Water Environment
Federation
Preserving & Enhancing
the Global Water Environment
IN-SITU
BIOREMEDIATION
OF
Contaminated Subsurface Media
Edited by
J. Kevin Rumery
jack w keeley
Barbara Wilson
John Wilson
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IN-SITU BIOREMEDIATION OF
CONTAMINATED SUBSURFACE MEDIA
Edited by
J. Kevin Rumery'
Jack W. Keeley1
Barbara Wilson1
John Wilson2
Robert S. Kerr Environmental Research Laboratory
U.S. Environmental Protection Agency
Ada, Oklahoma 74820
July 1993
1 Dynamac Corporation, Robert S. Kerr Environmental Research Laboratory,
Ada, Oklahoma
2 Robert S. Kerr Environmental Research Laboratory, Ada, Oklahoma
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ACKNOWLEDGEMENT
This paper is based on a report entitled, "In-Situ Bioremediation of Ground Water and Geological
Material: A Review of Technologies," by Robert D. Norris, Robert E. Hinchee, Richard Brown, Perry L.
McCarty, Lewis Semprini, John T. Wilson, Don H. Kampbell, Martin Reinhard, Edward J. Bouwer,
Robert C. Borden, Timothy M. Vogel, J. Michele Thomas and C.H. Ward. The full report has been
published by the U.S. EPA and is available from NTIS.
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CONTENTS
SECTION 1— INTRODUCTION I
1.1 Background I
1.2 Scope 1
SECTION 2—THE SUBSURFACE ENVIRONMENT 3
2.1 Geology/Hydrology 3
2.1.1 Porosity, Permeability, and Hydraulic Conductivity 4
2.2 Subsurface Heterogeneity 4
2.3 Contaminant Distribution 6
2.3.1 Solubility 6
2.3.2 Sorption 6
2.3.3 Volatility 7
2.3.4 NAPLs 7
2.4 Critical Site Limitations to Bioremediation 7
2.4.1 Hydraulic Conductivity 7
2.4.2 Subsurface Heterogeneity 8
2.4.3 NAPLS and Residual Saturation . 9
SECTION 3—THE BIOREMEDIATION CONCEPT ] 1
3.1 Fundamental Principles 11
3.2 Natural or "Intrinsic" Bioremediation 12
3.2.1 Using Natural Biodegradation for Site Remediation 12
3.2.2 Case Studies—Natural Bioremediation 14
3.3 Bioremediation: Enhancing Natural Biodegradation 15
3.3.1 Nutrients 16
3.3.2 Terminal Electron Acceptors 17
3.3.2.1 Oxygen, Hydrogen Peroxide 18
3.3.2.2 Alternate Electron Acceptors—Nitrate, Sulfate, Carbon Dioxide 19
3.3.2.2.1 Nitrate and Denitrifying Systems 21
3.3.2.2.2 Iron(III) Reducing Systems 21
3.3.2.2.3 Sulfate and Sulfate Reducing Systems 22
3.3.2.2.4 Fermentative/Carbon Dioxide Reducing Systems 22
3.3.2.2.5 Mixed Electron Acceptor Systems 22
3.3.3 Introduced Microorganisms 23
3.3.3.1 Microbial Transport 24
3.3.3.2 Factors Affecting Bacterial Transport and Survival in the Subsurface 24
3.3.3.2.1 Matrix 24
3.3.3.2.2 Properties of Microbes 25
3.3.3.2.3 Operational Factors ,..25
3.3.3.2.4 Survival in the Subsurface 25
3.3.4 Cometabolic Substrates 26
SECTION 4—TYPES OF IN SITU BIOREMEDIATION SYSTEMS 27
Delivery of Nutrients and Electron Acceptors .27
4.1 In-situ Bioremediation of Soil and Ground Water 27
4.2 Air Sparging 29
4.3 Bioventing 31
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SECTION 5—CONTAMINANT BIODEGRADABILITY 35
5.1 Laboratory Testing 35
5.2 Field Testing: Pilot and Full Scale 35
5.3 Petroleum Hydrocarbons 35
5.3.1 Fuels 36
5.3.2 PAHS (Coal Tar, Creosote, Refinery Wastes) 38
5.4 Chlorinated Solvents 38
5.4.1 Aerobic Biodegradation of CAHs 40
5.4.2 Anaerobic Biodegradation of CAHs 41
SECTION 6—MONITORING AND PERFORMANCE EVALUATION
FOR BIOREMEDIATION SYSTEMS 44
6.1 Sampling Programs 44
6.2 Indicators 45
6.3 Meeting Treatment Goals 45
6.4 Bioremediation Limits 46
6.4.1 Concentration 46
6.4.2 Metals 47
6.4.3 Mass Transfer 47
SECTION 7—STATE OF BIOREMEDIATION TECHNOLOGY 49
7.1 Natural Attenuation 49
7.2 In-situ Bioremediation 49
7.2.1 Anaerobic Bioremediation—Alternative Electron Acceptors 49
7.3 Air Sparging 51
7.4 Bioventing 51
SECTION 8—SUMMARY 53
REFERENCES 54
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SECTION 1
INTRODUCTION
1.1 Background
Due to high costs associated with excavation and incineration for treating hazardous wastes and
materials contaminated with those wastes, alternative innovative technologies have been sought. One
innovative technology, in-situ bioremediation of contaminants in the subsurface, has been the focus of
considerable research in the past decade. A variety of approaches have been developed, differing mainly
in the mechanism whereby essential nutrients and electron acceptors are delivered. Factors such as site
hydrogeology and contaminant nature and distribution often control the success of in-situ bioremediation
efforts.
The use of biooxidation for environmental purposes has been practiced for decades, with
biological processes used in wastewater treatment since the early 1900s. Activated sludge and fixed-film
growth systems are commonly used for treating municipal wastewater and industrial wastes. This use of
biological degradation of organic compounds has generated a wide body of information regarding
biodegradability of specific compounds and classes of chemicals, nutrient and electron acceptor
requirements, and oxidation mechanisms. Land treatment processes for municipal wastewater and
sludges, as well as petroleum refinery wastes, have also been practiced for several decades and have
generated additional information on nutrient requirements, degradation rates, and other critical parameters
affecting biological oxidation (Overcash and Pal, 1979).
In the 1970s, several studies sponsored by the American Petroleum Institute were conducted
using the method developed by Richard L. Raymond, then at Sun Tech.,to biologically degrade
hydrocarbons in aquifers (Bauman, 1991). This method involved the recovery of ground water, treatment
using an air stripper tower and subsequent reinjection following amendment with nitrogen and
phosphorus sources (Raymond et al., 1976). Many of these early tests were conducted prior to the
enactment of federal and state mandated clean-up levels. As a result, these tests demonstrated that in-situ
bioremediation could reduce the levels of petroleum hydrocarbons in an aquifer, but did not document an
ability to reach ground-water quality standards in today's regulatory environment.
In the mid 1980s, there were few companies with experience in bioremediation of aquifers or
soils. Since that time, many companies have utilized bioremediation technologies, although claims of
experience are frequently overstated. Acceptance of in-situ bioremediation as a remediation technology
by the public and various regulatory agencies has been generally favorable over the last seven or eight
years and has improved significantly over the last two or three years with the support of the U.S. EPA,
many state agencies, and favorable publicity in trade journals and the popular press. As an in-situ
technology that is viewed as a natural process resulting in destruction rather than relocation of
contaminants, in-situ bioremediation meets many of the objectives of state and federal agencies.
1.2 Scope
It is the intent of this report to provide the Teader with a detailed background of the technologies
available for in-situ bioremediation of contaminated soil and ground water. The document has been
prepared for scientists, consultants, regulatory personnel, and others who are associated in some way with
the restoration of soil and ground water at hazardous waste sites. The presentation provides the most
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recent scientific understanding of the processes involved with bioremediation of soil and ground-water, as
well as a definition of the state-of-the-art of these technologies with respect to circumstances of their
applicability and their limitations.
A number of bioremediation technologies are discussed, along with the biological processes
driving those technologies. In addition to discussions and examples of developed technologies, the report
also provides insights to emerging technologies which are at the research level of formation, ranging from
theoretical concepts, through bench scale inquiries, to limited field-scale investigations. Although a wide
range of contaminants are potentially biodegradable, bioremediation systems have been most successfully
applied to petroleum hydrocarbons (fuels and refinery wastes), wood preserving wastes (creosote), and
chlorinated solvents (TCE). Therefore, the major focus of this report is limited to those contaminant
groups.
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SECTION 2
THE SUBSURFACE ENVIRONMENT
The applicability and success of in-situ bioremediation processes are primarily determined by the
geology and hydrology of sites. The physical and chemical nature of contaminants, as well as their
distribution in the subsurface, are also critical in determining whether in-situ bioremediation is successful.
In addition, the nature and performance of existing microbial populations are vital to successful
application of bioremediation technologies. Therefore, an understanding of key subsurface elements
impacting biodegradation processes is presented in the following section.
The physical subsurface environment can be visualized as having two major compartments.
These include: (1) a stationary, solid phase; and (2) a fluid or transient phase comprised of liquids and/or
gas in the void spaces between the solids (Figure 1). The solid phase consists of inorganic materials such
as clay, silt, sand, and gravel; and organic materials such as humic and fulvic acids. The physical and
chemical properties of the solid phase control the transport and transformation characteristics of
contaminants in this environment, as well as our ability to apply remediation activities. The void, or pore,
spaces are the primary pathway for movement of the fluid phases and, therefore, are the pathway for
contaminant transport in the subsurface environment. These spaces may be filled with either air
(unsaturated) or water (saturated). If non-aqueous phase liquid (NAPL) contaminants (such as gasoline)
are present, the pore space may be occupied by a mixture of air, water, and NAPL. Thus, contaminants in
the subsurface may partition between four major compartments (Huling and Weaver, 1991): (1) air
phase—vapor in the pore spaces; (2) adsorbed phase—sorbed to subsurface solids; (3) aqueous phase—
dissolved in water; and (4) NAPL phase—non-aqueous liquids. These compartments are illustrated in
Figure 2.
2.1 Geology/Hydrology
The geology and hydrology of the subsurface at a contaminated site play critical roles in
determining contaminant distribution; further, they determine whether or not in-situ bioremediation is
feasible.
Mass Balance Conceptual
Framework for the Subsurface
Fluid Phase
Water
Gas
Organic
Inorganic
Solid Phase
Figure 1. Conceptual framework for major subsurface compartments (USEPA, 1992a).
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Water
Air
DNAPL
Figure 2. Subsurface phases into which contaminants may partition (Huting and Weaker, 1991).
2.1.1 Porosity, Permeability, and Hydraulic Conductivity
Two major parameters which describe the subsurface in terms of water and contaminant
transport are porosity and permeability. Porosity refers to the amount of pore, or void, space present in
a specific volume of subsurface material. This parameter indicates the amount of storage that is available
in soil or aquifer materials as a function of particle size and texture. The porosities of several subsurface
materials are given in Figure 3. For example, clay can hold more water and chemicals than gravel
because of its greater porosity. Permeability, on the other hand, indicates the relative ease with which
fluids move through subsurface material, including water, with the nutrients and oxygen required for
enhanced in-situ bioremediation. Permeability is a function of the pore size. Materials with smaller pore
sizes, although more porous, often exhibit low permeability. As shown in Figure 4, clay has much more
porosity than gravel, yet gravels can be orders of magnitude more permeable than clay. Hydraulic
conductivity is a commonly used hydrogeologic measure of permeability and is often a limiting factor in
applying in-situ bioremediation. Contaminated subsurface materials with high porosity (storage) and low
hydraulic conductivities are poor candidates for in-situ bioremediation.
2.2 Subsurface Heterogeneity
A frequently overlooked characteristic of subsurface environments is the inherent variability, or
heterogeneity, of various layers. During site characterizations, a 'conceptual model' of the subsurface is
developed, often based on relatively little data. These models often overlook the presence of silt or clay
'lenses' which may not have been detected during geological investigations. Since ground water and soil
vapors will follow the path of least resistance in response to force (hydraulic or pneumatic gradients),
areas in the subsurface with high permeability will become preferential flowpaths. Regions in the
subsurface with lowest permeabilities, such as clays and silts, will remain contaminated. Even geologic
layers differing only in grain sizes, such as a fine sand layer below a coarse sand layer, may have
significantly different permeability. These subsurface heterogeneities play critical roles in contaminant
transport as well as in-situ remediation.
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Porosity
Silt
Gravel
Sand
Clay
—, | 1 1 1 1
10% 20% 30% 40% 50% 60%
Figure 3. Porosity of subsurface materials (USEPA, 1992a).
Permeability
Clay
Silt
Sand
Gravel
¦ % 'S • s»s
P /I
MA'A'.S'.V.S'SlS'A
10
! , , 1 1 1
-10 J0-g ](^6 10-4 10-2 1Q0 ]02
cm/sec
Figure 4. Permeability of subsurface materials (USEPA, 1992a).
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2.3 Contaminant Distribution
It is important to understand where contaminants are distributed among the various subsurface
compartments shown in Figure 2. This phase partitioning of contaminants is dependent upon a number of
factors, including the physical/chemical nature of the contaminants as well as that of the subsurface
environment. This distribution is exemplified in Figure 5, where contaminants are shown to be
associated with the vapor phase in the unsaturated zone, a residual phase, or dissolved in ground water.
Four major characteristics determine which subsurface compartment a contaminant will be most
likely to partition into and, subsequently, to what extent in-situ bioremediation will be applicable to the
contaminant. These are discussed briefly below:
2.3.1 Solubility
Solubility in water plays a critical role in transport and biodegradation of contaminants. Those
contaminants which are very water soluble will partition to the aqueous compartment in the subsurface.
Contaminants in the water phase, both in bulk liquid moving through pores (saturated) and in water films
surrounding particles (unsaturated), are most exposed to action of subsurface microbial communities.
Because of this, more water soluble contaminants are generally more biodegradable.
2.3.2 Sorption
Contaminants may partition to the subsurface solid phase. This process is referred to as sorption.
Sorption results when the contaminant interacts with either natural organic matter, usually humic and/or
fulvic acids, associated with soils or aquifer materials, or directly with the mineral surface. The degree
and 'tightness,' or reversibility, of sorption depends on the chemical structure of both the contaminant and
the organic matter, as well as the amount of organic matter present. The overall effect of sorption is to
Water
Table
Sourc
Vapor Phase
Residual „
Saturation
Dissolved
Contaminants
Capillary
Fringe
Ground Water
Figure 5. Possible contaminant locations in the subsurface (shaded areas).
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retard the movement of contaminants in the subsurface. Contaminants which are tightly sorbed to
subsurface materials such as clays are often not bioavailable and are resistant to biodegradation.
Examples of such contaminants include high molecular weight poly aromatic hydrocarbons such as benzo
(a) pyrene.
2.3.3 Volatility
Contaminants which have high vapor pressures partition into the air occupying the pore spaces
between particles. Once in the gas phase, vapors may migrate in response to air pressure gradients or
gravity, depending on the density of the vapor relative to air.
2.3.4 NAPLs
As noted above many contaminants do not mix with water and exist as distinct non-aqueous
phase liquids (NAPLs) in the subsurface. These may be either lighter or heavier than water, thus resulting
in the terms LNAPL (light non-aqueous phase liquids) and DNAPL (dense non-aqueous phase liquids).
When NAPLs exist as a continuous body of immiscible phase, they migrate vertically as a bulk liquid,
leaving behind isolated globules of material trapped in the pore spaces by capillary forces. This residual
saturation phase remains as a significant continuous source of contamination lo ground water. For
example, gasoline trapped at residual saturation in an aquifer may occupy up to 50 percent of the pore
space. Components of the gasoline such as benzene, toluene, and xylene partition or "bleed" into the
transient water and vapor phases and therefore serve as long-term sources of contamination. If sufficient
NAPL exists as bulk liquid, it will migrate to the water table. LNAPLs such as petroleum hydrocarbon
fuels will spread out into a floating 'lens' on the water table. Fluctuations in the water table level may
spread or 'smear' the lens, resulting in a greater volume of contaminated subsurface material over time
(Figure 6). DNAPLs such as chlorinated solvents and creosote will continue to migrate downward
through the aquifer until a relatively impermeable layer is reached, leaving behind more isolated globules
as residual saturation. Free-phase product may collect as pools or ponds in depressions on top of these
impermeable layers. Clay or silt lenses often act as such impermeable layers, leading to unpredictable
horizontal migration if the surface is tilted. The DNAPL may then follow the slope of the surface until
the lens ends, after which it continues downward to the next impermeable barrier. Typical patterns of
DNAPL migration are shown in Figures 7 and 8. Finding and quantifying DNAPLs in the subsurface is
difficult, and remediation technologies are just emerging.
2.4 Critical Site Limitations to Bioremediation
2.4.1 Hydraulic Conductivity
As noted above, permeability of the subsurface describes the ease with which fluids, either water
or air, can be moved through contaminated soils and aquifers. Since water is often the primary
mechanism for introducing amendments such as nutrients during in-situ bioremediation, successful
application of the technologies depends on the ability to move water into and through the subsurface.
Therefore, ground-water flow rate and flow paths are critical to the design and performance of in-situ
bioremediation systems. The ground-water flow must be sufficient to deliver the required nutrients and
oxygen (or other electron acceptors) according to the demand of the organisms, and the amended ground
water should sweep the entire area requiring treatment. This is a critical point in that it is often the
hydraulic conductivity of the ground-water system itself or the variability of the aquifer materials which
limits the effectiveness of in-situ technologies or prevents its utility entirely. A suggested target for in-
situ remediation technologies is a hydraulic conductivity of at least 10 * cm/sec (100 ft/yr). Soils or
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Static Water Table
Falling Water Table
Tank
Rising Water Table
Residual hydrocarbons (above water table)
Free liquid hydrocarbons
Residual hydrocarbons (below water (able)
Source: Modified from Schwille, 1984.
Figure 6. Smearing of hydrocarbon lens during water table fluctuations (API, 1989).
Legend
~
¦
u
aquifers with hydraulic conductivities less than this value are poor candidates for in-situ bioremediation,
due to extremely slow delivery rates of electron acceptors and nutrients.
2.4.2 Subsurface Heterogeneity
Subsurface heterogeneity can seriously limit the effectiveness of in-situ bioremediation.
Discontinuous layers or lenses of clay and silt may serve as long term sources of contamination, due to
their high storage and low permeability. Preferential ground water flow paths often develop in response
to slight variations in permeability, resulting in only partial remediation of aquifers. Sites with a high
degree of geological complexity are therefore often poor candidates for in-situ bioremediation.
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DNAPL
Residual
Ground-Water
Flow
DNAPL
, Pool
Figure 7. DNAPL migration patterns in homogeneous aquifers (Huling and Weaver, 1991).
DNAPL
Residual
Dissolved
Contaminant*
Ground-Water
Flow
¦"* Residual
Saturation
ONAPL Pool
SAND
AQUIFER
CLAY
Figure 8. DNAPL migration patterns in heterogeneous aquifers (Huling and Weaver, 1991).
2.4.3 NAPLS and Residual Saturation
The presence of either LNAPL or DNAPL as free-phase product in the subsurface presents a
major limitation to in-situ bioremediation systems for several reasons. First, if pore spaces are completely
saturated with the non-aqueous phase, water is excluded. Therefore, ground water amended with
nutrients and essential electron acceptors cannot come into contact with the bulk of the contamination,
and bioremediation is limited. Second, although the rate of mass transfer of contaminants from the NAPL
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phase to the ground water may be relatively slow, high enough concentrations of contaminants can
develop in ground water near the NAPL to cause toxicity to microbes essential to the bioremediation
process. Last, even if the first two limitations could be circumvented, the nutrient and electron acceptor
requirements to biodegrade large quantities of contaminants would not be feasible. Therefore, every
effort needs to be made to remove as much NAPL as possible before attempting bioremediation.
NAPLs present at residual saturation represent an often unrecognized source of contaminants.
Historically, in-situ bioremediation addressed only contaminants dissolved in ground water; and many
sites have been declared "clean" based on reduction of concentrations in ground water. However, NAPLs
at residual saturation have much higher surface area in contact with ground water; and ground water can
move much more freely through subsurface zones where pore spaces are not completely saturated with
NAPL. Therefore, bioremediation systems should address not only the contaminant present in ground
water, but also this significant remaining source.
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SECTION 3
THE BIOREMEDIATION CONCEPT
3.1 Fundamental Principles
Biodegradation of contaminants by microbial populations occurs to some degree without human
intervention in most ecosystems. These natural biodegradation mechanisms—along with abiotic
processes such as hydrolysis, dispersion, volatilization, and sorption—are collectively know as "natural
attenuation." As illustrated in Figure 9, biodegradation of contaminants requires nutrients, especially
nitrogen and phosphorous. In addition, microbial degradation of contaminants requires a terminal
electron acceptor. Oxygen is the required electron acceptor in many populations of microbes degrading
contaminants, and such populations are referred to as aerobic. In contrast, there are significant
populations of microbes in the subsurface which utilize electron acceptors other than oxygen. Biological
processes which occur in the absence of oxygen are referred to as anaerobic. There is considerable
evidence that both of these general types of microbial metabolism are operative and potentially applicable
in many contaminated subsurface ecosystems.
Biological treatment, whether of excavated soils, aquifer solids, or unsaturated subsurface
materials, uses microorganisms to convert harmful chemical species to less harmful chemical species in
order to effect remediation of a site or a portion of a site. The microorganisms are generally bacteria but
can be fungi. The terms "natural attenuation," "natural bioremediation," and "passive bioremediation"
are used somewhat interchangeably to describe the use of unassisted natural biodegradation processes for
site remediation. Abiotic processes also contribute to this "intrinsic" attenuation capacity of the
subsurface. Intrinsic bioremediation is therefore distinctly different from, and should not be confused
with, technologies which actively adjust subsurface conditions to maximize biodegradation rates.
Organic Carbon \
(Contaminants of Concern)
Contaminant modified
but not completely
broken down
Substrate Uptake
Mechanisms
CELL
Enzyme Systems
Nutrients
Mineralization
Substrate
Transformation
Electron Acceptors
• Oxygen • Nitrogen Contaminant
' Nitrate • Phosphorus broken down to
• Sulfate • Trace Metals CO2 and H2O
'Others
Figure 9. Concept overview; major components of biodegradation.
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Engineered bioremediation refers to active processes designed to maximize biodegradation of
contaminants through a variety of mechanisms, most involving addition of essential nutrients and electron
acceptors. In-situ bioremediation refers to technologies designed to treat contaminated soils, ground
water, or aquifer materials in place, with a minimum of excavation and other site disturbances.
3.2 Natural or "Intrinsic" Bioremediation
Microbial populations are capable of adapting to and degrading contaminants, and indigenous
bacteria that can degrade a variety of organic compounds are present in nearly all subsurface materials
(Borden, in press). The ability of microorganisms to degrade a wide variety of hydrocarbons is well
known. In an early review, Zobell (1946) identified over 100 microbial species from 30 genera that could
degrade some type of hydrocarbon. Since then, numerous studies have shown that hydrocarbon-
degrading microorganisms are widespread in the environment and occur in fresh and salt water, soil, and
ground water. Litchfield and Clark (1973) analyzed ground-water samples from 12 different aquifers
throughout the United States that were contaminated with hydrocarbons. These workers found
hydrocarbon-utilizing bacteria in all samples at densities up to 1.0 x 106 cells per ml. After a gasoline
spill in Southern California, McKee et al. (1972) found 50,000 hydrocarbon degrading bacteria per ml or
higher in samples from wells containing traces of gasoline, while a noncontaminated well had only 200
organisms per ml. Jamison et al. (1975) reported naturally occurring biodegradation of high octane
gasoline in ground water. Research at sites contaminated with wood-preserving wastes (Lee and Ward,
1984; Wilson et al., 1985) demonstrated that an adapted population of creosote-degrading
microorganisms was present within the contaminated zone, but not in the uncontaminated regions, of the
aquifer. Other studies correlated creosote biodegradation with the availability of dissolved oxygen (Lee
and Ward, 1984). In a more recent study, Ridgeway et al. (1990) identified 309 gasoline-degrading
species of bacteria from a shallow coastal aquifer contaminated with unleaded gasoline.
Ongoing research has shown that an aquifer's intrinsic assimilative capacity depends on the
metabolic capabilities of the native microorganisms, the aquifer hydrogeology and geochemistry, and the
contaminants involved. In many aquifers, conditions will not be perfect for natural bioremediation; and
less than optimal biodegradation will occur. The extent of aerobic biodegradation is controlled by the
amount of contamination released, the rate of oxygen transfer into the subsurface, and the background
oxygen content of the aquifer. When large amounts of contamination enter the subsurface, it overwhelms
the capacity of an aquifer to assimilate them, and extensive contamination may persist for long times and
distances downgradient from sources. When hydrogeologic conditions, such as clay confining layers or
naturally-occurring organic deposits, reduce the rate of oxygen transfer into the subsurface, the
assimilative capacity of the aquifer may be lower. Anaerobic biodegradation may be inhibited by low pH,
low buffering capacity, absence of appropriate electron acceptors (nitrate, iron, etc.), or the presence of
small amounts of oxygen. Heterogeneous conditions within the aquifer may prevent mixing and allow a
portion of the plume to migrate rapidly. If this occurs, the extent of biodegradation may be less than
would be expected for more uniform conditions.
3.2.1 Using Natural Biodegradation for Site Remediation
As the cost of performing site remediation continues to increase, interest in using "intrinsic
attenuation" also has increased. Intrinsic bioremediation is capable of treating contaminants aerobically
in the vadose zone, and at the margins of plumes (Figure 10) where oxygen is not limiting due to
reoxygenation.
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Oxygenated-Uncontaminated
Ground Water
Flow ~
1 Aerobic Margin
Anaerobic Core
Aerobic Margin
Flow
Oxygenated-Uncontaminated
Ground Water
Figure 10. Plan view - hydrocarbon plume undergoing natural biodegradation (Borden, in press).
When it works, natural bioremediation is capable of completely containing a dissolved
hydrocarbon plume. While there are few well-documented cases where this has occurred, there is a great
deal of anecdotal evidence that suggests that natural bioremediation can be effective in containing
dissolved hydrocarbon plumes. Typically, greater than 90% of all underground tanks are used to store
gasoline and other petroleum fuels. Yet a study by the California Department of Health Services (Hadley
and Armstrong, 1991) found that by far the most common ground-water contaminants were chlorinated
solvents, not petroleum constituents. These results suggest that the petroleum contaminants are being
removed to below detection limits before reaching water supply wells. Some sites have shown that
anaerobic bioremediation processes also occur naturally and can significantly reduce contaminant
concentration on aquifer solids and in ground water. Benzene, toluene, ethylbenzene, and xylene can be
removed anaerobically in methanogenic or sulfate-reducing environments. Highly chlorinated solvents
can undergo reductive dechlorination in anaerobic environments, and evidence exists that this is occurring
in-situ at several sites.
While there are no truly typical sites, it may be helpful to consider a hypothetical site where a
small release of gasoline has occurred from an underground storage tank (Figure 11). Rainfall infiltrating
through the hydrocarbon contaminated soil will leach some of the more soluble components including
benzene, toluene, and xylenes. As the contaminated water migrates downward through the unsaturated
zone, a portion of the dissolved hydrocarbon may biodegrade. The extent of biodegradation will be
controlled by the size of the spill, the rate of downward movement, and the existence of appropriate
environmental conditions. Dissolved hydrocarbons that are not completely degraded in the unsaturated
zone will enter the saturated zone and be transported downgradient within the water table where they will
be degraded by native microorganisms to an extent limited by available oxygen or other subsurface
conditions. The contaminants that are not degraded will move downgradient under anaerobic conditions.
As the plume migrates, dispersion will mix the anaerobic water with oxygenated water at the plume
fringes. This is the region where most natural aerobic degradation occurs.
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Gas
Oxygen
Exchange
Clean GW
Anaerobic Core
. Aetohic Marquis. . .
Mixing
Aerobic - Unconummated Ground Water
Figure 11. Profile of natural biodegradation of a typical UST release (Borden, in press).
One of the major factors controlling the use of natural bioremediation is the acceptance of this
approach by regulators, environmental groups, and the public (Borden, in press). Although natural
unaided bioremediation imposes little or no costs other than monitoring and the time for natural processes
to proceed, at sites where this approach is strongly opposed the cost may actually be higher than
conventional technologies. For example, in some states responsible parties could be allowed to request a
reclassification of contaminated ground water to a nonwater supply use. Although in such cases the
responsible party would not be required to actively remediate the site, no one has ever filed such a request
due to the perception that legal, administrative, and site characterization costs would be excessive and the
probability for success would be low. Implementing a natural bioremediation system differs from
conventional techniques in that a small portion of the aquifer is allowed to remain contaminated. This
results in the necessity of obtaining a variance from existing regulations. Consequently, some type of risk
evaluation will usually be required when natural bioremediation is considered.
Another barrier to acceptance is that, currently, there are no reliable methods for predicting the
effectiveness of natural bioremediation without first conducting extensive field testing (Borden, in press).
This is often the primary reason why this alternative is not seriously considered when evaluating remedial
alternatives. Without some reasonable assurances of success, responsible parties are not willing to risk
the large sums of money required for legal, administrative, and site characterizations costs.
32.2 Case Studies—Natural Bioremediation
One of the earliest studies of natural bioremediation was conducted at the United Creosoting
Company site in Conroe, Texas, by a team of researchers from the Robert S. Kerr Environmental
Research Laboratory (U.S. EPA) and the National Center for Ground Water Research. Early work (Lee
and Ward, 1984; Wilson et al., 1985) demonstrated that an adapted population of creosote degrading
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microorganisms was present within the contaminated zone but not in the uncontaminated regions of the
aquifer. Later studies correlated creosote biodegradation with the availability of dissolved oxygen (Lee
and Ward, 1984). These results were used to develop and calibrate the computer model, BIOPLUME, to
simulate hydrocarbon transport and aerobic biodegradation within the aquifer (Borden and Bedient, 1986;
Borden et al., 1986). Model results indicated that removal of the contaminant source would be sufficient
to contain the hydrocarbon plume and that active remediation by pump and treat would not be required.
Microbiologists from the U.S. Geological Survey have studied two different creosote
contaminated aquifers where anaerobic degradation of organic compounds has been observed. Field
studies at a contaminated aquifer in St. Louis Park, Minnesota, showed that methane production was
occurring in zones within the aquifer that had been contaminated with creosote (Godsy et al., 1983).
Later studies demonstrated that the presence of anaerobes (denitrifiers, iron reducers, sulfate reducers and
methanogens) was highly correlated with the presence of creosote. More recent work at an abandoned
creosote plant in Pensacola, Florida, has shown that a wide variety of organic compounds present in the
aquifer were undergoing methanogenic biodegradation and that transport distances in the aquifer could be
correlated with biodegradation rates observed in laboratory microcosms (Troutman et al., 1984; Goerlitz
et al., 1985).
Monitoring at petroleum contamination sites suggests that biotransformation of petroleum-related
compounds under methanogenic conditions may be more common than has generally been assumed.
Ehrlich et al. (1985) observed elevated numbers of sulfate-reducing and methanogenic bacteria in a jet
fuel contaminated aquifer. Evans and Thompson (1986) and Marrin (1987) monitored methane
concentrations in soil gas to map subsurface hydrocarbon contamination. In a study of soil gas
concentrations near underground storage tanks, Payne and Durgin (1988) found elevated methane
concentrations at over 20% of the 36 sites surveyed. Methane gas production can be so rapid that safety
problems occur at some sites. Hayman et al. (1988) had to develop a special apparatus to remove the
large quantities of methane generated from a fuel spill at the Miami, Florida, airport.
Hult (1987b) observed the production of large volumes of methane in the unsaturated zone
immediately below a crude oil spill at the U.S. Geological Survey research site in Bemidji, Minnesota. At
this same site, Eganhouse et al. (1987) observed a two order of magnitude decrease in alkylbenzene
concentration over a downgradient travel distance of 150 m. This decrease was accompanied by elevated
concentrations of aliphatic and aromatic acids in the ground water (Baedecker et al., 1987). The acids
included benzoic, methylbenzoic, trimethylbenzoic, toluic, cyclohexanoic, and dimethylcyclohexanoic.
These are the same acids identified by Grbit-Galit and Vogel (1987) as intermediates in anaerobic
degradation of alkylbenzenes. Ground-water and sediment analyses demonstrated that methanogenic
biodegradation caused a drop in pH and a rise in bicarbonate concentrations in the ground water. The
actual drop in ground-water pH appears to have been limited by dissolution of carbonate minerals (and
possibly aluminosilicates) (Siegel, 1987).
3 J Bioremediation: Enhancing Natural Biodegradation
The primary factors limiting natural biodegradation in the subsurface are most often a lack of
essential nutrients (i.e., nitrogen, phosphorous), appropriate electron acceptors (oxygen, nitrate, others),
and appropriate environmental conditions (pH, redox potential). Most approaches to engineered
bioremediation involve different ways to deliver these materials to the contaminated subsurface, where
existing microbial populations can then utilize them while degrading contaminants. For example; in-situ
bioremediation refers to treatment of soils or aquifer materials that are left in place. Air sparging,
bioventing, and ground-water recirculation with addition of nutrients and oxygen or are all
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examples of in-situ bioremediation. The major differences are in the approach taken to deliver electron
acceptors, and the subsurface location targeted for that delivery. Air sparging injects air into
contaminated aquifers just below the water table. Bioventing utilizes soil vacuum extraction techniques
to sweep air through contaminated soils above the water table. Ground-water recirculation targets the
saturated zone, and uses water as the carrier for nutrients and electron acceptors. In land treatment,
nutrients can be mixed with soils during tilling or added with irrigation water and oxygen is introduced
during the tilling process.
Considerable interest exists in the use of introduced microorganisms in bioremediation.
Microbes used in such "bioaugmentation" may come from a variety of sources; these include isolation of
specific contaminant degraders from sites, laboratory selection of strains with superior degrading
capabilities, or production of genetically engineered microbes (GEMs). Use of these exogenous, or
outside, organisms to inoculate the subsurface is generally not considered to be necessary or successful
for most in situ bioremediation efforts. However, in situations such as bioreactors or fresh spills of
contaminants, distinct applications for use of bioaugmentation; and many advances are being made in
biotechnology for degradation of industrial wastes under those circumstances. Use of GEMs is currently
controversial, and releases of genetically engineered strains to the environment are strictly regulated. The
following section provides information on nutrient, electron acceptors, introduced microorganisms, and
other amendments (cometabolic substrates) which are used for enhanced bioremediation.
3.3.1 Nutrients
While a variety of other minerals such as iron, magnesium, and sulfur are required by
microorganisms, the primary elements which are necessary for microbial growth include carbon (C),
hydrogen (H), oxygen (O), nitrogen (N), and phosphorous (P). The basic premise of contaminant
biodegradation is that microbes obtain carbon and energy from organic contaminants and, in the process,
convert or transform those materials to simpler compounds. The carbon and energy taken from
contaminants are used during growth, along with other elements, to make new cell components.
Adequate amounts of nitrogen, phosphorous, and electrons are necessary for microbial metabolism and
growth to effectively degrade contaminants. These nutrients must be available to the microbes in: (Da
useable form; (2) appropriate concentrations; and (3) proper ratios (Dragun, J., 1988). Nitrogen and
phosphorous are generally not present to a great degree in most contaminants, especially petroleum
hydrocarbons primarily composed of carbon and hydrogen. In addition, these elements, although present,
may not be plentiful in subsurface soils and aquifers. The other minerals are needed in trace amounts, and
adequate amounts are normally found in most ground waters and subsurface materials. On the other
hand, there is a considerable surplus of carbon in contaminated subsurface environments. Therefore,
biodegradation of contaminants only proceeds until available N and P supplies are depleted.
Several approaches have been used to supply nutrients. Commercial agricultural fertilizers have
been applied in solid form or mixed with irrigation water. Types of fertilizers which have been successful
include urea-phosphate, N-P-K mixtures, and ammonium and phosphate salts (Atlas, 1984). Raymond et
al. (1975, 1978) mixed a blend of approximately equal amounts of ammonium chloride and sodium
orthophosphate with ground water which was injected into the contaminated zone of an aquifer. Some
practitioners have begun using potassium salts to reduce swelling potential in clays and to
tripolyphosphates which will solubilize rather than precipitate iron, calcium, and magnesium (Brown and
Norris, 1988). Oleophilic fertilizers, which dissolve into petroleum hydrocarbons, have also been used to
stimulate biodegradation of crude oil on beaches in Prince William Sound following the Exxon Valdez oil
spill (Pritchard and Costa, 1991).
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A variety of C:N:P ratios have been used for bioremediation applications. Atlas and Bartha
(1973) reported optimum biodegradation rates of oil in seawater using an oleophilic fertilizer with a
C:N:P ratio of 100:10:1. For aerobic treatment, optimal concentrations of ammonia or nitrate nitrogen are
in the range of 2 to 8 pounds per 100 pounds of organic material, while inorganic phosphorus
requirements are about one-fifth of this (McCarty, 1988). C:N:P ratios of 120:10:1 have been suggested
for bioremediation of contaminated soils (USEPA,1989).
These ratios and others found in the literature should only be used as general guidelines. Specific
nutrient requirements are not easily predicted since a number of factors other than amount of contaminant
may affect the amount of nutrients needed. Sorption, precipitation, and ion exchange of nutrients by
geologic materials can substantially increase the amount of nutrients that have to be introduced in order to
distribute nutrients across the contaminated zone. Adsorption may be modest in clean sands but may
consume most of the nutrients in silts and clays, especially if the solids have a high natural organic
content. If all of the hydrocarbon mass were converted to cell material, the nutrient requirements based
on the mass of hydrocarbon to be consumed would be approximated by a ratio of carbon to nitrogen to
phosphorous of 100:10:1. However, the literature reports significant variations in this ratio, depending on
site-specific conditions such as type of contaminant, soil type, microbial population structure, and which
metabolic pathway the microbes use to biodegrade the contaminant. For instance, if not all of the
hydrocarbon is converted to cell material, but is mineralized to carbon dioxide and water, nutrient
requirements may be less than this ratio.
Several bioremediation companies market proprietary nutrient mixtures as part of their services.
However, these are not necessary in order to provide nitrogen and phosphorous. In fact, care should be
exercised in application of nutrient mixtures with "unknown" ingredients which may have unexpected
interactions in the subsurface. Undesired effects include swelling of clays and precipitation of calcium,
iron, and magnesium which are all important nutrients.
Two approaches have been used to select appropriate nutrient application rates: (1) addition of
nutrients based on the amount of organic carbon present to achieve arbitrarily selected C:N:P ratios and,
(2) experimental determination of most effective nutrient mixtures during site-specific treatability studies.
The first approach relies on experience obtained under varying conditions and information from published
accounts. Although this sort of information is an adequate starting point, the second approach generally
avoids surprises caused by site-specific variations in nutrient demands.
3.3.2 Terminal Electron Acceptors
Biodegradation is essentially a series of oxidation-reduction reactions where the contaminant is
oxidized (donates electrons) or reduced (accepts electrons). In the process, the contaminant is
transformed to simpler molecules and the amount of potential energy in the contaminant is decreased
(Figure 12). A variety of compounds act as electron acceptors in metabolic pathways. The most critical
of these to bioremediation processes are the terminal electron acceptors (TEA), which are usually
obtained from the environment outside the cell.
Microbes in nature utilize a variety of compounds as terminal electron acceptors; including
oxygen (O ), nitrate (N03), iron and manganese oxides (e.g., Fe(OH)3, MnO^), sulfate (SOp, and carbon
dioxide (C02). Which particular compound is used depends on redox conditions and the type of bacteria
present. Aerobic bacteria can only use molecular oxygen (Cy, while anaerobic bacteria can use other
compounds such as NCP, SO^", Fe(OH)jf or CCK Some contaminants are only transformed under
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Redox Couples in Biodegradation
AH2
Electron
Acceptor
B
Substrate
Potential
Energy Level
Substitute
(Electron Donor)
A
oxidized
B
reduced
A
BH2
r
Figure 12. Redox couples in biodegradation.
aerobic conditions, while others require strongly reducing anaerobic conditions, and still others are
transformed in both aerobic and anaerobic environments.
Microorganisms preferentially utilize electron acceptors that provide the maximum free energy
during respiration. In aquifers contaminated with biodegradable organic compounds, electron acceptors
tend to be used successively in order of decreasing oxidation-reduction (redox) potential and free energy
yield. Oxygen is the most preferred electron acceptor because it has the highest redox potential and
provides the most free energy to microorganisms during electron transfer (Figure 13). The redox
potentials of nitrate, Mn(IV) and Fe(III) oxides (MnO and FeOOH, respectively), sulfate, and carbon
dioxide are lower. As a result, they yield less energy during substrate oxidation and electron transfer
according to the order listed in Figure 13. This sequence applies to pH 7 and should be valid for most
field conditions where the appropriate microorganisms occur.
The importance of microbial reactions involving Mn(IV) and Fe(III) to organic contaminant
biotransformations is unknown. Sulfate and carbon dioxide are the least preferred because
microorganisms gain the least energy from these reactions. However, these latter compounds comprise
the alternate electron acceptors available for development of anaerobic bioremediation technologies.
33.2.1 Oxveen. Hydrogen Peroxide
The most common bioremediation approach is based on aerobic processes. As currently
practiced, conventional in-situ biorestoration of petroleum-contaminated soils, aquifer solids, and ground
water relies on the supply of oxygen to the subsurface to enhance natural aerobic processes to remediate
the contaminants. It has been recognized that the rate at which oxygen can be introduced by sparging air
in a ground-water injection well limits the effectiveness of the technology. Since the amount of oxygen
which can be added to water from air is limited (8-10 mg/1), other sources of oxygen have been used.
These include pure oxygen and hydrogen peroxide. Use of pure oxygen in place of air can increase the
rate of introduction of oxygen five-fold.
Hydrogen peroxide is commonly used as a method of introducing oxygen (Brown et al., 1984).
This liquid, which decomposes to oxygen and water, is completely soluble in water. Because of this
using hydrogen peroxide can, theoretically, provide oxygen 5-50 times faster than could sparging air or
pure oxygen into injection wells and should result in shorter remediation times. However, the efficiency
of delivering oxygen by this method has been quite variable even when favorable results were obtained
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Aerobic
(Oxygen as r
Electron Acceptor) *
> 0.5
Typical Primary f
Substrates { ~
(Electron Donors)
-0.5
02 + 4H+ + 4e
2NO"3 + 12H+ + 10e
2H20
N2 + 6H20
Mn02(s) + HCOj + 3H+ + 2e"
MnC03(s) + 2H20
y FeOOH(s) + HCO'3 + 2H+ + e* -4
FeC03(s) +2H20
SOI * 9H+ + 8e* -* HS" + 4H2°
^ C02 + 8H+ + 86- -» CH4 + 2H20 V
2C02 + 8H+ + 8e* -» CH3COOH + 2H20
^ 2H++ 2e* H,
Figure 13. Redox potential and energy yields of electron acceptors and donors important in
bioremediation (Bouwer, in press).
from laboratory screening tests (Lawes, 1991; Huling et al„ 1990; Hinchee and Downey, 1988; and
Flathman et al., 1991). Hydrogen peroxide breaks down rapidly in contact with soils and organic
contaminants. In aquifers, the released oxygen may migrate rapidly out of the saturated zone, resulting in
poor delivery of oxygen to contaminated zones. Further, as microbial populations decrease as a function
of decreasing food source (the contaminants), tolerance toward hydrogen peroxide may also decrease.
Practical considerations, including toxicity towards bacteria and precipitation of iron and phosphate, limit
hydrogen peroxide concentrations to 100 to l,000ppm. In addition, too much oxygen may stimulate
biofouling of well screens and aquifers. As a result, hydrogen peroxide may not be the most appropriate
oxygen source for many sites.
3.3.2.2 Alternate Electron Acceptors—Nitrate. Sulfate. Carbon Dioxide
Rapid aerobic degradation requires an ample supply of nutrients and oxygen, good mixing, and a
high microbial mass. These conditions are often difficult to maintain in aquifers (Wilson, B. et al., 1986;
Lee et al., 1988). Water is a poor mass transfer medium for 02 due to the low water solubility of CK
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Because of this, degradation of relatively small amounts of hydrocarbons requires that large amounts of
water come in contact with the aquifer solids. The complete oxidation of 1 mg of hydrocarbon
compounds requires 3.1 mg of O (Hutchins and Wilson, 1991). Thus, for the bioremediation of 1 kg of
aquifer material containing 10 g/icg hydrocarbon compounds, a minimum of 3.1 m3 of oxygenated water
containing 10 mg/'l 02 must be supplied. Furthermore, at many sites there may be a very high abiotic
oxygen demand due to hydrogen sulfide (H2S), iron (Fe2+), or other readily oxidizable compounds. This
may make it difficult to increase the reduction potential into the aerobic range (> + 0.82 volts).
Therefore, anaerobic conditions are expected to persist within aerobically treated aquifers, especially in
relatively impermeable zones and zones farther away from the injection wells.
Bioremediation using electron acceptors other than oxygen is potentially advantageous for
overcoming this difficulty in supplying oxygen for aerobic processes. When oxygen is consumed faster
than it can be supplied, anaerobic microorganisms may grow using alternate electron acceptors. Nitrate,
sulfate, and carbon dioxide are attractive alternatives to oxygen because they are more soluble in water,
inexpensive, and nontoxic to microorganisms. Anaerobic microbial processes can be significant in
oxygen-depleted subsurface environments that are contaminated with petroleum-based compounds and/or
chlorinated solvents. The dark shaded area in Figure 14 illustrates the location of contaminants that may
be remediated by introduction of alternate electron acceptors. In the absence of molecular oxygen,
microbial reduction reactions involving organic contaminants increase in significance as environmental
conditions become more reducing. In this environment, some contaminants are reduced by a biological
process known as reductive dehalogenation. In reductive dehalogenation reactions, the halogenated
compound becomes the electron acceptor. In this process, a halogen is removed and is replaced with a
hydrogen atom. Reductive dehalogenation is a significant anaerobic degradation process which has been
shown to work on such recalcitrant compounds as polychlorinated biphenyls (PCBs), organochlorine
pesticides (DDT, toxaphene), and chlorinated solvents (PCE.TCE).
Anaerobic degradation of aromatic hydrocarbons was initially identified at field sites (Reinhard et
al., 1984) and in microcosm studies (Wilson et al., 1987) and has now been demonstrated in the
Water
Table
Source
Vapor Phase.
T
Residual
Saturation
Ground Water
Capillary
Fringe
Dissolved
Contaminants
Figure 14. Contaminant locations treatable with alternate electron acceptors.
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laboratory using a number of redox conditions and electron acceptors, including reduction of nitrate,
iron(III) and manganese(IV) oxides, sulfate, and carbon dioxide. In contrast to aromatic hydrocarbons,
aliphatic hydrocarbon degradation without oxygen has not been reported. The feasibility of alternate
electron acceptors other than nitrate for bioremediation has not been documented at field scale but has
been widely studied at laboratory scale. The other possible alternate electron acceptors (i.e., iron(III),
sulfate, and COp have been found in systems that may be classified as passive bioremediation, such as in
landfill leachate plumes (Reinhard et al., 1984) and at spill sites (Lovley et al., 1989). Nitrate has been
used as an electron acceptor for bioremediation of benzene, toluene, ethylbenzene, and xylenes in ground
water and on aquifer solids.
3.3.2.2.1 Nitrate and Denitrifying Systems
Microbial populations capable of using nitrate as an electron acceptor are widespread in the
environment. Denltrification results when bacteria utilize nitrate and convert it to N2, which then leaves
the subsurface system as a gas. Nitrate salts are much more water soluble (92 g/1 as sodium nitrate) than
02 (10 mg/1). Approximately 50 times more reducing equivalent can be introduced into an aquifer using
saturated sodium nitrate solution rather than a saturated oxygen solution. Under denitrifying conditions,
oxidation of monoaromatic compounds has been demonstrated in a number of systems (e.g., Kuhn et al.,
1988; Mihelcic and Luthy, 1988; Altenschmidt and Fuchs, 1991; Ball et al., 1991; Evans et al., 1991a;
Evans et al., 1991b; Flyvbjerg et al., 1991; Hutchins et al., 1991a; Evans et al., 1992).
When biodegradation of benzene, toluene, ethylbenzene and xylenes (BTEX) mixtures was
tested under denitrifying conditions, degradation tended to be sequential, with toluene being the first
substrate to be degraded, followed by p- and m-xylene, ethylbenzene and o-xylene. Benzene does not
seem to be degraded (Kuhnet al., 1988; Evans et al., 1991a; Evans etal., 1991b; Hutchins et ai., 1991a)
although in one study Major et al. (1988) reported removal under conditions thought to be denitrifying.
Hutchins et al. (1991a) reported longer lag times and slower degradation rates in core material
contaminated with JP-4 aviation fuel than in uncontaminated core material. Using an enrichment culture
and ethylbenzene as the substrate, Ball et al. (1991) have shown that single aromatic substrates can be
degraded rapidly (within hours) and that nitrate reduction to nitrogen gas proceeds through nitrite.
Similar findings were reported by Evans et al. (1991a, 1991b) for toluene. Ball et al. (1991) also
demonstrated that composition and preparation of the growth medium can affect the observed
transformation rates. Degradation of PAHs (naphthalene) under denitrifying conditions has also been
reported (Mihelcic and Luthy, 1988).
3.3.2.2.2 Iron(UI) Reducing Systems
Once available oxygen and nitrate are depleted, subsurface microorganisms may use oxidized
ferric iron [Fe(III)] as an electron acceptor. Microorganisms have been identified that can couple the
reduction of ferric iron with the oxidation of aromatic compounds including toluene, phenol, p-cresol and
benzoate (Lovley and Lonergan, 1990; Lovley et al., 1989). Large amounts of ferric iron are present in
the sediments of most aquifers and could potentially provide a large reservoir of electron acceptor for
hydrocarbon biodegradation. This iron may be present in both crystalline and amorphous forms. The
forms that are most easily reduced are amorphous and poorly crystalline Fe(III) hydroxides, Fe(III)
oxyhydroxides, and Fe(III) oxides (Lovley, 1991).
Lovley and Lonergan (1990) have isolated an iron-reducing bacterium capable of degrading
toluene, /vcresol and phenol. Relative to other anaerobic processes, Fe(III) reduction has a very
unfavorable substrate to electron acceptor ratio. Transport of the dissolved iron Fe(II) from the aquifer
21
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could cause secondary problems such as clogging and fouling of the aquifer. Furthermore, the supply of
large amounts of colloidal iron(III)oxide or soluble Fe(III)citrate (Lovley et al., 1989) to an aquifer has
not been tested. To develop bioremediation strategies based on iron reduction, a better understanding of
occurrence, nutritional requirements, growth conditions and metabolism of iron-reducing bacteria must be
developed.
3.3.2.2.3 Sulfate and Sulfate Reducing Systems
Bioremediation using sulfate as the electron acceptor involves oxidation of aromatic
hydrocarbons by sulfidogenic organisms coupled with reduction of sulfate to hydrogen sulfide (Edwards
et al., 1991; Haag et al., 1991; Beller et al., 1992; Edwards et al., 1992). As in some denitrifying systems,
degradation under sulfate reducing conditions is also sequential; with toluene being the preferred
substrate, followed by />-xylene and with o-xylene degraded last (Edwards et al., 1991,1992).
Ethylbenzene and benzene were not degraded under the conditions of the experiment. In a follow-up
study, Edwards and Grbifc-Galic (1992) observed benzene degradation in the absence of all other aromatic
substrates. After a lag time of 30 days under strictly anaerobic conditions, these authors observed
mineralization of benzene and suspected sulfate to be the electron acceptor. Accumulation of HS' may
inhibit the process, however, and is a problem that remains to be resolved.
3.3.2.2.4 Fermentative!Carbon Dioxide Reducing Systems
Under very reduced conditions, anaerobic organisms utilize carbon dioxide as an electron
acceptor and produce methane (CH4). These conditions are therefore known as methanogenic. Under
methanogenic/fermentative conditions, several aromatic hydrocarbon compounds, including benzene and
toluene, have been shown to transform into C02 and methane (Grbi6-Gali£ and Vogel, 1987).
Biotransformation under these conditions was studied with toluene or benzene as the only carbon source.
Biotransformation began after a three-month lag time and was complete after 60 days of incubation.
Since this ground-breaking study, several other aromatic substrates have been shown to be degraded
under methanogenic conditions, including styrene, naphthalene and acenaphthalene (Grbifc-Galic, 1990),
as well as benzothiophene, a sulfur-containing heterocyclic compound (Godsy and Grbit-Galic, 1989).
Fermentation/methanogenic degradation could be used as a passive bioremediation technology
and is likely to be an ongoing process at many sites where the geochemical conditions have evolved
naturally, without human intervention. Reliable assessment of the process is difficult under field
conditions since mass balances are difficult to establish. Indications for the process are the occurrence of
methane in combination with characteristic intermediates such as aromatic acids (Reinhard et al., 1984;
Wilson et al., 1987; Baedecker and Cozzarelli, 1991).
3.3.2.2.5 Mixed Electron Acceptor Systems
Few laboratory studies have examined mixed electron acceptor systems, although they are
likely to be common at field sites. Terminal electron acceptors such as sulfate and carbon dioxide are
likely to co-occur naturally, either within the same aquifer compartment or spatially separated into
adjacent compartments. For instance, at the sites where denitrifying conditions were investigated, O was
frequently present in the nitrate feed water. Both electron acceptors were consumed, but the effect of' the
oxygen on the overall process was not determined. Werner (1985) proposed that if and nitrate are
present simultaneously, is used for the first oxidation step to produce partially oxygenated products
and nitrate is then used for mineralization of the oxidation products.
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Different electron acceptors and products of aromatic degradation processes can react with each
other in a number of biological and chemical reactions. Beller et al. (1992) have studied the link between
sulfate reduction to sulfide and iron(III) reduction to iron(II) by a sulfate-reducing enrichment culture.
Ferric iron appeared to reoxidize hydrogen sulfide in an abiotic process and/or lower the inhibitory effect
of hydrogen sulfide. Toluene was the sole carbon and energy source, but other substrates were not tested.
Anaerobic bioremediation has been tested only in a very few cases and is still considered
experimental. For instance, in a review of 17 sites contaminated with hydrocarbon fuels and oils (Staps,
1990), hydrogen peroxide was used as the electron acceptor at seven sites, air at five, combinations of
nitrate-ozone and nitrate-air at one site each, and nitrate alone was used only at three sites. Much
available information has been developed in laboratory studies; however, the applicability of these results
to field conditions remains to be studied. Anaerobic transformation rates can be slow and lag times long
and unpredictable, except for transformation in denitrifying systems which can be fast. In spite of slow
rates, anaerobic bioremediation could play a significant role in the future mainly because the principal
factor limiting aerobic bioremediation, the difficulty of supplying oxygen to the subsurface, is
circumvented.
The combination of an anaerobic process followed by an aerobic process has promise for the
bioremediation of highly chlorinated organic contaminants. Generally, anaerobic microorganisms reduce
the number of chlorines on a chlorinated compound via reductive dechlorination. Aerobic
microorganisms are more capable of transforming compounds with fewer chlorinated substitutes. With
the removal of chlorines, oxidation becomes more favorable than does reductive dechlorination.
Therefore, the combination of anaerobic and aerobic processes has potential as a control technology for
chlorinated solvent contamination.
3.3.3 Introduced Microorganisms
Inocula of microorganisms have been widely used for bioremediation of hazardous waste sites.
However, there is little documentation of the efficacy of this process; and important questions still persist
about the environmental responsibility of adding nonindigenous, exogenous microorganisms.
Microorganisms have been added to samples of soil and water in the laboratory and field to enhance
biodegradation of hydrocarbons; however, the results of these studies have been mixed. Atlas (1977)
stated in a review on stimulated petroleum biodegradation that seeding will not be necessary in most
environments because of the ubiquity of hydrocarbon-degrading organisms. Although hydrocarbon-
degrading organisms may be ubiquitous, the problem with natural bioremediation of these compounds is
that the rate of biodegradation is often too slow {Thomas and Ward, in press). Nutrient addition and
agents that render the compounds more bioavailable may enhance these rates. However, inoculation may
be important in environments in which the population of hydrocarbon-degrading organisms is too low or
absent, or the environment is too harsh. In the latter case, the added organisms must be able to tolerate
the extreme conditions. In addition, inoculation may be beneficial in the biodegradation of the high-
molecular-weight polycyclic aromatic hydrocarbons, which are recalcitrant (Bossert and Bartha, 1986). If
seeding is considered as a method for hydrocarbon remediation, a mixture of microorganisms will be
required. Zajic and Daugulis (1975) found that multiple species were required to degrade the complex
composition of crude oil. Ball et al. (1991) tested inocula from different sources for the potential to
degrade BTEX compounds. They found that microorganisms with the ability to degrade aromatic
hydrocarbons are not ubiquitous. Sewage seed that contains a diverse population of microorganisms, for
instance, did not adapt to the aromatic compounds tested.
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Operations in which seed organisms are added lo enhance contaminant biodegradation in the
subsurface usually involve treating contaminated ground water in an aboveground bioreactor, and then
reinjecting the treated water into the subsurface. The treated ground water that is reinjected contains
adapted microorganisms from the bioreactor or is amended with contaminant-degrading organisms to
enhance in-situ biodegradation (Ohneck and Gardner, 1982; Quince and Gardner, 1982a, b; Winegardner
and Quince, 1984; Flathman and Githens, 1985; Flathman and Caplan, 1985, 1986; Flathman et al., 1985;
Quince et al., 1985).
3.3 J. 1 Microbial Transport
Seed microorganisms have been added to the subsurface to aid in contaminant biodegradation;
however, the role of the added microorganisms has never been differentiated from that of the indigenous
microflora (Lee et al., 1988; Thomas and Ward, 1989). For added organisms to be effective in
contaminant degradation, they must be transported to the zone of contamination, attach to the subsurface
matrix, survive, grow, and retain their degradative capabilities. There are a number of phenomena which
affect the transport of microbes in the subsurface including grain size, cracks and fissures, removal by
sorption in sediments high in clay and organic matter, and the hydraulic conductivity. Many other factors
affect the movement of microorganisms in the subsurface, including their size and shape, concentration,
flow rate, and survivability.
The concept of microbial movement through the subsurface was First addressed as early as the
mid 1920s for microbial enhanced oil recovery (MEOR). At that time, Beckmann (1926) suggested that
microorganisms that produce emulsifiers or surfactants could be transported into an oil-bearing formation
to recover oil that remains after a well has stopped flowing. The addition of microorganisms to oil-
bearing formations to enhance oil recovery by biosurfactant or biogas production has since been
investigated and appears promising (Bubela, 1978). At about the same time, research on the transport of
microorganisms through the subsurface environment was being conducted to determine the effectiveness
of on-site wastewater disposal systems (i.e., pit latrines, septic tanks, land disposal of sewage) in
removing pathogens (Caldwell, 1937, 1938). More recently, the concept of transporting microorganisms
with specialized metabolic capabilities for subsurface bioremediation has been proposed (Lee et al., 1988;
Thomas and Ward, 1989) In recent years, research has been directed toward the introduction of
microorganisms to soil and ground water to introduce specialized metabolic capabilities, to degrade
contaminants which resist the degradative processes of indigenous microflora, or when the subsurface has
been sterilized by contaminants. One of the first studies that addressed microbial transport through
subsurface materials for the purpose of contaminant degradation was published by Raymond et al. (1977).
These investigators reported that heterotrophs and hydrocarbon-degrading bacteria penetrated and were
detected in the effluent of 1.45 x 31 cm columns packed with unconsolidated sands having effective
hydraulic conductivity (K) values ranging from 3.38 x 10 3 to 1.9 x 10"' cm/sec, which were run at a flow
rate of about 30 ml/h (Darcy flow 18/ cm/hr). Microorganisms also penetrated and were detected in the
effluent of 3.8 x 10 cm sandstone (consolidated) cores, with hydraulic conductivities ranging from 1.8 x
10 5 to 7.2 x 10 5 cm/sec, through which water was passed under unknown pressure. In a separate
experiment, it was determined that the added microorganisms were utilizing the gasoline.
3J-J.2 Factors Affecting Bacterial Transport and Survival in the Subsurface
3.3.3.2.1 Matrix
Matrix properties that affect microbial transport in the subsurface include hydraulic
conductivity, mineralogy, and sediment structure. Hydraulic conductivity has been the most studied
24
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parameter affecting transport through porous media; however, the results of laboratory studies in which
samples of porous media were packed to homogeneity may produce underestimates of microbial transport.
The use of intact cores will provide the full range of pore sizes present in situ for microbial transport
through available macropores. Marlow et al. (1991) reported that extent of transport of a yeast,
Rhodotorula sp., after 10 pore volumes through sand columns with hydraulic conductivity values of 5.59 x
10"2 and 1.37 x 10"1 cm/sec was about 2 and 50%, respectively, of the initial number of cells added (1 to 2
x 105 cells/ml). Fontes et al. (1991) investigated the effects of grain size, bacterial cell size, ionic strength
of the transporting fluid, and heterogeneities of the medium on microbial transport and found that grain
size was the most important variable.
3.3.3.2.2 Properties of Microbes
The properties of the microorganisms that may affect transport include size, aggregating
tendencies, shape, condition, and motility and chemotaxis. The results of studies designed to investigate
which cell characteristics are most important are mixed. Cell size is important in that transport will be
limited or prevented for cells that are bigger than the average pore size; however, cells that tend to
aggregate, even if they are small, will not be good candidates for transport. For microorganisms that form
spores, the spore, which is smaller than the vegetative stage, may be transported more efficiently.
Microorganisms that are in a starved state usually are smaller, and produce less extracellular
polysaccharide, which allows the organism to attach to surfaces. Thus the reduced size and stickiness of
the cells should enhance transport.
3.3.3.2.3 Operational Factors
The operational factors that will affect microbial transport include cell concentration, flow rate,
and the ionic strength of the transporting fluid. The results of studies designed to investigate the effects of
cell density on transport have been mixed. The effects of cell density on transport may be organism- and
site-specific. Microbial filtration and clogging of the matrix will be of concern. A direct relationship
exists between flow rate and microbial transport. Finally, there is an inverse relationship between the
ionic strength of the transporting fluid and microbial transport. Microorganisms tend to sorb to surfaces
under conditions of high ionic strength. Therefore, more cells will be transported in a fluid of low ionic
strength. The results from a number of studies suggest that in-situ bioremediation of the subsurface is
usually limited to formations with hydraulic conductivities of 10"4 cm/sec (100 ft/yr) or greater to
overcome the difficulty of pumping fluids through contaminated formations.
3.3.3.2.4 Survival in the Subsurface
Little information is available concerning the survivability of introduced microorganisms in the
subsurface (Thomas and Ward, in press). Transported organisms must not only reach the zone of
contamination but must compete with the indigenous microflora for nutrients, escape predation, retain
their biodegradative capabilities, and often tolerate extremes in pH, temperature, and other environmental
variables (Thomas and Ward, in press). Hardly anything is known about environmental factors and
survivability in the subsurface environment. The same factors that affect survival of microorganisms in
the surface soil and water environments will affect the survivability in the subsurface. These factors
include substrate concentrations, pH, temperature, and the presence of toxicants, predators, and alternate
substrates. By extrapolation from experience with surface water and soil, predation will probably be the
most important factor limiting the survival and activity of introduced microorganisms.
Although specialized microorganisms that have been cultured using selective enrichment
techniques can be used in environmental applications, those developed using genetic engineering
25
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techniques cannot be released into the environment for commercial purposes without prior government
approval (Pimentel et al., 1989). Genetically engineered microorganisms for use in such operations as
MEOR, bioremediation of Superfund sites, extraction and concentration of metals, and production of
specialty chemicals, may be regulated under the Environmental Protection Agency's Toxic Substances
Control Act, Section 5 (Clark, 1992). The use of microorganisms with specialized capabilities to enhance
bioremediation in the subsurface is an undemonstrated technique. However, research has been conducted
to determine the potential for microbial transport through subsurface materials, public health effects, and
microbial enhanced oil recovery. The use of introduced microorganisms has proven most successful in
surface bioreactors when treating ground water in closed-loop systems.
Since the study published by Raymond et al. (1977), which indicated that microorganisms can be
transported and enhance degradation of hydrocarbons in a column packed with sand, no one has
conclusively demonstrated that inoculation of the subsurface enhances bioremediation in the laboratory or
field (Thomas and Ward, in press). However, a significant amount of research tends toward working with
organisms that are easy to culture, and whose genetics are well understood. Little consideration has been
given to developing organisms with good transport properties and survival traits. These include ability to
survive in the subsurface environment, escape predation, and biodegrade the contaminants under in-situ
conditions. The best opportunities involve development of inocula that can degrade mixed wastes, that
have increased tolerance to toxicants, and that produce bioemulsifiers and biosurfactants to increase their
access to oily phase contaminants.
3.3.4 Cometabolic Substrates
Organic compounds can be biotransformed by microorganisms through two basically different
processes: (I) use as a primary substrate, and (2) cometabolism (McCarty and Semprini, in press). In the
first process, biodegradation occurs when the organism consumes an organic compound as a primary
substrate to satisfy its energy and organic carbon needs. Cometabolism, on the other hand, is the
fortuitous transformation of an organic compound by enzymes or cofactors produced by organisms for
other purposes. Here, the organisms obtain no obvious or direct benefit from the transformation. Indeed,
it may be harmful to them.
Cometabolism may occur under either aerobic or anaerobic conditions. For cometabolism to
occur, an active population of microorganisms having the cometabolizing enzymes or cofactors must be
present. Cometabolism requires that an appropriate primary substrate for growth and maintenance of
these organisms must also be present. The primary substrate must often be added to the aquifer, along
with an electron acceptor such as oxygen or nitrate for its oxidation. This is an aspect that adds greater
complexity and cost to cometabolic bioremediation. Cometabolic substrates which have been
investigated include methane, propane, acetate, toluene, and relatively low molecular weight PAHs such
as naphthalene. Contaminants which have been shown to be degraded through cometabolic processes
include chlorinated aliphatic solvents, high molecular weight polyaromatic hydrocarbons, PCBs, and
some pesticides. However, cometabolism has not been used extensively for treatment of organic wastes;
knowledge of practical application of these processes and factors affecting performance is quite limited.
26
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SECTION 4
TYPES OF IN SITU BIOREMEDIATION SYSTEMS
Delivery of Nutrients and Electron Acceptors
As noted earlier, bioremediation technologies differ primarily in the method used to deliver
nutrients, electron acceptors, and other amendments to contaminants in the subsurface. Development of
innovative delivery systems has been the primary focus of research and still remains as the major
challenge in successful application of in-situ bioremediation.
Both subsurface environment (geology and hydrology) and nature and distribution of
contaminants ultimately determine which delivery approach is most appropriate. Accurate information
concerning contaminant nature and distribution, site hydrogeology and geochemistry, and predominant
microbial populations is a vital prerequisite for selection and design of bioremediation technologies.
The specific electron acceptors and nutrients which are selected for delivery depend on the type
of microbial metabolism to be stimulated. The amounts of these materials to be delivered should be
determined with well-designed treatability studies coupled with adequate site characterization.
4.1 In-situ Bioremediation of Soil and Ground Water
In-situ bioremediation technology for the decontamination of soil and ground water contaminated
with petroleum-derived hydrocarbons involves the stimulation of naturally occurring microorganisms that
are capable of degrading the organic contaminants (Atlas, 1981; Lee et al., 1988). In-situ bioremediation
systems for aquifers typically consist of extraction points, such as wells or trenches, and injection wells or
infiltration galleries (Figure 15).
Contaminant locations most often treated with in-situ bioremediation are shown by the dark
shaded area in Figure 16. In most cases, extracted ground water is treated prior to the addition of oxygen
and nutrients, followed by subsequent reinjection. Ground-water treatment has frequently consisted of an
air-stripper tower or activated carbon, but may incorporate an oil/water separator, a biological treatment
unit, an advanced oxidation unit, or combinations of treatment units. When recovered ground water
contains more than a few ppm of biodegradable substances, treatment is likely to be required by
regulations, and is efficient from a process economics perspective. When the recovered ground water
contains low levels of readily degradable constituents, they will generally be degraded within a short
distance of the injection point and will not add significantly to oxygen and nutrient requirements.
Ground-water flow rate and flow paths are critical to the design of in-situ bioremediation
systems. The ground-water flow rate must be sufficient to deliver the required nutrients and oxygen
according to the demand of the organisms. Amended ground water should sweep the entire area requiring
treatment, and the recovery wells should capture the injected ground water to prevent migration outside
the designated treatment zone. In order to ensure that adequate control can be maintained over the ground
water, usually only a portion of the recovered ground water is reinjected. After appropriate treatment, the
other portion is discharged by an acceptable method. The hydraulic conductivity of the ground-water
system, or variability of the aquifer materials, often limits the effectiveness of in-situ technologies or
entirely prevents their use. A suggested target for in-situ remediation technologies is a hydraulic
conductivity of at least 10"4 cm/sec (100 ft/yr). The design of a ground-water recirculation system is best
27
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Ground-Water
Treatment
Oxygen
Source
Nutrients
Injection Well
Recovery Well
Monitoring Well
Soil Contamination
Ground-Water Gradient
Figure 15. In-situ bioremediation utilizing ground-water extraction and reinjection.
Water
Table
Source
Vapor Phase
Capillary
Fringe
Residual
Saturation
Ground Water
Dissolved
Contaminants
Figure 16. Contaminant locations treated with in-situ bioremediation.
28
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done using a ground-water flow model to evaluate optimum placement of injection/extraction wells
(Falatico and Norris, 1990). Such models allow several design concepts to be evaluated and can be more
effective than laboratory treatability studies to determine feasibility. In addition, modeling results can be
used to make midcourse modifications in operations. For most sites, a two-dimensional analytical flow
model will be sufficient. Models which incorporate biodegradation may also be used as an aid in design
and to evaluate bioremediation system performance.
In-situ bioremediation systems are often integrated with other remediation technologies, either
sequentially or simultaneously. For example, if free-phase hydrocarbons are present, a recovery system
should be used to reduce the mass of free-phase product prior to the implementation of bioremediation.
In-situ vapor stripping can be used to both physically remove volatile hydrocarbons and to provide
oxygen for bioremediation. These systems can also reduce levels of residual phase hydrocarbons as well
as constituents adsorbed to both unsaturated soils and soils which become unsaturated during periods
when the water table is lowered.
4.2 Air Sparging
Air sparging is the injection of air under pressure directly into a saturated formation below the
water table to create a transient air-filled space by displacing water from the soil matrix (Figure 17).
Air sparging effectively removes contamination below the water table. Air sparging can
successfully treat VOCs and petroleum hydrocarbons in ground-water aquifers through direct
volatilization and biodegradation. Air sparging is a remediation technology applicable to contaminated
Vapor Extraction Air Sparger
Monitoring
Probe
Monitoring
Probe
Well
Well
Vent
Radius
Sparge
Radius
Vent Radius - f(Vacuum)
Sparge Radius - f(Depth)(Pressure)
,£oil_Pgrti£lj
O •
• o
Contaminated
Soil
Transient Air
Filled Porosity
Figure 17. Diagram of air sparging system (Brown, in press),
29
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aquifer solids and vadose zone materials (Figure 18, dark shaded region). This is a relatively new
treatment technology which enhances biodegradation by increasing oxygen transfer to the ground water
while promoting the physical removal of organics by direct volatilization. Air sparging has been used
extensively in Germany since 1985 (Hiller and Gudemann, 1988) and was successfully introduced in the
United Slates in 1990 (Brown, et al., 1991; Marley et al., 1990; Middleton and Hiller, 1990).
When air sparging is applied, the result is a complex partitioning of contaminants between the
adsorbed, dissolved, and vapor states. Also, a complex series of removal mechanisms are introduced,
including the removal of volatiles from the unsaturated zone, biodegradation, and the partitioning and
removal of volatiles from the fluid phase. The mechanisms responsible for removal are dependent upon
the volatility of the contaminants. With a highly volatile contaminant, for example, the primary
partitioning is into the vapor phase; and the primary removal mechanism is through volatilization. By
contrast, contaminants of low volatility partition into the adsorbed or dissolved phase; and the primary
removal mechanism is through biodegradation. Figure 19 illustrates the relative importance of
volatilization and biodegradation as a function of product volatility.
One of the problems in applying air sparging is controlling contaminant migration. In either
bioventing or ground-water extraction, the systems are under control because contaminants are drawn to
the point of collection. However, air injected into saturated formations may travel in unpredictable
directions and distances. This can mobilize volatile contaminants to the vadose zone or can accelerate
migration of dissolved contaminants by creating locally higher ground-water gradients. Vertical barriers
to upward migration may trap air and produce lateral spread. VOC-laden air streams can rapidly migrate
through the vadose zone to low pressure regions such as basements, causing a vapor hazard. Physical
displacement of the water column can occur if too much pressure is used, increasing migration of ground
Water
Table
Sourc
Vapor Phase
Residual
Saturation
Dissolved
Contaminants
Capillary
Ground Water
Figure 18. Contaminant locations treatable with air sparging.
30
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SNNN\N\SN\\\\N\S\\\N\\\\
L\ SS\SNS\N\N\\\\\\\\\\\\\
>>S^J / // // // // // // // // // /
S\\\\\SSSNSSN\\S\\S
• •¦.~-¦J ~ / // // // // // // // // /
.• ^ N \ N \ \ w * "j * ./•• -•* \ \ S \ \ \
^<v \ S\\SNS\\NSS\\*\
• \ •.*. /¦ /~~~~~~~~//~~
f X. VVXVNV\\\V\\V%
» \ \\ •.'ts; / ~
¦ ,*.¦ .'.X s
v\•.*.•%~ / ssssssssss
/.'.'Tjv \ \ \\ \n\\v\\
/ // // // //
\\\\\\\\\
X ~
• \ \ \ \ \ N
* ~ X ~
\ S \ \ N
| / / X /
SL> \ X \
•.vSs/ / /
Biodegradation
100
0
No. 6 Fuel Waste
Oil Oil
Diesel Jet Fuel Mineral Gasoline
Spirits
Volatility-
Figure 19. Removal mechanisms as a function of product volatility.
water from the sparge area. To prevent these undesired effects, sparge systems should be operated in
conjunction with a vent system that effectively captures sparged gases, and at injection rates which cause
minimal disturbance of hydraulic gradients in the injections area.
As with any technology, there are limitations to the utility and applicability of air sparging. For
air sparging to effectively strip contaminants from ground water, the contaminant must be relatively
volatile and relatively insoluble. If air sparging is used to supply oxygen for biodegradation, the
contaminant should be soluble, relatively nonvolatile, and relatively biodegradable. The geological
characteristics of the site, most importantly heterogeneity, also limit the applicability of air sparging.
Changes in lithology can profoundly affect both the direction and velocity of air flow. If significant
stratification is present, sparged air could be held below an impervious layer and spread laterally, thereby
resulting in the spread of contamination (Figure 20).
As shown in Figure 21, highly permeable zones may lead to rapid and unpredicted movement of
air containing volatilized contaminants. Another constraint of concern is depth related. There is both a
minimum and maximum depth for a sparge system. A minimum depth of 4 feet, for example, may be
required to confine the air and force it to "cone-out" from the injection point. A maximum depth of 30
feet might be required from the standpoint of control. Depths greater than 30 feet make it difficult to
predict where the sparged air will travel.
4.3 Bioventing
Bioventing is the process of supplying air or oxygen to soil to stimulate the aerobic
biodegradation of contaminants (Hinchee, in press). Bioventing is a modification of the technology
variously referred to as soil vacuum extraction, vacuum extraction, soil gas extraction, or in-situ
volatilization. Bioventing systems are generally operated at lower air flow rates than soil vacuum
extraction systems to minimize extraction of volatile contaminants and maximize contact time between
soil microbes and injected air. Soil bioventing is applicable to remediation of contaminants of low
31
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Impervious Barrier
Contaminated Soil
ao o o
Dissolved Particles
Air/Contaminant Migration
Figure 20. Inhibited vertical air flow due to impervious barrier.
SN.V\\V\\\S\>\\NSN>.S>.%\NVSSN\V>
wwvs
\\KW»- *»• •
,\\\\\\\V\\\\\\\\\\\\\-.
ANWW\ ,
High Permeability Zon
Air/Contaminant Migration
<\N\\\AV\N\\N\\\\\\\\\\\N\N\\S\\\\N\\\\\N
^VVSSSSNSXSXSNSXVNVSVVVSVSNSSSXVVXNXNNNXNXSXSSNXNSNNNNNNVNXVVSSSS
\\\\\SSSS\NS\VSN
Figure 21. Preferential air flow through highly permeable zones (Brown, in press).
32
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volatility and can also reduce concentrations of volatile contaminants in off-gases, thus reducing the
amount of contaminants requiring off-gas treatment.
The primary advantage of using air instead of water to transport oxygen to the subsurface is that
the oxygen content of air is 21 %, or 21,000 parts per million. The amount of air required to satisfy high
oxygen demands during biodegradation is, therefore, much lower than the amount of oxygen-saturated
water. In addition, air is more easily moved through the subsurface, whereas hydraulic limitations when
using water may impede delivering necessary amounts of oxygen. Clean air may be injected directly into
the contaminated zone by injection wells or extracted with vacuum extraction wells. Figure 22 shows
two possible configurations of an in-situ bioventing system.
Bioventing is primarily applicable to contaminants in the vadose zone, but can also be applied to
saturated zones which have been dewatered (Figure 23, dark shaded region).
Laboratory research and field demonstrations involving bioventing began in the early 1980s, with
particular emphasis to the remediation of soil contaminated with hydrocarbons. A detailed history is
given by Hinchee (in press). Early on, researchers concluded that venting would not only remove
gasoline by physical means but would also enhance microbial activity and promote the biodegradalion of
gasoline (Texas Research Institute, 1980; 1984). The first actual field-scale bioventing experiments were
conducted by van Eyk for Shell Oil (Hinchee, in press). A series of experiments were conducted by Delft
Geotechnics to investigate the effectiveness of bioventing for treating hydrocarbon-contaminated soils,
and reported in a series of papers (Anonymous, 1986; Staatsuitgeverij, 1986; van Eyk and Vreeken, 1988,
1989a, and 1989b).
Surface
Emission Substrate
Sampler i
Pumps
Primary
Substrate &
Make l^p Air
Injection Well
Air & Primary
sstrate
, t Remjectioi
Well
Wti
Extraction
Well
Soil Gas Monitoring
Figure 22. Two configurations of bioventing. Left - air injection. Right - vacuum extraction with air
reinjection (Wilson and Kampbell, in press).
33
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Residual ^
Saturation
Vapor Phasfc.
^.Fringe
Capillary
Water
Table -
Ground Water
Dissolved
Contaminants
Figure 23. Contaminant locations treatable with bioventing.
Bioventing is potentially applicable to any contaminant that is more readily biodegradable
aerobicaliy than anaerobically. Although most applications have been to petroleum hydrocarbons,
applications to PAH, acetone, toluene, and naphthalene mixtures have been reported. In most
applications, the key is biodegradability versus volatility. If the rate of volatilization significantly exceeds
the rate of biodegradation, removal essentially becomes a volatilization process. In general, low-vapor
pressure compounds (less than 1 mm Hg) cannot be successfully removed by volatilization and can only
be biodegraded in a bioventing application. Higher vapor pressure compounds (above 760 mm Hg) are
gases at ambient temperatures and therefore volatilize too rapidly to be biodegraded in a bioventing
system. Within this intermediate range (1-760 mm Hg) lie many of the petroleum hydrocarbon
compounds of regulatory interest, such as benzene, toluene, and the xylenes, that can be treated by
bioventing.
In addition to the normal site characterization required for the implementation of this or any other
remediation technology, additional investigations are necessary. Soil gas surveys are required to
determine the amount of contaminants, oxygen, and carbon dioxide in the vapor phase; the latter are
needed to evaluate in-situ respiration under site conditions. An estimate of the soil gas permeability,
along with the radius of influence of venting wells, is also necessary to design full-scale systems,
including well spacing requirements, and to size blower equipment.
Although bioventing has been performed and monitored at several field sites, many of the effects
of environmental variables on bioventing treatment rates are still not well understood (Hinchee, in press).
In-situ respirometry at additional sites with drastically different geologic conditions has further defined
environmental limitations and site-specific factors that are pertinent to successful bioventing. However,
the relationship between respirometric data and actual bioventing treatment rates has not been clearly
determined. Concomitant field respirometry and closely monitored field bioventing studies are needed to
determine the type of contaminants that can successfully be treated by in-situ bioventing and to better
define the environmental limitations to this technology.
34
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SECTION 5
CONTAMINANT BIODEGRAD ABILITY
There is considerable information in the literature considering the biodegradability of organic
contaminants. However, extrapolation of this information to site-specific, full-scale bioremediation
systems is quite difficult. This requires that additional testing be done to establish that the contaminants of
concern actually will biodegrade under prevailing site conditions. Determination of biodegradability has
traditionally been done first at the laboratory scale, followed by bench- and field-scale studies.
5.1 Laboratory Testing
Laboratory tests can be used as screening tests to determine site feasibility, as treatability tests to
determine the rate and extent of biodegradation that might be attained during remediation, and as
engineering tests to provide design criteria (U.S. EPA, 1991b). Screening tests include pH and microbial
plate counts to determine if existing conditions are favorable to microbial growth. Respirometer tests,
which measure oxygen uptake but do not normally measure disappearance of the contaminant(s), provide
confirmation that the microbial population is metabolically active. These tests can be run under a number
of nutrient conditions to provide an indication of nutrient effects. Laboratory scale tests are often used as
"proof-of-concept" screens, rather than to obtain information to be used in designing full-scale systems.
Laboratory treatability studies are generally conducted with soil/ground-water slurries (flask
studies) or soils (pan studies). Several conditions are usually tested, including unmodified microcosms,
nutrient amended microcosms, and biologically inhibited conditions. These tests can measure the rate of
change of the constituents of concern as well as changes in pH and microbial populations. The tests
provide data on the rate and extent of conversion of contaminants. During bioremediation of
hydrocarbons in aquifers, the rate of degradation is usually controlled by the rate of supply of nutrient and
oxygen. Under these conditions, laboratory rate data do not extrapolate directly to the field. However,
laboratory data on the rate and extent of removals of hazardous constituents are important for the heavier
hydrocarbons, such as heavy crude oil, bunker oil, or coal gas tars. Removal of compounds from these
materials is often limited by the reaction kinetics of the microorganisms rather than the rate of supply of
some essential nutrient. The extent of biodegradation of oily phase hydrocarbons to microbial biomass or
metabolic end products is very site specific.
5.2 Field Testing: Pilot and Full Scale
Demonstration that a contaminant is biodegradable under laboratory conditions does not
automatically mean that bioremediation will be successful under full-scale field conditions. Many
variables such as site hydrology and geochemistry, contaminant nature and distribution, and predominant
environmental conditions may limit the extent of biodegradation. Because of these factors, pilot testing of
proposed bioremediation systems is almost always necessary.
5.3 Petroleum Hydrocarbons
As a class, petroleum hydrocarbons are biodegradable and can generally be mineralized, i.e.,
converted to carbon dioxide and water. The rate and extent of hydrocarbon biodegradation in the
subsurface will depend on several factors, including (1) the quantity and quality of nutrients and electron
35
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acceptors; (2) the type, number and metabolic capability of the microorganisms; and (3) the composition
and amount of the hydrocarbons. While virtually all petroleum hydrocarbons are biodegradable, the rate
and extent of biodegradation can be highly variable. Depending on environmental conditions,
biodegradation may be very rapid or very slow. The ease of biodegradation depends somewhat on the
type of hydrocarbon. As molecular weight increases, so does the resistance to biodegradation. The lighter,
more soluble members are generally biodegraded more rapidly and to lower residual levels than are the
heavier, less soluble members. Moderate to lower molecular weight hydrocarbons (C to C alkanes,
single ring aromatics) appear to be the most easily degTadable hydrocarbons (Atlas, 1^88). "flius,
monoaromatic compounds such as benzene, toluene, ethylbenzene, and the xylenes are more rapidly
degraded than the two-ring compounds such as naphthalene, which are in tum more easily degraded than
the three-, four-, and five-ring compounds. The same is true for aliphatic compounds where the smaller
compounds are more readily degraded than the larger compounds. Branched hydrocarbons degrade more
slowly than the corresponding straight chain hydrocarbons.
Because petroleum hydrocarbons are frequently found in the presence of other organic
constituents, it is necessary to consider the degradability of other classes of compounds. Non-chlorinated
solvents used in a variety of industries are generally biodegradable. For example, alcohols, ketones,
esters, carboxylic acids and esters (particularly the lower molecular weight analogs) are readily
biodegradable but may be toxic at high concentrations due to their high water solubilities. It is reasonable
to expect that some aerobic biodegradation of chlorinated solvents will occur in the presence of petroleum
hydrocarbon blends, particularly those containing appreciable amounts of toluene. This is, however, a
very site-specific phenomenon and one for which there is not enough documentation to make reliable
predictions. Further, many chlorinated solvents can inhibit biodegradation of petroleum hydrocarbons.
Site remediation is usually concerned with commercial blends of petroleum hydrocarbons, such
as gasoline and other fuels. As for individual compounds, the lighter blends are more readily degraded
than the heavier blends. The extent of conversion that is likely to occur is greatest for lower molecular
weight constituents. Gasolines contain primarily low to moderate molecular weight compounds and can
be biodegraded to low levels under many conditions. For gasoline, the extent of conversion is largely
limited by the efficiency and completeness of the distribution of nutrients and an electron acceptor.
Heavier products such as Number 6 fuel oil or coal tar, however, contain many higher molecular weight
compounds such as five-ring polyaromatic compounds. These mixtures degrade much more slowly than
gasoline and, as a result, significantly lower rates and extent of biodegradation should be anticipated.
5.3.1 Fuels
Gasoline is a complex mixture of hydrocarbons blended from various refinery products. The
specific composition is variable and depends on source of petroleum, individual refinery process streams
used for blending, and customer specifications. The primary components of gasolines are volatile
hydrocarbons which boil at temperatures below 200° C (390° F). Hydrocarbons in this boiling range have
4-12 carbon atoms in their molecular structure. In refinery terms these compounds are derived from light
ends, referring to both the relatively low boiling points and the low molecular weight which typify these
compounds. The major components of blended gasolines are branched-chain paraffins, cycloparaffins,
and aromatic compounds (Cline et al., 1991). Composition of a typical leaded and unleaded gasolines is
shown in Table 1. Gasolines also may contain a number of additives such as dyes, antiknock agents, lead
scavengers, anti-oxidants, metal deactivators, octane enhancers, corrosion inhibitors, and oxygenates.
Specific composition of gasoline will vary both regionally and seasonally. Daily changes in refinery
operation may cause variations in product composition.
36
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Table 1. Composition of Gasolines (from Cline et al., 1991)
Compound volume % In leaded fuel volume % in unleaded fuel
Normal/iso hydrocarbons
59
55
isopeniane
9-11
9-11
n-butane
4-5
4-5
Aromatic hydrocarbons
26
34
xylenes
6-7
6-7
toluene
6-7
6-7
ethylbenzene
5
5
benzene
2-5
2-5
napthalene
0.2-0.5
0.2-0.5
Olefins
10
5
Cyclic hydrocarbons
5
5
The most common dissolved hydrocarbons (benzene, toluene, ethylbenzene and xylenes) released
from gasoline spills are known to be readily biodegradable under aerobic conditions (Jamison et al., 1975;
Gibson and Subramanian, 1984; Thomas et al., 1990; Alvarez and Vogel, 1991). In addition, aerobic
hydrocarbon degrading microorganisms are very common in nature and have been recovered from
virtually all petroleum contaminated sites that have been studied (Litchfield and Clark, 1973). For most
petroleum sites, extensive studies to confirm the presence of BTEX degrading microorganisms are
probably not necessary. Jamison et al. (1975) found that the vast majority of gasoline components were
readily degraded by a mixed microbial population obtained from a gasoline contaminated aquifer.
Although many of the individual gasoline components would not support microbial growth as a sole
carbon source, they did disappear when gasoline dissolved in water was used as the substrate. This
suggests that a mixed microbial population may be necessary for complete degradation. In a study of the
catabolic activity of bacteria from an aquifer contaminated with unleaded gasoline, Ridgeway et al.
(1990) found that most isolates were very specific in their ability to degrade hydrocarbons. Although all
of the 15 hydrocarbons tested were degraded by at least one isolate, most organisms were able to degrade
only one of several closely related compounds. Toluene, p-xylene, ethylbenzene, and 1,2,4-
trimethylbenzene were most frequently utilized whereas cyclic and branched alkanes were least utilized.
In contrast to BTEX, there is much less information available on the biodegradability of many
fuel additives such as methyl tertiary butyl ether (MTBE), 1,2-dibromoethane (EDB), or 1,2-
dichloroethane (EDC). MTBE is of special concern because it is extremely water soluble, is used as an
oxygenate in fuels at levels up to 15%, and is not known to biodegrade. If persistence of fuel additives is a
concern, site specific studies may be needed to confirm the presence of microorganisms capable of
degrading these compounds and to estimate biodegradation rates.
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S3.2 PAHS (Coal Tar, Creosote, Refinery Wastes)
Polyaromatic hydrocarbons (PAHs) are present in heavier petroleum hydrocarbon blends and
particularly in coal tars, wood treating chemicals, and refinery wastes. PAHs have only limited solubility
in water, adsorb strongly to subsurface materials, and degrade at rates much slower than monoaromatic
hydrocarbons or most aliphatic and alicyclic compounds found in refined petroleum hydrocarbon
products. As a result, they often persist for long time periods even under ideal conditions (Lee, 1986;
Borden et al., 1989). Because of their low solubility and strong adsorption to solids, their availability for
degradation is often the limiting factor in treatment (Brubaker, 1991). For the heavier petroleum
hydrocarbons, especially PAHs, the limiting factors may be rate of solubilization, release from interstitial
pore spaces, or rate of degradation of these higher molecular weight constituents. They are more likely to
be biodegraded in mixtures with more soluble and thus more readily degradable hydrocarbons because the
more readily degradable species will support a larger microbial population (McKenna and Heath, 1976).
The higher molecular weight PAHs with high numbers of aromatic rings (e.g., benzo (a) pyrene) have the
slowest rates of biodegradation. This is due in part to very low solubility and high sorption tendency. In
addition, the complex molecular structure of these compounds makes them resistant to attack by a single
organism, and many of the 5- and 6-ring compounds are only biodegradable through cometabolic
processes. Unfortunately, many of these compounds are of regulatory concern because of cancer-causing
tendencies. Thus, the least biodegradable PAHs are often regulated to very low treatment goals.
5.4 Chlorinated Solvents
Chlorinated solvents and their natural transformation products represent the most prevalent
organic ground-water contaminants in the country. These solvents, consisting primarily of chlorinated
aliphatic hydrocarbons (CAHs), have been widely used for degreasing aircraft engines, automotive parts,
electronic components, and clothing. Because of their relative solubility in water and their somewhat poor
sorption to soils, they tend to migrate downward through soils, contaminating water with which they
come into contact. Being denser than water, their downward movement is not impeded when they reach
the water table, and so they can penetrate deeply beneath the ground-water table. CAHs have water
solubilities in the range of 1 g/1, or several orders of magnitude higher than the drinking water standards
for those that are regulated.
The major chlorinated solvents used in the past are carbon tetrachloride (CT), tetrachloroethene
(PCE), trichloroethene (TCE), and 1,1,1-trichloroethane (TCA). These compounds can be transformed by
chemical and biological processes in soils to form a variety of other CAHs, including chloroform (CF),
methylene chloride (MC), cis- and trans-1,2-dichloroethene (cis-DCE, trans-DCE), 1,1-dichloroethene
(1,1-DCE), vinyl chloride (VC), 1,1 -dichloroethane (DCA), and chloroethane (CA). These chemicals,
their solubilities in water, and drinking water maximum contaminant limits, if applicable, are listed in
Table 2. This is the group of chemicals generally to be addressed as a result of chlorinated solvent
contamination of ground water.
Fifteen years ago, many of these highly chlorinated organic compounds were considered
recalcitrant to biological degradation in the environment. Transformation products of the chlorinated
solvents then started to be found in ground waters, and this led to expanded efforts to determine the
chemical and biological processes responsible. It was found that most of the CAHs can in fact be
transformed by biological processes; but generally, the microorganisms responsible cannot obtain energy
for growth from the transformations. The transformations are most often brought about by cometabolism,
38
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Table 2. Common chlorinated aliphatic hydrocarbon (CAH) contaminants in ground
WATER (McCaRTY AND SeMPRINI, IN PRESS)
Water US.
Compound Formula Acronym Density Solubility Drinking
(mg/1) Water
MCL (fig/I)
Carbon tetrachloride
CCL,
CT
1.595
800
5
Chloroform
CHCLj
CF
1.485
8,200
100
Methylene chloride
CHjCLj
MC
1.325
13,000
-
1,1,1-trichloroethane
ch3ccl3
TCA
1.325
950
200
1,1-dichloroethane
CH3CHCL2
1,1-DCA
1.175
5,500
-
1,2-dichloroethane
ch2clch2cl
1,2-DCA
1.253
8,700
5
chloroethane
CHjCHjCL
CA
-
-
-
Tatrachloroethene
ccl-ccl2
PCE
1.625
150
5
Trichloroethene
chcl-ccl2
TCE
1.462
1,000
5
cis-1,2-dichloroethene
CHCL-CHCL
c/j-DCE
1.214
400
70
trans-1,2-dichloroethene
CHCL-CHCL
trans-DCE
1.214
400
100
1,1-dichloroethene
CHj-CCLj
1,1-DCE
-
-
7
Vinyl chloride
CH-CHCL
VC
-
-
2
through interactions of the CAHs with enzymes or cofactors produced by the microorganisms for other
purposes.
Microbially-mediated reactions of chlorinated solvents usually involve oxidation or reduction
reactions. Oxidation reactions are generally slower with highly halogenated compounds than with
compounds containing fewer halogen substituents, while the opposite is true for reduction reactions.
Oxidation reactions do not dehalogenate in the first, rate-limiting step, but in subsequent steps. Reduction
reactions normally include the dehalogenation of these solvents, producing less halogenated homologues.
The dechlorination occurs under anaerobic conditions and results in less chlorinated, and often aerobically
degradable, products. Engineered systems, or in-situ bioremediation, can effectively employ either
aerobic alone or sequential anaerobic/aerobic microbial processes to biodegrade chlorinated solvents.
There are now widespread efforts to take advantage of cometabolism for the transformation of
CAHs in ground water; but this is a much more complicated process than the usual biological treatment
processes that have been used for years, in which organic compound destruction is accomplished by
organisms that use the compounds as primary substrates for energy and growth. In cometabolism, other
chemicals must be present to serve as primary substrates to satisfy the energy needs of the
microorganisms. These substrates must be carefully selected to stimulate the production of the enzymes
that affect cometabolism of the CAHs.
Much has already been learned about cometabolism of CAHs. However, full-scale field
applications of this process are greatly limited; and there are virtually no sufficiently well-documented
full-scale applications at present that can be used to guide design and application or that can be used to
evaluate costs. Thus, any application of bioremediation for chlorinated solvent destruction in the field
must be considered as a research activity and should be evaluated as such. As with any new and untested
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process, failure to reach desired goals should be anticipated; and surprises can be expected. Nevertheless,
the understanding of the process is now at a stage where full-scale experimentation is desirable and
indeed is a necessity if biodegradation of chlorinated solvents is to become a reality rather than just a
laboratory curiosity.
5.4.1 Aerobic Biodegradation of CAHs
Several of the common chlorinated solvents (chlorinated ethanes and ethenes) can be degraded
under aerobic conditions (Norris, in press). Lightly chlorinated compounds such as chlorobenzene (U.S.
EPA, 1986), dichlorobenzene, chlorinated phenols and the lightly chlorinated PCBs are also typically
biodegradable under aerobic conditions. Highly chlorinated organic compounds are much more oxidized
than many natural organics. As such, these compounds do not provide much energy upon further
oxidation in aerobic environments. Most aerobic biodegradation processes start with a step that involves
the insertion of oxygen into a bond on the molecule. Due to the electrophilic nature of that oxygen
insertion, other electrophilic substituents (e.g., chlorine) hinder the reaction. Hence, the observation that
increasing chlorination within a homologous series often leads to a decrease in aerobic (oxidative)
biodegradation (Vogel et al., 1987). These more highly chlorinated analogs are more recalcitrant to
aerobic degradation but are more susceptible to degradation under anaerobic conditions.
Most of the research to date has described the microbial oxidations of mono- or dihalogenated
aliphatic compounds. The major exception to this is the work done on the oxidation of trichloroethylene
(TCE). Several different microbes or microbial enrichments have been shown to be capable of TCE
oxidation (Fogel et al., 1986; Nelson et al., 1986; Little et al., 1988) and chloroform oxidation (Strand and
Shippert, 1986). Apparently, the ease of oxidation increases as the number of halogens decreases. Hence,
dichloroethylene would be oxidized faster than TCE. Unfortunately, due to the nature of contaminant
release in the environment, mass balances are difficult to achieve; and no strong evidence for the
oxidation of halogenated solvents has been derived from actual hazardous waste sites.
Studies of the aerobic biodegradation of chlorinated compounds have illustrated several major
pathways of oxidation. These pathways resemble those for the nonchlorinated homologs. For example,
the oxidation of chlorinated ethylenes involves the formation of a chlorinated epoxide which degrades
rapidly in water. This is similar to the epoxide formed from ethylene. In both of these cases, the microbe
that degrades these compounds might require the addition of a natural nonchlorinated compound, or
cometabolite for growth and energy. The enzymes produced for degradation of that "normal" substrate
are also capable of degrading the pollutant (cometabolism). Certain contaminants such as toluene or
phenol, if present with the chlorinated species, can act as cometabolites. Although early work indicated
that some CAHs, particularly those with few chlorines on the molecule, were biodegradable by
microorganisms, knowledge that a broader range of CAHs can be oxidized aerobically through
cometabolism is rather recent. Wilson and Wilson (1985) showed for the first time that TCE may be
susceptible to aerobic degradation through use of soil microbial communities fed natural gas. These
methanotrophs use an oxygenase (methane monooxygenase or MMO) to catalyze the oxidation of
methane to methanol. MMO also oxidizes TCE fortuitously to form TCE epoxide (Little et al., 1988; Fox
et al., 1990), an unstable compound that undergoes abiotic chemical decomposition to yield a variety of
products, including carbon monoxide, formic acid, glyoxylic acid, and a range of chlorinated acids (Miller
and Guengerich, 1982). In mixed cultures as in nature, cooperation between the TCE oxidizers and other
bacteria occurs; and TCE is further mineralized to carbon dioxide, water, and chloride (Fogel et al., 1986;
Henson et al., 1989; Roberts et al., 1989; Henry and Grbit-Galifc, 1991a).
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Since the report of Wilson and Wilson (1985) on TCE cometabolism, much scientific research
addressing this phenomenon has been performed. The groups of aerobic bacteria currently recognized as
being capable of transforming TCE and other CAHs through cometabolism comprise not only the
methane oxidizers (Fogel et al., 1986; Little et al., 1988; Mayer et al., 1988; Oldenhuis et al., 1989; Tsien
et al., 1989; Henry and Grbit-Galic, 1990; Alvarez-Cohen and McCarty, 1991a,b; Henry and Grbic-Galic,
1991a,b; Lanzarone and McCarty, 1991; Oldenhuis et al., 1991), but also propane oxidizers (Wackett et
al., 1989), ethylene oxidizers (Henry, 1991), toluene, phenol, or cresol oxidizers (Nelson et al., 1986,
1987, 1988; Wackett and Gibson, 1988; Folsom et al., 1990; Harker and Kim, 1990), ammonia oxidizers
(Arciero et al., 1989; Vannelli et al., 1990), isoprene oxidizers (Ewers et al., 1991), and vinyl chloride
oxidizers (Hartmans and de Bont, 1992). These microorganisms all have catabolic oxygenases that
catalyze the initial step in oxidation of their respective primary or growth substrates and have potential for
initiating the oxidation of CAHs.
There is currently insufficient information on the relative advantages and disadvantages of the
different oxygenase systems to recommend definitively one over the other, but each may have its place.
Most research to date has been conducted with the methane oxidizers and the group of bacteria containing
toluene oxygenase, which can be induced with primary substrates such as toluene, phenol, and cresol. The
oxygenases for the above organisms are often nonspecific and fortuitously initiate oxidation of a variety
of compounds including most of the CAHs. The exceptions are highly chlorinated CAHs such as CT and
PCE. In general, oxygenases act on unsaturated CAHs such as TCE by adding oxygen across the double
bond to form an epoxide. With saturated CAHs, such as CF or TCA, a hydroxyl group is generally
substituted for one of the hydrogen atoms in the CAH molecule. Frequently, the resulting products from
CAH oxidation are chemically unstable and decompose as described above for TCE, yielding products
that are further metabolized by other microorganisms present in nature.
In other aerobic degradation pathways of lightly chlorinated compounds, microbes have been
shown to grow on the pollutant when it exists in sufficiently high concentration. Most of these
compounds are mono- or dichlorinated organics. A common reaction is the microbially mediated
substitution reaction where a hydroxyl group replaces a chlorine (Brunner et al., 1980). After this, the
compound is further oxidized; and the metabolites enter the anabolic and catabolic pathways of the
microbe. The possibility that these microbes would adapt to the use of chlorinated compounds as sources
of energy and carbon exists; but this might have limited engineering applications, as will be discussed
later. Selective pressure in natural environments will not be great if pollutant concentrations are relatively
low from a microbial adaptation point of view, even if these concentrations are high from a regulatory
point of view.
5.4.2 Anaerobic Biodegradation of CAHs
As noted above, many highly chlorinated compounds are resistant to aerobic degradation due to
their highly oxidized state. Examples of chlorinated organics resistant to aerobic degradation include (1)
tetrachloroethylene, which has not been observed to undergo epoxidation and (2) hexachlorobenzene,
which has all carbons occupied with chlorine substituents, allowing no site for hydroxylation. These
highly chlorinated organic compounds are not, however, resistant to anaerobic biodegradation (Vogel and
McCarty, 1985; Gibson and Suflita, 1986; Tiedje et al., 1987; Vogel and McCarty, 1987a; Vogel, 1988;
Freedman and Gossett, 1989; Bagley and Gossett, 1990; Nies and Vogel, 1990; Bhatnagar and Fathepure,
1991). Several studies provide evidence for anaerobic transformation of chlorinated solvents by pure
cultures of bacteria. The bacteria involved ranged from strict anaerobic microorganisms, such as
methanogens, sulfate-reducers, and Clostridia to facultative anaerobes such as Escherichia coli or
Pseudomonas putida. Reductive dechlorination was the predominant reaction pathway. Consequently, the
41
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chlorinated solvent biotransformation studies with environmental samples (mixed microbial cultures) and
pure bacterial cultures indicate that a broad variety of bacteria possess the enzymatic capability to
reductively dechlorinate the compounds. An electron donor, such as low molecular weight organic
compounds (lactate, acetate, methanol, glucose, etc.) or H2, must be available to provide reducing
equivalents for reductive dechlorination. Toluene was recently found to be a suitable electron donor for
the reductive dechlorination of PCE to DCE in anaerobic aquifer microcosms (Sewell and Gibson, 1991).
Anaerobic biotransformation of chlorinated solvents has been observed in field studies (Roberts
et al., 1982), in continuous-flow fixed-film reactors (Bouwer and McCarty, 1983b; Vogel and McCarty,
1985, 1987; Bouwer and Wright, 1988), and in soil (Kloepfer et al., 1985), sediment (Barrio-Lage et al.,
1986), and aquifer microcosms (Wilson, B. et al., 1986) under conditions of denitrification, sulfate
reduction, or methanogenesis. Table 3 illustrates commonly observed anaerobic biotransformations. The
initial step in the anaerobic biotransformation was generally reductive dechlorination. For example, CF
was produced from CT, and 1,1-dichloroethane (1,1-DCA) was produced from 1,1,]-TCA.
The transformations of PCE and TCE have been studied most intensely. General agreement exists
that transformation of these two compounds under anaerobic conditions proceeds by sequential reductive
dechlorination to dichloroethene (DCE) and vinyl chloride (VC), and in some instances, there is total
dechlorination to ethene or ethane. Of the three possible DCE isomers, 1,1-DCE is the least significant
intermediate. Several studies have reported that cis-1,2-DCE predominates over trans-1,2-DCE (Barrio-
Lage et al., 1986; Parsons et al., 1984; Parsons and Lage, 1985). CT, CF, 1,2-DCA, 1,1,1-TCA, and PCE
were partially converted to carbon dioxide during anaerobic biotransformations. Reductive dechlorination
of 1,1,1 -TCA and PCE occurred first, prior to mineralization to carbon dioxide. Most of the experiments
were conducted under methanogenic conditions. Several of the chlorinated compounds were also
transformed by similar pathways under conditions of denitrification and sulfate reduction (Table 3).
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Table 3. Anaerobic Transformation of Selected Chlorinated Solvents in Microcosms
and Enrichment Cultures Under Different Redox Conditions (Bouwer, in press)
Chlorinated
Redox
Transfor-
End
solvent*
condition*
mation*
Intermediate*
product
System
Refs"
CT
dn
+
CF
n.d.
biofilm reactor
d.e
sr
+
CF
n.d
biofilm reactor
e
me
+
CF
CO,
biofilm reactor/
c,e,n
aquifer material
CF
dn
__
__
biofilm reactor
d.e
sr
-
--
-
biofilm reactor
e
mc
+
n.d.
CO,
biofilm reactor
c,e
1,2-DCA
me
+
n.d.
CO,
biofilm reactor
c
1,1,1 -TCA
dn
__
__
biofilm reactor
d.e.j
sr
+
1,1-DCA
CA
biofilm reactor/
e.j
aquifer material
me
+
1,1-DCA
co2
biofilm reactor/
c.e.j.r
aquifer material
1,1,2,2-TeCA
me
+
-
1,1,2-TCA
biofilm reactor
c
HCA
ae
+
—
PCE
aquifer material
f
dn
+
n.d.
n.d.
biofilm reactor
e
sr
+
n.d.
n.d.
biofilm reactor
e
me
+
n.d.
n.d.
biofilm reactor
e
1,1-DCE
me
+
VC
n.d.
sediment/aquifer material
b,s
ciJ-1,2-DCE
me
+
VC
n.d.
sediment/aquifer material
b,s
/rans-l ,2-DCE
me
+
CA +VC
n.d.
sediment/aquifer material
b,s
TCE
me
+
cis -1,2-DCE/
n.d.
aquifer material
m,n
trans- 1,2-DCEl ,2-DCE
n.d.
aquifer material
k,s
PCE
sr
+
TCE
cjj-I,2-DCE
sewage sludge
a
me
TCE
CO,
biofilm reactor
q
ethene
sewage sludge
8-h
cis + trans-
aquifer material
P
1,2-DCE
aquifer material/
i,o
cis-1,2-DCE
sewage sludge
c,m,n,l
n.d.
aquifer material
* Abbreviations stand for CT - caibon tetrachloride; CF - chloroform; DCA - dichloroethane; TCA - trichloroethane; CA -
chloroethane; TeCA tetrachloroethane; HCA - hexachloroethane; DCE - dichloroethene; TCE - trichloroethene; PCE -
tetrachloroethene.
" ae — aerobic; dn - denitrification; sr - sulfate reduction; me - methanogenesis.
c + - transformation observed; - - no transformation
d n.d. - not determined
' a - Bagley and Gossett, 1990; b - Barrio-Lage et al., 1986; c - Bouwer and McCarty, 1983a;d - Bouwer and McCarty, 1983b;
e - Bouwer and Wright, 1988; f ¦ Criddle et al., 1986; g ¦ DiStefano et al., 1991; h « Freedman and Gossett, 1989; i ¦
KSstner, 1991; j - Klecka et al., 1990; k - Kloepfer et al., 1985; I - Parsons and Lage, 1985; m - Parsons et al., 1984; n -
Parsons et al., 1985; o - Scholz-Muramatsu et al., 1990; p - Sewell and Gibson, 1991; q ¦ Vogel and McCarty, 1985; r -
Vogel and McCarty, 1987; s - Wilson et al., 1986a.
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SECTION 6
MONITORING AND PERFORMANCE EVALUATION FOR
BIOREMEDIATION SYSTEMS
One of the key elements in applying in-situ bioremediation systems is evaluation of performance.
Several approaches have been used: (1) monitoring contaminant concentration in ground water
(dissolved phase), (2) monitoring contaminant concentration in soils or aquifer material (residual
saturation), and (3) monitoring changes in levels of electron acceptors and nutrients within and around the
contaminated zone.
Demonstrating decreases in ground-water concentrations of contaminants of concern has been the
most common approach. Unfortunately, the solubility of petroleum hydrocarbons is low; and most of the
hydrocarbon mass is associated with the solids, not the dissolved phase. Since the mass of contaminant
trapped at residual saturation within the soil or aquifer is the source of any hydrocarbons which are
dissolved in ground water, quantities associated with aquifer solids are far more important than ground-
water concentrations. Therefore, site characterization should include delineation of the extent of
contaminant mass acting as a source. Performance can then be evaluated based on decreases in the total
amount of contaminant, rather than the fraction which has dissolved in ground water. This, however, is
difficult lo do, and is often considered to be outside of the scope of many investigations.
Demonstrating decreases in the mass or concentration of specific contaminants of concern in soils
or aquifers is a primary requirement of performance evaluation. As noted earlier, measuring ground water
concentrations of contaminants may not accurately indicate the extent of biodegradation of the source.
Obtaining core samples from the contaminated subsurface before, during, and after bioremediation efforts
is often the only means of doing this. Methods selected for analysis of samples should look for specific
compounds rather than groups of compounds. Nonspecific parameters such as Total Petroleum
Hydrocarbon (TPH) can measure components that are not of interest (e.g., asphalt particles), do not
measure the most volatile compounds, and can yield highly variable results as shown in studies where
split samples have been sent to different laboratories (Anonymous, 1992).
6.1 Sampling Programs
Comprehensive site characterization is critical to proper evaluation of performance of
bioremediation systems. The primary criteria for evaluating treatment is a decrease in the concentration of
a contaminant of concern to below a defined treatment goal, usually regulatory in nature. Due to site
heterogeneity and costs associated with sampling and analysis, uncertainties about total mass of
contaminants and variability in contaminant concentrations often exist. The range of initial contaminant
concentration is often not known. If the variability in contaminant concentration at the outset of the
treatment is not known, it will be impossible to determine if contaminant concentrations at the end of
treatment represent statistically significant decreases, or simply variation in the distribution of
contaminants. Site characterization information is often not this detailed.
Sampling programs to monitor performance are also often limited and not designed to obtain
statistically representative data. Although obtaining representative data is sometimes difficult, particularly
for sites with heterogeneous conditions and/or multiple sources, it is critical to design sampling programs
which collect adequate information to demonstrate biodegradation. Although more extensive sampling
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programs may increase the cost of bioremediation, it is essential to note that, without adequate data,
claims of success are often not defensible.
6.2 Indicators
One useful method for assessing biodegradation activity is to monitor changes in the
concentration of inorganic compounds within the aquifer (Borden, in press). Biodegradation of
contaminants will result in the removal of electron acceptors (oxygen, nitrate, and sometimes sulfate) and
release of waste products (carbon dioxide, reduced iron and methane) in areas where microorganisms are
most active. If field monitoring indicates depletion of electron acceptors (oxygen, nitrate and/or sulfate)
or production of waste products (carbon dioxide, soluble iron, or methane) within a plume or
contaminated interval, this is a good indication that one or more of the contaminants are being
biodegraded. The major limitation of this approach is that it is not possible to determine which specific
compounds are being degraded or to what extent.
Oxygen, carbon dioxide, and methane are easily measured in the field, especially in soil gas
surveys. Depletion of oxygen in contaminated zones is caused primarily by bacterial respiration.
Accumulation of carbon dioxide within and adjoining the contaminated areas is also indicative of
bacterial respiration. However, direct interpretation of carbon dioxide concentrations is sometimes
difficult because carbon dioxide can be released during dissolution of certain minerals (e.g., limestone).
Thus carbon dioxide can come from sources other than bacterial degradation. Methane is produced only
under anaerobic conditions and can be used to monitor extent of anaerobic degradation. However,
methane can exist as the result of biodegradation of naturally occurring organic materials, and care should
be taken to verify that methane detected is actually generated within the contaminated area.
Other indicators of biological activity include changes in redox potential and pH. Measurement of
redox potential is relatively simple and can provide a good qualitative indicator of the overall oxidation-
reduction status of the aquifer. Redox potential can be measured using a platinum electrode and a
standard pH meter. In locations where the redox potential is negative, the ground water is strongly
reduced, indicating significant bacterial decomposition. In areas where the redox potential is positive, the
ground water is oxidizing, indicating that the contaminant plume has not reached this point or that
bacterial degradation has not occurred. In most cases, redox potentials should not be used for precise
calculations but as a qualitative indicator of environmental conditions within and outside the contaminant
plume (Barcelona et al., 1989). The pH of the system can be monitored to evaluate the extent of bacterial
respiration. Many by-products of bacterial degradation of contaminants are organic acids and tend to
decrease the pH of the system. However, aeration and addition of nutrients, as well as certain types of
bacterial metabolism (i.e., denitrification), can also increase the pH and interpretation of pH data should
take these factors into account.
6.3 Meeting Treatment Goals
The ability of in situ bioremediation to meet relevant regulatory goals depends both on the
specified cleanup levels and the limits of the technology (Norris, in press). Endpoints can be mandated by
state or federal numeric criteria such as National Safe Drinking Water Act MCLs, or they can be based on
site specific risk assessments. Cleanup standards and acceptable timeframes can vary significantly
according to state, and the specific levels and schedules set for a given site often determine which
specific technology is employed. Regulations that set levels at or below analytical detection limits or site
background concentrations may preclude the use of bioremediation.
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Under ideal conditions, in-situ bioremediation can reduce petroleum hydrocarbon levels to
nondetectable levels (10 mg/kg) (Norris, in press). This is more easily obtained with the lighter blends in
permeable and homogeneous formations where placement of injection and recovery wells (galleries, etc.)
is unencumbered. Generally, for lighter petroleum blends, the hardest regulatory endpoint to meet is the
benzene limit (Norris, in press). Although benzene is highly biodegradable, MCLs for benzene (5 ppb) are
at least an order of magnitude lower than for other specific light hydrocarbon constituents. As a result, if
the benzene endpoint can be reached, the level for the other components will most probably be met as
well.
For heavier petroleum hydrocarbons, BTEX compounds may not be present in significant
concentrations to be of concern (Norris, in press). Typically, TPH will be the target analysis to be met.
Because it is a nonspecific analysis, using background TPH levels as the remediation goal can create
difficulties in interpreting data and lead to misleading conclusions regarding the performance of the
system. The heavier the petroleum mixture, the more probable it is that there will be residuals of very
slowly degraded components. These components tend to have low water solubilities, which can limit their
rate of degradation. If TPH is the only criterion, the measurements will not determine which petroleum
hydrocarbons components have gone untreated. Compounds that are not of environmental concern may
contribute to reported TPH values and thus complicate interpretation.
Polyaromatic hydrocarbons can be difficult to treat to the regulated levels. The MCLs for many of
these compounds are low because they are suspected carcinogens. The rate of release of PAHs from
subsurface solids may be too slow to support an active microbial population, and degradation rates may
be impractically slow. Fortunately, the degradability of these compounds is better in mixtures containing
lower molecular weight compounds found in many commercial petroleum products. Available data on the
limits of PAH degradation under in-situ bioremediation conditions are limited and contradictory, and thus
predictions of treatment limits are likely to be unreliable.
6.4 Bioremediation Limits
6.4.1 Concentration
The range of contaminant concentrations that are amenable to bioremediation depends on a
number of factors. These are illustrated in Figure 24.
High concentrations of contaminants can cause toxicity to microbes essential for biodegradation.
Toxicity can occur with petroleum hydrocarbons, some chlorinated solvents and with certain very soluble
compounds such as alcohols. The concentration at which a compound is toxic will be to some extent site
and contaminant specific, as microbial communities usually have substantial capacity to adapt. In
addition, mixtures of contaminants may exert toxicity where single contaminants would not. For example,
individual components of creosote (such as naphthalene) exert less toxicity by themselves than the
mixture. Some data on approximate concentrations at which specific compounds are toxic,
biodegradability, and other properties of environmental interest are available from several handbooks
(Montgomery, 1991; Montgomery and Wilkom, 1990; Howard, 1989 and 1990; and Verschueren, 1983).
However, it should be remembered that this information has been generated under a variety of conditions.
Therefore, determining the extent of contaminant toxicity generally requires site-specific measurements.
Very high concentrations of contaminants also may create very high oxygen and/or nutrient
demands. Meeting these demands might require excessively longer times and higher costs than other
technologies (Piontek and Simpkin, 1992), especially in aquifers with low hydraulic conductivity.
46
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Bioremediation
Concentration Limits*
1 ppb? 10 ppm?
manv variables
Low
100 ppm? ^ Rate dependent on ^
Poor enzyme induction
— may use other substrates
Toxicity
Inhibition
• Availability
— sorption to solids
- solubility limits • Must be considered when
selecting remedies based on cleanup criteria
Figure 24. General concentration limits for bioremediation.
Although the contaminant may be amenable to in-situ bioremediation, in these situations it may be more
practical to combine in-situ bioremediation with other technologies such as free-phase recovery, ground-
water sparging, and in-situ vapor stripping.
Low contaminant concentrations also present limitations to bioremediation. First, certain
concentrations of contaminant are required to induce biodegradation pathways in microbial populations.
Below these concentration thresholds contaminants are often not degraded. Second, at low concentrations
many compounds may not be available for microbial degradation. For example, high molecular weight
PAHs such as benzo (a) pyrene have low water solubility and high sorption tendencies. At lower
concentrations most of these types of contaminants will be bound to surfaces, and unavailable to
microbial populations. Therefore, at concentrations considerably higher than some cleanup goals,
biodegradation may either not occur at all or may stop before the target concentration is reached. Last, but
not least, contaminant concentrations have to be high enough to support growth of the degrader
populations. Contaminants concentrations which are above regulatory action levels may not be sufficient
to support such growth.
6.4.2 Metals
Bioremediation is not generally applicable to metals-contaminated sites, but may mobilize or
immobilize various metals. Generally, the presence of metals has little direct effect on the bioremediation
process. While some metals such as zinc or mercury can be toxic to bacteria, the microbial population
frequently adapts to the concentrations present. The effect of metals on the microbial population of a
specific site must usually be tested through specific treatability studies. Bioremediation technologies for
removal of metals from contaminated ground water are being developed but are still considered as
emerging.
6.4.3 Mass Transfer
When large quantities of oily phase or viscous materials such as the heavier fuel oil blends are
present, the flow of water through the contaminated zone may be impeded, preventing the delivery of
nutrients and/or electron acceptors. In these situations, in-situ bioremediation is generally not feasible.
47
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The concentration at which this occurs will vary with soil type, but generally will be above 20,000 mg/kg.
In aquifers where NAPLs occur, both free-phase product and residual saturation may not be available for
microbes to biodegrade. Research is in progress investigating the use of microbes which produce
biosurfactants to overcome some of these barriers, but application of these in bioremediation systems is
still considered to be in the developmental stages.
48
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SECTION 7
STATE OF BIOREMEDIATION TECHNOLOGY
7.1 Natural Attenuation
At present, there are no well-documented full-scale demonstrations of natural bioremediation,
although there has been some limited research into the processes that control the natural biodegradation of
dissolved hydrocarbon plumes (Borden et al., 1986; Barker et al., 1987; Franks, 1987; Hult, 1987a;
Chaing et al., 1989; Wilson et al, 1992). In addition, there is almost no operating history to judge the
effectiveness of natural bioremediation. Early attempts at aquifer remediation focused on using
conventional remediation techniques to remove or permanently immobilize contaminants at the highest
priority sites. At many low priority sites, regulators have assumed that natural bioremediation would be
adequate to control migration of dissolved contaminants. However, these sites have typically not been
monitored sufficiently to determine if this approach is actually effective or to identify those factors that
influence the efficiency of natural bioremediation. The primary repositories of expertise on natural
bioremediation are in universities, in industry, the U.S. Environmental Protection Agency (U.S. EPA) and
the U.S. Geological Survey (U.S.G.S.).
7.2 In-situ Bioremediation
In-situ bioremediation technologies are in varying stages of development. Although there has
been considerable experience with laboratory and field-scale systems, information on full-scale
application of bioremediation is still relatively limited. The technology, in general, is applicable and
practicable. In-situ bioremediation of petroleum hydrocarbons is fairly well established. However,
considerable expertise is required along with detailed knowledge of site hydrogeology, contaminant
nature and distribution, and desired microbial metabolism. In addition, difficulties in evaluating
performance hamper efforts to establish the success of various applications.
Information on commercial application of in situ bioremediation has recently been compiled by
the USEPA Office of Research and Development (USEPA, 1992b). Detailed information concerning
these case studies is maintained on the ATTIC (Alternative Treatment Technology Information Center)
electronic bulletin board database. Vendors supplied information on 132 case studies. Full-scale systems
were reported for 86 of these studies. Approximately 85% of the systems reported were for petroleum
related, wood-preserving wastes, or solvents.
7.2.1 Anaerobic Bioremediation—Alternative Electron Acceptors
The demonstration of nitrate-based bioremediation in the field is limited; therefore, the use of this
alternate electron acceptor for bioremediation must be viewed as a developing treatment technology.
Table 4 summarizes results of selected field studies where denitrification was tested as a means to
remove aromatic hydrocarbon contamination. The Traverse City and the Rhine River Valley studies
involved actual contamination sites. The Borden experiment involved injection of a mixture containing
benzene, toluene, and xylene isomers (BTX) in one experiment and gasoline in another. In general, BTX
compounds were found to disappear within the nitrate amended zone. Interpretation of these data,
however, was complicated by a number of factors, especially the co-occurrence of nitrate and 02, and the
lack of complete characterization of the organic substrates. Nitrate removal exceeded the expected
49
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Table 4. Field Studies Where Denitrification Has Been Evaluated (Reinhard, in press)
Study Site
and Authors
Contamination
and Conditions
Major Implication for
In-Situ Bioremediation
Traverse City, MI,
Hutchins and Wilson, 1991
JP-4 fuel,
NaN03: 62 mg/I,
02: 0.5 to 1 mg/1
(1) Removal of benzene, toluene, m,p-xylene;
Recalcitrance of o-xylene.
(2) Nitrate removed exceeded stoichiometric
amount of BTEX removal.
(3) Partitioning of compounds into the water phase
appears to be a major factor determining
compound removal.
Borden, Ontario,
Berry-Spark et al., 1988
Gasoline and
BTX; Oxygen
and nitrate
(1) BTX transform more slowly when gasoline
is present than in systems where BTX
are the only substrate.
(2) In systems containing both 02 and nitrate,
aerobic and (facultative) denitrifying
organisms appear to cooperate.
Seal Beach, Reinhard
et al. 1991
Gasoline contaminated
ground water feed,
NO,- (6 mg/1)
90% nitrate removal in mixed nitrate/sulfate
system, aromatics removal toluene>
p-\y lene>o-xy lene>benzene.
Rhine Valley, FRG,
Werner, 1985
Fuel oil (?)
Aerated water (02),
NOj (>300 mg/1),
P04 (>0.3 mg/1),
NH/ (>1.0 mg/1)
(1) Removal fastest for benzene, slower
for toluene, slowest forp-xylene.
(2) Oxygen suspected to be electron acceptor
initiating the transformation.
amount based on the substrates analyzed, and this was attributed to the dissolved organic carbon in
ground water (Berry-Spark et al., 1988).
Bioremediation of chlorinated solvents using alternate electron acceptors is a developing
treatment technology that is mostly being investigated at the laboratory scale. Limited field experience
exists on stimulation of anaerobic biotransformation for control of chlorinated solvents. One field study
demonstrating this technology was conducted at the Moffett Field Naval Air Station, Mountain View,
California (Semprini et al., 1991). This site was used earlier to study in-situ restoration of chlorinated
aliphatics by methanotrophic bacteria (Roberts et al., 1990). Reducing conditions were promoted in the
field in a 2 square meter test zone by stimulating a consortium of denitrifying bacteria, and perhaps
sulfate reducing bacteria, through the addition of acetate as primary substrate (25 mg/L). The aquifer
contained both nitrate (25 mg/L) and sulfate (700 mg/L). CT was continuously injected at a concentration
of 40 ng/L, and between 95 and 97 percent CT biotransformation was observed in the 2-m test zone with
stimulated anaerobic growth. CF was an intermediate product and represented 30 to 60 percent of the CT
transformed. Other halogenated aliphatics were biotransformed but at slower rates and lower extents of
removal. Removals achieved for Freon-11, Freon-113, and 1,1,1-TCA ranged between 65 to 75 percent,
10 to 30 percent, and 11 to 19 percent, respectively.
50
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A second field demonstration was conducted at a chemical transfer facility in North Toronto
(Major et al., 1991). The aquifer at this site was contaminated with organic solvents (methanol, methyl
ethyl ketone, vinyl and ethyl acetate, and butyl acrylate) and PCE. Samples of the aquifer material were
amended with PCE plus acetate/methanol. Over a 145-day incubation period in the laboratory, PCE was
dechlorinated to TCE, then cis-1,2-DCE, VC, and in many instances, to ethene. From these results the
investigators hypothesized that the methanol present in the contaminated site serves as a primary substrate
for complete dechlorination of PCE by anaerobic microorganisms. In-situ anaerobic bioremediation
appears to be occurring at the site without addition of chemicals.
The possible formation of toxic metabolites has been the major impediment to the development of
practical anaerobic bioremediation in the field for cleanup of chlorinated solvent contamination. The
intermediates commonly observed, such as cj'j-1,2-DCE, trans-1,2-DCE, VC, 1,1-DCA, and CF, also
pose a threat to public health. For anaerobic bioremediation to be useful, chlorinated solvents must be
biotransformed to nonchlorinated, environmentally acceptable products. Some recent laboratory studies
have demonstrated that this is possible and help provide impetus to further develop anaerobic biological
processes for bioremediation.
7.3 Air Sparging
Air sparging is a potential means of extending the advantages of vapor extraction technology and
bioventing to the saturated regime. With air sparging, air is injected under pressure below the water table
creating a transient air-filled porosity. This enhances biodegradation as well as volatilization of petroleum
hydrocarbon contaminants from the soil and ground water. The net result is a rapid and significant
decrease in contaminant levels.
There is very little information available concerning effectiveness of air sparging at full scale in
the United States. Air sparging is an emerging technology for the treatment of ground water contaminated
with volatile organic compounds. It is being used to increasingly greater extents to treat petroleum
hydrocarbon contaminated ground-water aquifers, overcoming the limitation of SVE for treating saturated
zone contaminants and improving the efficacy of bioremediation. The benefits and limitations of this
technology are still being defined both in field application and research.
7.4 Bioventing
Bioventing has been performed and monitored at several field sites contaminated with middle
distillate fuels, mainly JP-4 jet fuel (Dupont et al., 1991; Miller et al., 1991; van Eyk and Vreeken, 1991;
Urlings et al., 1991). Several large scale studies have been conducted at United States Air Force bases
which have contributed significantly to knowledge concerning optimum design and operations. Results
from these field demonstrations indicate that bioventing may be a feasible option for in situ
biodegradation of residual fuel contaminants not amenable to recovery by SVE alone. Methods to reduce
vapor extraction rates and maximize vapor retention times in the soil are compatible with enhancing
biodegradation reactions through moisture management. Use of bioventing should minimize
volatilization, potentially eliminate the need for off-gas treatment, and maximize in situ utilization of
oxygen. It is estimated that various forms of bioventing have been applied to more than 1,000 sites
worldwide; however, little effort has been given to the optimization of these systems.
The effects of environmental variables on bioventing treatment rates are still not well understood.
In-situ respirometry at additional sites with drastically different geologic conditions has further defined
environmental limitations and site-specific factors that are pertinent to successful bioventing. However,
51
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the relationship between respirometric data and actual bioventing treatment rates have not been clearly
determined. Additional field respirometry and closely monitored field pilot bioventing studies at the same
sites are needed to determine what types of contaminants can be successfully treated in situ by bioventing
and what the environmental limitations are. Studies to date clearly show that many preconceptions
regarding the factors that control bioventing rates can be wrong. For example, active respiration at a
subarctic site at Eielson AFB near Fairbanks, Alaska, suggests that good rates of in-situ hydrocarbon
degradation can occur at locations that are continually subjected to a cold environment. Failure to
accelerate biodegradation rates by adding nitrogen fertilizer to biovented soils that contain low nitrogen
levels indicates that nutrient addition at some sites may not be required. Also, fine-grained moist clayey
soils have been readily aerated and showed aerobic respiration, indicating that bioventing is feasible at
times in soils having low permeabilities. Other low permeability sites have not proven amenable to
bioventing, and better procedures to evaluate sites are needed.
Vapor phase biodegradation occurs and can take place in situ. The question of how soil sorption
and partitioning of volatile organic compounds into soil air affects biodegradation rates was addressed
earlier by McCarty (1987). This question needs further attention as the movement of the vapor phase in
soils is complex and dependent on changing soil environmental conditions. Bioventing rates need to be
determined under varying vapor extraction rates since an important purpose for bioventing is to
biodegrade the vapor within the soil profile. The minimal soil aeration levels that provide for high
degradation rates must be determined under different soil conditions. Interaction of the vapor phase with
soil particles and microorganisms in the uncontaminated soil profile needs further research in both the
laboratory and in the field. The primary problems encountered with bioventing are:
• Accurate estimate of emissions—One of the key variables in bioventing cost and design is the
need for offgas treatment. Regulators typically require an estimate of emission rate before
permitting a facility without offgas treatment.
• Time required for remediation—Bioventing may require two or more years for remediation. At
many sites this can be a problem.
• Determination of effectiveness on nonpetroleum hydrocarbons—Many sites are contaminated
with a mixture of chemical wastes, and little is known of the effectiveness of bioventing on
nonpetroleum hydrocarbons.
• Regulatory acceptance—With this technology, as with many emerging technologies, obtaining
regulatory acceptance can be difficult.
52
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SECTION 8
SUMMARY
In-situ bioremediation technologies have been developed in response to cost and technical factors
associated with excavation and incineration of contaminated soils and aquifers. These technologies differ
primarily in the mechanisms of delivering nutrients and electron acceptors to contaminated subsurface
soils and aquifers. Currently these delivery systems include ground water recirculation, air-sparging
below the water table, and venting of unsaturated (vadose zone) soils. The primary limitations of in situ
bioremediation technology are site geology and hydrogeology, which limit delivery rates and capacities.
Therefore, comprehensive site characterization is vital to successful design and implementation of in-situ
bioremediation systems. Without detailed information concerning site hydrogeology, and nature and
distribution of contaminants, appropriate designs cannot be selected. Performance evaluation for
bioremediation also depends on adequate site characterization.
Although a wide range of contaminants are potentially biodegradable, bioremediation systems
have been most successfully applied to petroleum hydrocarbons (fuels and refinery wastes), wood
preserving wastes (creosote), and chlorinated solvents (TCE). Petroleum hydrocarbons and wood-
preserving wastes are primarily biodegraded under aerobic conditions, using oxygen as an electron
acceptor. Bioremediation of these contaminants has focused on mechanisms for overcoming oxygen
limitations. Some petroleum hydrocarbons, principally aromatic components of fuels (BTEX), and
chlorinated solvents (PCE, TCE) have been shown to biodegrade under anaerobic conditions using
electron acceptors other than oxygen. Nitrate, sulfate, and carbon dioxide are attractive alternatives to
oxygen because they are more soluble in water, inexpensive, and nontoxic to microorganisms. The
demonstration of this technology in the field is limited, therefore, its use for bioremediation must be
viewed as a developing treatment technology.
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