&EPA
United States
Robert S. Kerr
EPA/600/9-89/073
Environmental Protection Environmental Research Laboratory August 1989
Agency Ada, OK 74820
Research and Development |
Bioremediation of
Contaminated
Surface Soils
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EPA-600/9-89/073
Bioremediation of Contaminated
Surface Soils
J.L. Sims, B.C. Sims, and J.E. Matthews
August 1989
This report was developed by the
Roberts. Kerr Environmental Research Laboratory
U.S. EPA, ORD
Ada, Oklahoma 74820
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Contents
1. Introduction
2. Overview of Soil Biodegradation and Other Soil Processed.
3. Waste and Soil Characterization
4. Microbial Factors Affecting Biodegradation
5. Treatability Studies for Determination of Bioremediation Potential
6. Integration of Information from Site Characterization and Treatability Studies.
7. Potential Applications and Limitations of Bioremediation Technology
8. Example of Bioremediation Potential for Polycyclic Aromatic Hydrocarbons (PAHs) in a Soil System
vlil
9. Implementation of Bioremediation at Sites Contaminated vj/ith Organic Wastes
10. Conclusions
11. References
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5
6
8
10
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12
18
20
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1. Introduction
Biological remediation of soils contaminated with organic
chemicals is an alternative treatment technology that can
often meet the goal of achieving a permanent clean-up
remedy at hazardous waste sites, as encouraged by the
U.S. Environmental Protection Agency (U.S. EPA) for
implementation of The Superfund Amendments and
Reauthorization Act (SARA) of 1986. Bioremediation is
consistent with the philosophical thrust of SARA, for it
involves the use of naturally occurring microorganisms to
degrade and/or detoxify hazardous constituents in the soil
at a contaminated site to protect public health and the
environment. Bioremediation of contaminated soils,
including applications and limitations, has been addressed
at several recent scientific meetings and conferences [1,2,
3,4]. With regard specifically to wood preserving
contaminated sites McGinnis et al. [5] have stated that
reliable, safe, economical bioremediation techniques using
soil systems are attractive and warrant thorough study and
evaluation. The use of bioremediation techniques in
conjunction with chemical and physical treatment
processes, i.e., the use of a "treatment train," is an
effective means for comprehensive site-specific remediation
[6].
Wilson [7] identified biological processes, including microbial
degradation, as a mechanism for attenuating contaminants
during transit through the vadose zone to the groundwater.
(The vadose zone is the region extending from the ground
surface of the earth to the upper surface of the principal
water-bearing formation [8]). On-site soil remedial
measures using biological processes can reduce or
eliminate groundwater contamination, thus reducing the
need for extensive groundwater monitoring and treatment
requirements [7, 9,10]. Lehr [11 ] also emphasized that
monitoring for attenuation of contaminants occurring in the
vadose zone provides information for understanding their
movement in and through the vadose zone and in the
groundwater.
On-site bioremediation of contaminated soils generally is
accomplished by using one of three types of systems:
(1) In situ;
(2) Prepared bed; or
(3) Bioreactor (e.g., slurry reactors) systems.
This discussion focuses on in situ and prepared bed
systems, which utilize the soil as the treatment medium, as
contrasted to bioreactor systems, in which contaminated
soil is treated in an aqueous medium.
An in situ system consists of treating contaminated soil in
place. Contaminated soil is not moved from the site. In
general, naturally occurring microorganisms are allowed to
treat the contaminants. Treatment often may be enhanced
by a variety of physical/chemical methods, such as
fertilization, tilling, soil pH adjustment, moisture control, etc.
In some instances, addition of supplemental populations of
adapted organisms may serve to enhance treatment.
In a prepared bed system, the contaminated soil may be
either (1) physically moved from its original site to a newly
prepared area, which has been designed to enhance
bioremediation and/or to prevent transport of contaminants
from the site; or (2) removed from the site to a storage area
while the original location is prepared for use, then returned
to the bed, where the treatment is accomplished.
Preparation of the bed may consist of such activities as
placement of a clay or plastic liner to retard transport of
contaminants from the site, or addition of uncontaminated
soil to provide additional treatment medium. Treatment may
also be enhanced with physical/chemical methods, as with
in situ systems.
2. Overview of Soil Bio-
degradation and Other Soil
Processes
Bioremediation of a soil contaminated with organic
chemicals is accomplished by degradation of specific
organic constituents, i.e., the "parent" compounds. The
term degradation may refer to complete mineralization of
the constituents to carbon dioxide, water, inorganic
compounds, and cell protein. The ultimate products of
aerobic metabolism are carbon dioxide and water.
However, biodegradation of a compound is frequently a
stepwise process involving many enzymes and many
species of organisms. Therefore, in the natural
environment, a constituent may not be completely
degraded, but only transformed to intermediate product(s)
that may be less, equally, or more hazardous than the
parent compound, as well as more or less mobile in the
environment. Under anaerobic conditions (i.e., in the
absence of oxygen), metabolic activities result in the
formation of incompletely oxidized simple organic sub-
stances such as organic acids as well as other products
such as methane or hydrogen gas.
The goal of on-site bioremediation is degradation that
results in detoxification of a parent compound to a product
or product(s) that are no longer hazardous to human health
and/or the environment. Information on degradation and
detoxification of a parent compound may be obtained using
chemical and bioassay analyses [12, 13,14]. Chemical
analysis and identification of intermediate products may
yield information about biochemical degradation pathways
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and products, but are often time consuming and expensive.
Bloassays may be used to demonstrate detoxification of
parent compounds and are usually less expensive and time
consuming. Before bioremediation Is implemented at a
contaminated site, degradation pathways for specific
constituents present and/or detoxification demonstrations
require investigation to ensure that environmental and
health protection can be achieved.
Degradation of most organic compounds in soil systems
may be described by monitoring their disappearance in a
soil through time. Disappearance, or rate of degradation, is
often expressed as a function of the concentration of one or
more of the constituents being degraded. This is termed the
order of the reaction and is the value of the exponential
used to describe the reaction [15]. Either zero or first order
power rate models are often used in environmental studies.
Zero order reactions are ones in which the rate of
transformation of an organic constituent is unaffected by
changes in the constituent concentration because the
reaction rate is determined by some other factor than the
constituent concentration. If a constituent C is transformed
to X, the rate of change of C is:
dC/dt = -k
On integration, the equation becomes:
(1)
(2)
where C,» concentration of constituent remaining at time t;
C0«initial concentration of constituent; and k = zero order
rate constant. A useful term to describe the reaction kinetics
is the half-life, t,,, which is the time required to transform
50% of the initiafconstituent:
Ct« C0/2, then
=0/2k
(3)
The first order rate model (Equation 4) is widely used
because of its effectiveness in describing observed results
as well as its inherent simplicity. Its use also allows
comparison of results obtained from different studies. In a
first order rate reaction, the rate of transformation of a
constituent is proportional to the constituent concentration:
dC/dt = -kC
(4)
where C = contaminant concentration (mass/mass); t =
time; and k = first order rate constant (1/time). After
integration of Equation 4 and rearrangement of the
integrated equation, Equation 5 may be used to graphically
determine the rate constant, k:
ln(C/C0) = -kt
(5)
where C,» concentration of constituent remaining at time t;
and C0 = initial concentration of constituent. A plot of ln(C/
C0) versus t is linear with a slope of -k. The rate constant k
is independent of the concentration of constituent, since the
slope Is constant overtime. To calculate the time required
to transform one-half of the initial constituent (C, = C/2), the
following equation is used:
In ((Co/2)/C) = -KL
which is equal to:
t1/2 = 0.693/k
where t1/2= half-life of the constituent.
(6)
(7)
First order kinetics generally apply when the concentration
of the compound being degraded is low relative to the
biological activity in the soil. However, very low
concentrations may be insufficient to initiate enzyme
induction or support maintenance requirements necessary
for microbial growth, even if the compound can be used as
an energy source [16].
A second model used to describe degradation in soils is the
hyperbolic rate model, which is similar to Michaelis-Menten
enzyme kinetics. This model is expressed as:
dC/dt = -k
(8)
where k, and k2 are constants. The constant k, represents
the maximum rate of degradation that is approached as the
concentration increases. This model simulates a catalytic
process in which degradation may be catalyzed by
microorganisms.
Often an organic compound that cannot be used as a sole
carbon and energy source for microorganisms is degraded.
Biodegradation of the compound does not lead to energy
production or cell growth. This biodegradation process is
referred to as cometabolism [17] or co-oxidation if the
transformation involves an oxidation reaction [18].
Cometabolism occurs when an enzyme produced by an
organism to degrade one substance that supports growth
also degrades another nongrowth substrate that is neither
essential for, nor sufficient to, support microbial growth.
The nongrowth substrate is only incompletely oxidized, or
otherwise transformed, by the microorganism involved,
although other microorganisms may utilize by-products of
the cometabolic process. Cometabolism may be a
prerequisite for the mineralization of many recalcitrant
substances found in the environment, such as polycyclic
aromatic hydrocarbons [19].
Measurement of physical abiotic loss mechanisms and
partitioning of organic constituents in a soil should be used
in conjunction with conventional degradation studies to
ensure that information generated from modeling
degradation represents only biological degradation of parent
compounds, and not other possible disappearance
mechanisms of the constituents in the soil system.
The soil is a complex system, consisting of four phases
(Figure 1): (1) soil gases; (2) soil water; (3) inorganic
solids; and (4) organic solids. Gases and water, which are
found in the pore spaces of a soil, together comprise about
50% (by volume) of a typical soil. An organic constituent,
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Gas >&
15-35% /|i
Water
15-35%
Inorganic
38-45%
Figure 1. Typical Volumetric Composition of Soil.
depending upon its solubility and its tendency to volatilize,
may be found in varying proportions in these two phases.
Pore sizes and continuity and relative proportions of water
and air in the pores are examples of factors that affect the
mobility of contaminants (both upward out of the soil and
downward to the saturated zone) in a specific soil.
Depending upon site-specific soil characteristics and
constituent-specific chemical and physical properties,
constituents in these two phases may be relatively mobile
or immobile.
Soil solids are comprised of organic and inorganic
components. The inorganic components are comprised of
sparingly soluble chemicals known as minerals, which are
primarily sand, silt, and clay particles in most soils. The
solids may contain highly reactive charged surfaces that
play an important role in immobilizing organic constituents
in a specific soil. Certain types of clays are especially high
in negative charges, thus exhibiting what is termed as a
high cation exchange capacity. Clays may also contain
positively charged surfaces and act as anion exchange
media for negatively charged constituents.
Soil organic matter also has many highly reactive charged
surfaces and may aid in retaining organic constituents in a
soil system. The term humus refers to the relatively stable
portion of soil organic matter that remains in soil after the
chemicals comprising plant and animal residues have
decomposed. Hydrophobia organic constituents may
partition from soil water into soil organic matter and thus
become less mobile in the soil system. Immobilization of
constituents may result in additional time for biodegradatlon
to occur. However, immobilization also could result in less
bioavailability to microorganisms. Research is required to
discover whether such immobilization constitutes adequate
treatment if the constituent is so tightly and irreversibly
bound that it poses no harm to human health and the
environment.
Soil solid/organic chemical interactions may be quite
complex. The structure of an organic constituent, as it
affects such properties as molecular volume, water
solubility, octanol-water partition coefficients, and vapor
pressure, determines the magnitude of sorption onto the
surfaces of a specific soil. The specific aspects of chemical
structure that affect sorption onto soil surfaces, as
summarized by Dragun [20], include:
(1) molecu lar size- in general, the larger the molecule,
the greater its tendency to exist in the adsorbed
state. This is attributed to multiple Van der Waal's
forces arising from many points of contact between
the soil surface and the adsorbed molecule;
(2) hydrophobicity or lipophilicity-hydrophobicity refers
to the preferential migration to and accumulation of
an organic chemical in hydrophobic solvents or on
hydrophobia surfaces such as soil organic matter,
in preference to aqueous solvents or hydrophilic
surfaces. In general, molecular groups comprised
of carbon, hydrogen, bromine, chlorine, and iodine
are hydrophobic groups, while molecular groups
containing nitrogen, sulfur, oxygen, and
phosphorus are primarily hydrophilic groups. The
net hydrophobicity of a molecule is determined by
the combined effects of hydrophobic and
hydrophilic groups that comprise the molecule;
(3) molecular charge- some organic chemicals contain
functional groups with permanent positive negative
or positive charges. These compounds will interact
with charged soil solids and adsorb onto soil
surfaces. Soils typically possess a significantly
greater number of negative surfaces than positive
ones, thus negatively charged organic anions may
be repelled by soil surfaces. Some organic
chemicals contain functional groups that may or
may not possess a positive or negative charge,
depending upon the acidity of the soil/water
system. The pKa of a chemical is a mathematical
description of the effect of acidity on the charge of
the chemical. The relative ratio of charged to
uncharged molecules at a pH level In a soil/water
system may be estimated and used in identification
of the effect of a molecular charge on the extent of
adsorption. For chemicals that possess both types
of functional groups, i.e., ones that can acquire a
positive charge and ones that can acquire a
negative charge, the isoelectric point (IP) may be
used to predict the effect of pH on the adsorption
of these chemicals. The IP is the pH at which the
organic chemical has zero charge. Above the IP,
the organic chemical has a net negative charge;
below the IP, the organic chemical has a net
positive charge. The IP represents a general
summation of the effects of the pKas of each
functional group in the molecule;
(4) organic molecular functional groups that undergo
hydrogen bonding- hydrogen bonding occurs when
a hydrogen atom serves as a bridge between two
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electronegative atoms. The hydrogen atom is
linked to one electronegative atom by a covalent
bond and to the other by an electrostatic bond;
(5) three-dimensional arrangement and interaction of
molecular functional groups- adsorption potential of
a chemical Is affected by intramolecular reactions
of adjacent molecular groups or fragments or
Interference with a particular adsorption mech-
anism caused by the presence of one or more
functional groups or molecular fragments; and
(6) molecular functional groups that undergo coordina-
tion bonding- coordination is the formation of a
weak bond between an organic molecule that
Is capable of donating electrons and adsorbed
cations that are capable of accepting electrons.
The net result is a partial overlap of orbitals and a
partial exchange of electron density. Coordination
can occur between organic chemicals and cations
in the water phase of a soil system as well as with
soil particle surfaces and with adsorbed cations.
Many chemical properties of a specific organic chemical are
the result of sums and Interactions of functional group
contributions to each specific property.
Other abiotic loss mechanisms In addition to surface
sorptlon/desorptlon reactions that may account for loss of
parent compounds include:
(1) hydrolysis- a chemical reaction in which an organic
chemical reacts with water or a hydroxide ion;
(2) substitution and elimination- reactions where other
chemicals in the soil react with an organic
chemical;
(3) oxidation- the reaction resulting in the removal of
electrons from a chemical. This removal generally
occurs by two different pathways: (a) heterolytic or
polar reactions, (an electrophilic agent attacks an
organic molecule and removes an electron pair
leading to the formation of an oxidized product); or
(b) homolytlc or free-radical reaction (an agent
removes only one electron to form a radical that
undergoes further reaction); and
(4) reduction- which Is a reaction that results In a net
gain of electrons [20].
Successful bloremediation depends upon a thorough
characterization and evaluation of the pathways of move-
ment and potential mechanisms of removal of organic
constituents at a specific site, as illustrated in Figure 2. To
assess the potential for use of bioremediation, the rate of
transport of the constituents may be compared to the rate
of degradation to determine if the rate of transport is signi-
ficant in relation to the rate of degradation.
VOLATILIZATION
MINERALIZATION
BIOMASS
t
SOIL INTERACTIONS
PHASES:
SOLID LIQUID GAS
Figure 2. Fate of Hazardous Contaminants In Soil
A means of predicting rate of transport of a constituent
through a soil system is to describe its mobility (or relative
immobility) by predicting its retardation. Retardation Is a
factor that describes the relative velocity of the constituent
compared to the rate of movement of water through the
soil, i.e.,:
R = VW/VC (9)
where R = retardation factor; Vw = average water velocity;
and Ve = average constituent velocity. A retardation factor
greater than one indicates that a constituent is moving more
slowly than water through a soil. A factor developed from a
transport model combined with a description of sorption
processes, as defined by a linear Freundlich isotherm [21,
22], can be calculated from the following equation:
= 1+(pKd/9)
(10)
where p = soil bulk density; Kd = soil water partition
coefficient, which describes the partitioning between the soil
solid phase and soil water; and 9 = volumetric moisture
content. This information can be used to manage a
contaminated soil system (i.e., through control of soil
moisture, changes in bulk density, or addition of
amendments to the soil that affect the soil water partition
coefficient) so that constituents can be "captured" or
contained within the system, thus allowing time for
implementation and performance of bioremediation
treatment techniques.
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3. Waste and Soil Charac-
terization
Interfacing "soil-based behavioral characteristics" of specific
organics with specific site and soil properties allows a
determination of potential for bioremediation of a site and
potential for contamination of other media, i.e., the ground
water under the contaminated area, the atmosphere over
the site or at the site boundaries, surface waters, etc.
Specific characteristics important for describing and
assessing the environmental behavior and fate for organic
constituents in soil are listed in Table 1. For each chemical,
or chemical class, the information required can be
summarized as: (1) characteristics related to potential
leaching, e.g., water solubility, octanol/water partition
coefficient, solid sorption coefficient; (2) characteristics
related to potential volatilization, e.g., vapor pressure,
relative volatilization index; (3) characteristics related to
potential biodegradation, e.g., half-life, degradation rate,
biodegradability index; and (4) characteristics related to
chemical reactivity, e.g., hydrolysis half-life, soil redox
potential [21].
An adequate site characterization, including surface soil
characteristics, subsurface hydrogeology, and micro-
biological characteristics, is the basis for the rational design
Table 1. Soil-Based Waste Characterization [21]
Chemical Class
Soil Sorption
Parameters
of a bioremediation system. Site constraints may limit rate
and/or extent of treatment of the contaminated vadose
zone; therefore, a thorough site characterization is
necessary to determine both the three-dimensional extent
of contamination as well as engineering constraints and
opportunities.
Important soil hydraulic, physical, and chemical properties
that affect the behavior of organic constituents in the
vadose zone are presented in Table 2. In this zone, water
primarily coexists with air, though saturated regions may
occur. Perched water tables may develop at interfaces of
layers with differing textures. Prolonged infiltration may
also result in saturated conditions. The vadose zone
usually consists of topsoils, typically three to six feet deep,
which are weathered geological materials, arranged in
more or less well developed profiles. Water movement in
the vadose zone is usually unsaturated, with soil water at
less than atmospheric pressure. Weathered topsoil
materials gradually merge with underlying earth materials,
which may include residual or transported clays or sands.
The topsoil differs from the material lying below it in that it is
more weathered, contains organic matter, and is the zone
of plant root growth. In some regions, the entire vadose
zone may be hundreds of feet thick and the travel time of
constituents to ground water hundreds or thousands of
years. Other regions may be underlain by shallow potable
aquifers that are especially susceptible to contamination
due to short transport times and reduced potential of soil
materials and processes for pollutant attenuation.
Soil Degradation
Parameters
Chemical
Properties
Acid
Base
Polar neutral
Nonpolar neutral
Inorganic
Freundlich sorption constants
(K.N)
Sorption based on organic
carbon content (K )
Octanol water partition co-
efficient (K)
Halffllfe (t1/2)
Rate constant (first order)
Relative biodegradability
Molecular weight
Melting point
Specific gravity
Structure
Water solubility
Volatilization Parameters
Chemical
Reactivity
Soil Contamination Parameters
Ainwater partition coefficient (KJ
Vapor pressure
Henry's law constant (1/KW)
Sorption based on organic carbon
content (Koc)
Water solubility
Oxidation
Reduction
Hydrolysis
Precipitation
Polymerization
Concentration in soil
Depth of contamination
Date of contamination
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Table 2. Site and Soil Characteristics Identified as
Important In In Situ Treatment [21]
(3) detection and measurement of metabolic
processes [15].
Site location/topography and slope
Soil type, and extent
Soil profile properties
boundary characteristics
depth
texture*
amount and type of coarse fragments
structure*
color
degree of mottling
bulk density*
day content
type of day
cation exchange capacity*
organic matter content*
pH*
Eh*
aeration status*
Hydraulic properties and conditions
soil water characteristic curve
field capacity/permanent wilting point
water holding capacity*
permeability* (under saturated and a range of
unsaturated conditions)
infiltration rates*
depth to Impermeable layer or bedrock
depth to groundwater,* including seasonal variations
flooding frequency
runoff potential*
Geological and hydrogeologlcal factors
subsurface geological features
groundwater flow patterns and characteristics
Meteorological and climatological data
wind velocity and direction
temperature
precipitation
water budget
'Factors that may be managed to enhance soil treatment
Microbiological characterization of a contaminated site
should be conducted to ensure that the site has a viable
community of microorganisms to accomplish biodegradation
of the organic constituents present at the site. Approaches
for estimating the kinds, numbers, and metabolic activities
of soil organisms include:
(1) determination of the form, arrangement, and
biomass of microorganisms in the soil;
(2) isolation and characterization of subgroups and
species; and
Examples of techniques to accomplish these activities
include direct microscopy of soil (e.g., fluorescent staining,
buried-slide technique), biomass measurement by chemical
techniques (e.g., measurement of ATP), measurement of
enzyme activity, and cultural counts of microorganisms
(e.g., plate counts, dilution counts, isolation of specific
organisms). Biotransformation studies that measure the
disappearance of contaminants or mineralization studies
that indicate complete destruction of contaminants to
carbon dioxide and water may be used to confirm the
potential for biodegradation of specific organic chemicals.
Specific techniques include batch culture and electrolytic
respirometer studies. Controls to detect abiotic
transformation of the contaminants and tests to detect toxic
effects of contaminants on microbial activity should be
included in the studies.
Information from waste and soil/site characterization studies
of a specific site and from laboratory evaluations of
biodegradation and immobilization potential of specific
constituents at the site may be integrated by the use of
predictive mathematical models. The resulting msithematical
description may be used to: (1) evaluate the effectiveness
of use of on-site bioremediation for treatment of the
contaminated soil; (2) develop appropriate containment
structures to prevent unacceptable waste transport from the
treatment zone; and (3) design performance monitoring
strategies.
4. Microbial Factors
Affecting Biodegradation
The upper layers of soil contain large numbers and diversity
of microorganisms. Biodegradation of organic constituents
is accomplished by enzymes produced by the
microorganisms. Since many enzymes are not released by
microbial cells, substances to be degraded must contact or
be transported into the cells. Enzymes are generally
specific in the substances they affect, so many types may
be required to complete biodegradation of organic
constituents. The production of enzymes is genetically
controlled, thus mutations and adaptations of the native soil
microbial populations can improve the ability of the
populations to degrade organic substances [23].
Microbial ecologists have identified ranges of critical
environmental conditions that affect the activity of soil
microorganisms (Table 3). Many of these conditions are
controllable and can be changed to enhance biodegradation
of organic constituents.
Water is necessary for microbial life, and the soil water
matric potential against which microorganisms must extract
water from the soil regulates their activity. (The soil matric
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Table 3. Critical Environmental Factors for
Mlcroblal Activity [15,21, 23]
Environmental Factor
Optimum Levels
Available soil water
Oxygen
Redox potential
PH
Nutrients
Temperature
25 - 85% of water holding
capacity; -0.01 MPa
Aerobic metabolism: Greater
than 0.2 mg/1 dissolved
. oxygen, minimum air-filled
pore space of 10%;
Anaerobic metabolism: O2
concentrations less than 1%
Aerobes and facultative
anaerobes: greater than
50 millivolts; Anaerobes:
less than 50 millivolts
5.5 - 8.5
Sufficient nitrogen,
phosphorus, and other
nutrients so not limiting
to miorobial growth
(Suggested C:N:P ratio of
120:10:1)
15-45°C(Mesophiles)
potential is the energy required to extract water from the
soil pores to overcome capillary and adsorptive forces). Soil
water also serves as the transport medium through which
many nutrients and organic constituents diffuse to the
microbial cell, and through which metabolic waste products
are removed. Soil water also affects soil aeration status,
nature and amount of soluble materials, soil water osmotic
pressure, and the pH of the soil solution [15].
Microbial respiration, plant root respiration, and respiration
of other organisms remove oxygen from the soil
atmosphere and enrich it with carbon dioxide. Gases diffuse
into the soil from the air above it, and gases in the soil
atmosphere diffuse into the air. However, oxygen
concentration in a soil may be much less than in air while
carbon dioxide concentrations may be many times that of
air. Even so, a large fraction of the microbial population
within the soil depends on oxygen as the terminal electron
acceptor in metabolism. When soil pores become filled with
water, the diffusion of gases through the soil is restricted.
Oxygen may be consumed faster than it can be replaced by
diffusion from the atmosphere, and the soil may become
anaerobic. Clay content of soil and the presence of organic
matter also may affect oxygen content in soil. Clayey soils
tend to retain a higher moisture content, which restricts
oxygen diffusion, while organic matter may increase
microbial activity and deplete available oxygen. Loss of
oxygen as a metabolic electron acceptor Induces a change
in the activity and composition of the soil microbial
population. Facultative anaerobic organisms, which can use
oxygen when it is present or can switch to alternative
electron acceptors such as nitrate or sulfate in the absence
of oxygen, and obligate anaerobic organisms become the
dominant populations.
Another soil parameter that describes the effect of the soil
environment on metabolic processes is the redox potential
of the soil [15]. Biological energy is obtained from the
oxidation of reduced materials. Electrons are removed from
organic or inorganic substrates to capture the energy that is
available during the oxidative process. Electrons from
reduced compounds are moved along respiratory or
electron transport chains composed of a series of
compounds. In an aerobic process, O2 acts as the terminal
electron acceptor. In some cases where Q2 is not
available, nitrate (NO3"), iron (Fe3*), manganese (Mn2*), and
sulfate (SO42-) can act as electron acceptors if the
organisms have the appropriate enzyme systems. A
measurement of the oxidation-reduction potential (redox
potential) of a soil provides a measurement of the electron
density of the system. As a system becomes reduced, O2
is depleted, and other substances are used as terminal
electron acceptors. There is a corresponding increase In
electron density, resulting in a progressively increased
negative potential. Redox potential is measured as Eh,
expressed in millivolts, or as PE, which is equal to -log [e~]
where [e-] is the concentration of negatively charged
electrons.
Oxygen levels in a soil system can be maintained by:
(1) prevention of saturation with water;
(2) presence of sandy and loamy soil materials
(excessive clay contents are undesirable);
(3) moderate tilling;
(4) avoidance of compaction of soil; and
(5) limited addition of additional carbonaceous
materials [23].
Soil pH also affects the activity of soil microorganisms.
Fungi are generally more tolerant of acidic soil conditions
(below pH 5) than are bacteria. The solubility of
phosphorus, an important nutrient in biological systems, is
maximized at a pH value of 6.5. A specific contaminated
soil system may require management of soil pH to achieve
levels that maximize microbial activity. Control of pH to
enhance microbial activity may also aid in the immobilization
of hazardous metals in a soil system (a pH level greater
than 6 is recommended to minimize metal transport).
Microbial metabolism and growth is dependent upon
adequate supplies of essential macro- and micronutrients.
Required nutrients must be present and available to
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microorganisms In: (1) a usable form; (2) appropriate
concentrations; and (3) proper ratios [20]. If the wastes
present at the site are high in carbonaceous materials and
low In nitrogen (N) and phosphorus (P), the soils may
become depleted of available N and P required for
blodegradation of the organic constituents. Fertilization
may be required at some contaminated sites as a
management technique to enhance microblal degradation.
Blodegradation of organic constituents declines with
lowering of soil temperature due to reduced microbial
growth and metabolic activity. Biodegradatlon has been
shown to essentially stop at a temperature of 0° C. Soils
exhibit a variation in the temperature of the surface layers,
both diumally and seasonally. Diurnal changes of
temperature decrease with depth of the soil profile. Due to
the high specific heat of water, wet soils are less subject to
large diurnal changes than dry soils [15]. Factors that affect
soil temperature include soil aspect (direction of slope),
steepness of slope, degree of shading, soil color, and
surface cover.
The environmental factors presented in Table 3, as well as
soli and waste characteristics, interact to affect microbial
activity at a specific contaminated site. Computer modeling
techniques are useful to attempt to describe the interactions
and their effects on treatment of organic constituents in a
specific situation.
5. Treatabijity Studies for
Determination of Biore-
mediation Potential
Treatability studies for sites contaminated with organic
wastes are used to provide specific information concerning
the potential rate and extent of bioremediation of surficial
soil and deeper vadose zone soils by providing information
on fate and behavior of organic constituents at a specific
contaminated site. Treatability studies can be conducted in
laboratory microcosms, at pilot scale facilities, or in the field.
To determine whether a specific site is suitable for
bioremediation, Information from treatability studies is
combined with information concerning site and waste
characteristics In order to determine potential applications
and limitations of the technology. Ultimate limitations to the
use of bioremediation at a specific site are usually related
to: (1) time required for cleanup, (2) level of cleanup
attainable, and (3) cost of cleanup using bioremediation.
Information from treatability studies also is used to prepare
an approach to the engineering design and implementation
of a bioremediation system at a specific site. An
engineering design to accomplish bioremediation at the site
Is generally based upon information from simulations (e.g.,
mathematical modeling) or estimates of pathways of
migration of chemicals. These simulations or estimates are
Table 4. Materials Balances and Mineralization
Approaches to Blodegradation Assessment
Biodegradalion Approach Process Examined
Materials balances
Mineralization
Recovery of parent compound
in the air, soil water, soil
solids (extractable)
Recovery of transformation
products in the air, soil
water, and soil solids
(extractable)
Production of carbon dioxide,
and/ or methane from the
parent compound
Release of substituent groups,
e.g., chloride or bromide ions
generated from treatability data and site/soil charac-
terization data in order to: (1) determine containment
requirements to prevent contamination of off-site receiver
systems; (2) develop techniques to maximize mass transfer
of chemicals affecting microorganism activity (addition of
mineral nutrients, oxygen, additional energy sources, pH
control products, etc.; removal of toxic products) in order to
enhance bioremediation; and (3) design a cost-effective and
efficient monitoring program to evaluate effectiveness of
treatment.
During the performance of a treatability study, biodegrada-
tion, detoxification, and partitioning (immobilization)
processes are evaluated as they affect the fate and
behavior of organic constituents in the soil.
To assess the potential for biological degradation at a
specific contaminated site, the use of treatability studies
incorporating materials balance and mineralization
approaches to determine the environmental fate and
behavior of the constituents in the specific soil is recom-
mended (Table 4). Rate of degradation is calculated by
measuring the loss of parent compound and the production
of carbon dioxide with time of treatment. Degradation rate
is often reported as half-life, which represents the time
required for 50 percent of the compound to disappear
based upon a first-order kinetic model.
Calculation of the rate of decrease of parent compound,
however, by itself does not provide complete information
concerning mechanisms and pathways by which organic
constituents are interacting with the soil environment [24].
Further information is necessary to understand whether a
constituent is simply transferred from one phase (e.g., solid
phase) to another (e.g., air phase) through a process of
interphase transfer, or is chemically altered so that the
-------
Influent
/"purge Gas
Sorbent
Tubas
Effluent Purge Gas
Figure 3. Laboratory Flask Apparatus Used for Mass Balance
Measurements.
properties of the parent compound are destroyed.
Evaluation of the fate of a constituent in a soil therefore also
requires identification and measurement of the distribution
of the constituent among the physical phases that comprise
the system as well as differentiation of the mechanisms by
which constituent may be chemically altered in a soil
system.
A laboratory flask apparatus that can be used as a
microcosm to measure interphase transfer and
biodegradation potential in a laboratory treatability study is
illustrated in Figure 3. The contaminated soil material is
placed in a flask, which is then closed and incubated under
controlled conditions for a period of time. During the
incubation period, air is drawn through the flask and then
through a sorbent material. Volatilized materials are
collected by the sorbent and are measured to provide an
estimate of volatilization loss of the constituents of interest.
At the end of the incubation period, a portion of the
contaminated soil is treated with an extracting solution to
determine the extent of loss of the constituents in the soil
matrix. This loss can be attributed to biodegradation and
possible immobilization in the soil materials. Selection of an
appropriate extracting solution is necessary to maximize
constituent recovery from the soil. Another portion of the
soil is leached with water to determine leaching potential of
remaining constituents. Abiotic processes involved in
removal of the parent compound are also evaluated by
comparing microbially active soil/waste mixtures with
mixtures that have been treated with a microbial poison,
e.g., mercuric chloride or propylene oxide. The use of a
procedure incorporating features illustrated by the use of
this microcosm is crucial in order to obtain a materials
balance of waste constituents in the soil system. Examples
of such protocols may be found in [14, 24, 25, 26]. A
certain amount of material is added to the soil, and tracking
the fate of the material as it moves through the multiple
phases of the soil system provides a materials balance.
Transformation refers to the partial alteration of hazardous
constituents into intermediate products. Intermediate
products may be less toxic or more toxic than the parent
compound, and therefore the rate and extent of
detoxification of the contaminated material should be
evaluated. Samples generated from the different phases of
the soil system in the microcosm studies can be analyzed
for intermediate degradation products and used in
bioassay studies to provide information concerning
transformation and detoxification processes.
Bioassays to quantify toxicity measure the effect of a
chemical on a test species under specified test conditions
[14]. The toxicity of a chemical is proportional to the
severity of the chemical on the monitored response of the
test organism(s). Toxicity assays utilize test species that
include rats, fish, invertebrates, microorganisms, and
seeds. The assays may utilize single or multiple species of
test organisms. The use of a single bioassay procedure
does not provide a comprehensive evaluation of the
toxicity of a chemical in the soil/organic chemical-impacted
system. Often a battery of bioassays is utilized that may
include measurements of effects on general microbial
activity (e.g., respiration, dehydrogenase activity) as well
as assays relating to activity of subgroups of the microbial
community (e.g., nitrification, nitrogen fixation, cellulose
decomposition). Bioassays utilizing organisms from
different ecological trophic levels may also be used to
determine toxicological effects. However, use of a single
assay as a screening test to identify relative toxicity
reduction in the environment is a common procedure
employed in treatability studies. Assays using
microorganisms are often used due to their speed,
simplicity, ease in handling, cost effectiveness, and use of
a statistically significant number of test organisms that is
required to detect the effects of potentially toxic materials
in the environment [27, 28].
Two microbial bioassays that have been used to evaluate
toxicity of wastes in soil systems are the Ames Salmonella
typhlmurium mammalian microsome assay and the
Microtox™ test system. The Ames assay is a measure of
the mutagenic potential of hazardous compounds [29, 30]
and has been widely used to evaluate environmental
samples [31, 32, 33, 34, 35]. A high correlation has been
shown between carcinogenicity and mutagenicity, where
about 90% of known carcinogens tested mutagenic in the
Ames assay [36]. Special strains of Salmonella
typhimurium that require histidine to grow are used to test
for mutagenicity. When plated on a histidine-free medium,
the only bacteria able to form colonies are those that have
reverted to the "wild" state and are able to produce their
own histidine. Without the addition of test chemicals, this
back mutation occurs at a rate specific to each strain type
(spontaneous reversion rate). The addition of chemicals
that are mutagenic increases the reversion rate. Several
dose levels of a chemical, mixture of chemicals, or an
environmental sample are added to obtain a dose
response. Some mutagens act directly on the bacterial
cells while others require activation by mammalian
-------
mlcrosomes. These mlcrosomes are generally obtained
from liver extracts of Aroclor 1254-induced rats (I.e., rats
Injected with the polychlorinated biphenyl (PCB), Aroclor
1254). The extract, referred to as the S-9 fraction, contains
enzymes that metabollcally convert certain chemicals to
active mutagens, simulating the activity that occurs in living
mammalian systems. Several strains of Salmonella
typhlmurlumhave been developed in order to detect
different types of mutagens. The recommended strains for
general mutagenicity testing include TA97, TA98, TA100,
TA102. TA97 and TA98 detect frameshift mutagens.
TA100 detects mutagens causing base-pair substitutions,
while TA102 detects a variety of mutagens not detected by
the other strains.
The Microtox™ assay is an aqueous general toxicity assay
that measures the reduction in light output produced by a
suspension of marine luminescent bacteria in response to
an environmental sample [37]. Bioluminescence of the test
organism depends on a complex chain of biochemical
reactions. Chemical Inhibition of any of the biochemical
reactions causes a reduction in bacterial luminescence.
Therefore, the Microtox™ test considers the physiological
effect of a toxicant and not Just mortality. Matthews and
Bulich [38] have described a method of using the
Microtox™ assay to predict the land treatability of
hazardous organic wastes. Matthews and Hastings [39]
described a method using the Microtox™ assay to
determine an appropriate range of waste application
loading for soil-based treatment systems. Symons and
Sims [40] utilized the assay to assess the detoxification of
a complex petroleum waste in a soil environment. The
assay was also included as a recommended bioassay in
the U.S. EPA Permit Guidance Manual on Hazardous
Waste Land Treatment Demonstrations [25].
Immobilization refers to extent of retardation of the
downward transport (leaching potential) and upward
transport (volatilization potential) of waste constituents.
Interphase transfer potential for waste constituents among
soil, oil (waste), water, air, and solid (organic and inorganic)
phases Is affected by the relative affinity of the waste
constituents for each phase, and may be quantified through
calculation of partition coefficients [25]. Partition
coefficients are calculated as the ratio of the concentration
of a chemical in the soil, oil, or air phase to the
concentration of a chemical in the water phase, and are
expressed as Ko (oil/water), Kh (air/water), and Kd (solid/
water). Calculation of retardation factors (Equations 9 and
10) also may be used to predict immobilization of
constituents In a soil system [22,41].
Either laboratory microcosm, pilot scale reactors, or field
plots may be used to generate treatability data. The set of
experimental conditions, e.g., temperature, moisture, waste
concentration, etc., under which the studies were
conducted should be presented along with experimental
results.
Treatability study results provide information relating to
rates and extent of treatment of hazardous organic
constituents when mass transfer rates of potential limiting
substances are not limiting the treatment. Treatability
studies usually represent optimum conditions with respect to
mixing, contact of soil solid materials with waste constituents
and with microorganisms, and homogeneous conditions
throughout the microcosm. Therefore, treatability studies
provide information concerning potential levels of treatment
achievable at a specific site. Under field conditions, the rate
and extent of bioremediation is generally limited by
accessibility and rate of mass transfer of chemical
substances (oxygen, nutrients, etc.) to the contaminated soil
as well as by mass transfer of the contaminants to the
microbial population and removal of microbial degradation
products.
6. Integration of Information
from Site Characterization
and Treatability Studies
Information from the performance of site characterization
and treatability studies may be integrated with the use of
comprehensive mathematical modeling. In general, models
are used to analyze the behavior of an environmental
system under both current (or past) conditions and
anticipated (or future) conditions [42]. A mathematical model
provides a tool for Integrating degradation and partitioning
processes with site/soil- and waste-specific characterization
for simulating the behavior of organic constituents in a
contaminated soil and for predicting the pathways of
migration through the contaminated area, and therefore
pathways of exposure to humans and to the environment.
Models may also be used to approximate and estimate the
rates and extent of treatment that may be expected at the
field scale under varying conditions. DiGiulio and Suffet [43]
have presented guidance on the selection of appropriate
vadose zone models for site-specific applications, focusing
on recognition of limitations of process descriptions of
models and difficulties in obtaining input parameters
required by these process descriptions.
The Regulatory and Investigative Treatment Zone Model
(RITZ Model, developed at the U.S. EPA Robert S. Kerr
Environmental Research Laboratory by Short [44]) is an
example of a model that has been used to describe the
potential fate and behavior of organic constituents in a
contaminated soil system [45]. The Ritz Model is based on
an approach by Jury [46]. An expanded version of RITZ, the
Vadose Zone Interactive Processes (VIP) model,
incorporates predictive capabilities for the dynamic behavior
of organic constituents in unsaturated soil systems under
conditions of variable precipitation, temperature, and waste
loadings [25, 47, 48, 49, 50]. Both models simulate vadose
zone processes, including volatilization, degradation,
sorption/desorption, advection, and dispersion [51].
10
-------
Once interphase transfer potential and pathways of escape
have been identified by treatability studies and simulation
modeling, containment requirements for the constituents of
interest at the site can be determined. If the major pathway
of transport is volatilization, containment with respect to
volatilization control is required. An inflatable plastic dome
erected over a contaminated site is a containment method
that has been used to control escape of volatile
constituents. Volatiles are drawn from the dome through a
conduit and treated in an above ground treatment system. If
leaching has been Identified as important, control of soil
water movement should be implemented. For example, if
contaminated materials are expected to leach downward
from the site, the contaminated materials can be
temporarily removed from the site, and a plastic or clay liner
placed under the site. When downward, as well as upward,
migration is significant, both volatilization and leaching
containment systems can be installed. Some hydrophobic
chemicals dp not tend to volatilize or to leach but are
persistent within the soil solid phase; therefore, containment
efforts may not be required.
A critical and cost-effective use of modeling is in the
analysis of proposed or alternative future conditions, i.e.,
the model is used as a management or decision-making
tool to help answer "what if" type questions [42]. Attempting
to answer such questions through data collection programs
would be expensive and practically impossible in many
situations. For example, information can be generated to
evaluate the effects of using different approaches for
enhancing microbial activity and for accelerating
biodegradation and detoxification of the contaminated area
by altering environmental conditions that affect microbial
activity.
Results of modeling also can aid in the identification of
constituents that will require treatment in the air (volatile)
phase, in the leachate phase, and in the solid (soil) phase.
Monitoring efforts therefore can be concentrated on
monitoring the appropriate environmental phase to evaluate
treatment effectiveness. If a comprehensive and thorough
evaluation of a specific contaminated system has been
conducted, not all chemicals need to be monitored in each
phase.
7. Potential Applications
and Limitations of
Bioremediation Technology
Existing information for constituents of interest at a specific
site/soil contaminated system should be collected as a first
step in the investigation of the application of bioremediation
as a potential treatment technology. Many organic
constituents from a wide range of chemical classes have
been shown to be amenable to biodegradation in laboratory
studies, using both single strains of microbial species or
consortia of microbial populations. Biodegradation has also
been demonstrated in both aqueous cultures or soil
microcosm studies. A summary of biodegradation and
disappearance rates for almost 300 chemicals has been
prepared by Dragun [20]. Examples of specific chemical
classes shown to be biodegradable include: amines and
alcohols [14]; polycyclic aromatic hydrocarbons (PAHs) [5,
12,13, 52, 53]; chlorinated and non-chlorinated phenols
[14]; chlorinated aromatic hydrocarbons [54];
polychlorinated biphenyls (PCBs) [55], halogenated aliphatic
compounds [50, 57]; pesticides [13, 47, 58, 59, 60, 61]; and
various hazardous substances [13, 62]. Industrial wastes
from petroleum refining, wood preserving, leather tanning,
coal gasification/liquefaction, food processing, pulp and _
paper manufacturing, organic chemical production, animal
production, munitions production, textile manufacturing,
pesticide manufacturing, and pharmaceutical
manufacturing, as well as municipal wastewaters, sludges,
and septage from septic tanks, have all been successfully
treated in land treatment systems [14,20].
RSKERL, as part of its responsibilities to manage research
programs to determine the fate, transport, and
transformation rates of pollutants in the soil, the
unsaturated and the saturated zones of the subsurface
environment, initiated a research program to develop
comprehensive screening data on the treatability in soil of
specific listed hazardous organic chemicals and specific
listed hazardous wastes. Research results have been
presented by Sims et al. [13], Loehr [14], and McGinnis et
al. [5]. A Soil Transport and Fate (STF) Data Base was
also developed for RSKERL [63]. The Data Base contains
quantitative and qualitative information on degradation,
transformation, partitioning among the soil phases, and
toxicity of hazardous organic constituents in soil systems. It
may be used as a tool for contaminated site assessment
and remediation activities. The Data Base provides input
data concerning degradation rates, partition coefficients,
and chemical property data for mathematical models
simulating the behavior and fate of chemical constituents in
contaminated surface and subsurface soils. The information
is also useful for providing assistance in determining
treatment potential at contaminated sites using in situ
techniques. Chemicals may be evaluated with respect to
the importance of natural processes in controlling
persistence and transport potential, and therefore, the
susceptibility to degradation or retardation within a
subsurface environment.
A report was prepared for the U.S. EPA evaluating the
effectiveness of soil treatment practices at Superfund sites,
EPA Office of Research and Development tests,
Department of Defense and Department of Energy studies,
state remediation efforts, private party studies, and vendor
demonstrations [64, 65]. Bioremediation was shown to
successfully treat many non-halogenated compounds, but
was less successful with halogenated compounds.
Removal efficiencies for non-halogenated aromatics,
heterocyclics and other polar compounds were greater than
95%. Halogenated aliphatic compounds were also
11
-------
successfully treated, with removal efficiencies averaging
98%; however, volatilization may have contributed to
observed losses. More complex halogenated and nitrated
compounds exhibited lower removal efficiencies, ranging
from 50 to 85%.
Even though a specific organic constituent has been shown
to blodegrade under laboratory conditions, whether or not it
will degrade In a specific soil/site system is dependent on
many factors [54]. Potential degradability requires
Investigation In site-specific treatability studies. Available
oxygen may be limiting In some cases, while other
compounds may require the presence of anaerobic
conditions. Other environmental conditions that may place
restrictions on biological activity include pH, temperature,
and moisture. Upon exposure to the soil environment, the
constituent may be biologically or chemically altered so as
to be rendered persistent and/or toxic in the environment.
The system may lack other nutrients required for microbial
activity. Other chemicals present may serve as preferred
substrates, or act to repress required enzyme activities.
High concentrations of metal salts may be inhibitory or toxic
to many microorganisms.
Most chemicals require the presence of a consortium of
microbial species for mineralization, some of which may not
be present at the specific site. Also, most organisms require
a period of acclimation to the constituent before metabolism
occurs. During this period, the level of constituent must be
high enough to promote acclimation without being toxic or
inhibitory. Prior exposure to the constituent or similar
constituents may help to shorten the acclimation period.
8. Example of Bioremed-
iation Potential for
Polycyclic Aromatic
Hydrocarbons (PAHs) in
a Soil System
To demonstrate the potential effectiveness of
bioremediation, results are presented for the semi-volatile
chemical class of compounds known as the polycyclic
aromatic hydrocarbons (PAHs). These compounds are of
environmental significance because of their recalcitrance
to biological degradation, their chronic toxic effects on
humans, and their widespread occurrence at contaminated
waste sites. Specifically, PAH compounds are associated
with oily wastes, such as wastes from petroleum refining
operations and wastes from the wood preserving industry.
The higher molecular weight PAH compounds are of
special concern, because they exhibit mutagenic,
carcinogenic, and teratogenic potential.
Table 5. Degradation of PAHs Present fn a Complex Oily Waste, Applied at 2% Oil and Grease
In Clay Loam Soli [66]
Compound C0*
days
95% Confidence Interval (t1/2)
(days)
Lower
Upper
Fluor-
anthene
Pyrene
351
283
15
32
0.966
0.884
13
26
18
41
Benzo(a)
anthracene
Benzo(g,h,
l)perylene
Indeno-
pyrene
86
8
5
139
1661
69
0.397
0.006
0.559
87
139
43
347
ND
139
Initial Concentration
Half-life (first order kinetics)
12
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Table 6. Effect of Manure and pH Amendments on
PAH Degradation in a Complex Waste
Incorporated into Soil [67]
Half-Life in Waste/Soil Mixture (days)
PAH Compound
Without
Amendments
With
Amendments
Acenaphthylene 78
Acenaphthene 96
Fluorene 64
Phenanthrene 69
Anthracene 28
Fluoranthene 104
Pyrene 73
Benz(a)anthracene 123
Chrysene 70
Benzo(b)fluoranthene 85
Benzo(k)fluoranthene 143
Benzo(a)pyrene 91
benzo(ghi)perylene 74
Dibenz(a,h)anthracene 179
lndeno(1,2,3-cd)pyrene 57
14
45
39
23
17
29
27
52
42
65
74
69
42
70
42
Table 7. Effect of Soil Moisture on PAH Degradation [67]
Half-life in Waste/Soil Mixture (Days)
Moisture
Anthra-
cene
Phenan-
threne
Fluoran-
thene
20-40% field capacity 43 61 559
60-60% field capacity 37 54 231
The degradation of PAH compounds in soils has been
demonstrated in laboratory treatability studies [66]. The
results presented in Table 5 for PAH compounds present
in a complex oily waste show that the half-lives for four of
the five compounds ranged from only 15 to 139 days.
However, the half-life for benzo(g,h,i)perylene, a higher
molecular weight PAH compound, was still quite long
(1661 days). McGinnis et al. [5] in a laboratory soil
treatability study of PAH compounds present in creosote
waste sludges also found that degradation of PAH was
dependent on molecular weight and number of aromatic
rings. PAHs with two rings generally exhibited half-lives
less than ten days, while three- ring compounds in most
cases exhibited longer half-lives, which were usually less
than one hundred days. Most of the four- or five-ring PAHs
exhibited half lives of one hundred days or more. The
results of these two studies suggest that means of
enhancing biological degradation of more recalcitrant PAH
compounds should be investigated.
When additional carbon and energy sources were provided
and soil pH was adjusted from 6.1 to 7.5. the half-lives of
PAH waste constituents present in a complex fossil fuel
waste added to a soil were decreased, as shown In Table 6
[67]. In this laboratory study using first-order kinetic
modeling of degradation, the use of manure as an
amendment and control of soil pH significantly decreased
the t,_ of the PAH constituents studied. For example, the
half-life of phenanthrene decreased from 69 to 23 days,
benz(a)anthracene from 123 to 52 days, and benz(a)pyrene
from 91 to 69 days.
The control of soil moisture also resulted in enhanced
biodegradation of PAHs, as shown in Table 7 [67]. Soil
moisture in this study was described in terms of percent of
field capacity. Field capacity is defined as the percentage
of soil moisture remaining in a soil after haying been
saturated and after free drainage has practically ceased.
Therefore, soils with moisture levels of 60 to 80% of field
capacity are wetter than soils with levels of 20 to 40% of
field capacity. At higher levels of soil moisture, the half-life
of the PAH constituents studied decreased. For example,
forfluoranthene, the half-life decreased from 559 days to
231 days. At a specific site where containment has been
achieved, the addition and control of soil moisture may be a
tool to accomplish faster degradation of the constituents.
An increase in soil temperature also can decrease the time
required to accomplish degradation, especially the loss of
lower molecular weight PAHs [68]. In a laboratory study,
for example, the half-life of fluorene decreased from 60
days to 47 days to 32 days at 10°, 20°, and 30° C,
respectively (Table 8). At a field site, soil temperature may
be difficult to control. However, if a cover is used at the site
to control the release of volatile materials, an increase in
soil temperature may also occur. Seasonal climatic changes
will affect the rate of degradation of organic constituents, as
well as geographical location of a specific contaminated
site.
If a soil has been exposed previously to similar or the same
type of contamination, the soil microbial population may
have become acclimated to the waste, and waste
degradation may occur at a faster rate. In a laboratory
study investigating the acclimation of a soil to a fossil fuel
waste, a greater reduction in concentration of all the waste
PAH compounds studied was achieved in 22 days in an
acclimated soil, compared to the reduction seen in 40 days
in an unacclimated soil (Table 9) [67]. These results show
that at a site that has been contaminated for a period of
time, the indigenous microbial population may become
acclimated to the presence of wastes, and techniques to
stimulate microbial activity may produce significant
degradation. Mixing of a small amount of a contaminated
soil that has developed an acclimated population with the
13
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Table 8. Percentages of PAH Remaining at the End of the 240 Day Study Period and
Estimated Apparent Loss Half Lives [68]
Compound
Acenaphthene
Fluorene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Benzo(a)pyrene
Dibenz(a,h)anthracene
Benzo(g,h,I)perylene
lndeno(1 ,2,3-c,d)pyrene
10°C
5
8
36
83
94
93
82
85
77
93
73
88
81
80
Percent of PAH
Remaining
20°C 30°C
0
3
19
51
71
89
71
88
75
95
54
87
76
77
0
2
2
58
15
43
50
86
62
89
53
83
75
70
Estimated Half Life (day)*
10°C 20°C 30°C
<60
60(+11/-10)
200(+40/-40)
460{+310/-140)
+
+
680(+300/-160)
980(+520/-270)
580(+520/-180)
910{+690/-270)
530(+1700/-230)
820(+1100/-300)
650(+650/-230)
600(+310/-150)
<10
47(+6/-5)
<60
260(+160/-70)
440(+560/-160)
1900(+6200/-800)
430(+110/-70)
1000{+900/-250)
610(+590/-200)
1400(+3300/-560)
290(+570/-120)
750(+850/-260)
600(+570/-190)
730(+1100/-270)
<10
32(-t-5/-3)
<60
200(+90/-30)
140{+40/-20)
210(+160/-60)
240(+40/-40)
730(+370/-180)
360(+150/-80)
910(+4400/-410)
220(+160/-60)
940(-f12000/-450)
590{+1800/-250)
630(+2500/-280)
t1/2 (95 percent confidence interval)
Least squares slope (for calculation of t1/2) = zero with 95% confidence.
Table 9. Acclimation of Soil to Complex Fossil Fuel Waste [67]
Unacclimated Soil
PAH
Compound
Naphthalene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz(a)anthracene
Chrysene
Benz(a)pyrene)
Initial Soil
Concentration
(mg/kg-dry wt)
38
30
38
154
177
30
27
10
Reduction in
40 days (%)
90
70
58
51
47
42
25
40
Acclimated Soil
Soil Concentration
after First Reappli-
cation of Waste (after
168 days incubation
at initial level)
(mg/kg-dry wt)
38
30
38
159
180
40
33
12
Reduction in
22days'(%)
100
83
99
82
86
70
61
50
14
-------
contaminated soil to be treated may also result In faster
cleanup of a site. Amendment of the soil with exogenous
microorganisms developed in laboratory batch cultures
would not be required.
One method to assess detoxification of waste constituents
in a soil system involves use of the Microtox™ assay [38].
The assay is used to measure acute toxicity of aqueous
solutions or water soluble fraction extracts. The Microtox™
system is a standardized instrumental-based system that
utilizes a suspension of marine luminescent bacteria
(Photobacterium phosphoreum) as bioassay organisms.
The bioassay organisms are handled like chemical
reagents. Suspensions of about one million bacteria are
Nunn clay loam
' a 2%OB&Graas»
• 414 Oil 4 Grease
30 60 90 MO ISO 180
TIME(d«y«)
4a
Kidman sandy loam
100-
• 2%Oils Great*.
30
60 90 120 150 180
TIME (d*y>)
4b
Figure 4. EC50 as a Function of Time for Two Soils and Three
Waste Loading Rates
"challenged" with additions of serial dilutions of an aqueous
sample or extract. Light output from each bacterial
suspension is measured before and after each addition of
sample. Results are presented as EC50 values, which are
defined as sample concentrations resulting in a 50%
decrease of light produced by the luminescent bacteria. High
EC50 values indicate lower toxicity than low values.
Detoxification of a contaminated soil system is indicated by
increased Microtox™ EC50 values approaching 100%. A
value of 100% is considered as non-toxic.
In a clay loam soil, a petroleum refinery waste was added to
soil at application rates of 2%, 4%, and 8% by weight of oil
and grease [40]. The results of the study are shown in Figure
4. Time of incubation is plotted on the x-axis, and EC50
values, as determined by the Microtox™ assay.on the y-axis.
At the 2% loading rate, the waste material was detoxified to
an EC50 value of 100 in a period of about 100 days (Figure
4a). At the highest level of contamination (8% loading rate),
the materials remained toxic, even after 1 80 days. In addition
to providing evidence of detoxification of waste constituents,
this study also showed the potential for enhancement of
biodegradation by mixing uncontaminated soil with
contaminated soil to produce a treatment medium with waste
contents at levels not toxic to microbial populations.
In a sandy loam soil amended with the same contaminated
material, a longer period (about 170 days) was required to
detoxify the 2% contamination level to an EC50 value of
100% (Figure 4b). Results of these studies show that mixing
of contaminated soils with uncontaminated soils can result in
detoxification. However, since the rate of detoxification may
be a function of soil type, these results also illustrate the site
specificity of bioremediation efforts and underscore the need
to perform site-specific characterization of the contaminated
area.
Another technique to assess detoxification of organic
constituents in a soil system is the use of the Ames
Salmonella typhlmurium assay [29, 30, 52]. The assay
utilizes the bacterium S. typhimuriumlo indicate the
presence of mutagenic constituents, which may include
—tr- TAOOwlmSa
_.£ — TA 100 without 89
Figure 5. Ames Assay Results for Waste; Soil Mixture
Immediately After Waste Incorporation Into Soil [67]
15
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000 2000 3000 4000 SOOO 6000 7000 8000
Ctose t/ig/plole)
Figure 6. Ames Assay Results for Waste; Soil Mixture After
Forty-two Days Incubation [67]
transformation products of parent compounds. Different
strains of S. typhlmurium are selected to indicate the
mechanism of mutagenlcity, i.e., point mutations or frame-
shift mutations. Mutagenicity is measured as a ratio of the
number of colonies that grow in the presence of a test
sample (e.g., chemical, mixture of chemicals, or extract of
an environmental sample) to the number of colonies in the
absence of the test sample. Since growth occurs in
proportion to mutagenic potential, growth will be greater in
the presence of a mutagen, and will increase as the dose of
mutagen is increased. The increase in growth in response
to dose is depicted graphically in dose-response curves.
The minimum mutagenic ratio (ratio of number of colonies
that form in the presence of a test sample to the number of
colonies on a control growth plate) is 2.0. Therefore, a
sample exhibiting a mutagenic ratio greater than 2.0 is
considered to possess mutagenic properties.
A study to evaluate detoxification of mutagenic potential of
a complex fossil fuel waste containing PAH compounds
treated In a soil system was conducted utilizing the Ames
assay [67]. Mutagenic ratios for S. typhimurium strain TA98
(a test strain used to detect f rameshift mutagens such as
PAHs) with metabolic activation (to simulate mammalian
metabolism by the addition of a mammalian liver extract
(referred to as the S9 fraction)), and without metabolic
activation (without the addition of the S9 fraction) were
determined immediately after waste incorporation and after
42 days of incubation of the waste in the soil. Results as
shown in dose response curves showed that the mutagenic
ratios decreased from about 4.5 and 7.0 at the highest dose
levels tested immediately after waste incorporation (Figure
5) to borderline mutagenic levels (i.e., mutagenic ratios of
about 2) after 42 days of treatment (Figure 6). For a
different S. typhimurium strain, (TA100, a test strain used to
detect mutagens causing base-pair substitutions), no dose-
response effects or mutagenic activity were measured
during the study.
Results of a pilot scale field study have also demonstrated
that bloremediation of PAH contaminated soils is a
technology that can result in significant cleanup of
contaminated soils [69]. A coal gasification waste was
thoroughly mixed into a soil at a one-half acre site.
Sampling of soil cores was performed at 10 feet intervals
across 100 feet rows. Data presented in Table 10 are
composite values from the sampling efforts. In all cases,
concentrations of the PAH compounds in the soil were
greatly reduced after 91 days. Data quality was poorer at
the 91-day sampling period, as measured by the coefficient
of variation (CV), which is the mean value measured (AVG)
divided by the standard deviation (SD). The poorer data
quality was attributed to increased analytical difficulties
when levels of constituents near detection limits are
measured.
The fate and environmental impact of transformation
products is an area of bioremediation that needs more
consideration. In a laboratory study, the transformation of a
14C radio-labelled PAH compound, 7,12- dimethylbenz(a)-
anthracene, in a sandy loam soil was investigated for a 28-
day incubation period [70]. At time 0, 62% of the applied
parent compound was recovered from the soil, which
represents the extraction efficiency of the test (Table 11).
After 28 days of incubation, only 20% of the parent
compound was recovered. Since recoveries in control
reactors poisoned with mercuric chloride were not
significantly different over the 28-day incubation period,
biological treatment was the proposed mechanism of
compound removal from the soil system. Table 11 also
shows that the decrease in parent "C was accompanied by
an increase in the metabolite 14C fraction. The appearance
of transformation products increased from 4% of the total
14C applied at time 0 to 53% after 28 days. None of the
radiolabeiied carbon appeared as CO2 in this study, but 12
to 17% of the radiolabeiied material was associated with the
solid phase of the soil during the incubation period. The
mass balance for the study ranged from 78 to 90%
recovery of the applied radiolabeiied carbon. Therefore, the
appearance, toxicity, fate, and behavior of a metabolite
fraction may need to be evaluated on a site-specific basis.
The environmental significance and fate and behavior of
many transformation products of PAH constituents, as well
as transformation products from many other organic
constituents, are not yet known. Therefore, incorporating
detoxification assessment into a bioremediation plan is
recommended to evaluate these concerns.
16
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Table 10. Field Results for Soil Treatment of PAHs in Coal Gasification Wastes [69]
C.*(MQ/fl)
Compound
AVG SD C
Naphthalene 186 68
Acenaphthene 729 276
Phenanthrene 78 28
Benz(a)
anthracene 86 42
Dibenz(a,h)
anthracene 52 36
*C0 = Initial Soil Concentration
Table 11. Transformations of ("C) DMBA byN
14C appearii
Soil extract
Time, DMBA,
days parent compound Metabolites
0 62(69) 4(6)
14 26 43
28 20(60) 53(11)
* Poisoned (control) data in parentheses.
C After 91 days (Ma/9)
/(%) AVG SD CV(%)
37 3 1.8 61
38 1 1.8 157
36 2.6 0.6 23
49 2 0.8 38
69 ND
cLaurin Sandy Loam Soil* [70]
ig in each fraction, percent
Soil
Residue CO., Total
12(13) 0(0) 78(88)
16 0 85
17(16) 0(0) 90(87)
17
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Table 12. Wood Preserving Sites Where Bioremedlatlon has been Proposed
for Soil or Lagoon Sediments [71]
Site Name
State (U. S. EPA Region)
Proposed Remediation
LA. Clark and Son
Brown Wood Preserving
Burlington Northern
(Brainard)
North Cavalcade Street
United Creosoting Company
Baxter/Union Pacific
Burlington Northern
Somers)
Libby (Champion
International)
Koppers, Co.
J.H. Baxter
VA(lll)
FL(IV)
MN(V)
TX (VI)
TX (VI)
WY (VIII)
MT (VIII)
MT(VIII)
CA (IX)
CA (IX)
Bioremediation
Bioremediation
Landfarm
Bioremediation
In Situ remediation
Bioremediation
Landfarm
In Situ Bioremediation
and Landfarm
Bioremediation
Bioremediation
9. Implementation of
Bioremediation at Sites
Contaminated with Organic
Wastes
A recent survey conducted for the U.S. EPA concerning the
use of bioremediation at sites with soils contaminated with
wood-preserving wastes identified ten sites that currently
plan to use bioremediation techniques to cleanup
contaminated soils and sediments (Table 12) [71]. Sims
observed a wide range of variability In target clean-up
levels. A wide variability also was observed in criteria for
selecting target levels (maximum contaminant levels
(MCLs) based on drinking water standards vs. negotiated
levels vs. risk assessment-based levels) and in selection of
soil phases that must meet target levels (solid phase,
leachate phase, and/or air phase). Target levels were
determined on a site-specific basis.
An example of a bioremediation plan for a facility identified
In the survey was presented by Lynch and Genes [72]. On-
site treatment of creosote-contaminated soils from a
shallow, unllned surface impoundment was demonstrated
at a disposal facility for a wood-preserving operation in
Minnesota. The contaminated soils contained creosote
constituents consisting primarily of PAHs at concentrations
ranging from 1,000 to 10,000 ppm. Prior to implementation
of the full scale treatment operation, bench-scale and
pllotscale studies simulating proposed full-scale conditions
were conducted to define operation and design
parameters. Over a four-month period, 62% to 80%
removal of total PAHs were achieved in all test plots and
laboratory reactors. Two-ring PAH compounds were
reduced by 80-90%, 3-ring PAHs by 82-93%, and 4+-ring
PAHs by 21-60%.
The full-scale system involved preparation of a treatment
area within the confines of the existing impoundment. A
lined waste pile for temporary storage of the sludge and
contaminated soil from the impoundment was constructed.
All standing water from the impoundment was removed,
and the sludges were excavated and segregated for
subsequent free oil recovery. Three to five feet of "visibly"
contaminated soil was excavated and stored in the lined
waste pile. The bottom of the impoundment was stabilized
as a base for the treatment area. The treatment area was
constructed by installation of a polyethylene liner, a
leachate collection system, four feet of clean backfill, and
addition of manure to achieve a carbon:nitrogen ratio of
50:1. A sump for collection of stormwater and leachate
and a center pivot irrigation system were also installed.
The lined treatment area was required because the natural
soils at the site were highly permeable. A cap was also
needed for residual contaminants left in place below the
liner. Contaminated soil was periodically applied to the
treatment facility and roto-tilled into the treatment soil. Soil
moisture was maintained near field capacity with the
irrigation system. During the first year of operation,
greater than 95% reductions in concentration were
obtained for 2- and 3-ring PAHs. Greater than 70% of 4-
and 5-ring PAH compounds were degraded during the first
year. Comparison of half-lives of PAHs in the full-scale
facility were in the low end of the range of half-lives
reported for the test plot units. Only two PAH compounds
were detected in drain tile water samples, at
concentrations near analytical detection limits.
Bioremediation of a Texas oil field site with storage pit
backfill soils contaminated with styrene, still bottom tars,
18
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and chlorinated hydrocarbon solvents was demonstrated on
a pilot scale [73]. The remediation efforts also included
chemical and physical treatment strategies. The pilot scale,
solid-phase biological treatment facility consisted of a
plastic film greenhouse enclosure, a lined soil treatment bed
with an underdrain, an overhead spray system for
distributing water, nutrients, and inocula, an organic vapor
control system consisting of activated carbon absorbers,
and a fermentation vessel for preparing microbial inoculum
or treating contaminated leachate from the backfill soils.
Soils were excavated from the contaminated area and
transferred to the treatment facility. Average concentrations
of volatile organic compounds (VOCs) were reduced by
more than 99% during the 94 day period of operation of the
facility; most of the removal was attributed to air stripping.
Biodegradation of semivolatile compounds reduced aver-
age concentrations by 89% during the treatment period.
A solid-phase treatment system to remediate petroleum
contaminated soil at a hazardous waste site in California
was described by Ross et al. [6]. The treatment process
involved stimulating the existing microbial population in the
soil to degrade petroleum hydrocarbon contaminants. A
biotreatability evaluation prior to full-scale operation
demonstrated that the existing microorganisms in the soil
could degrade the petroleum hydrocarbons, but that the
nutrient levels in the soil were not sufficient to maintain
growth and support complete degradation of the
hydrocarbon contaminants. With adequate nutrients,
hydrocarbons decreased from 3500 ppm to less than 100
ppm in 4 weeks in bench scale studies. The degradation
process exhibited biphasic kinetics, likely due to the fact the
petroleum hydrocarbons were a mixture of a lighter diesel
fuel and more recalcitrant waste motor oils. The full scale
facility, which began operation in 1988, consists of a four
acre treatment site that has had 30 inches of contaminated
soil applied to the surface. Bioremediation of the top 15
inches was proceeding by the addition of nutrients, daily
tilling and maintenance of adequate soil moisture levels.
When the first 15 inches of contaminated soil have been
remediated to the target cleanup level of 100 ppm, it will be
removed and the second 15 inches will be treated. During
the first four weeks of operation, the average concentration
of petroleum hydrocarbons was reduced from 2,800 to 280
mg/kg. The rate of hydrocarbon biodegradation measured
in the field was consistent with the rate measured in the
laboratory.
A solid phase treatment system to clean up pesticides in
soil contaminated as a result of a fire at a chemical storage
facility was also described by Ross et al. [6]. Water used
to extinguish the fire carried large amounts of insecticides
and herbicides into the soil beneath the warehouse facility.
Laboratory biotreatability studies showed that moderately
contaminated soils (90 mg/kg of 2,4-D) could be treated in a
soil treatment system to meet regulatory criteria (total
MCPA and 2,4-D = 10 mg/kg), while highly contaminated
soils (2,4-D concentrations greater than 200 mg/kg)
required treatment in a soil/water slurry bioreactor. A five
acre soil treatment area was constructed with an
engineered clay liner 12 inches thick and a drainage system
to control water movement. Ten thousand cubic yards of
soil contaminated with a complex mixture of herbicides and
insecticides, including 2,4-D, alachlor, trifluralin, carbofuran,
and MCPA, were spread on the treatment bed to an
average depth of 15 inches. During operation, soil
conditions were optimized for biological activity by daily
tilling and by maintenance of
soil moisture content between 8% and 15% by weight.
During three months of operation, the combined 2,4-D
and MCPA concentrations decreased from 86 ppm to 5
ppm.
Brubaker and Exner [74] reported on two case histories that
involved microbial degradation of chemical contaminants to
remediate chemical spills. Both sites also involved other
remediation tools in addition to microbial remediation,
emphasizing the need to examine complementary and
synergistic remediation techniques. At the first site, residual
contamination from a formaldehyde spill was treated using
chemical oxidation with hydrogen peroxide, followed by
microbial "polishing" to complete the remediation. A
commercial inoculum of microorganisms acclimated for
formaldehyde degradation and a nutrient solution were
mixed in an aeration tank and then sprayed on the site.
Water was collected in a sump and recycled through the
aeration tank. Treatment effectiveness was measured by
reduction of concentration of formaldehyde in the aeration
tank. After 25 days, concentrations had dropped from over
700 mg/l to less than 1 mg/l. At the second site, a gasoline
leak from an underground storage tank was remediated
with enhanced bio reclamation techniques, which consisted
of addition of nutrients and hydrogen peroxide as an oxygen
source. A series of injection and recovery wells were used
to recycle water through the site. Soil samples showed a
decrease in volatile fuel hydrocarbons from an average of
245 ppm at the initiation of the bioreclamation process to
0.8 ppm after 200 days.
Bioremediation of a site contaminated with PCBs, which
have generally been considered resistant to biodegradation
in the environment, has been demonstrated at a drag-
racing track in New York [55]. Laboratory treatability studies
using contaminated soils from the sites inoculated with pure
resting cell cultures of PCB-degrading organisms that had
been isolated from environmental samples showed
substantial PCB biodegradation, up to 51% of the PCBs
present in three days. Follow-up laboratory studies were
conducted using only 3-4% of the number of cells used in
the earlier studies, lower moisture content, lower
temperatures, and no shaking or aeration of the reaction
mixtures. PCB degradation was not observed until 30 days
after the initiation of the study. In an undisturbed soil
sample inoculated three times weekly with the PCB-
degrading microorganisms, 50% of the PCBs in the top 1
cm of soil was degraded in 15 weeks. Only 10%
degradation was seen at depths below 1 cm. When a
duplicate of the undisturbed soil experiment was mixed at
three months with continued inoculation, the redistributed
soil again exhibited the highest degradation7ate at the
surface. In experiments where soils were inoculated three
times weekly and mixed after each application, 35% of the
PCBs were degraded after 23 weeks at all depths. This
degree of degradation represents a greater amount of PCB
19
-------
destruction since the RGBs were degraded throughout the
whole sample and not just at the surface. Thus mixing was
Identified as an important site management variable.
Preliminary results at a field scale test site at the drag-
racing track Indicated significant PCB degradation after
eight to ten weeks.
10. Conclusions
Consideration of bioremediation for remediation of a site
contaminated with organic constituents requires a detailed
site, soil, and waste characterization that must be
conducted In order to evaluate the potential application of
the technology at the site and to demonstrate the feasibility
of the approach. A sound and thorough engineering
remediation plan developed at the onset of the project will
allow cost-effective and efficient use of resources for
Implementation of site clean-up. The use of treatability
studies and simulation modeling are also necessary
components of the bioremediation plan so that necessary
data to evaluate potential use and to identify pathways of
migration are collected in a cost-effective manner.
Bioremediation of sites contaminated with organic
chemicals is a promising technology, especially if it is
Incorporated in a remediation plan that uses an integrated
approach to the cleanup of the complete site, i.e., a plan
that Involves the concept of a "treatment train" of physical,
chemical, and/or biological processes to address
remediation of all sources of contaminants at the site.
11. References
1. Omenn, G.S. (ed.). 1988. Environmental
Biotechnology- Reducing Risks from Environmental
Chemicals through Biotechnology. Plenum Press, New
York, NY, 505 pp.
2. Engineering Foundation. 1988. Proceedings,
Conference on Biotechnology Applications in
Hazardous Waste Treatment. Engineering Foundation
Conferences, Longboat Key, Florida, October 31 -
November4.
3. AWMA/EPA. 1989. Proceedings of the Internal!.
Symposium on Hazardous Waste Treatment:
Blosystems for Pollution Control. Air and Waste
Management Association and U.S. Environmental
Protection Agency, Cincinnati, Ohio, February 20-23.
4. U.S. EPA. 1989. Bioremediation of Hazardous Waste
Sites Workshop. CERI-89-11, U.S. Environmental
Protection Agency, Cincinnati, OH.
5. McGinnis, G.D., H. Borazjani, L.K. McFarland, D.F.
Pope, and D.A. Strobel. 1989. Characterization and
Laboratory Soil Treatability Studies for Creosote and
Pentachlorophenol Sludges and Contaminated Soil.
EPA/600/2-88/055, Robert S. Kerr Environmental
Research Laboratory, U.S. Environmental Protection
Agency, Ada, OK.
6. Ross, D., T. P. Marziarz, and A.L. Bourquin. 1988.
Bioremediation of hazardous waste sites in the USA:
Case histories, pp. 395-397. In: Superfund '88, Proc.
9th Natl. Conf., Hazardous Materials Control Research
Institute, Silver Spring, MD.
7. Wilson, L.G. 1983. Monitoring in the vadose zone: Part
III. Ground Water Monitoring Review (Winter}:155-166.
8. Everett, L.G., E.W. Hoylman, L.G. McMillion, and L.G.
Wilson. 1982. Vadose zone monitoring concepts at
landfills, impoundments, and land treatment disposal
areas. In: Management of Uncontrolled Hazardous
Waste Sites. Hazardous Materials Control Research
Institute, Silver Spring, MD.
9. Wilson, L.G. 1981. Monitoring in the vadose zone: Part
I, Storage changes. Ground Water Monitoring Review
(Fall):32-41.
10. Wilson, L.G. 1982. Monitoring in the vadose zone: Part
II. Ground Water Monitoring Review (Winter):31-42.
11. Lehr, J.H. 1988. The misunderstood world of
unsaturated flow. Ground Water Monitoring Review
(Spring):4-6.
12. Sims, R.C., J.L. Sims, D.L. Sorensen, W.J. Doucette,
and L.L. Hastings. 1986. Waste/Soil Treatability
Studies for Four Complex Industrial Wastes:
Methodologies and Results. Vol. 1 and 2. EPA/600/6-
86/003a and b, Robert S. Kerr Environmental Research
Laboratory, U.S. Environmental Protection Agency,
Ada, OK.
13. Sims, R.C., W.J. Doucette, J.E. McLean, W.J.
Grenney, and R.R. Dupont. 1988. Treatment Potential
for 56 EPA Listed Hazardous Chemicals in Soil. EPA/
600/6-88-001, Robert S. Kerr Environmental Research
Laboratory, U.S. Environmental Protection Agency,
Ada, OK.
14. Loehr, R. 1989. Treatability Potential for EPA Listed
Hazardous Wastes in Soil. EPA/600/2-89/011, Robert
S. Kerr Environmental Research Laboratory, U.S.
Environmental Protection Agency, Ada, OK.
15. Paul, E.A., and F. E. Clark. 1989. Soil Microbiology and
Biochemistry. Academic Press, Inc., San Diego, CA.
16. Rittmann, B.E., and P.L. McCarty. 1980. Model of
steady-state biofilm kinetics. Biotec. Bioeng. 22:23-43.
20
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17. Horvath, R.S. 1972. Microbial co-metabolism and the
degradation of organic compounds in nature. Bacteriol.
Rev. 36:146-155.
18. Perry, J.J. 1979. Microbial cooxidation involving
hydrocarbons. Microbiol. Rev. 43:59-72.
19. Keck, J., R.C. Sims, M. Coover, K. Park, and B.
Symons. 1989. Evidence for cooxidation of poly-
nuclear aromatic hydrocarbons in soil. Water Res. (In
press).
20. Dragun, J. 1988. The Soil Chemistry of Hazardous
Materials. Hazardous Materials Control Research
Institute, Silver Spring, MD.
21. Sims, R.C., D.L. Sorensen, J.L. Sims, J.E. McLean, R.
Mahmood, and R.R. Dupont. 1984. Review of In Place
Treatment Techniques for Contaminated Surface Soils.
Volume 2: Background Information for In Situ
Treatment. EPA/540/2-84-003a, Municipal
Environmental Research Laboratory, U.S.
Environmental Protection Agency, Cincinnati, OH.
22. Mahmood, R.J., and R.C. Sims. 1986. Mobility of
organics in land treatment systems. J. Environ. Eng.,
Am. Soc. Civil Eng. 112:236-245.
23. Huddleston, R.L., C.A. Bleckmann, and J.R. Wolfe.
1986. Land treatment biological degradation processes.
pp. 41-61. In: R.C. Loehr and J.F. Malina, Jr. (eds.)
Land Treatment: A Hazardous Waste Management
Alternative. Water Resources Symposium No. 13,
Center for Research in Water Resources, The
University of Texas at Austin, Austin, TX.
24. Park, K.S., R.C. Sims, R.R. Dupont, W.J. Doucette,
and J. E. Matthews. 1989. Fate of PAH compounds in
two soil types: Influence of volatilization, abiotic loss,
and biological activity. Environ. Toxicol. Chem. (In
press).
25. U.S. EPA. 1986. Permit Guidance Manual on
Hazardous Waste Land Treatment Demonstrations.
EPA-530/SW-86-032, Office of Solid Waste and
Emergency Response, U.S. Environmental Protection
Agency, Washington, DC.
26. U.S. EPA. 1988b. Interim Protocol for Determining the
Aerobic Degradation Potential of Hazardous Organic
Constituents in Soil. U.S. EPA Scientific Steering
Committee, Biosystems Technology Development
Program, and Soil Treatment Processes Committee,
Robert S. Kerr Environmental Research Laboratory,
U.S. Environmental Protection Agency, Ada, OK.
27. Liu, D., and B.J. Dutka (eds.). 1984. Toxicity Testing
Procedures using Bacterial Systems. Marcel Dekker,
Inc., New York, Inc.
28. Dutka, B.J., and G. Bitton. 1986. Toxicity Testing using
Microorganisms. CRC Press, Inc., Boca Raton, FL.
29. Ames, B.N., J. McCann, and E. Yamasaki, E. 1975.
Methods for detecting carcinogens and mutagens with
the Salmonella/mammalian-microsome mutagenicity
test. Mutation Res. 31:347-364.
30. Maron, D.M., and B.N. Ames. 1983. Revised methods
for the Salmonella mutagenicity test. Mutation Res.
113:173-215.
31. Sims, R.C., J.L. Sims, and R.R. Dupont. 1984. Human
health effects assays. J. Water Pollut. Control Fed. 56:
791-800.
32. Sims, R.C., J.L. Sims, and R.R. Dupont. 1985. Human
health effects assays. J. Water Pollut. Control Fed. 57:
728-742.
33. Sims, R.C., J.L Sims, and R.R. Dupont. 1986. Human
health effects assays. J. Water Pollut. Control Fed. 58:
703-717.
34. Sims, R.C., J.L. Sims, and R.R. Dupont. 1987. Human
health effects assays. J. Water Pollut. Control Fed. 59:
601-614.
35. Sims, R.C., J.L. Sims, and R.R. Dupont. 1988. Human
health effects assays. J. Water Pollut. Control Fed. 60:
1093-1196.
36. McCann, J.R., R. Choi, E. Yamasaki, and B.N. Ames.
1975. Detection of carcinogens as mutagens in the
Salmonella/microsome test: Assay of 300 chemicals.
Proc. Natl. Acad. Sci. 72:5135-5139.
37. Bulich, A.A. 1979. Use of luminescent bacteria for
determining toxicity in aquatic environments, p. 98-106.
In: L.L Markings and R.A. Kimerle, eds. Aquatic
Toxicology. ASTM 667, Amer. Soc. for Testing and
Materials, Philadelphia, PA.
38. Matthews, J.E. and A.A. Bulich. 1984. A toxicity
reduction test system to assist predicting land
treatability of hazardous wastes, pp. 176-191.
In: J.K. Petros, Jr., W.J. Lacy, and R.A. Conway,
eds., Hazardous and Industrial Solid Waste
Testing: Fourth Symposium STP-886, American So-
ciety of Testing and Materials, Philadelphia, PA,
39. Matthews, J.E. and L. Hastings. 1987. Evaluation of
toxicity test procedure for screening treatability
potential of waste in soil. Toxicity Assessment: An
Internatl. Quarterly 2: 265-281.
40. Symons, B.D. and R.C. Sims. 1988. Assessing
detoxification of a complex hazardous waste, using the
Microtox™ bioassay. Arch. Environ. Contamination
Toxicol. 17: 497-505.
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