremediation mechanisms are described below:
Phytostabilization
Phytoaccumulation/phytoextraction
Phytotransformation/phytodegradation
Hydraulic control
Rhizodegradation
Rhizofiltration.
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Phytovolatilization
Evapotranspi ration
H2O + VOCs
Phytoac c u m u lat i o n/
Phytoext faction
Uptake
Rhizo(sphere) Degradation
HYTOREMECHATON CDB
Figure 3-20. Phytoremediation Schematic
Phytostabilization is the use of plants to increase sequestration of contaminants (heavy metals and
hydrophobic organics) in soil. Soil sequestration occurs as plants alter water flux and reduce contaminant
mobility. Plants and microbial enzymes bind contaminants into soil (humification). Plants also
incorporate free contaminants into plant roots (lignification) and prevent wind and rain erosion.
Phytoaccumulation/Phytoextraction uses specific plant species to absorb unusually large amounts of
metals. This mechanism is typically used for remediation of soils and groundwater contaminated with
heavy metals (Pb, Cd, Zn, As, Cu, Cr, Se, and U). Uptake plants may be harvested and later ashed.
Phytotransformation/Phytodegradation is the process where plant enzymes completely mineralize or
partially break down contaminant compounds such as herbicides (atrazine, arochlors), aromatics (BTEX),
chlorinated aliphatics (TCE), nutrients (NO3, NH4, PO4), and ammunition wastes.
Hydrologic Control (organic pumps)/phytovolatilization is the use of plants to control the migration of
contaminants in groundwater by exploiting their natural hydraulic properties. This application can result
in chemical uptake and transpiration of chemicals through the leaf tissue.
Rhizodegradation is the process whereby plant roots excrete sugars, acids, and alcohols that
microorganisms in the rhizosphere utilize for food. Through biostimulation or cometabolism,
microorganisms subsequently degrade organic contaminants such as pesticides, aromatics, and PAHs.
Rhizoflltration utilizes plant root systems developed hydroponically to remediate contaminated waste
streams that are directed through the root mass; alternatively, plant root systems may be encouraged to
develop in situ in contaminated saturated zones to uptake contaminants from contaminated waste streams
79
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or leachate beds. Rizofiltration typically is used for heavy-metal-contaminated water and landfill
leachates.
Each of these mechanisms has been demonstrated with varying levels of success. Most sites involve
more than one mechanism, and most studies, particularly field studies, cannot easily distinguish between
the different mechanisms that contribute to overall site remediation. Of the mechanisms described above,
phytostabilization, phytoaccumulation/phytoextraction, phytotransformation/phytodegradation, and
rhizodegradation may be used for soils treatment. However, while hydrologic control and rhizofiltration
are confined to groundwater or surface water wastes, they also may contribute to lessening the amount of
contaminant infiltration through soils.
3.2.2.1 Advantages and Limitations
Phytoremediation can be a cost-effective treatment because it is an in-situ technology that is passive and
solar driven. It can be employed for a wide variety of environmental contaminants and provides an
aesthetically pleasing site appearance. It is a relatively unobtrusive application with site soils remaining
in place, and, unlike chemical treatment technologies, will not alter soil characteristics.
The technology is limited to the ability of the plants of choice to be subjected to abnormal environments
where contaminant concentrations must remain below the toxicity threshold to maintain a healthy plant
population. For applications that involve contaminant uptake through plant root systems or for
applications that utilize the rhizosphere, roots must be able to reach to the vicinity of contamination,
which at some sites and for some plants may be limited to surface soils. Several growing seasons may be
needed so that the plant(s) of choice may reach a level of maturity required for the remedial process.
There is growing concern about the use of non-native species for phytoremediation. Native species or
species that will not threaten the existing biodiversity in the vicinity of the site should be used. This
restriction could limit the optional selection of plants to meet site-specific cleanup requirements.
3.2.2.2 Technology Performance
Unfortunately, insufficient field data are available to evaluate phytoremediation performance. For those
field studies that have been completed, phytoremediation has not been an effective remedial strategy for
soils under most circumstances. While phytoremediation seems to be a good technique for hydraulic
control of groundwater, it does not appear to be an effective stand-alone technology for cleanup of
contaminated soils. At sites where there is no danger of contaminated soils impacting the population,
phytoremediation may be a good long-term option for eventual cleanup. The fact that phytoremediation
has not worked well to date to remediate soils limits its use in the field at full scale, and keeps
phytoremediation in the emerging technology category of this report. The following four issues generally
must be addressed to gain regulatory acceptance of phytoremediation at a site (Rock and Sayre, 1999):
Provide site-specific evidence for the effectiveness of phytoremediation using site soils and
contaminants; this includes laboratory and/or greenhouse evidence
Adequate containment of contaminated soils must be assured; plants must establish themselves
to a point where they contain/degrade the contaminants of interest.
Site monitoring must address the fate of contaminants in soils and plants; monitoring the efficacy
of an innovative treatment such as phytoremediation tends to be more extensive than is required
for more conventional technologies
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If phytoremediation is attempted, but does not succeed, an adequate backup plan with a high
chance of success must be in place.
To implement phytoremediation at field scale, the current database must be expanded to show
contaminant-specific field rates and extent of activity of the various phytoremediation mechanisms. More
information is needed that outlines specific data on plants and their interactions with contaminants under
different climatic conditions. In addition, regulators need a standardized monitoring system. Currently,
there is no consensus from the scientific community as to what parameters should be considered critical
measurements (Rock and Sayre, 1999).
Numerous research teams, private companies, and government agencies have been working to advance
phytoremediation for soil applications. A variety of contaminant applications have been investigated over
recent years. One of the more extensively studied forms of phytoremediation is phytoextraction. Private
companies and industry have implemented phytoextraction in the field. One such company, formally
known as Phytotech Inc., dedicated its practice to the application of lead remediation using Brassica
juncea to achieve high metal accumulation rates. In a Brownfield demonstration in Trenton, NJ,
Phytotech used B. juncea combined with soil amendments to reduce the average soil surface lead
concentrations by 13%. The target soil concentration of 400 mg/kg was achieved in approximately 72%
of the treated area in one cropping season (Blaylock et al., 1999). Another pilot-scale demonstration was
conducted to extract uranium from contaminated soil at a former DOE facility in Ashtabula, OH (no data
available) (Zodrow, 1999). Additional site-specific case studies can be found in Appendix B.
Bench-scale research has shown that plants are capable of enhancing the mineralization rate of a variety
of herbicides, such as atrazine and methoxychlor (Burken and Schnoor, 1996; Kruger et al., 1997). Soils
in the rhizosphere of Kochia plants were shown to degrade atrazine by 62% after 36 days over that of
nonvegetated controls. Other rhizospheric soils capable of atrazine degradation included mush thistle,
catnip, foxtail barley, witchgrass, and lambsquarter (Elsevier, 1996). Others have shown the ability of
plants to facilitate the degradation of other hazardous organic compounds in the rhizosphere including
2,4-D, Diazinon, PAHs, and TCE (Anderson et al., 1993; Shann, 1995).
Researchers at Kansas State University conducted greenhouse studies to determine the effect of the
rhizosphere on PAH-contaminated soil. The rate of PAH removal was compared in rhizosphere soil,
nonrhizosphere soil, and sterile soil. Non-rhizosphere soil contained no roots but was spiked with root
exudate to mimic the carbon contribution to the rhizosphere that normally occurs in whole root systems.
The results showed that after 180 days soils with plants had more than twice the mineralization rate than
soils without plants, indicating that the presence of plants is a necessary part of the phytoremediation
process. Simple exudation of organic compounds did not fully mimic the presence of roots (Banks et al.,
1996).
Others have found that vegetation can be actively used to promote microbial degradation of TCE (Walton
and Anderson, 1990; Brigmon et al., 1999). Significantly higher levels of TCE-degrading cultures were
found in the rhizosphere of TCE-contaminated soils than in nonvegetative contaminated soils. This
increase in TCE degradation in the rhizosphere indicates that plant root/microbial interaction had a
significant role in enhancing TCE removal.
Researchers have found that red mulberry (Morus rubra L.) roots release phenolics into soils and that
these compounds can serve as carbon sources for the growth of PCB-degrading bacteria (Hedge and
Fletcher, 1996).
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3.2.3 Sequential Anaerobic/Aerobic Treatment
For soils, anaerobic biodegradation involves the dechlorination of halogenated hydrocarbons, such as
PCBs, dioxins and furans, and pesticides. Anaerobically, these compounds undergo microbially catalyzed
reductive dechlorination, which removes chlorine substituents from the more highly chlorinated
congeners and replaces them with hydrogen atoms. Reductive dechlorination is a well-understood
process that has been studied extensively in the laboratory for soils and sediments contaminated with
PCBs, chlorinated pesticides, and CAH. For PCBs, it has seen limited field application due to limitations
in the extent of PCB dechlorination, and because destruction removal efficiencies usually are insufficient
to meet treatment goals for PCBs. However, by shifting the congener distribution to less chlorinated
analogs, the overall toxicity of the mixture is typically reduced and the mixture becomes more susceptible
to aerobic degradation. For pesticides and herbicides, anaerobic/aerobic sequential treatment faces
similar problems with respect to achieving cleanup goals. For CAHs, anaerobic or ex situ processes have
difficulty competing with more conventional technologies like SVE.
In general, PCB reductive dechlorination preferentially removes chlorines from the meta and para
positions and replaces them with hydrogen atoms, resulting in substantial reductions in carcinogenicity
and "dioxin-like" toxicity, which tends to coincide with meta-chlorinated PCBs. In addition to lowering
the overall toxicity of PCB-contaminated materials, the tendency of the PCB mixture to bioaccumulate is
also reduced, as discussed previously. For example, 2-chlorobiphenyl and 2,2-dichlorobiphenyl display
an approximately 450-fold decrease in the tendency to bioaccumulate in fish compared with tri- and tetra-
chlorinated PCBs (Abramowicz and Olson, 1995).
3.2.3.1 Advantages and Limitations
The primary advantage of anaerobic/aerobic processing for soils contaminated with recalcitrant
compounds like PCBs is that it may be used to address soils contaminated with relatively highly
chlorinated PCB congeners that cannot be degraded aerobically. Another potential advantage may be its
applicability to soils with high organic and/or moisture contents, where oxygen depletion and anaerobic
conditions can be promoted readily.
The primary limitation of anaerobic/aerobic processing for soils is that PCBs degrade slowly and
biodegradation of PCBs has not yet been demonstrated to consistently meet treatment goals, usually in the
range of 50 mg/kg or less. Because this technology has yet to be proven consistently reliable, it is not
often used for full-scale treatment. If used, it must be preceded by a site-specific, field-scale treatability
study to demonstrate its effectiveness.
3.2.3.2 Technology Performance
Several approaches have been attempted to enhance the microbially catalyzed reductive dechlorination of
PCBs. Researchers have attempted to stimulate dechlorination by amending microcosms with carbon
substrates (e.g., fatty acids). Although in some cases this resulted in shortened lag times or increased
initial rates of dechlorination, the overall extent of PCB dechlorination was not significantly increased
(Abramowicz and Olson, 1995). Others have attempted to stimulate dechlorination by adding individual
polychlorinated or polybrominated congeners to microcosms. The process is designed to selectively
enhance populations of organisms that can use the supplied congener as an electron acceptor. In one
instance, this strategy reduced 79% of hexa- through nonachlorobiphenyls in sediments contaminated
with Aroclor 1260; the resulting dechlorination products were predominately tri- to pentachlorobiphenyls
(Abramowicz and Olson, 1995). This approach may not be applicable to in-situ remediation efforts due
to the potential regulatory resistance encountered at the prospect of adding polyhalogenated biphenyls to a
contaminated site, but it may be appropriate for ex-situ applications.
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The use of anaerobic biodegradation as a stand-alone treatment for PCB-contaminated sites would likely
be hampered by regulatory treatment goals that are generally based on mass removal, not toxicity
reduction. Although anaerobic dechlorination may provide greater toxicity reduction than aerobic PCB
biodegradation, it is less likely to be used at a contaminated site because it does not produce the same
level of PCB mass removal. The discrepancy in mass removals can be partially explained by the fact that
a mass reduction of only 34.45 g is observed for each mole of chlorine atoms removed via reductive
dechlorination, while a 233.45 g mass reduction (about seven times more) is observed for each mole of
trichlorobiphenyl destroyed aerobically. To optimize the reduction of both PCB toxicity and mass
removal, researchers have begun investigating the utility of following anaerobic biodegradation with an
aerobic biodegradation step.
Sequential anaerobic-aerobic biodegradation is a two-step process in which soils contaminated with
large-molecular-weight chlorinated PCBs are first incubated anaerobically to reductively dechlorinate the
more heavily chlorinated compounds. The anaerobic incubation is followed by aerobic incubation
intended to degrade the resulting mass of lower-chlorinated congeners. In principle, this process should
permit a significantly greater PCB mass removal compared to anaerobic or aerobic processes alone
because the anaerobic process is confined primarily to relatively highly chlorinated PCBs (i.e., PCBs with
four or more chlorines), while the aerobic process is confined to the relatively low chlorinated PCBs
(i.e., PCBs with three or fewer chlorines).
Evans et al. (1996) conducted a study of sequential anaerobic-aerobic treatment of PCB-contaminated soil
collected from a capacitor bank at a power substation; PCB concentrations were approximately
100 mg/kg, and the congener pattern resembled weathered Aroclor 1248. A 19-week anaerobic
incubation period resulted in a 50% reduction of the raeto-substituted chlorines (from 1.5 to 0.75 meta-
chlorines/biphenyl), but little change in the />ara-chlorination pattern, yielding mostly para- and ortho-
substituted PCBs. (Longer incubations of up to 60 weeks did not yield significantly more dechlorination
than the 19-week period.) Inoculation with Pseudomonas sp. LB400 and aerobic incubation for 19 weeks
followed the anaerobic incubation period. During the aerobic incubation period, all dichlorobiphenyls
and a large fraction of the tri-, tetra-, and pentachlorobiphenyls were degraded, resulting in a 70% overall
decrease in total PCB concentrations using sequential anaerobic/aerobic treatment. Without anaerobic
treatment, aerobic treatment alone resulted in a 67% PCB concentration reduction, leaving primarily tetra-
and pentachlorobiphenyls. The fact that the vast majority of congeners in Aroclor 1248 are susceptible to
aerobic degradation skews data in favor of the aerobic incubation for this particular aroclor. Nonetheless,
results showed that microcosms undergoing aerobic treatment alone contained a higher proportion of
penta- and hexachlorobiphenyls. Only 1% of the congeners in Aroclor 1248 contain six or more chlorine
atoms. Anaerobic dechlorination, on the other hand, is more effective for the more highly chlorinated
congeners. Thus, for some sites where lower-chlorinated PCB congeners predominate, it may be more
effective to undergo solely aerobic degradation.
Shannon et al. (1994) demonstrated a 9% decrease in the PCB mass concentration after 12 weeks of
anaerobic incubation, followed by an additional 72% decrease after aerobic incubation resulting in a total
mass reduction of 81%. However, this study did not compare results of sequential anaerobic/aerobic
treatment with aerobic-only treatment.
Although no full-scale field demonstrations of PCB bioremediation have been documented, a few
pilot-scale field demonstrations have been conducted with varying degrees of success. General Electric
(GE) conducted the first field-scale attempt to bioremediate PCBs at a former racing drag strip in New
York contaminated with Aroclor 1242. Initial PCB concentrations at the site ranged from 50 to 525 ppm
in a 5-m x 12-m area rototilled to a depth of 20 cm and inoculated with Pseudomoniasputida Strain
LB400, an aerobic PCB-degrading bacterium. The area was dosed with 200 L LB400 (2 x 109 cells/mL).
PCB biodegradation was first detected after 8 to 10 weeks, and the maximum reported PCB loss was
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approximately 25% in the top 3 cm of soil after 18 weeks; no degradation was observed in the control
plots that did not receive LB400. This reduction was only about half of what was expected based on
bench-scale treatability studies conducted in the laboratory using site soil. Researchers speculated that
environmental factors such as temperature and moisture content in the soil hampered treatment
effectiveness due to high summer temperatures at the site (McDermott et al., 1989). The heat dried the
soil and likely desiccated the bacteria; LB400 cell counts on the soil were virtually zero only 2 days after
inoculation during the hottest weather.
In 1991, GE attempted another PCB bioremediation field study with Hudson River sediments in an in-situ
bioslurry using caisson reactors with added inorganic nutrients, biphenyl, and oxygen; the study duration
was 73 days. The initial concentration of PCBs in the sediments was 39 ppm, and a 37 to 55% reduction
in PCB concentrations was observed; repeated inoculation with a purified PCB-degrading bacterium
failed to improve biodegradative activity. A possible explanation for the low destruction was low
bioavailability of the PCBs. The authors speculated that the resistant PCB fraction was in a sorbed state
and would have to diffuse through the organic matrix before it became desorbed and bioavailable
(Harkness et al., 1993).
PCB-contaminated sludge from the Madison, WI Metropolitan Sewerage District was applied to test plots
in eight combinations to study the effects of PCB sludge concentration, sludge loading rate, and sludge
application methods on performance (Gan and Berthouex, 1994). PCB concentrations ranged from 25 to
75 mg/kg for untreated soils, with approximately 85% of the untreated PCBs being 2-, 3-, 4-, and 5-
chlorinated PCBs. Most PCB congeners showed significant decrease in their soil concentrations over
time, although the more highly chlorinated PCBs were more persistent in the sludge than the lower
chlorinated PCBs. Biodegradation was thought to be the primary removal mechanism based on results of
active versus control plots, and analysis of leachate data from the plots. A simple first-order model was
used to describe the disappearance of PCBs in the surface soils of the sludge-amended farmland. The
half-lives of 2-C1, 3-C1, and 4-C1 congeners ranged from 7 to 11, 5 to 17, and 11 to 58 months,
respectively. A total of 24 congeners (mostly 6-, 7-, and 8-C1 PCBs) appeared to remain stable in the
soils.
3.2.4 Natural Attenuation
Natural attenuation is a remedy-of-choice that allows natural biological, chemical, and physical processes
(biodegradation, dispersion, diffusion, weathering, etc.) to slowly minimize or "attenuate" contaminant
concentrations. While the use of natural attenuation has become extremely popular in recent years for the
remediation of contaminated groundwater (Use of Monitored Natural Attenuation at Superfund, RCRA
Corrective Action, &UST Sites. Directive 9200.4-17P. EPA, 1999b), there is very little information in
the literature documenting the use of this technology for treatment of contaminated vadose-zone soils.
Most of the available literature on natural attenuation of soils pertains to metals, which are susceptible to
physical and chemical (i.e., abiotic) reactions with the soil matrix that can "stabilize" the metals through
sorption and precipitation reactions. These stabilization reactions render the metals less mobile and
ultimately less bioavailable. There is virtually no literature documenting the use of intrinsic
bioremediation as a remedy for vadose zone soils contaminated with organic contaminants. Figure 3-21
is a schematic presentation of how natural attenuation works in groundwater.
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Monitoring Wells
Transverse to the
Groundwater Flow Direction
Source
Area
Background
Monitoring
Well
Monitoring Wells
Aligned Along the
Plume Centerline
Downgradient
Extent of
Dissolved Plume
Direction of Groundwater Flow-
Explanation
O> Plume boundary
ฎ Monitoring well
Soil sample point
WOURAL ATTENUATION CDR
Figure 3-21. Groundwater Natural Attenuation Schematic
PAHs and other high-molecular-weight compounds undergo a variety of weathering processes, including
dissolution and dilution into the aqueous phase, sorption and sequestration, volatilization, and
biotransfbrmation. It is important to understand the mechanisms involved in those weathering processes
to understand the potential for natural attenuation of contaminated soils, and to understand the impact of
weathering processes on the risks posed by the contaminants. The popularity of natural attenuation for
groundwater contaminants is due in part to the fact that the most mobile contaminants tend to be the most
readily biodegradable contaminants with regard to petroleum hydrocarbons. Octanol-water partition
coefficient, Kow, has been shown to be negatively correlated with solubility and biodegradability;
chemicals with low Kow values tend to be transported through the soil and enter the groundwater, where
they may be degraded, whereas chemicals with high Kow values tend to remain sorbed onto soils (Brady et
al., 1999). Poorly soluble organic compounds are often resistant to intrinsic biodegradation by virtue of
the fact that it is difficult for them to readily reach reaction sites in the microbial cells (Haider, 1999). So
while poorly soluble compounds tend to be degraded slowly, they also tend not to be very mobile and
tend to remain associated with soils.
While weathering of soils has been investigated historically, little information is available regarding its
ability to remediate a site through natural attenuation. Contaminant weathering in soils can occur to
provide a permanent contamination reduction. Weathering includes such mechanisms as dilution,
volatilization, biodegradation, and sequestration. However, these mechanisms do not affect all
contaminants equally. PAH weathering tends to preferentially remove the lower-molecular-weight PAH
compounds that are more easily degraded or mobilized (two- or three-ringed compounds).
Another potential explanation for the paucity of documented soil-natural-attenuation case studies is the
phenomenon of decreasing bioavailability with time. Prolonged contact of organic contaminants with
soils causes sorbed chemicals to become increasingly less available for microbial biodegradation
(Alexander, 1994), presumably due to sequestration or penetration of the chemical into biologically
inaccessible sites (Adriaens et al., 1999; Linz and Nakles, 1997). Microorganisms are typically present in
pore sizes between 0.25 (im and 6 (im in diameter; the 0.25-(im diameter is the smallest size limit for
entry of bacteria, while sizes in excess of 6 (im allow for entry of protozoa that feed on bacteria (Adriaens
et al., 1999). Thus, sequestration may limit the microbial bioavailability of organic contaminants and the
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ability for natural attenuation to result in the permanent destruction and removal of these contaminants
from soils. Figure 3-22 shows that the amount of a contaminant mineralized by microorganisms is lower
in soils that have been in contact with the contaminant for a longer period of time. Thus, compounds with
the greatest tendency to sorb to the soil matrix tend to be biodegraded the slowest.
40
Figure 3-22. Illustration of Decreased Biodegradability of Aged Phenanthrene in Soil
(from Adriaens et al., 1999; originally from Hatzinger and Alexander, 1995)
In addition to being sequestered, certain organic chemicals such as pesticides can also form soil-bound
residues, which also limits their degradability by microorganisms. These complexes form as a result of
attachment of the compounds to reactive sites on the surface of organic colloids or by their incorporation
into the structure of humic and fulvic acids (Adriaens et al., 1999).
Limited microbial bioavailability should not be confused with reduced toxicity. The risk posed by sorbed
compounds may not decrease simply because they are less likely to desorb in the environment. Thus, the
sequestration of inorganic and organic contaminants may reduce the ability of microbes to biodegrade or
transform these contaminants in the environment, but may have little overall impact on the toxicity of the
contaminated medium.
Despite these limitations, natural attenuation may be a viable treatment technology for soil-bound
contaminants, but more research is required to demonstrate its effectiveness and to develop a protocol for
the investigation and demonstration of this technology. For metals, this may take the approach of
demonstrating that they have become irreversibly sorbed into crystal lattices and thus effectively isolated
from soil, groundwater, and biota, as suggested by Brady et al. (1999). For organic contaminants, the
study of natural attenuation would require the investigation of weathering phenomena such as dissolution,
adsorption, sequestration, and biodegradation.
3.3 EXISTING BIOREMEDIATION TECHNOLOGIES SUMMARY
This section reviews the following conventional bioremediation technologies for their effectiveness to
treat contaminated soils: land treatment, biopile treatment, composting, bioslurry reactors, enhanced
bioventing, and conventional bioventing. Emerging technologies including phytoremediation, sequential
anaerobic/aerobic treatment, natural attenuation, and anaerobic bioventing also were reviewed with
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respect to the state-of-knowledge required for their application in the field. In an effort to focus on actual
cost and performance information, and to gauge the extent of their application at full scale, a variety of
on-line and off-line data bases were reviewed for existing case studies. The summary of this case study
search is presented in Appendix B and was used to evaluate the specific technologies.
The cost of various conventional treatment technologies can vary greatly from site to site, depending on
site-specific contaminants and conditions. Table 3-20 shows the primary types of equipment used for the
conventional technologies discussed in Section 3.1, and Table 3-21 identifies factors that commonly
impact biological treatment costs negatively.
Table 3-20. Primary Equipment Used for Conventional Bioremediation Technologies
Technology Typical Equipment Types
Land Treatment
Containment pad
Leachate collection and management system
Tilling equipment
Leachate collection and treatment system
Biopile/Biocell
Soil pile support pad or container
Aeration pipes and blowers
Off-gas treatment equipment
Leachate collection and treatment system
Nutrient feed and chemical stabilizer system
Composting
Mixing equipment for organic additives
Organic additives and bulking agents
Soil pile support pad and aeration pipes and blowers for windrow
turning machine or composting reactor (depending on method)
Off-gas treatment equipment (for static pile or in-vessel treatment)
Slurry-Phase Bioreactors
(soil and sediment)
Soil/sediment mixing equipment
Bioreactor with aeration components
Clarifier
Off-gas treatment system
Nutrient feed and chemical stabilizer system
Bioventing
Air injection wells
Blowers
Soil-gas monitoring points
Off-gas treatment equipment (if required)
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Table 3-21. Factors That Tend to Increase Costs for Conventional Bioremediation Technologies
Cost Factor
Concentration of
contaminants
Presence of higher-
molecular-weight organics
Area or volume requiring
treatment
Depth of contamination
Complex geology
Low soil permeability
Presence of recalcitrant
contaminants
Presence of halogenated
organics
Comments
Higher concentrations increase the time needed for treatment.
Higher-molecular-weight organics (e.g., PAHs) tend to increase
treatment time.
Increased area or contaminated volume requires increased capital
expenditures. However, unit costs typically decrease with increased
volume to be remediated.
Deeper contamination increases the amount of well drilling required for
bioventing or excavation requirements for ex-si tu applications.
Complex interbedding of high and low permeability layers can be
difficult to treat, possibly increasing the density of wells needed or the
length of treatment time for bioventing or increased mixing requirements
and treatment time for ex-si tu applications.
Treating low permeability soils decreases the radius of influence of in
situ technologies, requiring an increase in the treatment density for in-
situ treatment.
Relatively recalcitrant contaminants may require increased treatment
time or alternative treatment strategies, or may preclude biological
treatment entirely.
Halogenated contaminants may require anaerobic pretreatment to reduce
the level of chlorination, and may require increased off-gas control
measures.
3.3.1 Applicability of Bioremediation for Contaminated Soils
To date, bioremediation's greatest successes have been with technologies that exploit the use of aerobic
processes for the biotreatment of organic contaminants, especially for petroleum hydrocarbons, which
have been shown to degrade under a wide variety of environmental conditions. Each of the aerobic
bioremediation processes have been shown to be capable of meeting cleanup goals for petroleum
hydrocarbons such as BTEX, TPH, and many PAHs. The use of composting is fairly well accepted for
more complex waste streams such as soils contaminated with explosives or larger-molecular-weight
PAHs. The following criteria must be met to ensure the efficacy of aerobic bioremediation to meet
treatment goals for contaminated soils:
The contaminant of interest must be able to be degraded (preferably mineralized) biologically; if
the contaminant is not mineralized, biotransformation should not result in the production of toxic
byproducts
Bacteria must be present that are capable of biodegrading the contaminant
If bacteria are not immediately available for contaminant degradation, a suitable acclimation
period should be provided, or the soils may be augmented with known contaminant-degrading
bacteria (both of these processes should be pilot tested before implementing at full scale)
Nutrients required for biodegradation must be readily available
The contaminant must be bioavailable; for hydrophobic contaminants that are strongly sorbed to
soils, bioavailability is likely to be a rate-limiting step toward complete remediation
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Soil chemical conditions, such as alkalinity and pH, must be suitable for bioremediation
A suitable electron acceptor must be abundantly available; for aerobic processes, this means
ensuring that oxygen is not limiting (in water, DO *2 mg/L; in soil gas O2 > 5%)
Adequate time for acclimation must be provided to promote bacterial growth for contaminant
degradation; this may involve weeks or months, depending on the contaminant and the media
chemical and microbiological characteristics; for some sites, preferential degradation of easily
degraded compounds may necessitate prolonged acclimation periods before the more recalcitrant
contaminants of interest are degraded.
While bioremediation is well understood and well accepted for the aerobic treatment of easily degraded
petroleum hydrocarbons, the challenge for bioremediation increases significantly for more recalcitrant
contaminants and/or more complex site conditions. Large-molecular-weight PAHs, PCBs, chlorinated
aromatic and aliphatic compounds, dioxins, pesticides and herbicides, and nitroaromatic compounds all
present unique obstacles for biological treatment, particularly with respect to the ability of bacteria to
degrade and ultimately mineralize these various contaminants. It is common for these compounds to be
biotransformed and not degraded or mineralized, often resulting in the production of potentially toxic
byproducts, some of which can be equally or more toxic than their parent compound.
The addition of bacteria cultivated for the degradation of a specific compound (bioaugmentation) has
been used to promote faster and potentially more complete degradation, but has seen limited use in the
field, most likely because it has not yet been proven to be a reliable strategy. The fate of bacteria added
to a complex soil medium is unpredictable. Furthermore, such bacteria may favor the degradation of
more readily degradable contaminants before they degrade the contaminant of interest.
There has long been interest in anaerobic processes to biotransform or degrade more recalcitrant
compounds, particularly halogenated and nitro-substituted compounds. Anaerobic processes are seeing
increased field applications and continue to be investigated in the laboratory and the field. Anaerobic
treatment systems tend to be more difficult to control and maintain than are aerobic systems, and, for
certain compounds, anaerobic systems catalyze the incomplete biotransformation of contaminants and
require subsequent degradation using aerobic bacteria. In some cases, such as the dechlorination of
highly chlorinated PCBs to lower chlorinated PCBs, dechlorination can have a significant detoxification
effect (NRC, 2001); yet even this may be insufficient for site restoration if cleanup requirements are
based on total PCB removal and not on a congener-specific treatment approach to detoxify soils. In other
cases, such as in the production of VC from chloroethene or chloroethane dechlorination, the daughter
product may be more toxic than its parent compound. Consequently, anaerobic bioremediation often may
require post treatment that may include aerobic or physical/chemical treatment processes.
Anaerobic/aerobic sequenced technologies have significant potential for remediation of recalcitrant
contaminants, but continued research is needed to bring them to the marketplace. Despite its obstacles,
anaerobic treatment has substantial promise for treating soils, either solely or in combination with aerobic
treatment, particularly if it can be employed in situ, as would be the case for in-situ anaerobic bioventing.
In addition to anaerobic bioventing, cometabolic bioventing is gaining interest as an aerobic approach to
treat certain halogenated aliphatic contaminants. Cometabolic bioventing provides another alternative for
vadose zone remediation and complete destruction of CAHs that do not require anaerobic conditions.
3.3.2 Bioremediation Advantages and Disadvantages
Perhaps the greatest advantages and attractions of bioremediation are that 1) contaminants can be
degraded to nontoxic byproducts and some can be mineralized, resulting in the complete removal of the
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contaminant from the environment; 2) bioremediation uses natural means to achieve contaminant
removal; and 3) bioremediation is perceived as relatively inexpensive and environmentally friendly. The
fact that bioremediation can detoxify contaminants using natural methods means that contaminants are
not displaced to another medium, such as activated carbon, that could require further treatment, and they
are not discarded in a landfill. A primary disadvantage of landfill disposal is that the original owner
retains ownership of the contaminated medium indefinitely. Thus, a technology such as bioremediation
that can detoxify or destroy contaminants on site is very attractive because of its ability to reduce or
eliminate liability for contaminants.
The perception that bioremediation is a relatively low-cost treatment alternative probably stems from the
fact that most bioremediation technologies rely on low-tech and low-cost methods. Aeration, for
example, rarely requires the use of pure oxygen and can be affected by simply blowing air through the
soil media using conventional blowers. While nutrient addition and bioaugmentation are difficult in situ,
ex-situ nutrients (and bacteria) can be added and mixed with soils during excavation and stockpiling or
during construction of the treatment system. Some of the most significant cost impacts on bioremediation
systems are sampling and O&M. Because biotreatment is often a relatively slow process, O&M costs
tend to escalate when treatment operations are prolonged from weeks to months, or to years. Minimizing
O&M through careful design and management can have a significant impact on reducing overall
treatment costs for biotreatment processes. While the reported costs for bioremediation range from less
than $50/cu yd to more than $l,000/cu yd, bioremediation treatment costs should be competitive with
physical/chemical treatment processes; bioremediation treatment alternatives that greatly exceed
physical/chemical treatment costs may be inappropriately targeting recalcitrant contaminants or
conditions not well suited for bioremediation.
The most significant disadvantage of bioremediation is that, in general, it is much less predictable than
most physical/chemical and thermal technologies, often necessitating pretreatment or field pilot-scale
studies. The reason for this is that the site-specific bacterial population, physical conditions, and nutrient
availability for the degradation of a specific contaminant cannot easily be predicted. For example, while
extensive research has been conducted to demonstrate the ability of bacteria to biodegrade PAHs, the
actual rate of PAH biodegradation and the potential for certain large-molecular-weight PAHs to be
recalcitrant to biodegradation cannot be predicted without treatability testing. Thus, at some sites PAH
degradation may be relatively rapid, requiring months of biotreatment, while at other sites PAH
degradation may require years or may not be achieved at all. Fortunately, treatability studies are practiced
widely and can easily be implemented at a relatively low cost. Their greatest disadvantage is that they
may take a long time and have the potential to significantly delay site restoration. Under a climate of
wanting to remove wastes rapidly, many site owners are unwilling to take time to conduct a treatability
test, which itself is unpredictable, and often prefer the more predictable results of physical/chemical or
thermal treatment alternatives.
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4.0 IMPLEMENTATION OF BIOREMEDIATION TECHNOLOGIES
Bioremediation testing and implementation follow similar steps for testing and implementation of any
remedial process, namely, site characterization, treatability testing for technology screening/selection, and
pilot-scale testing. While the specific data needs of bioremediation technologies differ, there are many
common parameters that affect the implementation of any remedial technology and these must be known
to effectively select an appropriate alternative. The following sections summarize the general approach
for implementing the bioremediation technologies described in this report
4.1 SITE CHARACTERIZATION
Before selecting any remedial approach, it is necessary to perform some level of site investigation to
characterize the contamination, delineate the horizontal and vertical extent of the contamination, define
the hydrogeologic conditions including the depth to the water table and the soil stratigrahy, and determine
key soil characteristics such as pH, moisture, texture, and permeability. Site characterization for
successful implementation of bioremediation technologies requires more investigation than do
physical/chemical processes because a wide range of parameters may impact the desired metabolic
activity of microorganisms in the soil environment. The parameters of importance depend on the type of
contaminant being treated, the microbial process being exploited, and the selection of an in-situ or ex-situ
approach. It is recommended that the following important screening parameters be measured in any site
investigation where bioremediation is being considered.
Contaminant type(s)
Depth to groundwater
Soil stratigraphy/texture
Contaminant distribution
pH and alkalinity
Organic matter
Nutrients (N, P, K)
Soil gas O2 and CO2
Electron acceptors
Soil moisture
Soil permeability
Microbial populations
1. Contaminant type(s): It is important to determine not only the contaminant(s) of concern (COC) from
a regulatory perspective, but also to know what contaminants and co-contaminants are present that
may affect the microbial degradation of the target compounds. Co-contaminants can range from
compounds that are preferentially degraded over the COCs and hence exert a demand for remedial
reagents, to compounds that are inhibitory to the microorganisms that carry out the desired
degradation reactions. An example of the latter case was observed at a wood preserving site in
Minnesota where the degradation of the PAHs of interest was preceded by degradation of the carrier
oil components in a creosote bioventing project (McCauley et al., 1999). This required the bioventing
system to be operated for 3 years before any significant reduction in COCs occurred.
2. Depth to groundwater: Knowing the depth to groundwater and the amount that the depth fluctuates
over time is important for designing any remedial approach and for screening technologies. Sites
with shallow water tables (<3 - 5 ft) can pose problems with subsurface delivery of remedial reagents
such as gas injection, and surface application of liquid solutions could be a concern due to
mobilization of contaminants to the aquifer. Depending on site specifics, shallow water table sites
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may be more suitable for ex-situ technology applications. Sites with deeper water tables are more
appropriate for in-situ bioremediation if subsurface reagent delivery is possible. Sites that exhibit
large water table fluctuations can pose challenges due to the cyclic saturated/unsaturated conditions
that can impede reagent delivery and the potential for continuous smearing of contaminants when
NAPL is present. In summary, knowing the depth to the water table is a crucial parameter for
selecting and designing a remedial approach, and the limitations of this parameter on the potential for
implementation of a specific technology are dependent on other site characteristics.
3. Soil stratigraphy/texture: Observations made during drilling or soil coring are used to define the
stratigraphy for the site. Knowing the stratigraphy is important for understanding how the
contamination is distributed in the soil and determining the potential for delivering remedial reagents.
Less permeable soils such as silts and clays tend to slow the migration of contaminants and often
contain more contaminants than the more permeable strata. The presence of more permeable strata
can facilitate delivery of reagents; a high degree of silt and clay content in the soil can preclude such
delivery. Contaminants held up in very tight soils are difficult to treat in-situ and depending on the
depth and volume needing treatment may be candidates for ex-situ treatment. Fracturing technologies
are being developed to "open up" tight soil formations to allow delivery of reagents.
4. Contaminant distribution: Defining the contaminant distribution in the subsurface is a prerequisite
for selecting any remedial approach. Knowing vertical and horizontal extents of the contamination
allows calculation of the volume of soil requiring treatment, the potential for, and extent of,
excavation for ex-situ treatment, and designing the placements for injection/extraction and monitoring
devices for in-situ treatment.
5. pH and alkalinity: Although microbial activity occurs over a wide pH range, the optimal pH for the
majority of the soil microbial activity exploited for contaminant destruction is usually in the range of
pH 6 to 8. The addition of certain reagents as well as enhancement of microbial activity can often
cause a significant pH change, which in turn can adversely affect contaminant degradation.
Measuring the pH and alkalinity provides a measure of how well the system is naturally buffered
against such changes. If the buffering capacity is too low, buffer addition might be required. Buffer
addition is more easily applied to ex-situ soil treatment as in-situ delivery often requires saturation of
the soil, which can result in a decrease in the effective porosity and impede reagent delivery and/or
cause problems with contaminant mobility.
6. Organic matter (Total Organic Carbon): Naturally occurring organic matter can affect degradation
performance by imparting an oxygen demand, serving as an electron donor, or decreasing the
effective bioavailability. Organic matter can interfere with the delivery of remedial reagents
including both vapor-phase and liquid nutrients. Typically, high organic concentrations are found in
shallower soils in areas with thick vegetation. Soils with high percentages of organic matter tend to
hold water. Distribution of reagents throughout the subsurface can be difficult, and ex-situ treatment
may be required.
7. Nutrients (N, P, K): Microbial activity in soils with low concentrations of essential nutrients can be
limiting, slowing the degradation process and extending the time required for treatment. In some
soils, nutrient concentrations can be limiting to the point that degradation does not occur. The
primary nutrients, nitrogen (N), phosphorus (P), and potassium (K), can easily be measured to
determine if nutrient-limiting conditions exist. Trace nutrients may be present in concentrations
below analytical detection capabilities, but may still be present in sufficient quantities. Treatability
studies are useful for evaluating nutrient limitations and can be designed to investigate various
nutrient addition scenarios to optimize degradation and minimize chemical costs. Although nutrient-
to-substrate ratios are often prescribed, care must be taken when adding nutrient salts so as not to
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decrease the soil water potential to the point where microbial activity is adversely impacted.
Nutrients are added easily to soils treated in ex-situ reactors; in-situ nutrient addition can be more
challenging.
8. Soil-gas oxygen and carbon dioxide: O2 and CO2 concentrations in soil gas provide a good
indication of the aerobic biodegradability of the contaminant, the presence/absence of indigenous
microorganisms that can degrade the contaminant, and the need for O2 delivery to promote the
degradation by those organisms. A decreased ratio in the percent O2 to the percent CO2 in the
contaminated soil compared to that in a non-contaminated background soil is a good indicator of
increased biological activity and that the increased activity is due to the presence of the contaminant.
Note, these ratios are not direct evidence that a specific compound of interest is being degraded if that
compound is the sole contaminant. The ratios are considered reliable indicators of the biodegradation
potential for the fuel hydrocarbon mixtures, fairly reliable for PAH mixtures such as creosote or MGP
wastes, fairly reliable for mixtures of other directly aerobically metabolizable compounds such as
certain halogenated benzenes and phenolics, and not so reliable for compounds that are only
aerobically degraded cometabolically such as certain chlorinated solvents and/or other
chloroorganics.
9. Electron acceptors: Biological degradation reactions involve the transfer of electrons. The molecule
from which the electron is removed is termed the "electron donor," and the molecule to which the
electron is transferred to is termed the "electron acceptor." Under different conditions, contaminants
can serve as either electron donor or acceptors. The aerobic degradation of benzene is an example of
the contaminant serving as the electron donor as electrons are transferred from the benzene to oxygen
resulting in the formation of CO2. It is important to know the electron acceptor concentrations in the
soil under these conditions to calculate the amount of O2 that needs to be added to promote the
desired degradation. Anaerobic degradation of PCE via dehalorespiration is an example of the
contaminant serving as the electron acceptor as electrons are transferred from hydrogen to PCE,
resulting in the reductive dechlorination of the PCE to the lesser-chlorinated ethenes. In the case of
PCE, H2 generally serves as the electron donor. H2 can be provided via direct gaseous H2 injection, as
in the case of anaerobic bioventing, or via the anaerobic fermentation of an organic substitute.
10. Soil moisture. Soil microorganisms, like all life forms, require water. This includes soil microbes
that grow attached to soil particles as well as those suspended in the water bound to soil particles and
within the interstitial spaces between soil particles. Typically, the water content of soil is measured
as soil moisture, and optimal moisture content for biological activity is reported to be between 60%
and 80% of field capacity. A more appropriate measure of water availability is the soil water
potential that takes into account the salinity of the water. Measuring the water potential is more
involved than measuring the soil moisture, but the data are more useful, especially when considering
adding nutrient salts.
11. Soil permeability: Soil permeability provides a measure of the potential for effective delivery of
remedial reagents to subsurface soils. Soil gas permeability is measured by injecting air at different
flow rates and measuring pressure changes at radial distances from the point of injection. Soil liquid
permeability can be measured by two methods, infiltration from the surface using a double ring
infiltrometer where a constant head is maintained in two concentric rings and the rate of infiltration is
measured in the center ring, or permeability in the subsurface using a modified slug test. The data
from these tests can be used as a direct indicator of the ability to deliver reagents and may exclude in-
situ treatment if soils are too tight. Soil gas permeability improves the reliability of designing well
spacing and placement for technologies such as bioventing. Liquid injection is subject to gravity flow
through preferential flow channels. This coupled with the fact that the diffusion of liquid reagents is
limited makes liquid delivery in the vadose zone more challenging than vapor-phase delivery.
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12. Microbialpopulations: Enumerating microorganisms from soils is by no means an exact science
because of the difficulty in removing the cells from the soil particles coupled with the fact that many
soil microorganisms are simply not culturable in the laboratory. Even though this may be a
shortcoming from a quantification standpoint, it is possible to enumerate various classes of
microorganisms as an indicator of the potential for promoting a desired biodegradative process.
General plate counts (i.e., total heterotrophs) provide a general measure of the ability of
microorganisms to survive in a soil, but provide no direct evidence of any biodegradation potential.
Comparing these counts to counts from soil collected from an uncontaminated location at the site is
sometimes used as indirect evidence and a very qualitative indicator of the impact of the contaminant
on the microbial population. More specific plating techniques can provide better evidence of the
potential for promoting microbial activity for a specific degradation process. This entails plating on
specialized media with the contaminant included with any other essential co-substrates, electron
donors, and/or other essential nutrients. These plating techniques can be used to isolate and
enumerate microorganisms with specific metabolic capabilities and/or requirements. Phospholipids
fatty acids (PLFA) analysis provides information on biomass concentration, metabolic activity level,
and what types of microbes are present in a soil sample. This technique can be used to determine
how environmental factors (temperature fluctuations, pollution, disturbances, etc.) affect a microbial
population. Fatty acid methyl ester (FAME) analysis provides a microbial fingerprint which can be
used to identify the bacterial strain present. Denaturing Gradient Gel Electrophoresis (DGGE)
analysis uses DNA sequencing to identify specific organisms present in a sample. By using the 16S
RNA gene and comparing DNA sequences using national databases, different bacterial species can be
identified. FAME and DGGE analyses can separate contaminant microbes and aid in positive
identification of specific microbes. Although microbial enumerations are often considered as an
optional analysis, proper technique can provide valuable information for technology selection and
design.
4.2 TECHNOLOGY SELECTION
Biotreatment technologies should be considered as viable remedial options for a wide range of
contaminants under a wide range of environmental scenarios. The individual technologies that should be
considered in the technology screening process are site specific as discussed below. The two governing
factors that would exclude consideration of bioremediation are short treatment time requirements and
contaminant biodegradability considerations. For the most part, bioremediation approaches can require
more time than the more aggressive thermal and/or physical/chemical treatment technologies. This might
not be true for in-situ applications where large volumes of soil are to be treated, but for smaller volumes
or for ex-situ applications, biologically-based processes are usually slower. What makes the biological
approaches attractive is that they:
Result in contaminant destructions, not simply phase transfer or removal for off-site disposal
Are often simple designs that require minimal energy input
Are less costly than most other alternatives
Allow on-site reuse of the soil.
Selection of a bioremediation approach is a stepwise process. The first step is to determine if an in-situ or
ex-situ application is required. Generally, in-situ technologies are preferred, but site constraints such as
soil permeability, depth to groundwater, and contaminant distribution may dictate the selection of an
ex-situ approach. Ex-situ technologies are preferred when the ability to deliver remedial agents is limited,
the remedial process includes reagents that could mobilize the contaminant (i.e., surfactants, buffers, or
liquid nutrients), and/or more exact control of the remedial process is required (i.e., temperature, pH,
moisture, nutrient concentration, or leachate control).
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4.3 TREATABILLITY STUDIES AND PILOT-SCALE TESTING
Biological processes are affected by complex interactions of many environmental variables, and the
potential for performance of any bioremedation technology is best assessed through treatability studies.
These studies can be conducted in the laboratory or at small scale in the field. The goals of the
treatability study are to determine if biological activity can be promoted to achieve the desired level of
treatment for a specific site soil and contaminant, and to determine the degradation kinetics to provide a
preliminary estimate of the time required to achieve treatment. Because treatability tests are relatively
inexpensive and they can provide information that is useful for screening and selecting technologies, it is
often recommended that they be conducted as part of any Remedial Investigation/Feasibility Study
(RI/FS) where bioremediation is being considered. During that time, treatability studies allow the
flexibility of investigating a range of environmental variables such as pH, alkalinity, temperature, and
nutrient addition, all of which can affect degradative performance and be controlled in the field when
using the appropriate technology. The data could prove useful for selecting ex-situ over in-situ
approaches.
Pilot-scale testing is conducted when technologies require site-specific design data for full-scale
implementation. Technologies that require more sophisticated engineering, such as bioslurry reactors,
require pilot-scale testing to collect data necessary to size reactor system components and determine the
energy input required to maintain the slurry in suspension. Less complex technologies can also benefit
from pilot-scale testing. For example, the standard practice for bioventing combines the treatability test
with the pilot-scale test to simultaneously collect data to calculate the soil-gas permeability and an initial
biodegradation rate. The soil-gas permeability data are used to determine vent well spacing, while the
initial biodegradation rate is used to determine an air exchange rate and to size the blowers. The initial
degradation rate can be used to make a preliminary estimate of the time that will be required to achieve
cleanup.
Both treatability and pilot-scale testing require the objectives of the tests to be clearly defined, the
experiments/tests to be designed and conducted to collect the data needed to achieve those objectives, and
a good quality assurance plan to ensure the validity of the data collected. Frequently, more detailed
scientific and engineering data are collected for both scale-up considerations and to better understand the
underlying microbial processes for system optimization.
4.4 FULL-SCALE DESIGN
Once treatability and/or pilot-scale testing is completed, full-scale designs can be completed that
incorporate the data collected and take site hydrogeological and other logistical constraints into
consideration. For many of the ex-situ bioremediation technologies, the volume of soil requiring
treatment, the time required/allowed for treatment, and space constraints dictate many aspects of the
full-scale design. For example, land treatment, biopile treatment, and composting are relatively simple
designs, but often require more residence time than the more complex biotreatment designs such as
bioslurry reactors and treating the same volume of soil in the same amount of time will require more
space.
In-situ technologies are screened and selected based on the results of the treatability tests, and full-scale
systems are designed based on hydrogeologic and contaminant distribution data collected during the site
investigation and/or pilot testing. The location of injection wells/points as well as monitoring devices
must take into account the soil permeability and the horizontal and vertical extent of the contamination.
The injection equipment (i.e., pumps, blowers, etc.) is designed based on the number and size of the
injection points and the kinetics of reagent utilization.
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4.5 PROCESS OPTIMIZATION
The treatment conditions defined by treatability and/or pilot-scale testing serve as the initial design
parameters for full-scale implementation. Optimizing the system involves tweaking the operational
parameters to decrease the costs of operation and/or the time required to achieve cleanup concentrations.
For example, aerobic technologies such as bioventing that promote contaminant degradation by providing
O2 achieve optimum performance when O2 concentrations in the soil gas are maintained above 5% to 8%.
At concentrations above this range, the degradation is zero order with respect to O2 so no enhanced
degradation is realized. Optimizing the system by decreasing the flow rate of the blower or operating in a
pulsed mode to maintain the O2 partial pressure at 5% rather than at ambient levels could reduce the
energy costs. Data generated during periodic soil-gas measurements and respiration testing can be used to
make such adjustments. An example of optimization of an anaerobic technology would involve adjusting
the feeding strategy and hence the addition of the electron donor to minimize the amount of donor that is
lost to methanogenesis. Ex-situ technologies might benefit from inoculating the soil with a small amount
of soil from the previous reactor run by shortening the time of acclimation and therefore the run time.
Controlling aeration to maintain the non-rate limiting O2 concentration can decrease energy input and
off-gas collection and treatment requirements. The process of system optimization requires periodic
system monitoring and making adjustments to system operation to compensate for any changes during the
treatment cycle.
4.6 CONTAMINANT DEGRADATION PLATEAUS
Biological degradation plateaus are temporary or semipermanent degradation endpoints that often occur
during the biotreatment of environmental contaminants. They occur for a variety of physical, chemical,
and microbiological reasons, usually resulting in residual contaminant concentrations that exceed target or
expected degradation endpoints. The major reasons for degradation plateaus include:
1. Preferential microbial degradation of compounds that are more easily biodegraded than the target
contaminants
2. Exhaustion of a cometabolic growth substrate
3. Limited contaminant bioavailability
4. Nutrient limitations
5. Predation of contaminant-degrading bacteria by opportunistic predators
6. Energetic limitations
7. Buildup of toxic intermediates.
This section briefly describes each of these biotreatment mechanisms and their potential for contributing
to the occurrence of biodegradation plateaus.
4.6.1 Preferential Degradation of Easily Degraded Compounds
Bacteria rely on environmental compounds for carbon, energy, and nutrients for growth. The microbial
degradation of organic compounds provides carbon necessary for growth, while oxidation/reduction
reactions that involve organic and inorganic compounds provide energy. As a general rule, bacteria
preferentially degrade the most energetically favorable compounds followed by compounds that are
decreasingly energetically favorable. In other words, bacteria first metabolize compounds that provide
the most carbon and/or energy for the least amount of work, provided that they have the necessary
enzymes and exist under the necessary environmental conditions to metabolize those compounds.
This phenomenon is often apparent in environmental media with contaminant mixtures, such as petroleum
hydrocarbons. In the presence of TPH, for example, and because biodegradation of PAHs and other
petroleum hydrocarbons tends to be more difficult with increasing molecular weight, bacteria
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preferentially degrade lower-molecular-weight compounds like short-chain aliphatics and BTEX,
followed by increasingly difficult-to-degrade compounds such as longer-chained aliphatic compounds and
lower-molecular-weight PAHs (2- or 3-ring PAHs), again followed by the most difficult-to-degrade
compounds like high-molecular-weight aliphatic compounds and the 4- to 6- ring PAH compounds. This
leads to reduced degradation rates for the more difficult-to-degrade, higher-molecular-weight compounds,
the apparent step-wise degradation of environmental contaminants, and the appearance of degradation lag
times or plateaus for the more difficult-to-degrade contaminants. These degradation plateaus may show
evidence of downward movement after bacteria exhaust the more easily degraded contaminants and begin
to degrade the more difficult-to-degrade compounds. Figure 4-1 is a conceptual illustration of atypical
contaminant plateau.
Engineered or Natural
Process Change
Potential Further
Degradation
Time
Figure 4-1. Contaminant Degradation Plateau
4.6.2 Exhaustion of a Cometabolic Growth Substrate
Aerobically, bacteria that grow on hydrocarbons typically initiate oxidation by incorporating molecular
oxygen into organic compounds by the action of enzymes known as oxygenases (Wackett and
Householder, 1989), which destabilize carbon-carbon bonds and render the organic molecule more
susceptible to degradation. In some cases, nonspecific oxygenases show activity for other compounds, a
process known as cometabolism. The oxidation of some environmental contaminants occurs through
cometabolism, which involves the degradation of a primary growth substrate and the fortuitous
degradation of the cometabolized contaminant. The best-known and most exhaustively researched
examples of cometabolism involve the aerobic cometabolic degradation of low-molecular-weight
chloroethenes (TCE, DCE, and VC) by methanotrophs (methane-degrading bacteria), utilizing the
methane monooxygenase enzyme. In addition to cometabolically degrading selective chlorinated
solvents, methane monooxygenase also has been shown to convert naphthalene to 1- and 2-naphthols
(Dalton et al, 1981).
A wide variety of organic compounds can serve as primary growth substrates or as cometabolically
degraded substrates. Although cometabolism has not been exploited as a bioremediation mechanism for
PAHs or PCBs to date, many scientists believe it is a relatively common phenomenon in the environment
and is likely responsible for a substantial portion of contaminant degradation at petroleum release sites.
PAH compounds, specifically the high-molecular-weight compounds that prove very difficult for bacteria
to metabolize, may be degraded cometabolically by nonspecific oxygenase enzymes of bacteria that
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degrade 2- and 3- ring PAHs. Evidence for this phenomenon with PAHs is beginning to surface at the
EPA's Bedford, IN site where multiple soil bioremediation technologies are being investigated.
At sites where cometabolism is primarily responsible for degrading a specific compound, degradation
plateaus occur when the primary growth substrate is exhausted. Assume, for example, that compound B
is cometabolically degraded along with compound A. The bacteria require compound A for carbon and
energy, which they do not obtain from compound B. If compound A is exhausted first, the non-specific
enzyme(s) responsible for compound B degradation will no longer be produced or activated, resulting in
the termination of compound B degradation. The result is a residual compound B concentration plateau.
Theoretically, compound B degradation will resume in the presence of more compound A, so that adding
compound A to the medium may be a potential strategy to further stimulate compound B degradation.
4.6.3 Limited Contaminant Unavailability
A prerequisite for the microbial degradation of any compound is that the compound be bioavailable. That
is, for bacteria to metabolize a substrate, the bacteria (or its enzymes) and substrate must come into direct
contact, and substrate mobility must permit its degradation. The adsorptive binding of hydrocarbon
compounds to soils can make them unavailable for biodegradation (Prince and Drake, 1999); this is
particularly true of the higher-molecular-weight compounds, which tend to have a higher affinity for
sorption to soils due to their higher degree of hydrophobicity. The hydrophobic nature of many high-
molecular-weight contaminants, including PAHs and PCBs, often renders these compounds unavailable
to bacteria. Furthermore, prolonged contact of organic contaminants with soils may cause sorbed
chemicals to become increasingly less available for microbial biodegradation (Alexander, 1994),
presumably due to sequestration, the migration of the contaminant into biologically inaccessible sorption
sites (Adriaens et al., 1999; Linz and Nakles, 1997). Contaminants with the greatest tendency to sorb to
the solid matrix tend to be biodegraded the slowest.
Contaminants in soils may be categorized in three fractions: a readily bioavailable fraction, a moderately
or slowly bioavailable fraction, and an unavailable fraction. The readily bioavailable fraction is
biodegraded with minimal lag and may include a mobile, aqueous phase of the contaminant, a NAPL
phase from which dissolution into the aqueous phase is not rate limiting, and a sorbed phase where
desorption into the aqueous phase also is not limiting due to relatively weak sorptive forces. The
moderately bioavailable fraction is represented by sorbed or NAPL phases that can desorb or dissolve into
the aqueous phase, but for which desorption/dissolution are rate-limiting steps toward biodegradation.
The unavailable fraction is represented by strongly sorbed or sequestered contaminants that for all
practical purposes are insoluble. It is the moderately bioavailable and unavailable fractions that have the
potential to create the appearance of degradation plateaus.
4.6.4 Nutrient Limitations
Nutrient limitations also can result in degradation plateaus. In addition to the need for carbon and energy,
microbial contaminant degradation also requires macronutrients like nitrogen and phosphorus and
micronutrients including metals and salts. Nutrient limitations can be determined through microcosm
testing or pilot testing by comparing nutrient-amended soils with unamended controls. At McClellan
AFB, for example, nitrogen limitations hindered propane degradation, which was being used to stimulate
cometabolic TCE degradation. Background nitrogen in groundwater was available at 5 mg/L; this
concentration was insufficient to support propane degradation when propane was added at 4% air (Lynch
et al., 2001; Tovanabootr et al., 2001). The result was a degradation plateau for TCE cometabolic
degradation. Because nitrogen could not easily be added to this in-situ application, propane gas
concentrations had to be reduced so that nitrogen requirements would not exceed the nitrogen availability
in groundwater.
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4.6.5 Predation of Contaminant-Degrading Bacteria
Predation of contaminant-degrading bacteria can theoretically reduce their population, resulting in a
degradation plateau. Environmental restoration technologies stimulate microbial contaminant degradation
by adding amendments, nutrients, or growth factors that optimize the environment for contaminant-
degrading bacteria. Creating such ideal environments for bacteria can result in the rapid proliferation of
contaminant-degrading bacteria, and a subsequent proliferation of predator microorganisms that live off
of contaminant-degrading bacteria. The result is a cyclic process where contaminants are degraded
relatively rapidly followed by a rapid decline in the biological population resulting in reduced degradation
rates. This may be followed by new bacterial growth and renewed contaminant degradation. In the
environmental restoration field, this phenomenon has been most apparent with methanotrophic
cometabolic processes in which the rapid and enormous proliferation of methanotrophic bacteria due to
methane and oxygen addition to the environment results in a proliferation of opportunistic
microorganisms that live off the methanotrophs. These opportunistic microorganisms lead to system
failures where the target cometabolic process ceases to work.
4.6.6 Energetic Limitations and Buildup of Toxic Intermediates
Degradation rate and extent is governed by microbial thermodynamics. Thermodynamic equations
depend both on the amount of energy that can be released by the contaminant degradation process as well
as the relative concentrations of degradation substrates and byproducts. Bacteria may be limited in their
ability to degrade contaminants to very low regulatory concentrations because they may be unable to gain
sufficient energy for growth at low contaminant concentrations. Contaminant-degrading enzymes may
have a relatively poor affinity for the contaminant of concern, also resulting in high concentration
endpoints.
Inhibition phenomena also can hinder contaminant degradation. Such inhibition can include substrate or
product inhibition, or competition among substrates for degradation. A well-known example of product
inhibition occurs during anaerobic fermentation of organic compounds. In an anaerobic consortium of
bacteria, syntrophic bacteria (syntrophs) comprise a unique group of bacteria that catalyze substrate
oxidations via interspecies hydrogen transfer. The syntrophs produce H2 via anaerobic fermentation of an
organic substrate, and require the removal of the H2 by methanogens, sulfate-reducing bacteria, or other
hydrogen-utilizing bacteria. Some syntrophs, including Syntrophomonas wolfei, are able to convert the
range of C-4 to C-8 compounds and some aromatic compounds such as benzoate and phenol to
short-chain fatty acids, particularly acetate, and to H2 and CO2 (Balows et al., 1992). Others
(e.g., S. sapovorans and C. bryantii) are able to use C-l 1 to C-18 compounds. Because fermentative
bacteria often operate at conditions very close to thermodynamic equilibrium, excess production of
acetate and H2 is thermodynamically unfavorable. Thus, the efficiency of the organic carbon fermentation
process depends on the efficient removal of acetate and H2by methanogens or sulfate reducers. The
involvement of syntrophic bacteria in the anaerobic degradation of organic contaminants requires a well
balanced, healthy biological system to maintain the contaminant degradation process and prevent the
appearance of degradation plateaus.
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5.0 FUTURE DIRECTION OF BIOREMEDIATION TECHNOLOGIES AND
BIOREMEDIATION RESEARCH
This report reviews the state-of-the-art of bioremediation, primarily focusing on organic contaminants.
Bioremediation has been and continues to be employed with relative confidence for a variety of waste
streams that are amenable to microbial degradation. However, there are numerous waste streams that
have proven difficult to remediate biologically. Continued research and development (R&D) is needed to
identify environmentally and economically marketable biotechnology approaches for remediating
recalcitrant wastes. This section focuses on future R&D needs to help identify contaminants, waste
streams, and treatment technologies that require further development, and to identify those developmental
requirements. The discussion of research needs is divided into three categories: 1) optimization of
technologies that are proven in the marketplace,
2) development of emerging technologies to bring them to the marketplace, and 3) fundamental R&D
approaches for waste streams for which there are no existing technologies.
5.1 OPTIMIZATION OF TECHNOLOGIES THAT ARE PROVEN IN THE MARKET PLACE
As summarized in Section 3, the aerobic degradation of TPH and of lower-molecular-weight aliphatic and
aromatic hydrocarbons is well understood and has been applied at hundreds of sites using bioventing,
land treatment, biopile treatment, composting, or bioslurry reactors. There is a wealth of information
available to scientists and engineers to aid in the design and implementation of these technologies. For
contaminants that are readily biodegraded aerobically, such as BTEX, low-molecular-weight TPH and
PAHs, and some wood-treating wastes, these aerobic biotechnologies can be used "off-the-shelf," and
laboratory or pilot-scale treatability tests can be minimized or even eliminated at most sites. In fact, EPA
has designated bioremediation as a "presumptive remedy" for wood-treating waste streams. Thus, the
regulatory approval for the aerobic treatment of such nonrecalcitrant contaminants can be readily
achieved, and treatment can be conducted with confidence that these contaminants can be removed to
meet treatment goals.
The confidence with which these technologies can be applied raises the question of where R&D can best
be applied for these technologies. The following research needs may be applied generally to the aerobic
technologies listed in this section.
1. Process improvements: Process improvements may include reduced O&M requirements, improved
contaminant destruction and removal efficiencies, enhanced contaminant destruction and removal
rates, and increased contaminant throughput rates.
2. Increased target contaminant range: While these technologies are available for readily degradable
contaminants, there is a wide range of relatively recalcitrant contaminants that may require
treatability testing before field implementation or for which these technologies may not yet be
applicable. Such contaminants include larger-molecular-weight PAHs, PCBs (particularly Aroclor
mixtures with a wide range of PCB congeners), dioxins and furans, some chlorinated solvents and
pesticides, and explosives.
3. Improved monitoring techniques: A significant amount of O&M is expended on process monitoring.
Improved monitoring may include the use of automated and/or remote sensors, improved statistical
approaches for monitoring to minimize sampling requirements, and improved analytical methods.
4. Nutrient requirements: Nutrient addition has the potential to enhance contaminant degradation rates,
and at some sites may be a limiting factor to contaminant degradation.
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5.2 DEVELOPMENT OF EMERGING TECHNOLOGIES TO BRING THEM TO THE
MARKETPLACE
R&D is still needed for emerging technologies to bring them to the marketplace for full-scale
implementation. Much of the microbiology of these technologies is relatively well understood, and while
there is a continuous need to better understand the microbiology, the engineering application of these
technologies remains the limiting factor for bringing them to the marketplace. Emerging technologies
include anaerobic treatment, anaerobic/aerobic sequencing, cometabolic bioventing, phytoremediation,
bioaugmentation, and natural attenuation.
5.2.1 Anaerobic Treatment and Anaerobic/Aerobic Sequencing
Anaerobic treatment technologies show increasing promise for treatment of relatively recalcitrant
contaminants including chlorinated solvents, chlorinated aromatics (including some pesticides and PCBs),
petroleum hydrocarbons, nitroaromatics (including explosives), and heavy metals (Fathepure and Tiedje,
1999). Anaerobic treatment offers a wide range of alternative electron acceptors (i.e., NO3", Mn4+, Fe3+,
SO4 , and CO2 for methanogenesis) under which biodegradation can occur, and there is increasing
evidence of contaminant degradation under these various electron acceptor conditions. While anaerobic
treatment holds great promise for bioremediation, its application at full scale has been limited. A primary
limitation of anaerobic treatment technologies is that they often result in biotransformation byproducts
that may be more harmful than their parent compounds. For example, PCE or TCA dechlorination can
result in the production of DCE isomers and VC; PCB dechlorination generally results in the
accumulation of lower-chlorinated byproducts; and anaerobic degradation of nitroaromatic compounds
generally results in their conversion to amino-aromatic byproducts. Thus, the use of anaerobic treatment
for many of these contaminants is likely to require subsequent aerobic treatment. The use of dual
treatment processes results in increased cost and O&M requirements. While anaerobic/aerobic treatment
trains have been investigated in the laboratory, this approach has not yet been tried at full scale.
Another reason why anaerobic treatment technologies have not been widely used is that they often target
relatively recalcitrant contaminants. Because aerobic treatment has been so effective in the treatment of
petroleum hydrocarbon contaminants, much of the remediation efforts over the past decade have focussed
on aerobic treatment technologies and on removing these easily degraded contaminants. Furthermore,
compared to chlorinated solvents, PCBs, energetics, and metals, the suite of petroleum hydrocarbon
contaminants tend to be much more easily accessed and targeted for treatment. For example, LNAPLs
are much more easily treated than DNAPLs, which in some cases cannot even be characterized; PAH-
contaminated soils are more easily handled than soils contaminated with PCBs and dioxins; and most
petroleum hydrocarbons tend to be associated with point sources resulting from accidental releases,
compared to nonpoint sources for energetic compounds and heavy metals. In light of the success of
aerobic bioremediation of petroleum hydrocarbons, there is an increasing portion of R&D funding that is
targeting more recalcitrant compounds, which by default will include anaerobic treatment technologies
because of the numerous compounds that degrade more readily under anaerobic conditions. Thus, it is
increasingly likely that anaerobic biological treatment will be used in the field in the near future.
Research needs for anaerobic treatment technologies include the following:
1. Technology maturity: Extensive literature is available on anaerobic processes, but little has been done
to date to bring technologies to the marketplace. Thus, a broad area of R&D opportunities exists for
expanding and transferring concepts developed in the laboratory to pilot-scale and, ultimately,
full-scale use. For example, there is extensive knowledge of the ability of anaerobic cultures to
reductively dechlorinate chlorinated solvents, but reductive dechlorination in the vadose zone has
seen limited field use. The primary limitation of enhanced anaerobic dechlorination resides in the
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ability to introduce a suitable electron-donating compound to the environment in the area of
contamination particularly for in-situ soils.
2. Control and treatment of anaerobic treatment byproducts: The potential for byproducts to form using
anaerobic processes introduces the possibility of the production of undesirable contaminants, even
though the parent compound may be transformed microbially. Development of treatment methods to
manage byproducts formed during anaerobic treatment is needed. The most likely biological
treatment scenario is the sequential use of anaerobic and aerobic treatment.
3. Nutrient requirements: As with aerobic technologies, nutrient addition has the potential to enhance
contaminant removal rates using anaerobic treatment approaches, and at some sites may be a limiting
factor to contaminant degradation.
4. Toxicity measurements: Methods for measuring soil toxicity and reductions in soil toxicity due to
biological treatment are needed. This is especially true for anaerobic PCB treatment. Certain
congeners, especially higher chlorinated congeners and those that have dioxin-like characteristics,
may be much more toxic than other congeners that are less chlorinated and do not behave like dioxins
(NRC, 2001). For example, EPA's Integrated Risk Information System (IRIS) uses two different
cancer slope factors for highly-declorinated PCB mixtures and lower-chlorinated PCB mixtures. A
cancer slope factor at 2.0 (mg/kg-day)"1 is stipulated for all PCB mixtures except those congeners with
more than 4 chlorines, which comprise less than 0.5% of the total PCB mass (EPA/600/P-96/001F).
The dioxin-like characteristics are associated mostly with meta- and />ara-chlorinated PCBs and the
more highly chlorinated PCBs (four or more chlorines). Meta- and/>ara-chlorinated compounds are
the PAHs most susceptible to reductive dechlorination (NRC, 2001; Bedard & Quenger, 1985;
Quensen et al., 1998). Thus, dechlorination has the potential to significantly detoxify soils, even
though it does not reduce the molar concentration of PCBs in soils.
5. Metals stabilization: Anaerobic technologies may be used to stabilize certain metals in soils.
Stabilization occurs when these metals are transformed as hydroxide or sulfide precipitates under
anaerobic conditions. The long-term stability of these precipitates after the soils become aerobic is an
important factor in using anaerobic treatment for metals and warrants further investigation.
5.2.2 Phytoremediation
Phytoremediation is gaining increased use for soils remediation. Phytoremediation research needs
include the following:
1. Contaminant fate and transport: The fate and transport of contaminants during phytoremediation
plays an important role in its application. Contaminants may be biotransformed in the rhizosphere
due to increased biological activity. If plants transpire contaminants, they may be released into the
atmosphere unaltered or they may go through various levels of transformation by plant enzymes.
Alternatively, plants may simply act as a hydraulic barrier to rainwater infiltration and the dissolution
of contaminants.
2. Field implementation studies: Field studies of phytoremediation are needed to demonstrate this
technology at pilot or full scale. A variety of field demonstrations are in their infancy. As these
demonstrations mature, more information about the efficacy of phytoremediation will be forthcoming.
3. Types of plants that can be used for phytoremediation: Fundamental studies on phytoremediation
continue to be required to identify plants that can be used for this technology and contaminants that
are amenable to phytoremediation.
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5.2.3 Bioaugmentation
Bioaugmentation has been a potential tool for environmental engineers and scientists for a decade, but has
seen limited application. The primary obstacles to bioaugmentation that require further R&D include:
1. Bioaugmentation performance: For many sites, cultivation of indigenous bacteria is more cost
effective than growing and concentrating foreign bacteria for augmentation. Additional comparison
of the performance of augmented soils vs. unaugmented soils is needed to better understand the
potential improvements made by bioaugmentation.
2. Culture stability: The stability of externally cultivated bacteria is not well understood, particularly
when cultivated bacteria are introduced into foreign environments after augmentation.
3. Delivery methods: Delivery methods for bacteria are not well developed, particularly for in-situ
vadose zone applications. R&D to produce improved delivery mechanisms is needed to ensure
efficient delivery in the area of contamination.
5.2.4 Natural Attenuation
Natural attenuation may be a viable treatment technology for contaminated soils, but more research is
needed to demonstrate its effectiveness and develop methods to validate its use. Research needs include
the following:
1. Sorption/precipitation of metals: For metals, studies on irreversible sorption or precipitation are
needed to ensure the stability of metals in the environment.
2. Contaminant weathering: Weathering of organic contaminants should be investigated to identify the
fate and transport of these contaminants in soils.
3. Protocol development: The development of analytical tools to assess the efficacy of natural
attenuation is needed to establish a uniform, engineered approach. To date, an approach to assess
natural attenuation of soils has not been established.
5.3 FUNDAMENTAL RESEARCH AND DEVELOPMENT NEEDS FOR RECALCITRANT
WASTE STREAMS
A variety of relatively recalcitrant compounds and waste streams have proven difficult to biodegrade,
resulting in their persistence in the environment. More fundamental research is needed to develop
potential biological treatment approaches that can address these environmental concerns.
1. Dioxin biological treatment: Under reduced conditions, dioxins (PCDD or PCDF) have been
demonstrated to dechlorinate to potentially less harmful daughter products. Aerobically, the lesser-
chlorinated dioxins may be biodegraded to form chlorinated salicylates, catechols, or phenols, which
may require further anaerobic or aerobic biotransformation. Thus, the sequence of biological
reactions may be relatively complex for dioxins and requires further investigation.
2. Biotreatment of other relatively recalcitrant contaminants: Fundamental research continues to be
needed for the variety of relatively recalcitrant contaminants in the environment for which there are
few treatment alternatives. Among the list of recalcitrant contaminants are the dioxins, identified
above, PCBs, large-molecular-weight PAHs, chlorinated solvents and aromatic compounds, and
energetic (nitroaromatic) compounds.
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3. Bioavailability: In addition to the complexity of dioxin biotransformation pathways is the fact that
dioxins, while ubiquitous in the environment, usually exist at sub-parts-per-million concentrations.
Their low concentrations, low aqueous solubilities, and high sorption affinity make them relatively
unavailable (low bioavailability) for microbial degradation. Research is needed to develop
bioremediation approaches for low-concentration or low-bioavailable contaminant waste streams, and
to better understand the biological fate of these contaminants in the environment.
5.4 SUMMARY OF RESEARCH NEEDS
Biological treatment of soils has been remarkably successful over the last decade when applied
to low-molecular-weight petroleum compounds and other contaminants that are easily biodegraded.
Today, the environmental restoration field is at a critical juncture. The biological treatment of easily
degraded contaminants is relatively well understood and accepted, but a large number of contaminants
remain for which there are no readily available technologies and for which biological treatment remains
challenged.
It is helpful to remember that about a decade ago it was thought that aroclor mixtures were among the
most recalcitrant contaminants in the environment due to their low bioavailability and high degree of
chlorination (Adriaens et al., 1999). However, since the late 1980s, it has been well documented that
PCB toxicity can be reduced through reductive dechlorination and can be completely mineralized through
the combination of anaerobic and aerobic treatments. Reports of new and previously undocumented
biotransformation pathways for recalcitrant contaminants continue to appear in the literature. Examples
include recent reports of the anaerobic degradation of benzene and PAHs under sulfate-reducing
conditions (Coates et al., 1996, 1997), anaerobic oxidation of DCE and VC (Bradley and Chapelle, 1996,
1997), and the ability to stimulate anaerobic PCB dechlorination by the addition of surrogate
polybrominated biphenyl compounds to soils or sediments (Bedard et al., 1998) and the complete
dechlorination of PCBs (Bedard and van Dort, 1998). These studies not only have important implications
for the biodegradation of the specific contaminants on which they reported, but also indicate that new
biodegradation pathways and mechanisms continue to be discovered and provide an optimistic future for
the biodegradation of environmentally persistent contaminants.
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APPENDIX A: DATABASE REPORT SUMMARY TABLES
and
APPENDIX B: DATABASE REPORTS
for
APPLICATION, PERFORMANCE, AND COSTS OF BIOTREATMENT
TECHNOLOGIES FOR CONTAMINATED SOILS
on
CONTRACT NO. 68-C-00-185
TASK ORDER NO. 13
Submitted to
U.S. ENVIRONMENTAL PROTECTION AGENCY
NATIONAL RISK MANAGEMENT RESEARCH LABORATORY
CINCINNATI, OHIO
Jennifer Goetz
Project Officer
Richard C. Brenner
Task Order Manager
BATTELLE
505 King Avenue
Columbus, OH 43201-2693
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